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. 

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Copyright © Taylor & Francis Ltd 1995 

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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 

Preface vii 

List of Contributors ix 

PART ONE Bioavailability and metabolic aspects of industrial 

chemicals 

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. Toxicokinetics and Biodisposition of Industrial 

Chemicals 

N.P.E.Vermeulen R.T.H.van Welie B.M.de 

Rooij J.N.M.Commandeur 

3. Metabolic Activation of Industrial Chemicals 

and Implications for Toxicity 

G.J.Mulder 

4. Sizing Up the Problem of Exposure 

Extrapolation: New Directions in Allometric 

Scaling 

D.B.Campbell 

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 


P.Sagelsdorff 

72 

PART THREE Pulmonary toxicology of industrial chemicals 90 

7. Studies to Assess the Carcinogenic Potential of 

Man-Made Vitreous Fibers 

T.W.Hesterberg G.R.Chase R.A.Versen 

R.Anderson 

91 

1 

2 

12 

36 

44

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 

9. Pulmonary Hyperreactivity to Industrial 

Pollutants 

J.Pauluhn 

10. Mechanisms of Pulmonary Sensitization 

I.Kimber 

11. Occupational Asthma Induced by Chemical 

Agents 

C.A.C.Pickering 

PART FOUR Biomarkers and risk assessment of industrial 

chemicals 

12. Biomarkers and Risk Assessment 

K.Hemminki 

13. Extrapolation of Toxicity Data and Assessment 

of Risk 

N.Fedtke 

14. Molecular Approaches to Assess Cancer Risks 

A.S.Wright J.P.Aston N.J.van Sittert 

W.P.Watson 

15. Evaluation of Toxicity to the Immune System 

H.-W.Vohr 

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 

117 

129 

138 

149 

157 

158 

167 

180 

197 

207 

PART FIVE Mechanisms of toxicity of industrial chemicals 222 

17. Peroxisome Proliferation 

B.G.Lake R.J.Price 

223 

v

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 

19. Endocrine Toxicology of the Thyroid for 

Industrial Compounds 

C.K.Atterwill S.P.Aylward 

20. Testing and Evaluation for Reproductive 

Toxicity 

A.K.Palmer 

238 

255 

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 

25. Low Dose of a Genotoxic Carcinogen does not 

‘Cause’ Cancer; it Accelerates Spontaneous 

Carcinogenesis 

W.K.Lutz 

26. Controversial Mechanistic and Regulatory 

Issues in Safety Assessment of Industrial 

Chemicals—an Industry Point of View 

H.-P.Gelbke 

355 

356 

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 CIBA- 

GEIGY 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 


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 


Box 50, Elkton Road, Newark, DE 19714–0050, USA 

Kari Hemminki 

CNT, Karolinska Institute, Novum, S-141 57 Huddinge, Sweden 

Thomas W.Hesterberg 


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 



Ian Kimber 

Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, 

Cheshire SK10 4TJ, UK

Beverly M.Kulig 



Brian G.Lake 

BIBRA International, Woodmansterne Road, Carshalton, Surrey, SM5 

4DS, UK 

Jan H.C.M.Lammers 



Werner K.Lutz 

Universität Würzburg, Institut für Toxikologie, Versbacher Strasse 9, 

D-97078 Würzburg, Germany 

Aart C.Monster 



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 

xi

xii 

Vera Rogiers 

Vrije Universiteit Brussel, Department of Toxicology, Laarbeeklaan 

103, B-1090 Brussels, Belgium 

Ben M.de Rooij 




Peter Sagelsdorff 



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 


Willem S.Sloot 



Walter Sterzel 

Henkel KGaA, TTB-Toxikologie, Geb. Z33, D-40191 Düsseldorf, 

Germany 

Helmut Thomas 


Basel, Switzerland. Current address: Ciba-Pharmaceuticals, Stamford 

Lodge, Wilmslow, Cheshire SK9 4LY, UK 

Kornelis J.van den Berg 



Yves Vandenberghe 


B-1090 Brussels, Belgium 

Antoine Vercruysse 


B-1090 Brussels, Belgium 

Nico P.E.Vermeulen 


Vrije Universiteit Amsterdam, De Boelelaan 1083, H1081 NL-V 

Amsterdam, The Netherlands

Richard A.Versen 


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 


David B.Wahrheit 


Box 50, Elkton Road, Newark, Delaware 19714–0050, USA 

William P.Watson 


Hansjörg Weideli 

CIBA-GEIGY Ltd, R-1002.2.59, PO Box, CH-4002 Basel, Switzerland 

Ronald T.H.van Welie 




Frederik A.de Wolff 


Centre, Meibergdreef 15, 1105 Amsterdam, The Netherlands 

Alan S.Wright 


xiii

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


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.


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 m 2 

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


is usually no longer than a few minutes. The exposed area is usually in the 

order of 20 cm 2 . 

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 cm 2 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 cm 2 , 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 t max. 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 

cm 2 . In fact the result should be extrapolated to the total surface of the 

human skin, which is about 2 m 2 . 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).


Figure 1.2 Elimination of 1,1,1-trichloroethane and tetrachloroethene by expired 

air after dermal exposure to the liquid of 27 cm 2 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 cm 2 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.

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 


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 cm 2 to 1.4 mmol l −l air). 

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.


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.


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).

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 

N.P.E.VERMEULEN ET AL. 13 

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

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 


Figure 2.2 Disposition and biological effects of xenobiotics subdivided into three 

phases. 

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


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


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 (k el) and the 

urinary half-life of elimination (t 1/2) 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 (C p), the t 1/2 derived from the 

urinary elimination rate constant is identical to the t 1/2 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 t 1/2 (Table 2.1). 

No accumulation will occur when the intervals between the exposure


Figure 2.3 Schematic representation of first order kinetics of (A) the plasma 

concentration (C p) 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=–k el/2.303 and t 1/2=0.693/k el. 

Figure 2.4 Urinary excretion of a hypothetical metabolite during 3 days of 

intermittent exposure: t 1/2

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.


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

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 


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. 

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


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).

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-acetyl- 

S-(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


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 Z- 

DCP-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 E- 

DCP-MA. There is still a lack of knowledge about the magnitude of the


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 

individuals a 

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 E- 

DCP. 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


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.

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 

B 6C 3F 1 and male Swiss mice exposed to TRI by inhalation. The effect was 

not observed in male B 6C 3F 1 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 

B 6C 3F 1 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


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

a References: see review Goeptar et al., 1995b. 

n.d.: not determined.


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 B 6C 3F 1 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 B 6C 3F 1 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-


Figure 2.11 Possible routes of metabolism of S-(1,2-dichlorovinyl)glutathione (1,2- 

DCV-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-DCV- 

Cys (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


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


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

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. 

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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.

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 

G.J.MULDER 37 

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

38 METABOLIC ACTIVATION OF INDUSTRIAL CHEMICALS 

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

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 

G.J.MULDER 39 

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


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

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 

G.J.MULDER 41 

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


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 

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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.

G.J.MULDER 43 

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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. 

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of liver cancer associated with human exposures to chloroform using PbPK 

modeling, Toxicol. Appl. Pharmacol., 105, 443–59. 

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air, Regul. Toxicol. Pharmacol., 9, 175–85. 

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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

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·W 0.75 , liver blood flow from 

37·W 0.85 , blood weight from 0.055·W 0.99 , and respiratory rate from 0. 

019·W 0.26 . 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 

D.BRUCE CAMPBELL 45 

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.

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·W 0.67 

(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 


Figure 4.1. Allometric scaling of Fotemustine clearance compared with the body 

weight in various species. 

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


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).

(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.


Table 4.1 Comparison of exponential values for b with MLP corrected clearance 

(CL u ·MLP=aW b ) 

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

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 a oximately 300 ml min 1 ppr 

− with a 95 per cent confidence,


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


a Campbell DB, 1993 unpublished data. 

b CL=812 ml min−1


Table 4.3 Interspecies comparisons of human clearance predictions expressed as 

percentage from observed clearances using maximum life potential corrections 

a MLP=22 years. 

b MLP=35 years. 

c Percentage of compounds with a predicted human clearance greater than 100% of 

that observed. 

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

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. 

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PART TWO 

Reactive industrial chemicals

5 

Metabolism of Reactive Chemicals 

PETER J.van BLADEREN 1,2 and BEN van OMMEN 1 

1 TNO Toxicology, Zeist 

2 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

P.J.VAN BLADEREN AND B.VAN OMMEN 61 

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

62 METABOLISM OF REACTIVE CHEMICALS 

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


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


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


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


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-


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 bis- 

DNA-adducts have been identified (Humphreys et al., 1991). 1,2- 

Dibromochloropropane 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


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 isothiocyanates 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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6 

Methods for the Determination of Reactive 


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.

P.SAGELSDORFF 73 

Figure 6.1 Examples of electrophilic compounds. The arrows indicate the suspected 

electrophilic centres (from Lutz, 1979).

74 METHODS FOR THE DETERMINATION OF REACTIVE COMPOUNDS 

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).

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


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).

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


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). 


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

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 [ 32 P]- or [ 33 P]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.


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.

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 

Lack of bioavailability of 3,3′-dichlorobenzidine from 

diarylide pigments 


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

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


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

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 10 6 

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 10 6 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


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 [ 32 P]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 

[ 32 P]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 10 8 

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 [ 32 P]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.


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).


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.

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. 

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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 Al 2O 3 and 

SiO 2 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

92 CARCINOGENIC POTENTIAL OF MAN-MADE VITREOUS FIBERS 

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-

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 

T.W.HESTERBERG ET AL. 93 

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


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

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.


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

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


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)

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

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter


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

inhalation 


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 


Table 7.3 Combined summary of lung pathology findings: chrysotile and RCF1 

from RCF MTD study in rats; and RCF multidose study in rats 



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

rats 

a WHO fibers=fibers with length/diameter ≥3, length >5 µm, and diameter 5 µm, and a diameter 10 µm in length in the lung were similar for each 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.


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

study in rats 



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


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.


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

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


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


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.,


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.

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


a Lung fibers: for humans, NIOSH A rules; for rats, total fibers (all fibers length/diameter >3:1). 

b 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.


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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.BUTTERICK 2 and STEVEN R.FRAME 1 

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

118 PULMONARY TOXICITY STUDIES WITH MAN-MADE ORGANIC FIBRES 

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).

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 

D.B.WARHEIT ET AL. 119 

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


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 3.7×10 7 (±7.4×10 6 ) fibres/ 

lung for the high dose group and 1.3×10 7 (±4×10 6 ) 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 7.6×10 7 (±1.9×10 7 ) fibres per lung for 

the high dose group and 4.8×10 7 ( ±2.1×10 7 ) 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


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


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


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).


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.

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

rats 

a p


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

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 


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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.

130 PULMONARY HYPERREACTIVITY TO INDUSTRIAL POLLUTANTS 

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

methodological 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 

J.PAULUHN 131 

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).


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

J.PAULUHN 133 

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


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 (Q s/Q t) anddecrease in forced expiratory flow rates 

(MMEF) as well as mucus products (sialomucins), polymorphonuclear

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 

J.PAULUHN 135 

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, Q s/Q t: venous admixture, PMN: 

polymorphonuclear cells, Eos: eosinophilic granulocytes, Sialomucins: total sialic 

acid (after hydrolysis). 

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


isocyanate, damage of the airways is characterized by a steep concentrationresponse 

curve. Based on acute 45-min exposure of rats the threshold for 

respiratory tract irritation is approximately 1 mg m −3 . Exposures equaling 

this concentration were tolerated without exposure-related effects, whether 

exposure occurred singly for 45-min or repeatedly for 2 weeks. Marginal 

effects were observed at 4 mg m −3 , all effects, including mortality, were 

produced at 7 mg m −3 . This demonstrates that selection of appropriate 

exposure concentrations appears to be most critical in the rat model. The 

assessment of diagnostic sensitivity of the methods used to probe damage 

to the respiratory tract demonstrated that respiratory function data, blood 

gas measurements, and BALF analysis facilitate a meaningful interpretation 

of the effects observed and are important adjuncts to common inhalation 

toxicological studies on rats to describe quantitatively the diseased state of 

the lung. 

The guinea-pig model is experimentally less demanding and therefore can 

suitably be used as a screening test for respiratory sensitization, as far as 

the limitations of this model are taken into account. Studies on guinea-pigs 

demonstrate that elicitation of respiratory hypersensitivity is 

challengeconcentration dependent and that the concentrations used should 

slightly exceed the threshold concentration for irritation to maximize the 

magnitude of the response. However, sensitization by inhalation may 

increase the susceptibility to irritant stimuli and thus confounds the 

selection of the most appropriate concentration for challenge. The 

combined approach of evaluating several breathing parameters, e.g. 

respiratory rate, flow- and volume-derived parameters, during both the 

hapten (free or conjugated) and the ACh challenge provides a promising 

method to distinguish specific and nonspecific hypersensitivity responses. 

Furthermore, it is critically important to assess the respiratory irritant 

potency of the compound under investigation. For potent irritant 

substances such as volatile isocyanates, challenge with the haptenprotein 

conjugate minimizes the likelihood to confound specific hypersensitivity 

responses with those evoked merely by irritation. Taking all imponderable 

factors into consideration, it appears that the guinea-pig intradermalinduction 

inhalation-challenge protocol is adequately susceptible to identify 

potent respiratory tract sensitizers. However, if the airway inflammation 

related features of asthma are the endpoints of primary interest other 

animal models appear to be more appropriate. 

References 

BARNES, P.J., BARANIUK, J.N. and BELVISI, M.G., 1991a, Neuropeptides in the 

respiratory tract (Part II). Am. Rev. Respir. Dis., 144, 1391–9.

J.PAULUHN 137 

BARNES, P.J., CHUNG, K.F., PAGE, C.P., 1991b, Pharmacology of asthma, 

Chapter 3, Inflammatory Mediators in Page, C.P. and Barnes, P.J. (Eds.), pp 

54–106. Handbook of Experimental Pharmacology, Berlin, Heidelberg, New 

York: Springer-Verlag. 

BOTHAM, P.A., HEXT, P.M., RATTRAY, N.J., WALSH, S.T. and WOOD- 

COCK, D.R., 1988, Sensitisation of guinea-pigs by inhalation exposure to lowmolecular-weight 

chemicals, Toxicol. Lett., 41, 159–73. 

BRIATICO-VANGOSA, G, BRAUN, C.J.L., COOKMAN, G., HOFMANN, T., 

KIMBER, I., LOVELESS, S.E., MORROW, T., PAULUHN, J., SØRENSEN, T. 

and NIESSEN, H.J., 1993, Respiratory Allergy. ECETOC Monograph No. 19. 

BROOKS, S.M., WEISS, M.A. and BERNSTEIN, I.L., 1985, Reactive airway dysfunction 

syndrome (RADS). Persistent asthma syndrome after high level irritant 

exposure, Chest, 88, 376–84. 

HAYES, J.P., DANIEL, H.R., TEE, R.D., BARNES, P.J., NEWMAN-TAYLOR, 

A.J. and CHUNG, K.F., 1992, Bronchial hyperreactivity after inhalation of 

trimellitic anhydride dust in guinea-pigs after intradermal sensitization to the 

free hapten, Am. Rev. Respir. Dis., 146, 1311–14. 

KAY, A.B., 1991, Asthma and inflammation, J. Allergy Clin. Immunol., 87, 893– 

910. 

KIPS, J.C., CUVELIER, C.A. and PAUWELS, R.A., 1992, Effect of acute and 

chronic antigen inhalation on airway morphology and responsiveness in 

actively sensitized rats, Am. Rev. Respir. Dis., 145, 1306–10. 

NIELSEN, G.D., 1991, Mechanisms of activation of the sensory irritant receptor by 

airborne chemicals, Crit. Rev. Toxicol, 21, 183–208. 

PAULUHN, J., 1994, Test methods for respiratory sensitization in use of 

mechanistic information in risk assessment, EUROTOX Proceedings, Arch. 

Toxicol., suppl. 16, 77–86. 

PAULUHN, J. and EBEN, A., 1991, Validation of a non-invasive technique to 

assess immediate or delayed onset of airway hypersensitivity in guinea-pigs, J. 

Appl. Toxicol, 11, 423–31. 

PAULUHN, J. and MOHR, U., 1994, Assessment of respiratory hypersensitivity in 

guinea-pigs sensitized to diphenylmethane-4,4'-diisocyanate (MDI) and 

challenged with MDI, acetylcholine or MDI-albumin conjugate, Toxicology (in 

press). 

VIJAYARAGHAVAN, R., SCHAPER, M., THOMPSON, R., STOCK, M.F. and 

ALARIE, Y., 1993, Characteristic modifications of the breathing pattern of 

mice to evaluate the effects of airborne chemicals on the respiratory tract, Arch. 

Toxicol, 67, 478–90.

10 

Mechanisms of Pulmonary Sensitization 

IAN KIMBER 

Zeneca Central Toxicology Laboratory, Macclesfield 

Introduction 

A wide range of chemicals is known to cause allergic contact dermatitis. It 

is apparent, however, that chemicals also have the potential to provoke other 

forms of allergy and of growing concern is pulmonary sensitization. 

Examples of chemicals identified as human respiratory allergens are listed 

in Table 10.1. Respiratory allergic hypersensitivity is characterized by 

pulmonary reactions which occur normally in only a proportion, and 

frequently in only a small proportion, of exposed individuals. In those who 

are sensitized, respiratory reactions can be provoked by atmospheric 

concentrations of the causative chemical allergen which were tolerated 

previously and which are without effect in the non-sensitized population 

(Newman Taylor, 1988). Almost invariably there is a latent period between 

the onset of exposure and the development of respiratory symptoms such 

as asthma and rhinitis. 

By definition, allergy, including sensitization of the respiratory tract, 

results from the stimulation of specific immune responses by the causative 

agent. Although it is assumed frequently that effective allergic sensitization 

of the respiratory tract results largely or wholly from inhalation exposure, 

this is not necessarily the case. Allergic reactions manifest in a particular 

organ commonly result from the local provocation by the inducing agent of 

a systemically sensitized individual. There is no reason to suppose that the 

quality of immune response necessary for sensitization of the respiratory 

tract may not result from exposure to the chemical allergen at a different 

site. Consistent with this is evidence that occupational respiratory allergy 

may be caused by dermal contact with the chemical (Karol, 1986; Nemery 

and Lenaerts, 1993). Furthermore, it has been reported that respiratory 

rate changes can be provoked by inhalation exposure of guinea pigs 

sensitized previously by either topical or subcutaneous treatment with the 

same chemical (Karol et al., 1981; Rattray et al., 1994). Despite the fact 

that, in practice, pulmonary sensitization may not be caused exclusively by

Table 10.1 Chemicals identified as human respiratory allergens 

I.KIMBER 139 

inhalation of the chemical allergen, it is likely that this is an important 

route of exposure in the occupational setting. 

It is well established that respiratory sensitization caused by protein 

aeroallergens is effected by IgE antibody. This class of antibody in man is 

homocytotropic and is able to associate, via specific membrane receptors, 

with mast cells, including mast cells in the respiratory tract. Following 

subsequent exposure of the sensitized individual to the same allergen, mast 

cell-bound IgE is cross-linked and this, in turn, results in mast cell 

degranulation and the release of both preformed and newly-synthesized 

mediators which provoke acute inflammatory reactions. In the case of 

sensitization of the respiratory tract caused by chemicals, however, an 

invariable association with the presence of specific IgE antibody has failed 

to emerge. Although IgE antibody specific for all recognized chemical 

respiratory allergens has been demonstrated, it is not uncommonly the case 

that individuals displaying symptoms of pulmonary hypersensitivity have 

been reported to lack demonstrable IgE. This may suggest that 

immunological processes independent of IgE antibody may play a decisive 

role in the induction of respiratory sensitization. An alternative explanation 

is that inappropriate or insensitive techniques have been employed for 

serological analysis and that IgE antibody may be associated more 

commonly than suspected previously with chemical respiratory allergy. In 

this context it is relevant that it has been found that positive skin prick 

tests can be provoked in patients sensitized to acid anhydrides who, on the 

basis of radioallergosorbent tests (RAST), were found to lack measurable 

levels of serum IgE antibody (Drexler et al., 1993). Despite the absence of 

formal confirmation that there exists a universal causal relationship 

between specific IgE and pulmonary hypersensitivity induced by chemicals,

140 MECHANISMS OF PULMONARY SENSITIZATION 

it remains likely that this class of antibody is responsible, in at least the 

majority of cases, for the acute onset symptoms associated with respiratory 

allergy (Karol et al., 1994). 

The induction and regulation of IgE responses 

IgE antibody responses are subject to a variety of immunoregulatory 

control mechanisms. Chief among these are the stimulatory and inhibitory 

actions of cytokines which serve to influence the induction and duration of 

IgE responses. It has been found in mice that interleukin 4 (IL-4) is 

necessary for the initiation and maintenance of IgE antibody production 

(Finkelman et al., 1988b). The essential role for this cytokine in IgE 

responses has been emphasized further by studies of mice homozygous for 

a mutation that inactivates the gene for IL-4. These animals lack detectable 

serum IgE and fail to mount IgE responses (Kuhn et al., 1991). Importantly, 

in mice which produce constitutively high levels of IL-4, significantly 

elevated concentrations of serum IgE are evident (Burstein et al., 1991). A 

balance to the promotional influence of IL-4 is provided by interferon 

(IFN- ), a cytokine which exerts an inhibitory affect on IgE responses 

(Finkelman et al., 1988a). The reciprocal antagonistic activity of these 

cytokines is not restricted to the mouse, IL-4 and IFN- have been found to 

regulate human IgE production (Del Prete et al., 1988; Pene et al., 1988). 

The cytokines which influence the integrity of IgE responses are the 

products of discrete subpopulations of T helper (Th) cells, lymphocytes 

characterized by possession of the CD4 membrane determinant. It has been 

found in both mouse and man that there exists a functional heterogeneity 

among Th cells. Two major populations, designated Th 1 and Th 2, have 

been described (Mosmann and Coffman, 1989; Romagnani, 1991). It is 

believed currently that these subsets represent the most differentiated forms 

of Th cells and develop from less mature precursors as the immune 

response evolves (Mosmann et al., 1991). The major functional distinction 

between Th 1 and Th 2 cells resides in the spectrum of cytokines which they 

elaborate (Mosmann and Coffman, 1989). The cytokine products of 

murine Th 1 and Th 2 cells are displayed in Table 10.2. 

It has been reported previously that chemicals known to cause 

respiratory hypersensitivity in man induce in mice immune responses 

characteristic of Th 2 cell activation, stimulate the production of specific IgE 

antibody and cause an increase in the serum concentration of IgE. 

Conversely, chemical allergens considered not to cause respiratory 

sensitivity, but which are nevertheless able to induce skin sensitization, 

elicit instead Th 1-type responses. In the latter case, immune responses are 

characterized by comparatively high levels of IgG2a antibody (an isotype 

known to be upregulated by IFN- ) and the absence of specific IgE 

(Dearman and Kimber, 1991, 1992; Dearman et al., 1991, 1992a,c,d,

Table 10.2 The cytokine products of murine Th 1 and Th 2 cells 

From: Mosmann and Coffman (1989). 

I.KIMBER 141 

1994). The implication is that certain chemicals favour the development of 

Th 2 cells which will then synthesize and secrete IL-4 and thereby encourage 

IgE antibody responses and mast cell sensitization. The converse is that 

other classes of chemical allergen preferentially stimulate Th1 cells and IFNproduction. 

Such conditions will be nonpermissive for IgE antibody 

production and cell-mediated immune responses, including contact 

sensitization, will be favoured instead. A selective stimulation by different 

classes of chemical sensitizers of divergent Th responses may provide an 

explanation at the cellular level for the observation that chemicals vary 

with respect to the nature of allergic reactions that they will elicit 

preferentially in man. The stimulation by chemical allergens of 

differentiated Th cell responses may have implications for allergic disease 

other than the regulation of IgE antibody. It is known for instance that 

IL-3, IL-4 and IL-10, all of which are products of murine Th 2 cells 

(Table 10.2), act as mast cell growth factors or cofactors (Smith and 

Rennick, 1986; Thompson-Snipes et al., 1991). Moreover, IL-5 is a growth 

and differentiation factor for eosinophils (Yokota et al., 1987) and serves 

to regulate the accumulation of these cells at the site of allergeninduced 

hypersensitivity reactions in the respiratory tract (Gulbenkian et al., 1992). 

It has been found recently that the cytokines IL-3 and IL-4 also enhance the 

secretory activity of mast cells following activation (Coleman et al., 1993). 

Antagonistic and inhibitory influences of Th cell products may also affect 

the elicitation of allergic reactions. It has been found that IFN- not only 

suppresses the secretory function of mast cells (Holliday et al., 1994), but 

also antagonizes the antigen-induced infiltration of eosinophils into the 

respiratory tract of sensitized mice (Iwamoto et al., 1993). Contact allergic 

reactions may in theory be regulated by Th 2 cytokines. It has been shown 

that IL-4 and IL-10 act in concert to inhibit Th 1 cell function and to


depress cell-mediated immunity (Powrie et al., 1993) and that Il–4 is able 

to reduce significantly the severity of contact allergic reactions in mice 

(Gautam et al., 1992). 

Taken together the available data suggest that the selective stimulation 

of Th cell responses and the consequent balance created between Th 1- and 

Th 2-derived cytokines will have an important impact on both the induction 

and elicitation stages of allergy. It is perhaps not surprising, therefore, that 

there is increasing evidence for selective Th responses in human allergic 

disease. Clones of T lymphocytes specific for aeroallergens such as house 

dust mite and grass pollen, which cause IgE-mediated respiratory allergic 

reactions in susceptible individuals, have been shown to elaborate Th 2 

cytokines, but not IFN- (Parronchi et al., 1991). A predominance of the 

Th 2-type cells has been found at sites of skin reactions in atopic individuals 

(Kay et al., 1991) and increased numbers of IL-4 + T lymphocytes have been 

identified in the nasal mucosa in allergen-induced rhinitis (Ying et al., 1994). 

By contrast, human immune responses to nickel, a common cause of 

allergic contact dermatitis, are characterized by the selective activation of 

Th 1-type cells. Allergen-specific T lymphocyte clones isolated from the 

peripheral blood of patients sensitized to nickel have been found to secrete 

only low or undetectable amounts of IL-4 and IL-5, but high levels of IFN- 

(Kapsenberg et al., 1991). 

Although the relative contribution of Th 1 and Th 2 cells during immune 

responses, and in particular the relative availability of IL-4 and IFN- , is 

likely to play a predominant role in the regulation of IgE antibody, other 

factors may be relevant. Not least, the priming of Th 1 cells for the 

production of IFN- may in turn be dependent upon the action of another 

cytokine, interleukin 12 (IL-12) (Manetti et al., 1994; Morris et al., 1994; 

Schmitt et al., 1994). It has been demonstrated also that CD8 + T 

lymphocytes exert an important immunoregulatory influence on IgE 

responses (Kemeny et al., 1994; Renz et al., 1994), possibly via downregulation 

of CD4 + Th 2 cell development (Noble et al., 1993). 

It is clear that conditions outwith the immune system also influence the 

magnitude of IgE responses. Certainly genetic predisposition is an 

important, although poorly understood factor. In addition, there have been 

suggestions that cigarette smoking and exposure to certain environmental 

pollutants may result in increased IgE levels and may also serve to 

aggravate asthma (Zetterstrom et al., 1981; Muranka et al., 1986; 

Wardlaw, 1993). 

Cell-mediated immune responses in chemical respiratory 

allergy 

The elicitation of chemical respiratory hypersensitivity may be associated 

with both immediate-onset and late phase reactions. While IgE antibody

and local degranulation of mast cells may be necessary for acute 

symptoms, late asthmatic responses appearing some hours following 

exposure are characterized by an infiltration of mononuclear cells and 

increased numbers of leucocytes in bronchoalveolar lavage fluid. Chronic 

inflammation is an important component of asthma and, in addition to 

mononuclear cell accumulation, is characterized by mucus production, the 

destruction and sloughing of airway epithelial cells and subepithelial 

fibrosis secondary to collagen deposition. Eosinophils, acting together with 

infiltrating T lymphocytes, play a pivotal role in chronic bronchial 

inflammation (Corrigan and Kay, 1992). It is apparent also that the 

generation of eosinophilia in the respiratory tract is influenced markedly by 

Th cell products. As described previously, IL-5 effects the accumulation of 

eosinophils at the site of hypersensitivity reactions in respiratory tissues, 

while IFN- , secondary to an inhibition of CD4 + cell infiltration, 

antagonizes this process (Gulbenkian et al., 1992; Iwamoto et al., 1993). It 

may prove that the cell-mediated immune processes relevant to the 

development of respiratory hypersensitivity and asthma are also a function 

of Th cell heterogeneity. Certainly the stimulation of Th 2 cell activation 

will have profound effects on all stages of respiratory allergy. The 

infiltration of such cells into sites of encounter with inducing allergen, a 

process perhaps facilitated by vasodilation resulting from mast cell 

degranulation, will provide a local source of cytokines such as IL-4 and 

IL-5. Mast cell secretory activity will be potentiated by the former and 

eosinophil accumulation triggered by the latter. That Th 2 cells do in fact 

accumulate in the area of immediate-type hypersensitivity reactions is 

supported by the studies of Kay et al. (1991) who demonstrated that the 

cells infiltrating lesional skin at the sites of late phase cutaneous reactions 

in atopic patients produce IL-3, IL-4, IL-5 and GM-CSF, but not IFN- . 

Practical applications 

I.KIMBER 143 

In the course of investigations designed to examine the characteristics of 

immune responses induced in mice by chemical sensitizers it was found 

that only those materials known to cause respiratory hypersensitivity in 

man provoked in mice a substantial increase in the serum concentration of 

IgE; a phenomenon thought to reflect the selective stimulation of Th 2 celltype 

responses by this class of allergen. It was observed also that contact 

allergens known or suspected not to cause occupational respiratory 

hypersensitivity failed to result in similar changes in serum IgE levels 

(Dearman and Kimber, 1991, 1992; Dearman et al., 1992a,d). The 

differential ability of chemical respiratory and contact allergens to 

stimulate changes in the concentration of serum IgE in mice forms the basis 

of a novel approach to the identification of chemicals which have the 

potential to cause sensitization of the respiratory tract. This method, the


mouse IgE test (Dearman et al., 1992b, Kimber and Dearman, 1993) is 

being evaluated currently in the context of internal and inter-laboratory 

validation studies. 

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11 

Occupational Asthma Induced by Chemical 

Agents 

C.A.C.PICKERING 

Wythenshawe Hospital Manchester 

Introduction 

Occupational asthma may be defined as variable airways narrowing 

causally related to exposure in the working environment to airborne dust, 

gases, vapours or fumes (Newman Taylor, 1980). This definition includes, 

therefore, both immunological and nonimmunological causes of asthma in 

the workplace. Immunological causes of asthma in general demonstrate a 

latent period between exposure and the development of symptoms. Once 

sensitisation has occurred airway responses may be seen at very low levels 

of exposure. Both high and low molecular weight agents may cause 

sensitisation. Irritantinduced occupational asthma characteristically follows 

within 24 h of a usually, single, high level exposure to an irritant substance 

and has been named reactive airways dysfunction syndrome (Brooks et al., 

1985). 

The number of chemical agents causing occupational asthma is 

extensive. As new, highly reactive, chemicals are developed these numbers 

are likely to grow. Low molecular weight chemicals may act as haptens, 

reacting with body protein to form a complete antigen to which specific 

antibodies are formed. 

The incidence of occupational asthma in most countries is not known 

with any great accuracy, there are considerable variations in reporting 

systems between countries. Since many individuals with occupational 

asthma change jobs without a specific diagnosis being established, the 

published figures of incidence will be significant underestimates of the true 

incidences. In Japan (Kobayashi et al., 1973), the prevalence of 

occupational asthma amongst adult male asthmatics is said to be about 

15%. In the UK a new reporting system has recently been established— 

Surveillance of Work-related and Occupational Respiratory Disease Project 

(SWORD). Newly diagnosed cases of workrelated respiratory disease are 

reported monthly by consultant chest and occupational physicians. 

Between 1989 and 1991, 631 cases of chemically induced occupational 

asthma were reported, of these 53% were associated with exposure to

150 OCCUPATIONAL ASTHMA INDUCED BY CHEMICAL AGENTS 

isocyanates. A similar system (SHIELD) has been established in the UK in 

the West Midlands (Gannon et al., 1993). They reported an incidence of 43 

new cases per million workers per year. Specific occupational incidences 

varied from 1833 per million paint sprayers to 8 per million clerks. Again 

more than half the cases of asthma were attributed to isocyanates. 

The initial diagnosis of occupational asthma is based on a workers 

history of respiratory symptoms improving on days away from work and 

when on holiday. At the onset of occupational asthma this pattern is 

usually present but continued exposure to the allergen leads to increasing 

airway reactivity. Their symptoms may then persist over weekends and be 

triggered by nonspecific factors outside the workplace such as exhaust 

fumes, aerosol sprays and perfumes. The difficulties of relying on the 

history alone in the diagnosis of occupational asthma has been well 

documented (Malo et al., 1991). A series of 162 hospital referrals to two 

expert physicians were initially categorised, on the basis of their histories, 

into highly probable, probable, uncertain, unlikely or absent occupational 

asthma. The diagnosis was then established by bronchial provocation 

testing and or serial measurements of lung function. The predictive value of 

a physician’s assessment of occupational asthma being highly probable or 

probable was only 63%. This improved to 83% in the groups in whom 

occupational asthma was assessed as being unlikely or absent. 

The early identification of work-related symptoms and their subsequent 

investigation in the workplace is important. Only rarely, when very acute 

episodes of workplace asthma are described, should lung function 

measurements at work be avoided. While pre- and post-shift measurements 

of lung function may identify a work-related effect, late asthmatic 

responses occurring in the evening after leaving work are frequently seen in 

chemically induced forms of occupational asthma. Serial measurements of 

lung function made every 2 h from waking to sleeping, both on working 

days and on days away from work, using a peak flow meter, will identify 

these late responders. The sensitivity of this type of investigation in 

establishing a diagnosis of occupational asthma is about 80% (Burge, 1982), 

this falls to 46% once the worker is started on specific treatment for his 

asthma, again emphasising the importance of early identification and 

investigation of work-related respiratory symptoms. 

Currently more than 140 low molecular weight chemicals have been 

reported to induce occupational asthma (Butcher and Salvaggio, 1986). 

The majority of these chemicals induce asthma by mechanisms which have 

yet to be identified. In a minority of instances specific IgE antibodies to the 

implicated chemical have been identified. 

Bronchial challenge tests with chemicals which are non-IgE dependent 

usually induce either an isolated late asthmatic response or a biphasic or 

dual asthmatic response. The IgE dependent responses induce immediate or 

dual asthmatic responses.

The most common chemical causes of occupational asthma include the 

isocyanates and the acid anhydrides. This chapter will examine these two 

groups in more detail. 

Isocyanates 

C.A.C.PICKERING 151 

The polyisocyanates and their oligomers are the most important cause of 

chemically induced asthma. These organic compounds are synthesised by 

the reaction of amines with phosgene. There are a number of related 

compounds the most important of which are 2,4- and 2,6-toluene 

diisocyanate (TDI), methylene diphenyldiisocyanate (MDI), hexamethylene 

diisocyante (HDI), napthalene diisocyanate (NDI), isophorone diisocyanate 

(IPDI), and polyisocyanates derived from HDI and MDI. 

The incidence of occupational asthma due to diisocyanates varies widely. 

It is influenced by the type of compound and its vapour pressure. TDI and 

HDI are highly volatile at room temperature, whereas MDI has to be 

heated to above 60°C to volatilise. It is thought that approximately 5% of 

an exposed working population will develop occupational asthma after 

exposure to TDI (Diem et al., 1982). Because of the known respiratory 

problems associated with exposure to isocyanates with high vapour 

pressure properties, new isocyanate compounds with low vapour pressure 

properties have been developed particularly for use in the paint spraying 

industry. Recent studies however continue to demonstrate significant levels 

of occupational asthma despite the use of recommended respiratory 

protection (Seguin et al., 1987, Welinder et al., 1988). Bronchial 

provocation studies with HDI- and MDI-derived polyisocyanates have 

confirmed their ability to cause occupational asthma. Airborne iso-cyanate 

prepolymers appear to be able to induce asthma to the same or greater 

frequency as isocyanate monomers. 

High exposures to isocyanate vapours, such as occur in a major 

industrial spillage, cause acute rhinitis, lacrymation, cough and wheezing 

leading to subsequent sensitisation. In some individuals this type of 

exposure induces persistent asthma—reactive airways dysfunction 

syndrome (RADS). Respiratory sensitisation may occur at very low levels 

of exposure. Pepys et al. (1972) described a boat builder who became 

sensitised to TDI at exposure levels of between 0.00173 and 0.0018 ppm. 

Similarly White et al. (1980) reported respiratory symptoms and the 

development of IgE antibodies to TDI, in machinists manufacturing carseat 

covers exposed to levels of TDI of between 0.0003 and 0.003 ppm. It is 

more usual, in the author’s experience, for the sensitised individual to 

provide a history of short lived peak exposures to isocyanates which have 

clearly been above the current threshold limit value. These intermittent 

relatively high level exposures may be important in the sensitisation


process. Once sensitised, a worker may have his symptoms initiated by very 

low exposure levels of isocyanates. 

Diisocyanate asthma is usually but not always associated with the 

presence of nonspecific bronchial hyperreactivity. The majority of workers 

who develop occupational asthma remain symptomatic requiring regular 

treatment permanently after cessation of exposure (Allard et al., 1989). 

The duration of exposure with symptoms before diagnosis has a major 

influence on recovery patterns. In a group of 43 isocyanate workers with 

occupational asthma, those who had fully recovered were exposed with 

symptoms for 1.6 years, those who had improved, 2.8 years and those who 

had not improved, 5.4 years (Pisati et al., 1993). The resolution or 

improvement in occupational asthma takes place over a 2 year period after 

cessation of exposure, symptoms still present at 2 years should be regarded 

as permanent. Most epidemiological studies have not identified any specific 

risk factors including atopic status, smoking or nonspecific bronchial 

hyperreactivity. 

The laboratory identification of specific antibodies to diisocyanates has 

proved of very limited value. Diisocyanate specific IgE is demonstrable in 

only 10–20% of sensitised individuals and have also been identified in 

individuals with no history of asthma (Butcher et al., 1983). Similarly 

specific IgG antibodies to diisocyanates have been described in workers 

both with and without evidence of disease. 

At the present time the recommended long-term exposure limit (8 h 

TWA reference period) for diisocyanates is 0.02 mg m −3 and the short-term 

exposure limit (10 min reference period) is 0.07 mg m −3 in the UK. There is 

discussion at the present time as to whether levels should be lower in order 

to prevent the development of diisocyanate asthma. However since most 

workers with diisocyanate airways disease describe exposures in excess of 

the current recommended exposure levels the prevalence of occupational 

asthma in a workforce without such exposures is not known. There need to 

be improvements in hygiene control to prevent these peak exposures to 

isocyanates. 

Acid anhydrides 

The acid anhydrides are a group of low molecular weight chemicals used as 

curing agents in the production of epoxy and alkyd resins and in the 

production of plasticisers such as dioctyl phthalate. Acid anhydrides exert 

diverse effects on man both as sensitisers, irritants or both. The most 

frequently used anhydrides, all of which have been described causing 

occupational asthma, are phthalic anhydride (PA), trimellitic anhydride 

(TMA), tetrachlorophthalic anhydride (TCPA) and maleic anhydride 

(MA). In addition himic anhydride and pyromellitic dianhydride (PMDA) 

have been described as causing asthma.

The direct toxicity of anhydrides involves irritation of the mucus 

membranes and skin which may result in eye lesions, epistaxis, pulmonary 

congestion, haemoptysis and skin burns (Venables, 1989). 

Occupational asthma is most frequently reported due to PA, less 

commonly to TMA, TCPA and MA and finally there are single case reports 

of asthma due to HA, HHPA and PMDA (Venables, 1989). A second type 

of response to acid anhydrides has also been described and is termed the 

‘late respiratory systemic syndrome’ (LRRS). This is characterised by the 

development of influenzal type symptoms, fever, generalised acheing and 

malaise, late in the working shift or in the evening after work. These 

symptoms may occur in isolation or in association with asthma. It is not 

clear whether this response is immunologically mediated or a nonspecific 

response to high levels of anhydride exposure. Lastly, exposure to TMA, 

probably at high exposure levels, has been described as causing severe 

pulmonary haemorrhage requiring both blood transfusion and mechanical 

ventilation (Rivera et al., 1989). 

Serum IgE and IgG antibodies to acid anhydrides have been identified. 

IgE antibodies appear to be more specifically associated with occupational 

asthma. Howe et al. (1983) reported seven cases of TCPA asthma all of 

whom had IgE antibody to TCPA, compared with 8% of 300 exposed 

workers without TCPA asthma; 29% of this exposed nonasthmatic 

population had IgG antibodies to TCPA. 

The exposure levels of acid anhydrides that initiate sensitisation are 

poorly understood. TMA at levels of 1.7–4.7 mg m −3 (Zeiss et al., 1977) 

and 0.007–2.1 mg m −3 (Bernstein et al., 1983; McGrath et al., 1984) have 

been described causing occupational asthma. PA at 0.03–15 mg m −3 has 

also been reported as causing asthma (Wernfors et al., 1986). As in other 

forms of occupational asthma, the early identification of cases of acid 

anhydride induced asthma and their removal from exposure is of prime 

importance. 

Reactive airways dysfunction syndrome 


Reactive airways dysfunction syndrome (RADS) or irritant-induced asthma 

was first described in 1985 (Brooks et al., 1985). The criteria used in 

diagnosis include a high level exposure to an irritant fume, vapour or smoke, 

the development of respiratory symptoms within minutes or hours of 

exposure, in an individual with no previous history of respiratory symptoms, 

with persistence of symptoms and physiological abnormalities for more 

than 1 year. A variety of different chemical exposures have been described 

inducing this syndrome including: chlorine (Moore and Sherman, 1991), 

glacial acetic acid (Kern, 1991), hydrochloric acid (Promisloff et al., 1990) 

and miscellaneous chemical exposures (Brooks et al., 1985). A comparison 

between cases of occupational asthma and RADS (Gautrin et al., 1994)


suggest that cases with RADS are left with less airway reversibility than 

occupational asthmatics. This would be consistent with the pathological 

findings (Boutet et al., 1993) in RADS, with more severe basement 

membrane thickening and bronchial wall fibrosis than is present in 

occupational asthma. 

The development of occupational asthma in any individual has 

potentially serious consequences both in terms of persisting disability, 

possible unemployment and loss of income (Gannon et al., 1993). It is 

incumbent on management to ensure safe working conditions with 

adequate control and regular monitoring of atmospheric levels of chemical 

agents. 

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PART FOUR 

Biomarkers and risk assessment of 


12 

Biomarkers and Risk Assessment 

KARI HEMMINKI 

Karolinska Institute, Huddinge 

Introduction 

Many chemical carcinogens cause covalent DNA-binding products, 

adducts, which may induce mutations or other types of DNA damage in 

important growth-controlling genes or loci resulting in aberrant cellular 

growth and cancer (Harris, 1991; IARC 1992; Hemminki, 1993). Human 

exposure to compounds such as polycyclic aromatic hydrocarbons (PAH) 

can be determined, for example, by ambient air, biological or DNA adduct 

monitoring. The usefulness of a method for the determination of DNA 

adducts in human biomonitoring requires high sensitivity because the levels 

of adducts are low. Here the primary focus is on the assessment of 

exposure using the above indicators in industries where high exposure to 

PAHs occur, such as iron founding, coke production, aluminium 

production, garage work and engine overhauling with exposure to used 

lubricating oils. 

Biomonitoring of PAH exposure 

Literature on the application of DNA adduct studies in humans is extensive 

(Beach and Gupta, 1992; IARC, 1993, 1994; Hemminki et al., 1993a; 

Hemminki, 1994). A large majority of the 32 P-postlabelling studies on 

human samples focus on tobacco smoking, occupational exposures and 

cancer chemotherapy patients. Most occupational exposures studied relate 

to complex mixtures, including polycyclic aromatic hydrocarbons (PAHs). 

In exposure to complex mixtures multiple radioactive spots (called diagonal 

radioactive zones, DRZ) are detected. The adduct spots cannot be 

definitively identified nor quantitated. As it has turned out that for many 

adducts labelling is not completed, even among structural analogues such 

as PAHs, the adduct levels measured are likely to be underestimates 

(Segerbäck and Vodicka, 1993).

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

K.HEMMINKI 159 

Notes: 

a Total white blood cells. 

b Lymphocytes. 

c Not given. 

d No data. 

The types of occupational groups studied by postlabelling include 

foundry, coke oven and aluminium workers, roofers, garage and terminal 

workers, car mechanics and chimney sweeps. All these groups have had an

160 BIOMARKERS AND RISK ASSESSMENT 

increased risk of lung cancer. As, however, epidemiological studies relate to 

exposure a few decades earlier the risks of present exposures can only be 

predicted. The levels of aromatic adducts are elevated in white blood cells 

or lymphocytes in many of these groups. The reported total aromatic 

adduct levels usually range between 1 and 10 adducts per 10 8 nucleotides. 

Long-lived lymphocytes tend to have higher levels of adducts than shortlived 

granulocytes (Savela and Hemminki, 1991; Grzybowska et al., 1993). 

As a rule of thumb, it can be assumed that in a steady-state (i.e. long term 

exposure) lymphocytes contribute to the level of adducts overwhelmingly. 

Because they represent about 25 percent of the DNA in blood, the 

relationship between total white blood cell (WBC) and lymphocyte DNA 

adducts should be about 1:4, granulocytes only contributing to the amount 

of DNA denominator. Yet one has to be cautious in the comparison of 

results between various assays even within a laboratory as the results may 

‘drift’ with time. 

The levels of white blood cell/lymphocyte aromatic adducts from 

workers in several industries, as measured by postlabelling, and as 

compared to ambient air concentrations of benzo (a) pyrene (BP) and 1hydroxypyrene 

levels are presented in Table 12.1. A boxplot presentation 

of the adduct levels of bus maintenance and truck terminal workers is 

shown in Figure 12.1 (Hemminki et al., 1994). The differences that were 

statistically significant from the controls were, in addition to the groups of 

maintenance and terminal workers, garage workers and diesel forklift 

drivers. 

There does not seem to be a direct relationship between exposure and 

adduct levels. Electrode, coke and aluminium workers, exposed up to 

several 100 ng m −3 concentrations of BP, do not differ from the control 

more than foundry workers, exposed to less than 1/10 of the cited levels. 

The apparently higher level of adducts in the aluminium and electrode 

former workers (and controls) as compared to the other measurements, is 

due a method applied earlier with higher amounts of radioactive ATP. The 

later assays were carried out in small volumes but high concentrations of 

ATP (Hemminki et al., 1993b; Szyfter et al., 1994). An increased level of 

lymphocyte adducts has also been found in garage and truck terminal 

workers, with estimated exposures of about 10 ng m −3 (Hemminki et al., 

1994). This would imply that the detection limit of the postlabelling 

method in humans exposed to PAHs lies somewhere between 1 and 10 ng 

m −3 BP. Whether diesel exhaust is a particularly potent inducer of adducts 

remains to be demonstrated. The differences between the exposed and the 

controls are statistically significant among foundry workers, all bus 

maintenance personnel and garage workers as a subgroup, all truck 

terminal workers and the diesel forklift drivers in particular. Coke 

workers differed significantly from the local controls in summer when

environmental pollution was low and the adduct levels in the controls were 

about 1/10 of their level in the winter (Grzybowska et al., 1993). 

Adducts and other endpoints 


Figure 12.1 A boxplot of the white blood cell DNA adduct levels (per 10 8 

nucleotides) among bus maintenance and truck terminal workers and controls 

(Hemminki et al., 1994). 

It has become customary to include many types of endpoints to 

biomonitoring studies. The foundry study cited in Table 12.1 belongs to 

the most versatile of them. Exposure is measured by ambient air and 1hydroxypyrene 

monitoring (Santella et al., 1993). DNA adducts are 

assayed for by postlabelling and immunoassay. Plasma albumin PAH 

adducts are measured. Hypoxanthin guanine phosphoribosyl transferase 

(HPRT) and glycophorin A mutations are assayed for in lymphocytes and 

erythrocytes, respectively (Perera et al., 1993, 1994). Single-stand breaks in 

DNA and three types of cytogenetic parameters, chromosomal aberrations, 

sister chromatid exchanges and micronuclei, are analysed, in addition to 

genotyping of drug metabolising enzyme genes. Sampling of workers was 

repeated in four consecutive years, each at the same time of the year. As the 

last sampling was in the end of 1993, it will take some time before the 

complete data set will be available for analysis. 

In some published work from this data set an increase in DNA adducts 

and mutation frequency in the HPRT and glycophorin A genes was 

reported (Figure 12.2). Yet unreported results appear to show an increase


Figure 12.2 Total white blood cell DNA adducts, determined by immunoassay, 

HPRT and glycophorin A mutations in foundry workers, exposed to various levels 

of BP (Perera et al., 1993). 

in singlestrand breaks while none of the cytogenetic parameters are 

elevated in the foundry workers. 

Adducts and metabolic genotypes 

The modulation of environmental carcinogenesis by host polymorphism in 

genes for xenobiotics metabolising enzymes is currently under extensive 

investigation. It was initially sparked by findings linking certain 

phenotypes of drug metabolism to cancer risk (Seidegård et al., 1986; 

Nebert, 1991). The enzymes of interest in the context of exposure to PAHs 

include cytochrome P450 CYP1A1 and glutathione transferase GST, 

involved in the activation and inactivation, respectively, of PAHs. By 

restriction enzyme mapping two allelic forms, ml and m2, and two other 

closely linked codons for isoleucine (Ile, linked to ml) and valine (Val, 

linked to m2) can be defined, where m2 and valine represent the rare 

mutant genotypes, associated with the inducibility of the enzyme activity 

(Hayashi et al., 1991). Polymorphism in GST1 involves the presence or the 

absence of the gene (Nakachi et al., 1992). The null genotype lacks the 

enzyme completely. 

Among chimney sweeps there was an association of the rare, inducible 

CYP1A1 genotype ml/m2 with low adduct levels in white blood cell DNA 

(Ichiba et al., 1994). In the same study an increased level of adducts was 

noted in the GST1-individuals. The level of DNA adducts appeared to be 

related to both the GST and CYP1A1 genotype (Figure 12.3). Analysis of 

micronuclei in chimney sweeps resulted in no differences between 

individuals of either CYP1A1 ml/m2, m2/m2 or Ile/Val genotypes nor of 

GST1 + or − genotypes (Carstensen et al., 1993). There was however a

correlation between white blood cell DNA adducts and micronuclei and it 

was stronger among the GST-individuals (Ichiba et al., 1994). 

How important is the role of metabolic phenotype or genotype as a 

predictor of cancer risk remains to be established. However it would seem 

prudent to assume some role as long as there is significant exposure to a 

carcinogen, metabolism of which is regulated by polymorphic genes. It 

would be important to note that the question can only be addressed if both 

of these conditions are met. In much of the published literature there are 

uncertainties regarding the active agents and their metabolic routes in the 

tissues studied. Adjustment for a metabolic phenotype or genotype, when 

justified, may increase the precision in the measurement. 

Risk assessment 


Figure 12.3 Total white blood cell DNA adducts, measured by postlabelling, 

according to CYP1A1 and GST1 genotype (Ichiba et al., 1994). Controls, 

Sweeps. 

Monitoring of DNA adducts in occupational setting has mainly been 

applied to workers exposed to PAHs. In the case of 32 P-postlabelling 

increases in the level of adducts has been noted at exposures around 10 ng 

BP m −3 or slightly below. This is close to the detection limit that can 

conveniently be attained with personal monitoring or by measuring urinary 

1-hydroxypyrene. As the adduct measurements also reflect some aspects of 

metabolism and DNA repair, they extend the scope of exposure


measurements to host factors that may underly individual susceptibility to 

cancer. 

It has become increasingly common to try and incorporate other 

endpoints to DNA adduct studies. These include metabolic parameters, 

discussed above, protein adducts, cytogenetic parameters and point 

mutations. Examples include ethylene oxide exposed workers (Tates et al., 

1991) and foundry workers (Perera et al., 1993, 1994; Santella et al., 

1993). In both studies several parameters were elevated. The study on 

chimney sweeps illustrated how the intermediary endpoint may increase 

precision in the measurements (cf. Figure 12.3). The initial study showed 

no correlation between sweeping and micronuclei even though an 

adjustment was made for CYP1A1 and GST genotypes (Carstensen et al., 

1993). There was a moderate correlation between sweeping and white 

blood cell DNA adducts, and adducts and micronuclei. Both of these 

correlations were strengthened once GST genotype was considered (Ichiba 

et al., 1994). 

Increasing circumstantial evidence associates DNA adducts within 

groups to an increased risk of cancer (IARC, 1992; Hemminki, 1993). 

Many of the adduct studies have been carried out in occupational groups 

which have been at a risk of cancer based on epidemiological results. These 

studies may be old and relate to exposures decades ago. Even new 

epidemiological publications on cancer cannot accurately address 

exposures after about 1970. Simultaneously there have been large changes 

in technology and industrial hygiene, undermining the quantitative and 

sometimes even the qualitative findings of the old epidemiological studies. 

This is one justification for the biomonitoring studies. 

Another justification is on exposures where epidemiological studies have 

not been conducted or have provided inadequate results, in spite of 

suspicions raised by short-term or animal experiments. The International 

Agency for Research on Cancer has pointed out this as one of the criteria 

to be used in the evaluation of carcinogenicity of chemicals (IARC, 1992). 

Styrene belongs to this group of industrial exposures, where 

epidemiological findings are equivocal but adduct data are available on 

workers (Vodicka et al., 1993). 


The research was supported by the Swedish Medical Research Council and 

Work Environment Fund.

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13 

Extrapolation of Toxicity Data and Assessment 

of Risk 

NORBERT FEDTKE 

Hüls AG, Marl 

Introduction 

Risk assessment provides a link between scientific research and risk 

management, or in other words, it is ‘a method for reaching public policy 

decisions’ (Silbergeld, 1993). Risk assessment includes the key elements 

hazard identification, dose-response assessment, and exposure assessment. 

These elements are integrated in a risk characterization step to predict 

adverse effects that may occur in a given population in a particular 

exposure situation, often based on the quantification of the likelihood of this 

occurrence. Risk management determines whether the particular exposure 

situation presents an acceptable or unacceptable risk and whether it is 

necessary to reduce the risk by reducing the exposure. Whereas risk 

management has to account for public health, socio-economical factors, 

technical feasibility, social perceptions, governmental policy and political 

consequences, risk assessment should be based on scientific principles. 

Since for the majority of industrial chemicals no or only limited human 

data exist, the question of how to extrapolate the data obtained from 

laboratory studies in experimental animals in order to predict the effects in 

humans has become one important aspect in risk assessment. The final 

goal is either to determine a level of exposure at which there is no reasonable 

doubt that an adverse effect will not occur in man or to define the risk 

associated with this exposure level. The use of mechanistic information to 

provide linkages between exposure, dose to tissue, and biological responses 

may assist in some of the steps necessary in the process of species 

extrapolation. Especially the use of physiologically based pharmacokinetic 

models (PBPK) for some aspects of risk assessment has been promoted to 

reduce the uncertainty associated with the current default methods. PBPK 

modeling is explained in general terms and a recently developed PBPK model 

for 2-butoxyethanol is provided as an example to illustrate the use of 

kinetic and mechanistic data in risk assessment.

168 EXTRAPOLATION OF TOXICITY DATA AND ASSESSMENT OF RISK 

Legislative background in the European Union 

Although the process of risk assessment is not new for the individual 

member states of the European Union (EU) and representatives of industry 

and government have been assessing risk for human health and the 

environment for decades, EU legislation for existing 1 and new 2 substances 

did not formally require a systematic risk assessment up to 1992. The 

situation has changed with the Seventh Amendment of the Directive on the 

Classification, Packaging and Labelling of Dangerous Substances (EEC, 

1992) and the existing substances regulation (EEC, 1993a), which address 

risk assessment of new and existing substances, respectively. 

For new chemicals, the general principles of risk assessment are defined 

in Commission Directive 93/67/EEC (EEC, 1993b). In addition, the 

Directorate General XI of the European Commission has issued a series of 

draft guidance documents for use by the competent authorities appointed 

by the member states. These documents provide the technical details for the 

risk assessment of new substances mainly by defining the testing strategies 

for individual toxic endpoints. For existing chemicals, a guidance 

document has been drafted. However, these technical guidance documents 

provide only little information on how to extrapolate laboratory data to 

humans. It has to be assumed that the extrapolation principles in the EU 

member states will be based on historical approaches used by authorities in 

other countries. 

Approaches to risk assessment 

The final goal of species extrapolation is to define a dose or dose rate 

which produces no adverse effects in humans. The estimation of a human 

no-effectlevel may include: 

– determination of the appropriate animal species for extrapolation to 

man, 

– determination of the most critical effect(s) and the target organ(s), 

– determination of the no-observed-adverse-effect level(s) (NOAEL) of 

this effect(s), often followed by 

– extrapolation of the NOAEL(s) from subacute or subchronic to chronic 

exposure (time extrapolation), 

– extrapolation of effects observed in a high-dose region of a dose 

response curve to a low-dose region, 

1 Listed in the European Inventory of Existing Commercial Substances 

(EINECS). 

2 Not listed in EINECS.

N.FEDTKE 169 

– extrapolation of effects from one route of exposure to another route, 

and 

– extrapolation of effects observed in a rather homogeneous animal 

population to a heterogeneous human population (interspecies 

extrapolation), which also has to take into account the existence of 

subgroups regarded as more sensitive as the rest of the population 

(intraspecies extrapolation). 

Essential for all procedures used in health risk assessment is the 

determination of the so-called critical effect. The critical effect may be 

defined as the adverse effect judged to be most appropriate as the basis for 

the risk assessment. Hence, the first step is the review of all available data 

on a chemical and the assessment of the adequacy of the database for the 

determination of the critical effect. On the basis of the critical effect, 

toxicants may be divided into two classes characterized by: 

– a threshold of response, i.e. the adverse effect on health is not expressed 

until the chemical, or the ultimately toxic metabolite, reaches a 

threshold dose or dose rate in the target tissue, or 

– no threshold of response, i.e. there is no threshold exposure level below 

which effects will not be expressed. This implies that there is some risk at 

any level of exposure. Examples are genotoxic carcinogens or germ cell 

mutagens. 

Based on these classes two general approaches to health risk assessment 

have been used. 

The first approach involves the use of ‘safety factors’ applied to the 

NOAEL or the lowest-observed-adverse-effect level (LOAEL) of a 

threshold effect determined in experimental animals (safety factors are 

recently referred to as ‘uncertainty’ or ‘assessment’ factors). The magnitude 

of the uncertainty factors varies between the regulatory bodies that are 

concerned with risk assessment, but usually they take into account the 

interspecies extrapolation (default factor 10) and intraspecies extrapolation 

(default factor 10). The magnitude of the default factors appears to be 

based more on the conventional use of the decimal system than on 

scientific reasons and have been proposed first by Lehman and Fitzhugh 

(1954) for the derivation of acceptable daily intakes (ADIs) for food 

additives. Additional uncertainty factors may be used for extrapolation to 

chronic exposure from subacute or subchronic exposure, adequacy of the 

database, extrapolation of a LOAEL to a NOAEL and severity of effects. 

The resulting overall uncertainty factor often reaches values of 1000 or 

higher, which is an indication of the imprecision of the derived tolerable 

intake. Refined extrapolation procedures using subdivisions of the default


factors or different default factors have recently been published (Lewis et 

al., 1990; Renwick, 1991, 1993). 

The second approach (for non-threshold effects) also relies mainly on 

default assumptions for dose-response extrapolation and cross-species 

extrapolation. Especially cancer risk assessment has been the subject of 

much debate and there are a number of extrapolation methods reviewed 

recently by Park and Hawkins (1993) and Hallenbeck (1993). The default 

methodology in the . has been summarized by Frederick (1993). In 

principle, the risk assessment is based on a chronic rodent bioassay 

conducted at or near the maximum tolerated dose (MTD). The lifetime 

constant dose rates and the tumour incidence data for the individual dose 

groups are used to determine the dose response by fitting the data with a 

computer program. The linearized multistage cancer model (LMS) is often 

used to perform this step. The LMS model extrapolates the rodent tumor 

data observed at the MTD to a dose with a predefined risk and the 95 per 

cent upper bound on the dose-response curve is calculated. The interspecies 

extrapolation to humans is performed by a correction factor based on body 

weight or body surface. Subsequently, the dose is determined that 

corresponds to a maximum allowable calculated upper bound on risk. The 

resulting number does not describe the actual human risk under low-level 

environmental exposure, but provides an upper bound to human risk that 

is assumed not to be exceeded. The actual risk may be in the range between 

0 and the upper bound. In the process described, the dose is defined as 

administered dose or inhaled concentration. As a result, the lowdose 

extrapolation does not take into account non-linearities in tissue dosimetry 

and response. In addition, the interspecies extrapolation is performed using 

a default approach that does not account for mechanistic species 

differences. 

Use of PBPK models in risk assessment 

General description 

Physiologically based pharmacokinetic (PBPK) models have been used 

increasingly over the past decade to improve several aspects of the 

assessment of risk associated with human exposure to chemicals. Examples 

are PBPK models for styrene (Ramsey and Andersen, 1984; Csanády et al., 

1994), dichloromethane (Andersen et al., 1987), 1,4-dioxane (Reitz et al., 

1990a), chloroform (Reitz et al., 1990b), ethyl acrylate (Frederick et al., 

1992), methanol (Horton et al., 1992) and 1,3-butadiene (Johanson and 

Filser, 1993). Recent reviews of the use of PBPK models in risk assessment 

have been published by several authors (Frederick, 1993; Travis, 1993; 

Wilson and Cox, 1993; Andersen and Krishnan, 1994).

N.FEDTKE 171 

PBPK models are based on the blood and tissue solubility of chemicals, 

their metabolism in various tissues and the physiology of the organism, 

thus incorporating the specific physiological description of animal species as 

well as specific physico-chemical descriptions of agents. Uptake, 

distribution, metabolism and excretion are described in physiologically 

realistic compartments (tissue groups) using computer simulation. The 

compartments are linked in parallel, represent the actual mammalian 

architecture, and include tissues such as lung and arterial blood, fatty 

tissue, poorly perfused tissues (muscles, skin), richly perfused tissues 

(brain, kidneys, heart, endocrine gland, gastro-intestinal tract), liver as the 

main metabolizing tissue, and mixed venous blood. The compartments are 

connected by arterial and venous blood flow and are characterized by a set 

of mass balance differential equations. The rate constants that describe the 

flow of material between the tissue groups and the rate of change in the 

chemical concentration of each compartment are proportional to blood 

flow, tissue solubility and compartment volumes. The basic mathematical 

description of a PBPK model for a volatile compound has been provided by 

Ramsey and Andersen (1984) and additional details may be found in 

appendices of manuscripts dealing with the development of PBPK models. 

Estimation of the constants used in PBPK models may be based on the 

literature in the case of physiological parameters such as ventilation rates, 

cardiac output, blood flow to tissues and tissue volumes. The EPA has 

compiled reference values for these parameters and their scaling (Arms and 

Travis, 1988). Chemical specific parameters such as blood and tissue 

solubilities may be determined from in vitro preparation (Sato and 

Nakajima, 1979; Gargas et al., 1989). The biochemical constants for 

metabolism may be derived from in vitro studies (Reitz et al. 1988; 

Carfagna and Kedderis, 1992; Johanson and Filser, 1993), in vivo 

toxicokinetic studies (Potter and Tran, 1993; Frederick et al., 1992) or in 

the case of volatile substances from gas uptake studies (Gargas et al., 1986, 

1990; Filser, 1992). 

Since the tissue groups have a defined biological meaning, scaling of the 

associated parameters between species is possible since many of the 

parameters used are correlated to body weight. Cardiac output, alveolar 

ventilation rate and V max are scaled by the 3/4 power of body weight 

whereas K m is assumed to be constant across species. However, the 

substitution of the physiological parameters with the appropriate values 

characteristic for the species of interest is preferred. 

The development of PBPK models is an iterative process involving 

comparison of the model simulations with experimental data and 

refinement of the estimates when the model fails to accurately predict the 

kinetic behaviour. Different exposure scenarios can be used to predict the 

concentrations of the parent chemical or its metabolites in the blood or the 

tissues, which are the target of toxic effects. The level of glutathione


depletion in hepatic and extrahepatic tissues (D’Souza et al., 1988; 

Frederick et al., 1992; Krishnan et al., 1992), kinetic interactions of parent 

compounds in mixed exposures (Tardif et al., 1993) or the amount of 

adducts formed by macromolecular binding (Krishnan et al., 1992) are 

predictions that may also be generated by PBPK modeling. As a result of 

the simulations, quantitative information on the internal dose of a 

chemical or its metabolites in the target tissue is obtained and can replace 

the administered dose conventionally used in risk assessment. After 

validation of the PBPK models in experimental animals, human PBPK 

models can be developed either by allometric scaling of the physiological 

and biochemical parameters or preferably using the actual human 

parameters. Following the prediction of the target tissue dosimetry in 

humans, the appropriate dose surrogates are related to the effect of interest 

and quantitative species differences are determined. This information 

provides the possibility to base the species extrapolation on scientific data 

instead of on arbitrarily assigned default factors and as a consequence the 

uncertainty of the extrapolation procedures applied in conventional risk 

assessment may be reduced. 

Description and use of the PBPK model for 2butoxyethanol 

2-Butoxyethanol (BE) is a widely produced glycol ether used as a key 

ingredient in water- or solvent-based coatings, industrial and consumer 

cleaning products, and as solvent in a variety of products. Haemolysis was 

identified as most sensitive indicator of BE-induced toxicity in several 

species of laboratory animals and has received the most attention as a 

critical effect for human risk assessment (ECETOC, 1985, 1994). The 

experimentally determined subchronic NOAEL for the rat is 25 ppm. The 

major metabolite of BE is 2-butoxyacetic acid (BAA) which has been 

identified as the metabolite responsible for the haemolysis of red blood 

cells in in vitro and in vivo studies (Bartnik et al., 1987; Ghanayem et al., 

1987; Ghanayem, 1989). Changes in the deformability of rat erythrocytes 

appear to precede haemolysis upon treatment with BAA. Treatment of 

human erythrocytes with BAA did not induce changes in deformability 

(Udden and Patton, 1994; Udden, 1994). The observed species differences 

may be due to differences in the lipid composition of erythrocyte 

membranes, differences in membrane proteins associated with anion 

transport processes, or differences in the erythrocyte cytoskeleton (Udden, 

1994; Udden and Patton, 1994). Humans are most likely to be exposed to 

BE by the dermal or inhalation routes due to the widespread use of BE in 

cleaning products. Assessment of the risk resulting from BE use has to 

account for these routes of exposure and the formation of BAA as the 

active metabolite. In order to assist in the risk assessment, PBPK models

N.FEDTKE 173 

were developed that describe the uptake, metabolism and disposition of BE 

and BAA (Johanson, 1986; Corley et al., 1993, 1994; Shyr et al., 1993). 

The model of Corley et al. (1993, 1994) is a refinement of Johanson’s 

model (1986) and consists of two submodels. The first submodel describes 

the uptake and disposition of BE and consists of the tissue compartments 

rapidly perfused organs, slowly perfused organs, fat, skin, muscle, 

gastrointestinal tract, and liver as the metabolizing tissue. The BE 

submodel allows uptake via the inhalation and dermal routes and in 

addition provides the possibility of uptake via IV infusion and the 

gastrointestinal tract in order to validate the model with laboratory data. 

The second submodel tracks the disposition of BAA in the same tissue 

compartments, but the kidney was removed from the rapidly perfused 

organs as separate tissue to allow for the excretion of BAA metabolites. 

The two submodels are linked together by the metabolism of BE to BAA 

via a saturable enzymatic pathway catalyzed by alcohol and aldehyde 

dehydrogenases in the liver. Competing pathways (BE conjugation and 

BE O-dealkylation) are lumped together and described by an additional 

enzymatic pathway with Michaelis-Menten kinetics. The model assumes 

that BAA is bound to proteins in blood and is eliminated by a saturable 

process in the kidneys. The rate of BAA elimination by the kidneys is 

described as the sum of glomerular filtration rate of BAA and the acid 

transport of BAA assuming that no reabsorption occurs. The biochemical 

constants determined experimentally in the rat were scaled to humans by 

(body weight) 0.7 . In the validation process, the model successfully described 

a wide variety of rat and human data from different laboratories using 

several routes of administration. 

BAA was predicted to be formed more rapidly in rats compared with 

humans, but to be eliminated slower in humans than in rats. In summary, 

higher maximum concentrations of BAA in blood (C max) and also higher 

areas under the BAA concentration-time curves (AUC) were predicted for 

rats than for humans, especially as the vapour concentration was 

increased. For the purpose of dose-response and interspecies extrapolation, 

BAA-C max and BAA-AUC were used as estimates of the internal dose 

surrogate; C max can be related directly to the in vitro haemolysis studies 

with BAA and is responsive to the dose-rate. The in vitro studies performed 

(Bartnik et al., 1987; Ghanayem et al., 1987; Ghanayem, 1989; Udden, 

1994; Udden and Patton, 1994) suggest that approximately 0.2 mM BAA 

is required to produce slight haemolysis of rat red blood cells. At about 2 

mM BAA nearly complete haemolysis was observed. The model predicts 

for nose-only exposure that these concentrations are reached in the rat at 

BE exposure concentrations of about 100 ppm and 800 ppm for 6 h, 

respectively, which is consistent with observations in vivo (Tyler, 1984; 

Sabourin et al., 1992).


For human red blood cells, the minimum BAA-concentration necessary 

to induce slight haemolysis is about 40 times higher compared with rats, 

i.e. 8 mM. The model predicts for human nose-only exposure that the C max 

of BAA in blood is slightly lower than the value observed in rats at a BE 

exposure concentration of about 100 ppm for 6 h and is only about 50 per 

cent of the BAA rat blood concentration at 800 ppm. In any case, the 

minimum toxic concentration of approximately 8 mM BAA in human 

blood is not achieved. 

The AUC has a time component which is important since haemolysis is 

not an instantaneous response (Udden, 1994; Udden and Patton, 1994). 

With respect to the AUCs for BAA, the model predicts that the values for 

rat and human blood are similar up to a BE exposure concentration of 

about 500 ppm. Higher BE concentrations cause higher AUCs for BAA in 

rat blood than in human blood. Thus, the model predicted a BAA-AUC in 

man at 22 ppm BE vapour exposure that was similar to the BAA-AUC in 

rats achieved at 25 ppm BE vapour exposure, the established subchronic 

NOAEL. 

The simulation of the dermal BE uptake assumed that 10 per cent of the 

body surface of rats and humans were exposed for 6 h to BE solutions in 

water (5–100 per cent) and that no losses of BE occurred from the dosing 

solution. The simulation predicted C max blood concentrations for BAA in 

rats that were highest (about 3 mM) for a 40 per cent BE solution. For 

humans, BAA-C max was predicted to reach about 1.3 mM for the same BEconcentration. 

Predicted BAA-AUCs were about twofold higher in rats 

compared with humans. Under these worst-case assumptions, no BE 

concentration is expected to achieve BAA concentrations in human blood 

that would cause haemolysis. 

ECETOC (1994) used the described PBPK model for BE and BAA 

disposition in combination with mechanistic data obtained by in vitro 

experiments to recommend an occupational exposure limit for BE: 

– BAA-AUC was used as the internal dose surrogate and 22 ppm BE 

vapour was predicted to cause a BAA-AUC in human blood similar to 

the BAA-AUC in rats exposed to a BE-concentration of 25 ppm 

(identified as the subchronic rat NOAEL). At 22 ppm BE vapour, the 

BAA-C max (33 µM) is predicted to be several hundredfold below the 

BAA concentration that causes pre-haemolytic effects in human red 

blood cells (8 mM). 

– ECETOC did not use an uncertainty factor for intraspecies 

extrapolation, since the in vitro studies indicated no increased sensitivity 

of red blood cells from individuals regarded as susceptible to haemolytic 

effects such as older persons, persons with hereditary spherocytosis or 

sickle cell disease (Udden, 1994; Udden and Patton, 1994).

– An uncertainty factor for time extrapolation (subchronic to chronic 

exposure) was also not applied, since the red blood cell haemolysis was 

regarded as a transient phenomenon observed predominantly on the 

first few days of exposure thus indicating that longer exposure would 

not have resulted in a lower rat NOAEL. 

– Although there is some uncertainty about the actual magnitude of the 

contribution of dermal uptake to the total uptake during BE vapour 

exposure, ECETOC concluded that even under worst-case conditions 

the BAA concentrations achieved are not sufficient to cause haemolysis 

in man and there is no need for the adjustment of the predicted human 

NOAEL for route. 

In conclusion, an occupational exposure limit of 20 ppm (8 h TWA) was 

recommended, also taking into account all other effects that may be 

associated with BE-exposure. This value is similar to the rat NOAEL of 25 

ppm for the most sensitive parameter, i.e. haemolysis, and was derived 

using scientific data instead of applying default factors to the rat NOAEL, 

a procedure which would have overpredicted the human risk associated 

with BE-exposure. 

Conclusion 

N.FEDTKE 175 

The use of PBPK models and mechanistic data in risk assessment tends to 

reduce the uncertainties in comparison with default methodologies by 

replacing the administered dose with the delivered dose and also tends to 

reveal uncertainties concealed in default methodologies (Wilson and Cox, 

1993). However, there are also limitations in the development of PBPK 

models. One limitation is that the mechanism of the toxic effect has to be 

known, otherwise the replacement of the external dose by internal dose 

surrogates is not possible. In addition, extensive validation of the model is 

necessary in order to replace default approaches in risk assessment. For the 

time being, the development of PBPK models appears to be restricted to 

high production chemicals where the existing data base allows 

identification of an accepted mechanism of toxic action and validation of 

the model. Concern has been expressed that the use of point estimates in 

PBPK modelling instead of ranges of biologically plausible values leads to 

an increase in the uncertainty (Portier and Kaplan, 1989). However, a 

recent study from the Delivered Dose Work Group of the American 

Industrial Health Council came to the conclusion that incorporation of 

‘pharmacokinetic information in a risk assessment,…, leads to both a more 

accurate estimate of risk and a better specification of the true uncertainty’ 

(Wilson and Cox, 1993). A detailed discussion of the sources of 

uncertainties is also provided in this reference.


If the data base is sufficient, PBPK models provide scientific credibility to 

interspecies extrapolation, extrapolation across routes of administration, 

extrapolation from high-dose to low-dose and intraspecies extrapolation. 

Recent concepts link the original tissue dose concept of PBPK models to 

biologically based tissue response models, thus relating the delivered dose 

via the mechanism of action to the toxic response and developing 

integrated biological models (Conolly et al., 1988; Moolgavker et al., 

1988; Cohen and Ellwein, 1990; Conolly and Andersen, 1991). Such 

approaches enable scientists to ask the right questions and to design new 

mechanistic studies that will lead toward the goal of a scientifically-based 

risk assessment. 

Acknowledgement 

The author thanks Richard A.Corley and the Glycol Ether Panel of the 

Chemical Manufactures Association for providing data on 2butoxyethanol. 

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14 

Molecular Approaches to Assess Cancer Risks 

ALAN S.WRIGHT, J.PAUL ASTON, NICO J.VAN SITTERT 

and WILLIAM P.WATSON 

Sittingbourne Research Centre, Sittingbourne 

Introduction 

Carcinogenesis is a complex process which is not yet fully understood. 

Nevertheless, it is generally accepted that carcinogenesis involves the 

accumulation of mutations in critical genes: proto-oncogenes and/or 

tumour suppressor genes. These mutations transform normal cells into 

‘initiated’ cells possessing the full complement of genetic changes necessary 

for malignancy (Figure 14.1). The critical mutations may result from 

exposures to radiation, to genotoxic chemicals or they may arise 

‘spontaneously’ as a consequence of miscoding errors during the normal 

replication of DNA. Concomitantly, mutations will also accumulate in 

other genes which, although not critical for cancer per se may, 

nevertheless, influence cellular character thereby contributing to the 

multifaceted nature of cancer. 

The precise nature and number of critical genetic changes required for 

initiation have not yet been established but will probably vary from case to 

case. Many researchers envisage a strict temporal sequence of genetic 

changes in carcinogenesis. However, it is probable that the critical 

mutations can occur in any sequence and at any time. Indeed, it is clear 

that one or more of the critical mutations can occur in parental cells. 

Transmission (inheritance) of these mutations either through the germ line 

or via somatic cell division increases the susceptibility of the progeny to 

carcinogens. Furthermore, it is important to note that each of the critical 

mutations necessary for malignancy may have a different cause. This 

potential for multiple causation has important implications in risk 

assessment (vide infra). 

Fully initiated cells may not automatically proliferate to form tumours. 

One possible explanation is that the surrounding normal cells restrain 

the initiated or latent cancer cell by providing essential growth regulators 

which are no longer produced by the initiated cell. Nevertheless, partially 

and fully initiated cells have a replicative and/or survival advantage over 

normal cells. Tissue injury caused by physical trauma, chemical agents or

Figure 14.1 Schematic representation of the carcinogenic process. 

A.S.WRIGHT ET AL. 181 

viruses may have a derestraining effect thereby triggering or facilitating the 

replication of partially or fully initiated cells to form benign tumours or 

malignant tumours. Increased functional demands may also serve to 

promote tumour development in affected tissues. 

It is clear that chemicals which promote tumour development are very 

important determinants of carcinogenesis. Indeed, promoting agents 

display a marked tendency for organotropism. Promoter action is, 

therefore, probably the most important determinant of the site of tumour 

development. Yet genotoxic chemicals which initiate the carcinogenic 

process are perhaps viewed with even greater concern. The reasons for this 

high level of concern hinge mainly on evidence that the mutagenic or 

initiating actions of genotoxic chemicals are additive, cumulative and 

essentially irreversible. Furthermore, in contrast to most other classes of 

toxic chemicals, including promoters operating via cytotoxic mechanisms, 

there is no theoretical reason or experimental evidence to support the view 

that mutagenic actions of genotoxic chemicals are thresholded. For these 

reasons even very low exposures of genotoxic chemicals are viewed with 

concern. These concerns have focused scientific and regulatory attention on 

a need to develop sound approaches to manage cancer risks—particularly 

low level risks associated with low exposures to genotoxic chemicals 

encountered in the occupational or environmental settings. Indeed, apart 

from clinical applications, high exposures to genotoxic chemicals cannot be 

countenanced.

182 MOLECULAR APPROACHES TO ASSESS CANCER RISKS 

Management of cancer risks (key requirements) 

The management of toxicological risks implies a capacity to control 

exposures within acceptable safety limits. Effective control is, therefore, 

dependent not only on the qualitative detection and identification of 

hazardous chemicals but also on a capacity to determine human exposure 

and to evaluate the health risks. This last requirement necessitates a 

knowledge of potency, i.e. quantitative human dose-response relationships. 

In the case of genotoxic chemicals, the relevant data reside in the very low 

region of the dose-response curve. 

The concept of acceptable risk is readily accepted when applied to the 

many classes of toxic chemicals operating by a thresholded mechanism 

indicative of the virtual absence of risk at sub-threshold doses. Absolute 

safety margins for genotoxic chemicals cannot be guaranteed (vide supra) 

leading to the adoption of conservative safety measures. Indeed, cursory 

analysis might suggest that quantitative risk data are not required for the 

effective management of cancer risks associated with genotoxic chemicals. 

Thus, it is generally accepted that human contact with carcinogens should 

be minimised. Purely qualitative identification of the hazard would permit 

the design of measures to limit human exposure and minimise carcinogenic 

impact. Indeed, a purely qualitative indication of genotoxicity can be an 

absolute deterrent to the development of new products. Nevertheless, 

certain exposures, e.g. to indigenous genotoxic chemicals and natural food 

components, are unavoidable. Furthermore, measures to reduce exposures 

to ‘avoidable’ genotoxic hazards, e.g. certain combustion products, key 

industrial base chemicals and intermediates, are often difficult and costly. 

Quantitative risk assessment is needed to prioritise these hazards and, most 

importantly, to determine safety margins. Certainly a failure to determine 

carcinogenic potency would lead to uncertainty about the adequacy of 

safety margins and, probably, to unnecessary measures to further reduce 

exposures. Thus, despite their additive and cumulative actions, even 

genotoxic chemicals can pose a negligible health risk. Of course, the 

definition of a negligible, i.e. acceptable, risk is a socio-political judgement 

which nevertheless has to be realistic in the case of unavoidable hazards 

and achievable in the case of avoidable hazards. 

Detection of genotoxic carcinogens 

Concerns about genotoxic hazards have provided an incentive for the 

development of a broad range of rapid tests to detect intrinsic genotoxic 

activity or potential. The principal aim of these approaches is to predict 

carcinogenic activity or, more accurately, cancer initiating activity. The 

most widely used tests are the coupled microsomal-microbial mutation 

assays developed by Ames et al. (1973). However, such approaches are


viewed as too remote to be of value in estimating cancer risks. The trend is 

towards increasingly sensitive and precise technology—particularly generic 

methods with potential for direct application in humans. 

Advances in molecular biology have permitted the development of a new 

generation of point mutation assays based on DNA base mismatch 

technology (Thilly, 1991; Lu and Hsu, 1992). This technology has a 

precision far exceeding that of conventional biological methods and a 

sensitivity permitting direct applications in humans. The full potential of this 

technology has not yet been realised. However, it seems probable that 

detection levels will ultimately obviate a need for prior phenotypic 

selection: paving the way to universal application. Avoidance of phenotypic 

selection would represent a powerful advantage over existing 

methodologies by providing a much more direct and reliable route to 

determining overall background mutation rates and increments due to 

specific exposures of key relevance to cancer risk assessment (vide infra). 

The most prospective of the current assays are those designed to detect 

primary DNA damage. Among these procedures, 32 P-post-radiolabelling 

technology developed by Randerath et al. (1981) to detect DNA adducts is 

by far the most sensitive. The justification for application of such a 

prospective approach to detect exposure to genotoxic carcinogens hinges 

on the causal relationship established between genotoxic activity and 

cancer. In general, genotoxic character is conferred by possession of a 

centre(s) of electrophilic reactivity. This reactivity permits the chemical to 

undergo chemical reactions with nucleophilic centres in the target molecule 

(DNA). In many instances the electrophilic centre(s) is introduced into an 

inactive precursor chemical by metabolic activation. Primary products, e.g. 

DNA adducts, formed when genotoxic chemicals react with DNA are 

generally promutagenic (or lethal) and their occurrence leads to an 

increased risk of mutation and cancer. There is no known category of 

chemical which forms DNA adducts that can be excluded from this 

generalisation. Not all DNA adducts are strongly promutagenic. However, 

because electrophiles do not display absolute specificity in their reactions 

with nucleophiles, the detection of even a weakly promutagenic adduct, 

e.g. N 7 -alkyldeoxyguanosine, signals the formation of a more strongly 

promutagenic adduct, e.g. O 6 -alkyldeoxyguanosine. If follows that the 

detection of DNA adducts provides qualitative evidence of (human) 

exposure to a genotoxic carcinogen.


Identification of human carcinogens 

Classical epidemiological approaches 

Until very recently, epidemiological approaches to detect and identify 

environmental carcinogens were based exclusively on the analysis of 

tumour incidence and chromosome aberrations in human populations. 

However, the endpoints of these biological methods lack the intrinsic 

resolving power needed to dis criminate between different contributory 

factors. Indeed, it is only in instances of specific, high and, often, localised 

exposures that these methods have been effective in identifying specific 

causative agents. Nevertheless, the results of epidemiological studies 

indicate that chemicals, which may include both natural and xenobiotic 

compounds in food, drink or in the local or general environment, play a 

major and broad role in the aetiology of human cancer. The identification 

of these chemical factors is a major goal in cancer prevention. 

In vitro genotoxicity assays 

In addition to applications in screening prospective chemical products, in 

vitro genotoxicity assays, particularly the Ames test, provided the first 

practicable, systematic approach to identify environmental carcinogens. 

However, this approach places very heavy demands on the time and effort 

required to fractionate environmental samples and test individual 

compounds. More importantly, however, like the animal cancer studies 

these assays complement or have largely supplanted, the approach is not 

specifically targeted towards identifying and prioritising human hazards. 

For example, these short-term in vitro test do not provide direct evidence 

of human exposure or effects. 

Molecular epidemiology 

32 P-Post-radiolabelling technology for the analysis of DNA adducts 

provides the basis of a very sensitive and generic approach to detect 

exposures to genotoxic carcinogens. This technology has universal 

application and can be applied to detect DNA adducts formed in 

laboratory species or humans during exposures to both known and, as yet, 

unidentified genotoxic chemicals at the low concentrations encountered in 

the environment and the workplace. Elucidation of the chemical structures 

of adducts in human DNA would provide a basis for identifying the 

causative agents and their sources or origins. This possibility of identifying 

the chemical initiators of human cancer is an exciting prospect. 

Unfortunately, however, these adducts are present at very low abundances 

and this is a major obstacle to identification. Thus, the methods for

detecting DNA adducts are much more sensitive than the physicochemical 

methods needed for structural characterisation. A number of strategies 

have been adopted in attempts to solve this problem. 

Protein adducts 

Genotoxic chemicals that react with DNA also react with nucleophilic 

centres in proteins and may also undergo ‘spontaneous’ and enzymecatalysed 

reactions with glutathione leading to the excretion of the 

corresponding mercapturic acids. Insofar as the formation of protein 

adducts and mercapturic acids reflect the formation of the corresponding 

DNA adducts, their detection may also furnish evidence of exposure to a 

genotoxic carcinogen. 

The potential for reaction of genotoxic chemicals with proteins (and 

glutathione) is much greater than with DNA. Furthermore, human 

proteins, e.g. haemoglobin, are available in much larger quantities and are 

more accessible than human tissue DNA. These advantages have been 

exploited, particularly in the pioneering work of Ehrenberg’s group, to 

develop a range of procedures for the qualitative and quantitative analysis 

of protein adducts (Osterman-Golkar et al., 1976; Calleman et al., 1978; 

Ehrenberg and Osterman-Golkar, 1980). (For a review of the available 

methods see Skipper and Naylor, 1991.) The most powerful and generic 

approach is undoubtedly that developed by Törnqvist et al. (1986a). An 

initial purification or enrichment step is key to any successful method for 

the analysis of low levels of organic residues. The amino-groups of the Nterminal 

valine residues of the α-and 

β-chains of human haemoglobin are 

major targets for reaction with a broad range of genotoxic chemicals. 

Törnqvist achieved selective enrichment of adducted N-terminal valine 

residues of haemoglobin by devising a modified Edman degradation which 

resulted in the scission of adducted residues whilst leaving the nonadducted 

N-terminal valines intact. This procedure provides the basis for 

identifying the adducting moieties and their quantitation by GC/MS. 

Applications of this technology have furnished evidence of background 

exposures to a range of alkylating species. Protein adduct technology has 

the potential for considerable further refinement. The possibility of using 

immunoaffinity technology to enrich both known and unidentified protein 

adducts is currently being explored. 

DNA adducts and immunoenrichment 


The need for effective enrichment technology for DNA adducts is even 

more pressing than for protein adducts. Ideally, the enrichment procedure 

should be applied at the earliest possible stage of analysis. The procedure


should be rapid and mild in order to minimise the formation of artefacts. 

Currently, immunoaffinity technology holds the greatest promise. 

Immunoenrichment of DNA adducts necessitates antibodies possessing 

the appropriate specificities and affinities to permit selective binding of 

adducts at the very low abundances encountered in hydrolysates or enzyme 

digests of human DNA. The immune system does not normally respond to 

small molecules per se. However, the system can be induced to produce 

effective antibodies by immunising with the small molecule (hapten) 

coupled to a protein. Such treatment induces a spectrum of antibodyproducing 

cells, each producing a specific antibody. Most of these 

antibodies recognise various regions (epitopes) of the carrier protein while 

a few may specifically recognise and bind the small molecule of interest. 

Suitable antibody-producing cells can be selected and cloned to provide a 

permanent source of homogenous antibody (monoclonal antibody, Mab). 

Mabs can be raised against virtually any organic chemical although some 

lower molecular weight compounds (

adducts are not viewed as particularly promutagenic. Nevertheless the N 7 - 

atom of dG residues in DNA is a major target for adduction and the 

detection of N 7 -dG adducts signals the production of adducts at other 

(more critical) sites in DNA (vide supra). 

In certain instances, a class of adducting moieties may possess a common 

structural feature that can be exploited for immunoenrichment. For 

example, a Mab has been raised against the major DNA adduct of benzo(a) 

pyrene (r-7,t-8,t-9-trihydroxy-c-10-(N 2 -deoxyguanosylphosphate)-7,8,9, 

10-tetrahydrobenzo(a)pyrene) in a collaborative study with Dr Baan’s 

group. This Mab recognises DNA adducts formed by a broad range of 

polycyclic aromatic hydrocarbons (PCAs) including benzo(a)pyrene (BP), 

chrysene, benz(a)anthracene, 5-methylchrysene, picene and dibenz(a,h) 

anthracene. It seems probable that this Mab recognises the common 

trihydric alcohol structure produced when the reactive diol epoxides of 

each of these polycyclic compounds reacts with nucleophilic centres in 

DNA or other macromolecules. The fact that the Mab does not bind the 

corresponding fluoranthene adduct is consistent with the spatial 

environment of the hydroxyl groups in fluoranthene-DNA adducts which 

is completely different from those generated from the other PCAs employed 

in this study. 

The performance of the Mab raised against the major BP-DNA adduct in 

the enrichment of PCA-DNA adducts is being evaluated using the 

immobilised Mab coupled to cyanogen bromide-activated Sepharose 4B. 

Results obtained to date demonstrate that the immobilised Mab selectively 

adsorbs the major BP-DNA adduct from DNA hydrolysates at abundances 

below 1 adduct per 10 9 nucleotide units. Results with the other PCA-DNA 

adducts are not yet available. However, the results obtained with the major 

BP-DNA adduct underlines the potential of immunoenrichment technology 

in the qualitative and quantitative analysis of adducts. Furthermore, such 

results provide an incentive to pursue the development of class-specific 

antibodies in order to permit or facilitate the identification of the chemical 

initiators of human cancer. 

Mercapturic acids 


Qualitative analysis of mercapturic acids also provides a basis for 

identifying human exposures to genotoxic chemicals (vide supra). However, 

the available analytical procedures are complex and tend to lack specificity 

and sensitivity. During the last dozen years we have undertaken a number 

of studies aimed at developing compound- and class-specific antibodies to 

facilitate the analysis of mercapturic acids. 

Conventional approaches to generate antibodies to low molecular weight 

(MW) organic chemicals involves the covalent attachment of the small 

molecule (hapten) to a strongly antigenic protein, e.g. keyhole limpet


haemocyanin or bovine serum albumin, for the immunisation of mice. This 

strategy is usually effective in the case of strongly antigenic haptens, e.g. 

aromatic nitro compounds of PCA-DNA adducts. However, all of our 

attempts to use this approach to generate antibodies against relatively low 

MW and weakly antigenic mercapturic acids, e.g. S-(2-hydroxyethyl)-Nacetylcysteine, 

failed. Antibodies were generated but these were directed 

against the strongly antigenic carrier protein(s). 

Covalent binding to macromolecules is believed to provide the basis of 

allergic responses, e.g. skin sensitisation reactions, to small molecules and, 

possibly, a basis for the induction of auto-immune responses. Thus, the 

binding of the small molecule transforms normal proteins into ‘foreign’ 

proteins which trigger an immune response. Recently we have employed 

this principle in an attempt to direct the immune response specifically 

against mercapturic acid haptens by immunising mice with the haptens 

bound to a non-antigenic carrier protein, i.e. mouse serum albumin. 

Preliminary results indicate that this tactic has been successful. Overall the 

treatment induced fewer antibody-producing cells. However, the antibodies 

that were generated show high affinities and specificity toward model 

mercapturic acids including S-(2-hydroxyethyl) and S-phenylmercapturic 

acid. Studies are in progress to investigate the performance of these 

antibodies in an immunoenrichment mode. 

The preliminary results of our studies using non-antigenic protein 

carriers are very encouraging and have provided fresh insights which may 

assist in directing immune responses against the specific structural features 

of interest. Improvements in our ability to tailor the antibody will prove 

extremely valuable in optimising the properties of antibodies to meet 

specific needs, e.g. to enrich DNA, protein or mercapturic acid adducts for 

application in identifying the chemical initiators of human cancer and 

quantifying exposures to these agents. 

Cancer risk assessment 

Human exposure monitoring (determination of dose) 

The assessment of cancer risks posed by exposure to genotoxic chemicals 

has two components: determination of the dose and determination of the 

effect (increment in cancer incidence) caused by that dose. The introduction 

of the target dose concept by Ehrenberg in the early 1970s has provided the 

key to modern strategies to assess genotoxic risks (Ehrenberg, 1974, 1979; 

Ehrenberg et al., 1974). This new dose concept was developed to provide a 

measure of the critical dose, i.e. the dose of the ultimate genotoxic agent(s) 

penetrating to DNA. Target dose is much more relevant to risk assessment 

than is exposure dose. The determination of target dose automatically

compensates for individual or species differences in the operation of 

metabolic and biokinetic factors that control the quantitative (and 

qualitative) relationships between the exposure and the dose of the ultimate 

toxicant delivered to the target. Measurements of target dose may be 

applied to improve the extrapolation of risk data from experimental 

models to humans and may also provide improved definitions of risks to 

individuals. The determination of target dose in humans may be viewed, 

therefore, as an approach towards direct risk monitoring as well as a more 

relevant approach to monitor human exposures to genotoxic chemicals. 

Determination of target dose 

The determination of target dose raises numerous technical and theoretical 

problems. Target dose can be determined by measuring primary products, 

e.g. DNA adducts, formed when genotoxic agents react with DNA. The 

kinetics of formation and decay of these adducts must also be determined 

(vide infra) in order to transform measurements of amounts of adducts into 

estimates of target dose. Human tissue DNA is not readily accessible for 

monitoring purposes: surrogate dose monitors are required. There are 

numerous possibilities including the determination of adducts in white 

blood cell DNA or of the corresponding adducts in the haemoglobin of 

circulating erythrocytes. 

Such indirect approaches require validation. Haemoglobin is the most 

extensively studied surrogate, not only because of its accessibility and 

relative abundance but also because of the relative stability of haemoglobin 

adducts and the longevity of erythrocytes which permit retrospective 

estimates of dose received by the erythrocytes over a period of about 4 

months. Current evidence indicates that all electrophiles that undergo 

covalent reactions with DNA also react with haemoglobin. Furthermore 

the amounts of haemoglobin adducts are quantitatively related to the rates 

of formation of DNA adducts in the tissues. However the proportional 

relationships between the doses delivered to tissue DNA and to 

haemoglobin or to any other surrogate will vary from chemical to chemical 

and will have to be established using experimental models. 

Measurement of haemoglobin adducts 


Genotoxic chemicals undergo covalent reactions with a variety of 

nucleophilic centres in haemoglobin including the sulphydryl group of 

cysteine, the N 1 and N 3 atoms of histidine and the amino groups of Nterminal 

valine residues. Ehrenberg’s group (Osterman-Golkar et al., 1976; 

Calleman et al, 1978; Törnqvist et al., 1986a) has pioneered the 

development of methods to detect, identify and quantify adducts formed at 

each of these centres. A review of these and methods for the analysis of


‘labile’ adducts formed, for example, during exposure to aromatic amines 

(Green et al., 1984; Albrecht and Neumann, 1985) is beyond the scope of 

this paper (for reviews see Farmer, 1991; Skipper and Naylor, 1991). 

However, probably the most powerful and valuable approach was 

developed by Törnqvist et al. (1986a, b) who showed that adducts with the 

N-terminal valine residues of haemoglobin could be specifically enriched by 

scission in a modified Edman reaction followed by extraction. This 

enrichment procedure greatly facilitates sample analysis by GC/MS. 

Immunoassays are also being introduced as alternatives to physicochemical 

methods for the determination of protein adducts (Wraith et al., 1988). 

However, as in the case of DNA adducts, the biggest impact of 

immunotechnology on the analysis of protein adducts will probably be in 

the immunoenrichment of low levels of adducts for analysis by physicochemical 

methods. 

Determination of biological effects 

Tumour incidence 

The determination of target dose is essential for assessing cancer risks 

posed by low-level exposures to genotoxic chemicals. The other requisite is 

know ledge of the human dose-carcinogenic response relationships in the 

low-dose range. The lack of intrinsic resolving power of classical 

epidemiological methods (vide supra) prevents effective applications to 

detect small carcinogenic effects associated with low exposures to any 

particular genotoxic chemical. Furthermore, the detection limits of animal 

cancer studies fall short of ‘acceptable’ risk limits by three to four orders of 

magnitude (Wright, 1991). This poor sensitivity compels the use of high 

test doses in order to ensure that significant carcinogens do not go 

undetected. However, it is generally accepted that high doses of chemicals 

may induce tumours by non-specific mechanisms, e.g. via tissue injury and 

compensatory cell proliferation, that do not operate at low doses (Ames, 

1989; Wright, 1991). Many, if not all genotoxic chemicals induce cell 

injury at high (thresholded) doses. Clearly, extrapolation of such high dose 

risk data to the relevant low-dose range may, at the very least, lead to a 

gross overestimation of risk. 

Determination of mutagenic potency 

In considering the impact that a low-level exposure to a genotoxic 

chemical may have on cancer incidence, it is reasonable to suggest that the 

mutagenic propensity of the chemical, although of a low order, would 

nevertheless be the overriding risk factor. Thus, it is probable that any


intrinsic promoter activity of the chemical arising, for example as a 

consequence of cell injury, would be negligible at low exposures— 

particularly when viewed in the context of the overall promoter pressure 

exerted on the populations at risk. Thus, it is unlikely that the added 

cancer risk would be greater than and may approximate to the increment in 

critical mutations caused by the specific exposure (Figure 14.1). 

Experimental studies indicate that all known categories of genotoxic 

agents ranging from methylating agents to polycyclic aromatic compounds 

can induce the critical mutations leading to malignancy (Figure 14.1). 

Furthermore, at low doses, the possibility that exposure to a particular 

genotoxic chemical would induce more than one of the critical mutations in 

any particular cell is extremely remote. Indeed it is probable that each 

critical mutation is induced by a different agent or mechanism, i.e. 

chemical, radiation or ‘spontaneous’. In this sense, each critical mutation 

would have equal status, i.e. no single event would be any more or any less 

critical than any other to the final outcome. Accordingly, the increment in 

cancer risk would equate with the increment in any decisive, e.g. oncogeneactivating, 

mutation in any critical gene. The induction of such mutations 

is almost certainly a direct function of overall mutagenic activity of the 

chemical, i.e. linked to the number of mutational events rather than the type 

of mutations induced by a given dose of the chemical. At low exposures, 

therefore, the increment in cancer incidence due to a specific genotoxic 

agent would approximate to the small increase in the total mutational load 

caused by the exposure, i.e. relative to the overall background level of 

mutations due to all causes, multiplied by the overall cancer incidence in 

the population at risk. (The latter function introduces a measure of the net 

impact of promoter and anti-promoter pressure acting on initiated cells in 

the population at risk.) Of course risks may also be calculated on the basis 

of specific tumours and specific tissues. 

The determination of small increments in mutation associated with lowlevel 

exposures to genotoxic chemicals in human populations presents 

enormous technical problems not least due to the much larger and variable 

background of mutations due to all causes. Increasing the sensitivity of 

mutation assays per se (vide supra) is unlikely to improve the situation. 

High resolving power is also needed to discriminate between effects due to 

different contributory factors. Nevertheless, while direct approaches to 

assess absolute cancer risks posed by such low-level exposures may elude 

us we can nevertheless begin to determine relative cancer risks and 

prioritise genotoxic chemicals on the basis of experimental determinations 

of mutagenic potency and estimates of target doses resulting from 

environmental or occupational exposures. Thus, according to the foregoing 

the relative cancer risk posed by a low exposure to a genotoxic agent 

would approximate to the number of mutations induced per unit target 

dose×estimated human target dose. Once relative cancer risks have been


established, the determination of the ‘absolute’ risks for one genotoxic 

chemical would permit calculation of the ‘absolute’ risks for the others. 

Such ‘absolute’ risks will nevertheless vary from population to population 

dependent upon variations in promoter pressure. 

Determination of ‘absolute’ cancer risks 

Increments in human mutation caused by low-level exposures to genotoxic 

chemicals are essential for risk estimation but cannot be determined 

directly (vide supra). Such increments may be estimated using experimental 

models. However, unless the variations in background are of a very low 

order, it is unlikely that even the most sensitive of the emerging mutation 

assays will permit the measurement of small increments of mutation at 

low, e.g. environmental, exposures. Extrapolation to low doses will be 

required and must necessarily be conservative, i.e. linear extrapolation to 

the origin. 

In addition to ‘high’ dose-low dose extrapolation, it will be necessary to 

apply corrections for differences between the model and humans in the 

operation of systemic factors that govern the relationships between 

exposure and mutagenic effect. Estimates of target dose in the human 

population at risk and in the experimental model compensate for 

differences in metabolic and biokinetic factors that determine the 

relationships between exposure and the critical dose. In effect, the 

determination of target dose provides a measure of the rates of formation of 

the key (primary and critical) chemical lesions leading to mutation. The 

final stage in translating the experimentally-determined risk data to 

humans is to apply corrections for systemic factors that determine the 

progression of the key lesions into mutations. 

The equivalent radiation dose concept 

The principal systemic factors determining the progression of key chemical 

lesions in DNA into mutations are the rates and fidelities of DNA repair 

and replication (Wright et al., 1988). Ehrenberg and co-workers have 

suggested that the repair of primary DNA damage induced by low doses of 

radiation may be proportionate to that induced by low doses of genotoxic 

chemicals. They have further suggested that the determination of the 

relative mutagenic effectiveness or potencies of radiation and any 

particular genotoxic chemical may be of value in correcting for species 

differences in factors determining the progression of primary DNA damage 

into mutations. The model proposed by Ehrenberg (1980) is based on the 

determination of the dose-response curves for the induction of the same 

mutation in the same experimental system by low target doses of the test 

chemical and acute -radiation. A consistent ratio between the two curves,


which need not be linear, would indicate proportionality over the low dose 

range of interest and would permit the mutagenic potency of the chemical 

to be expressed in terms of radiation equivalents, i.e. the number of rads 

giving the same response or risk as a unit of chemical dose (expressed in 

terms of target dose, e.g. millimolar hour, mM h). The significance of such 

radiation dose-equivalents hinges on their possible extrapolative value. 

Thus, in order to be useful in assisting the translation of experimentallydetermined 

mutagenicity data, the rad-equivalence value for the test 

chemical must have a similar numerical value in both the test system used 

to determine mutagenicity and in humans. However, it is improbable that 

rad-equivalence values can be directly determined in humans. 

Rad-equivalence values for the induction of mutations have been 

determined for a number of intrinsically reactive monofunctional alkylating 

agents using a wide range of genetic endpoints in a variety of biological 

systems including bacteria, plants and mammalian species—the latter, 

mainly in vitro (Ehrenberg et al., 1974; Ehrenberg, 1976, 1979; Calleman, 

1984; Kolman et al., 1989). The rad-equivalence value for a given 

alkylating agent was approximately the same (within a factor of two) in 

each of the test systems. On the basis of such evidence Calleman et al. 

(1978) concluded that there was no reason to presume that a value for radequivalence 

established in these disparate systems would differ in humans. 

The best studied example is ethylene oxide. Currently a conjoint programme 

is underway at the Universities of Stockholm and Leiden to determine radequivalence 

values in rodents in vivo using a variety of endpoints including 

the clonal HGPRT mutation assay and induction of pre-neoplastic nodules 

in rat liver. Preliminary findings are encouraging (Ehrenberg, personal 

communication). 

Demonstration of their extrapolative value would justify application of 

rad-equivalence values to compute small increments in mutation induced in 

humans by low exposures (determined as target dose) to genotoxic 

chemicals. The determination of these increments is the basis of the risk 

model (vide supra) in which the increment in cancer risk due to a particular 

chemical in a population is viewed as approximating to the increment in 

mutation induced by the chemical. However, the fact that risk coefficients 

have not yet been established for radiation-induced mutations in humans 

precludes applications of rad-equivalence values to estimate mutational 

risks, e.g. small increments in mutation, in humans. Of course, Ehrenberg 

realised that a genotoxic chemical(s) may prove to be superior to radiation 

as a reference standard for estimating mutational risks. Indeed, the use of 

chemicals that are representative of classes of genotoxic chemicals, repair 

pathways, etc. is envisaged in this developing ‘equivalence’ strategy 

(Törnqvist and Osterman-Golkar, 1991). However, at the time the original 

strategy was formulated there was and still is no reliable data relating low 

level exposure to a genotoxic chemical and the attendant risk of mutation


or cancer in human populations. In contrast, risk coefficients have been 

established for the induction of cancer by low levels of -radiation in 

human populations within, approximately, a factor of 2. 

According to the basic presumptions (vide supra), the increased cancer 

risks in a human population caused by low-level exposures to radiation or 

to a genotoxic chemical are predominantly due to the mutagenic 

propensities of these genotoxic factors. It would, therefore, seem 

reasonable to suggest that cells that had been initiated by exposure to 

radiation or to genotoxic chemicals or combinations of factors (vide supra) 

would, nevertheless, be subject to the same general promoter and 

modulating influences acting on the population. On this basis, 

experimentally-determined rad-equivalence values for genotoxic chemicals 

may be used to convert target doses of the chemicals (determined in human 

populations receiving low-level exposures) into the equivalent doses of 

radiation (for the induction of mutation). The respective cancer risks 

associated with any particular monitored target dose of a genotoxic 

chemical may then be obtained by direct reference to the human cancer risk 

coefficients established for radiation. 

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15 

Evaluation of Toxicity to the Immune System 

HANS-WERNER VOHR 

Bayer AG, Wuppertal 

Introduction 

A number of years ago the new field of immunotoxicology was established. 

Very early on, calls were heard from various sides demanding that the 

development of new chemicals should also take into account the influence 

of these substances on the immune system (Dean, 1979; Luster et al., 1988; 

Trizio, et al., 1988). These demands led on the one hand to the initiation of 

a number of investigations and collaborative studies (ICICIS; BGA; US- 

NTP), on the other hand to thoughts by the authorities and industry on the 

introduction of guidelines (US-EPA, 1982, 1990; Sjoblad, 1988; ECETOC, 

1990; UK-DOH, 1991; Hinton, 1992; OECD, 1992ab. 

If we define immunotoxicology as the science of adverse effects of 

substances on the immune system we can say further that these side-effects 

can lead to either immunopotentiation or immunosuppression. The former 

can lead to induction of autoimmune reactions and to Type I-IV 

hypersensitivity reactions, the latter to reduced resistance to infection, 

development of cancer and also to autoimmune phenomena. 

On the basis of this definition, immunotoxicological investigations have 

already been carried out for years during the development of substances; 

namely with respect to DTH reactions (Type IV) in the guinea pig (Bühler, 

1965; Magnusson and Kligman, 1969). As an alternative to these tests in 

the guinea pig, the so-called local lymph node assay (LLNA) in the mouse 

according to Kimber et al. (1989) was developed and validated and has 

meanwhile been adopted as alternative test in the OECD guidelines (OECD, 

1992a, b; Botham et al., 1991). 

The development or selection of suitable tests for immunotoxicological 

screening and thus for incorporation in guidelines presents considerable 

problems. Most of the tests which have been proposed for 

immunotoxicological investigations and most knowledge and experience in 

immunology are based on mouse models. The standard animal in the early 

phase of toxicological testing, however, is the rat. Transference of the tests 

is not always easy, partly because of lack of suitable reagents.

198 EVALUATION OF TOXICITY TO THE IMMUNE SYSTEM 

The next problem is to find which tests can suitably be used—simply and 

without having to treat additional animals—for reliable identification of 

interactions with the immune system. Another question which has not yet 

been solved is relating to the dosages and the changes in immunological 

parameters which are still tolerable and at which times these should be 

determined. 

National and international collaborative studies 

Most of the guideline drafts favour a two-tiered to three-tiered approach 

for the screening of immunotoxic side effects, with the greatest disparities 

in respect of the proposals for tier I tests, ranging from just organ weights 

and a little more emphasis on the histology of the lymphatic organs, to a 

series of elaborate, supplementary function tests including, in some cases, 

satellite groups. 

Any discussion about basic tests is hampered by a lack of data from 

routine toxicological and/or epidemiological studies although a few years 

ago a number of collaborative studies were initiated, namely: ICICIS 

(international), US-NTP (international, USA (two)), BGA (Federal German 

Health Office, Germany), GEVI (France). The aim of all these studies is to 

check on different histological parameters and functional tests for their 

possibility to flag an immunotoxic potential of a compound in a routine 14 

or 28 day study (rats) in an interlaboratory trial. Two immunosuppressive 

standards (azathioprine and/or cyclosporin A) have been used so far. 

Although the experimental phase is finished the evaluation of the data is 

still underway. 

Table 15.1 summarises the immunotoxicological experiments and 

collaborative studies currently in progress. With a few exceptions all 

investigations are based on a 28-day gavage study in rats. These basic tests 

were supplemented by extended histopathology and functional tests. 

ICICIS 

The first substance to be investigated in the international collaborative 

study was azathioprine. However this first attempt suffered from lack of 

harmonisation between the models used by the 28 participants world-wide. 

This made the comparability and the evaluation of the results particularly 

problematic. 

The second substance which was tested by ICICIS in a more restrictive 

design with slightly fewer participants was cyclosporin A. The 

experimental section as well as the first evaluation of the results has now 

been completed. 

As the final evaluation—including statistics—of ICICIS is likely to take 

some time (years?) it will not be discussed further at this point.

Table 15.1 Immunotoxicology collaborative studies 

Fischer 344 study (Kimber-White) 

H.-W.VOHR 199 

An other interesting approach was done by Kimber-White and colleagues. 

In this study the duration of treatment was 14 days, Fischer 344 rats were 

used as experimental animals. Apart from the usual parameters the plaque 

assay (PFCA), mitogenic stimulation (ConA, LPS) and NK activity were 

measured.


No effects were found with respect to organ weights or cell counts 

analysis but there were marked effects in the PFCA. Here there was a good 

correlation between the laboratories and clearly dose-dependent effects 

were seen. Mitogenic stimulation was not and NK activity only slightly 

affected. 

Results of the BGA collaborative study 

In order to put the discussion on a somewhat sounder footing it is 

imperative to test the various models for detection of immunotoxicological 

potential in practice. For this reason Bayer AG is taking part in a 

collaborative trial initiated by the German Federal Health Office in Berlin 

(BGA study). In this collaborative study standard immunotoxic substances 

are investigated in parallel under restrictive conditions in several 

laboratories. On the other hand Bayer AG has introduced a set of 

functional immunological tests into the routine toxicological testing of 

agrochemicals in rats in order to test the informative value of these 

parameters in practice (Vohr, 1995). 

In the course of this collaborative study cyclosporin A was investigated 

as the first substance in a very well harmonised design. One reason for 

choosing cyclosporin A was to permit comparison with the ICICIS results. 

The study was based on OECD guideline 407 which was supplemented by 

a number of histopathological, haematological, clinical-chemical and 

functional parameters. The final evaluation, including pathology and 

statistics will take a few more months. Nevertheless a first look at the data 

showed that there were no marked effects with respect to either organ 

weights (lymphatic organs) or blood cells. On the other hand dose-related 

effects were found for many parameters in the functional tests. 

Examples are shown of these parameters and their changes at the various 

dosages. Both sexes show dose-related changes from the lowest dosage (1 

mg kg −1 ) upwards with respect to the surface markers of the 

immunocompetent cells. The PFCA and the measurements of IgA 

antibodies in serum also proved to be sensitive parameters. 

In immunotoxic investigations based on data obtained in rats treated 

with test substances (28-day test) we found that secondary influences often 

occur and that occasionally one single parameter is changed at the highest 

dosage. Apart from these non-specific effects, dose-related reactions were 

observed from the lowest or middle dose upwards (for example in the case 

of cytostatic substances). Such genuinely immunotoxic compounds were 

reliably identified by surface markers (like CD4/CD45R or PanB) and 

changes in the serum Ig-titres. Changes in other parameters such as cell 

counts and macrophage activity verified these findings.

Findings of the US-NTP study (Luster) 

The US-NTP study investigated 51 substances, 35 of which were declared 

immunotoxic, in a comprehensive test battery in mice for changes in 

functional parameters after 28-day administration of the substances. The 

correlations of each of these parameters with the given classification and 

with the results of host-resistance studies were calculated. The correlations 

after combinations of individual tests were also calculated (Luster et al., 

1992, 1993). 

The conclusion drawn from these investigations was that the 

immunotoxic potential of a substance can be determined relatively reliably 

by combination of 2–3 specific tests. The most powerful tests proposed by 

Luster et al. for such a combination include surface markers, NK test and 

PFCA. Serum titres of Ig —particularly IgA—were unfortunately not 

determined. 

With regard to the correlation with host resistance (HR) results it was 

found that if effects were shown in the HR model there were always effects 

on functional parameters, too. There were, however, also cases in which 

there were effects in the functional tests although the HR studies were 

negative. Although these investigations were carried out on mice and the 

choice and classification of the substances are not entirely undisputed, 

these findings are nevertheless confirmed by our own experience. 

Discussion and prospect 

H.-W.VOHR 201 

It is undoubtedly too early to make any judgement. However, it appears 

that apart from the histology—particularly the immuno-histology—a few 

additional parameters such as analysis of surface markers of subpopulations 

and serum titre assays of IgG, IgM and IgA are sufficient as screening 

indicators to show the possible immunotoxic potential of a substance. One 

of these is the PFCA, which presupposes, however, that satellite groups are 

used or that the authorities accept injection of SRBC as GLP treatment. 

Positive findings in a combination of these tests should then occasion 

further elucidation of the immunotoxic potential. 

In summary it can be concluded from the currently available results of the 

collaborative studies that the following criteria must be fulfilled if a 

substance is to be characterised as possibly immunotoxic. The substance 

must: 

1. Induce significant dose-related changes in one of the effective tests 

listed above, or 

2. induce significant changes in the highest dose group in a combination 

of 2–3 of these tests.


But also findings in lymphoid organs which are remarkable with respect to 

quality, severity or quantity of changes should be of sufficient relevance to 

warrant further assessment. 

Finally I would like to point out the problems involved in the evaluation 

of immunological changes. 

• In contrast to other data obtained in routine toxicology, historical data 

and routine experience with respect to the functional tests are so far 

almost entirely lacking. On account of this paucity of data a decision as 

to the right effect level for the functional tests can only be made on the 

basis of subjective judgement at present. The discussion of uniform 

criteria is only just beginning. 

• If effects only occur secondarily, e.g. on account of inflammatory 

processes, they are not specifically immunotoxic. It must be discussed 

whether such effects on the immune system, which is only doing its 

normal job in these cases, can be classed as immunotoxic at all. 

• The immune system can show an ‘oscillating’ response to a substance. A 

substance may have a stimulating effect in a low dose range and an 

inhibitory effect in a high range or vice versa. Such reactions are also 

dose-related effects. 

• Most immunotoxic investigations have so far used known 

immunosuppressive drugs. There are, however, little data—particularly 

from rat studies— on immunostimulant substances. Since the other 

unwanted side-effect apart from immunosuppression is immune 

potentiation, future collaborative investigations must urgently 

concentrate on such substances. Before this, no final evaluation and thus 

no recommendations on relevant tests can be made. 

For pharmaceuticals (Hinton, 1992), pesticides (US-EPA, 1990, 1993) and 

veterinary medicinal chemicals (EEC, 1991) final drafts or notes for 

guidance for the screening of the immunotoxic potential of a compound 

already exist. These draft proposals for immunotoxicity parameters for 

incorporation into new guidelines are shown in Tables 15.2 (FDA) and 

15.3 (US-EPA). 

For industrial chemicals advanced screening of lymphoid organs also 

with respect to functional parameters had been expected to be incorporated 

into the adopted OECD guideline No. 407 (1992). But recommendations 

made by van Loveren and Vos (1992) have not yet been taken into 

consideration for this update of OECD guideline 407. The proposal of 

these authors recommended more histopathology (gut associated lymphoid 

tissue), measurement of serum immunoglobulins, bone marrow cellularity, 

cytofluorimetry of spleen cells and measurement of NK cell activity. 

A task force ‘Immunotoxicology’ initiated by ECETOC has put forward 

proposals on hazard identification and risk assessment of immunotoxic

Table 15.2 Summary of immunotoxicity testing recommendations for direct food additives 

H.-W.VOHR 203 

Abbreviations: CBC=complete blood cell count; WBC=white blood cell count; Ig=Immunoglobin; NK=natural killer; 

IL-2=interleukin 2; SRBC=sheep red blood cells; TNP-LPS=trinitrophenol lipopolysaccharide. 

* Recommended for inclusion in basic testing. 

potential of a compound on the basis of the routine 28 day treatment of 

rats (Basketter et al., 1994; ECETOC, 1994). A central point in these


Table 15.3 Proposed amendment to the subdivision F guideline requirements to 

provide for an evaluation of the immunotoxicity of chemical pesticides 

proposals is a flow diagram for the evaluation of results from this basic 

study for hazard identification and conclusions drawn from it. This flow 

diagram, shown in Figure 15.1, could be helpful for the evaluation of 

results obtained from incorporation of a basic immunotoxicity test battery 

into studies with routinely treated animals. 

References 

BASKETTER, D. et al., 1994, The identification with sensitising or 

immunosuppressive properties in routine toxicology, Food Chem. Toxicol. (in 

press). 

BOTHAM, P.A. et al., 1991, Skin sensitization—a critical review of predictive test 

methods in animals and men Food Chem. Toxicol, 29, 275–86. 

BÜHLER, E.V., 1965, Delayed contact hypersensitivity in the guinea pig, Arch. 

Dermatol. 91, 171. 

DEAN, J.H., PADARATHSINGH, M.L. and JERRELS, T.R., 1979, Assessment of 

immunobiological effects induced by chemicals, drugs or food additives. I. Tier 

testing and screening approach, Drug Chem. Toxicol. 2, 5–17. 

ECETOC, 1990, Skin Sensitization Testing, Monograph No. 14, Brussels.

Figure 15.1 Flow diagram (taken from the ECETOC monograph: 

Immunotoxicity). 

H.-W.VOHR 205 

ECETOC, 1994, Immunotoxicity: hazard identification and risk assessment, 

Monograph No. 21, Brussels. 

EEC, 1991, Annex to directive 81/852/EEC—Note of guidance, 22.


HINTON, D.M., 1992, Testing guidelines for evaluation of the immunotoxic 

potential of direct food additives, Crit. Rev. Food Sci. Nutrit, 32, 173–90. 

KIMBER, I., HILTON, J. and WEISENBERGER, C. 1989, The murine local lymph 

node assay for identification of contact allergens: a preliminary evaluation of 

in situ measurement of lymphocyte proliferation, Contact Dermatit. 21, 215– 

20. 

LUSTER, M.I. et al., 1988, Development of a battery to assess chemical-induced 

immunotoxicity: National Toxicology Program’s guidelines for 

immunotoxicity evaluation in mice, Fundam. Appl. Toxicol. 10, 2–19. 

LUSTER, M.I. et al., 1992, Risk assessment in immuno-toxicology. I. Sensitivity 

and predictability of immune tests, Fundam. Appl. Toxicol., 18, 200–10. 

LUSTER, M.I. et al., 1993, Risk assessment in immuno-toxicology. II. Relationship 

between immune and host resistance tests, Fundam. Appl. Toxicol., 21, 71–82. 

MAGNUSSON, B. and KLIGMANN, A.M., 1969, The identification of contact 

allergens by animal assay. The guinea pig maximisation test, J. Invest. 

Dermatol., 52, 268. 

OECD, 1992a, Organisation for Economic Cooperation and Development, 

Guidelines for testing of chemicals No. 407, adopted 12 July, 1992. 

OECD, 1992b, Organisation for Economic Cooperation and Development, 

Guidelines for testing of chemicals—skin sensitisation, No. 406, adopted 17 

July, 1992. 

SJOBLAD, R., 1988, Potential future requirements for immunotoxicology testing 

of pesticides, Toxicol. Indust. Hlth, 4, 391. 

TRIZIO, D. et al., 1988, Identification of immunotoxic effects of chemicals and 

assessment of their relevance to man. Food Chem. Toxic., 26, 527–39. 

UK Department of Health, 1991, Proposed to update OECD Guideline 407. 

US-EPA, 1982, Code of Federal Regulations, Washington DC, 152–18, 152–24 and 

158–165. 

US-EPA, 1990, Draft immunotoxicity study screen for testing chemical pesticides. 

VAN LOVEREN, H. and Vos, J.G., 1992, Evaluation of OECD Guideline 407 for 

assessment of toxicity of chemicals with respect to potential adverse effects to 

the immune system. RIVM Report No. 158801001, Bilthoven: National 

Institute of Public Health and Environmental Protection. 

VOHR, H.-W., 1995, Experiences with an advanced screening procedure for the 

identification of chemicals with an immunotoxic potential in routine 

toxicology (a position paper). Toxicology, (in press),

16 

New Strategies: the Use of Long-term Cultures of 

Hepatocytes in Toxicity Testing and Metabolism 

Studies of Chemical Products Other than 

Pharmaceuticals 

VERA ROGIERS, 1 MAY AKRAWI, 2 SANDRA COECKE, 1 

YVES VANDENBERGHE, 1 ELIZABETH SHEPHARD, 2 IAN 

PHILLIPS 3 and ANTOINE VERCRUYSSE 1 

1 Vrije Universiteit Brussel, Brussels; 2 University College 

London, London; 3 University of London, London 

Introduction: metabolism and toxicity of chemical 

products are closely linked 

Lipophilic compounds are metabolized in the liver by phase 1 and phase 2 

reactions into more polar, more hydrophilic metabolites, which are usually 

less biologically active. Bioactivation, however, may occur, forming toxic 

species by phase 1, cytochrome P450 (CYP) dependent oxidation (e.g. 

epoxidation of C=C to reactive epoxide intermediates (Guengerich et al. 

1991), CYP dependent reduction (e.g. dehalogenation of CCl 4 toa free 

radical intermediate) (Timbrell, 1993) or even by phase 2 reactions (e.g. 

reactive episulphonium ion formation by glutathione conjugation of 

dibromoethane) (Van Bladeren, 1988; Coles and Ketterer, 1990; Timbrell, 

1993). 

The major process involved in the bioactivation of chemical carcinogens 

is their oxidation catalyzed by CYP enzymes. Thirty or more different CYP 

forms exist within each animal species (Nebert et al., 1989), each with at 

least some distinct elements of catalytic specificity and regulation. The role 

of some of these CYPs in the activation and detoxication of chemical 

carcinogens has already been determined. For example: 

– CYP2E1 is a major catalyst involved in the oxidation of benzene, 

styrene, CCl 4, CHCl 3, ethylene dichloride, vinylchloride, acrylonitrile, 

vinyl carbamate (Guengerich et al., 1991), ethanol (Perrot et al., 1989), 

dialkylnitrosamines (Yoo et al., 1988), isobutene (Cornet et al., 1991) 

and some other small molecules. 

– CYP1A1 is involved in the oxidation of polycyclic aromatic 

hydrocarbons (Shimada et al., 1989b). 

– CYP1A2 activates arylamines (Shimada et al., 1989a).

208 USE OF LONG-TERM CULTURES OF HEPATOCYTES IN TOXICITY TESTING 

– CYP3A4 is a major catalyst in the activation of aflatoxins, pyrrolizidine 

alkaloids and polycyclic hydrocarbon dihydrodiols (Shimada et al., 

1989a; Shimada and Guengerich, 1989). 

The balance between the rates of formation of reactive metabolites and 

detoxication will greatly determine the potential toxic response of a 

chemical. Variables, known to affect normal biotransformation, such as 

enzyme induction and inhibition can also change the bioactivation rate of 

chemicals. A clear example is the metabolism of halogenated biphenyls 

after treatment with arochlor 1254 (Borlakoglu and Wilkins, 1993). 

Consequently, the toxicity of chemical products often depends upon their 

specific biotransformation, the presence, absence, induction and inhibition 

of specific phase 1 and phase 2 enzymes involved in their metabolism. 

Towards an in vitro approach of risk assessment 

Nearly all toxicological studies on chemical products, including industrial 

chemicals, agrochemicals, pharmaceuticals, additives, materials in contact 

with food and cosmetics, have been carried out in vivo using experimental 

animals, in particular small vertebrates (News and Views, 1993). 

Important scientific, technological, ethical and economic considerations, 

however, justify the actual search for in vitro alternatives, replacing or 

improving existing in vivo methods and reducing the number of animals 

involved (Frazier and Goldberg, 1990; Roberfroid, 1991). 

Since hepatocytes can be isolated from different species (Guguen- 

Guillouzo et al., 1982; Green et al., 1986; van’t Klooster et al., 1992) 

including man (Guguen-Guillouzo et al., 1982; Rogiers, 1993), they can 

represent a powerful tool for short-term risk assessment studies when used 

as suspensions of freshly isolated cells (up to 3–4 h) or as short-term 

cultures (up to 2 days) (Klaassen and Stacey, 1982; Guillouzo, 1986). They 

can be useful in biotransformation, cytotoxicity, hepatotoxicity and 

genotoxicity studies, in species selection and in mechanistic studies 

(Blaauboer, 1994). Long-term cultures of hepatocytes (up to several weeks) 

represent a somewhat different approach in risk assessment. Such systems 

are of interest for the assessment of long-term toxicity of xenobiotics 

including the occurrence of enzyme induction, the effects on xenobiotic 

biotransformation, lipid peroxidation, accumulation of triglycerides, 

changes in glutathione content, interaction between compounds and the 

hepatoprotection afforded by certain molecules. Consequently, long-term 

cultures have been particularly applied in the development of 

pharmaceuticals, in pharmaco-toxicological studies (Guillouzo, 1986, 

1992; Rogiers and Vercruysse, 1993; Skett, 1995). As far as chemical 

products other than pharmaceuticals and in particular industrial chemicals 

and agrochemicals are concerned, the practical needs during development

are different. The compounds brought onto the European market are 

labelled and need to fulfill only the legal demands of the specific category 

to which they belong. 

Testing is only obligatory when they reach the market and the type of 

tests needed per category of chemicals is clearly outlined. In Europe legal 

categories consist of dangerous compounds (88/379/EEC, 93/18/EEC), 1a,b 

phytopharmaceuticals (91/414/EEC, 93/71/EEC), 2 biocides (93/C239/03), 3 

cosmetics (76/ 768/EEC, 93/35/5/EEC) 4a,b and food additives. Of the latter 

category the most comprehensive surveys are carried out by the Joint 

Expert Committee on Food Additives (JECFA) of the World Health 

Organization and the Food and Agriculture Organization of the United 

Nations (Conning, 1993). 

Less sophisticated in vitro studies have been performed on industrial 

chemicals and agrochemicals than is the case for pharmaceuticals. The only 

field in which isolated hepatocytes have already been incorporated into 

routine screening of industrial chemicals for regulatory purposes is in 

genotoxicity testing (Swierenga et al., 1991). The potentialities, however, 

of in vitro testing for these compounds, in particular of the use of long-term 

cultures, has not yet been explored in depth, although induction, inhibition, 

biotransformation, chronic toxicity, interaction between chemicals and 

mechanistic studies are of great interest for these compounds too. 

Human exposure to chemical products such as pesticides, eventually 

reaching the food chain as residues or of potential risk for workers and 

operators spraying the fields, is such an interesting research area. For 

example, Alachlor ® , a herbicide, of which millions of tons are used per 

year, has been classified as a potential human carcinogen (Leslie et al., 

1989) because of tumour formation in rats and DNA damage observed in 

isolated rat hepatocytes (Bonfanti et al., 1992). 

In vivo studies concerning its biotransformation in rat, mouse and 

monkey, however, pointed to the observation that Alachlor ® is metabolized 

via different pathways in rodents and monkeys, suggesting a lower risk for 

man than assumed (internal report Monsanto, 1988). In the future, this 

type of biotransformation study could easily be performed with short- and 

long-term cultures of hepatocytes derived from different species, including 

man, providing relevant human information without interspecies 

extrapolation. 

Long-term hepatocyte cultures 

V.ROGIERS ET AL. 209 

During culture, hepatocytes undergo phenotypic changes as a function of 

culture time affecting selectively components of phase 1 and/or phase 2 

biotransformation (Nakamura et al., 1983; Guillouzo, 1986; Mooney et 

al., 1992; Kocarek et al., 1993; Rogiers and Vercruysse, 1993). These 

changes are interpreted as dedifferentiation. In the literature, data have


been presented that part of this loss of functionality can be prevented by 

several factors including soluble medium factors, extracellular matrix 

components and cell-cell interactions (reviews Guillouzo et al., 1990; 

Rogiers, 1993). At present no ‘ideal’ long-term hepatocyte culture model 

exists but valuable alternatives are being developed which take into 

account some of the factors mentioned. A recent development consists of 

hepatocytes cultured in a collagen gel sandwich configuration (Dunn et al., 

1991; Lee et al., 1992). This promising system is claimed to maintain longterm 

differentiation probably due to the reinstatement of the cellular 

polarity of the hepatocytes as a function of the extracellular matrix (Dunn 

et al., 1991; Lee et al., 1992). To date, only few results concerning xenobiotic 

metabolism, are available (Koebe et al., 1994). Another recently introduced 

model substantially improved the maintenance of xenobiotic metabolism 

by culturing hepatocytes on a mixture of crude membrane fractions with 

collagen type 1, combined with the use of culture medium supplemented 

with aprotinin and selenium (Saad et al., 1993). 

Also co-cultures of hepatocytes with rat epithelial cells of primitive 

biliary origin represent a rather new and valuable tool in xenobiotic 

biotransformation research and testing (Bégué et al., 1984a). The model 

has been developed in order to mimic better the microenvironment of the 

liver cells in vivo. It is until now, the only long-term hepatocyte culture 

system of which enough biotransformation data exist. Co-cultures of 

hepatocytes retain to a great extent the morphological and biochemical 

characteristics of adult hepatocytes in vivo, including phase 1 and phase 2 

xenobiotic metabolism pathways (Guillouzo, 1986; Rogiers et al., 1992; 

Akrawi et al., 1993a). The following text reviews the actual knowledge 

concerning xenobiotic biotransformation in cocultured hepatocytes with 

emphasis on the authors’ own research. 

Phase 1 

reactions in co-cultured hepatocytes 

It has been claimed that as much as 100 per cent of the CYP content and Naminopyrine 

demethylation activity can be maintained in co-cultured rat 

hepatocytes with primitive biliary duct cells (Bégué et al., 1984b). Some 

other phase 1 enzymatic activities also appear to be maintained since drugs 

such as ketotifen (Le Bigot et al., 1987) and testosterone (Utesch, 1992) 

were metabolized by several pathways. Maier (1988) however, has 

reported that aldrin epoxidase activity underwent a significant decrease as 

a function of culture time. Niemann et al. (1991) was able to show that, as 

was also the case for mono-cultured rat hepatocytes, enzymatic activities 

belonging to the 3-methyl-cholanthrene-inducible family were better 

maintained than those belonging to the phenobarbital inducible family. In 

our own experiments with adult rat hepatocytes co-cultured with rat liver


epithelial cells (Rogiers et al., 1990b), it was found that a steady-state 

situation for at least 10 days was obtained in which the total CYP content 

and 7-ethoxycoumarin O-deethylase and aldrin epoxidase activities were 

maintained at 25,100 and 15 per cent, respectively, of their corresponding 

values in freshly isolated hepatocytes. Both the total CYP content and 7ethoxycoumarin 

O-deethylase activity, but not the aldrin epoxidase 

activity, were induced by exposure to phenobarbital (Rogiers et al., 1990b) 

or sodium valproate (Rogiers et al., 1988b). Furthermore, the combination 

of inducing agents with co-cultivation of rat hepatocytes with rat epithelial 

cell clones has recently been proposed as the best way for stabilization of 

the CYP system. This method is better than the use of a perfusion system 

or changing the extracellular matrix from collagen to matrigel (Wegner et 

al., 1991). The results mentioned above indicate a degree of selection in the 

ability of co-cultured hepatocytes to maintain the expression and 

inducibility of individual members of the CYP superfamily. From the work 

of Akrawi (Akrawi et al., 1993a), using mRNA analysis of co-cultured rat 

hepatocytes, it appeared that the abundance of CYP2B mRNAs declined to 

about 30 per cent of the initial value by 4 days but that thereafter it remained 

constant. The inducibility by phenobarbital (Akrawi et al., 1993a) and 

sodium valproate (Rogiers et al., 1992) was also maintained. These results 

were confirmed by Western blotting (Akrawi et al. 1993b). RNase 

protection assays using probes capable of distinguishing between CYP2B1 

and CYP2B2 mRNAs demonstrated that the relative abundance and 

inducibility of each of the mRNAs were the same in co-cultures as in vivo. 

Co-cultured hepatocytes also maintained the expression of the CYP1A2 

gene and of genes coding for two other components of the CYP-mediated 

monooxygenase, namely NADPH cytochrome P450 reductase and 

cytochrome b 5 (Akrawi et al., 1993a). Both components were inducible by 

valproate and phenobarbital (Akrawi et al., 1993b; Shephard et al., 1994; 

Rogiers et al., 1994). 

In addition, we have shown using Western blotting (Akrawi et al., 

1993b) and mRNA analysis (Akrawi et al., 1994; Shephard et al., 1994) 

that the expression of CYP4A and its specific inducibility (induced by 

valproate and not by phenobarbital) were maintained in co-cultured rat 

hepatocytes. By using antisense RNA probes that could discriminate 

between RNAs encoding different members of the CYP4A subfamily it was 

further demonstrated that CYP4A1, CYP4A2 and CYP4A3 were all 

induced by valproate, although to differing extents. None of these mRNAs 

was increased by phenobarbital (Akrawi et al., 1994; Shephard et al. 

1994). The results were very similar to those observed in vivo (Shephard et 

al., 1994). 

Flavin-containing monooxygenase (FMO), a less well-known phase 1 

biotransformation enzyme than the CYP system, is responsible for the 

oxygenation of drugs, pesticides and dietary components (Ziegler, 1980). It


may activate as well as deactivate a number of important molecules such as 

thioether-containing pesticides (Hajjar and Hodgson, 1980), 3,3′-dichlorobenzidine 

(Iba and Thomas, 1988), N-methyl-4-aminobenzene (Kadlubar 

et al., 1976) and others. The expression of FMOs is also maintained better 

and for a longer time in co-cultures of rat hepatocytes than it is in monocultures 

of these cells (Coecke et al., 1993). In co-cultures a steady state 

situation is obtained at a level of approximately 40 per cent of its initial 

value in freshly isolated hepatocytes (Coecke et al., 1993). 

Hormonal regulation of FMO is retained (Coecke et al., 1995a,b). It was 

shown that 17β-oestradiol significantly decreased FMO activity in cocultures 

of male rat hepatocytes which was not the case for testosterone 

and 5α-dihydrotestosterone. These data are in accordance with in vivo 

results (Coecke et al., 1995b). In addition, the thyroid hormone thyroxine 

and its metabolite L-triiodothyronine were found to cause a significant 

decrease in FMO activity suggesting a suppressive role in the regulation of 

FMO in rat liver. In vivo data on this subject are not available in the 

literature. 

Phase 2 

reactions in co-cultured hepatocytes 

The activity, expression and regulation of the different glutathione Stransferase 

(GST) isoenzymes in co-cultured rat hepatocytes, has been 

extensively studied by our group (Vandenberghe et al., 1988a,b, 1989, 

1990a,b, 1991). As far as the enzymatic activity is concerned, it was 

observed that the composition of the culture medium was of much less 

influence in co-cultures than was the case in mono-cultures (Vandenberghe 

et al., 1988b). In cocultures GST activity was maintained for a longer 

period and at a more stable level, comparable to the in vivo situation. This 

observation was confirmed later by Niemann et al. (1991) and Utesch and 

Oesch (1992), although the latter investigators reported a high variability 

in their results depending on the batch of epithelial cells. GST activity in cocultured 

rat hepatocytes was found to be increased or decreased by 

phenobarbital and valproate, respectively (Rogiers et al., 1988a; Rogiers et 

al., 1992). 

These results are in good accordance with previously obtained in vivo 

data (Rogiers et al., 1988a; Rogiers et al., 1992). They were also confirmed 

at the protein level using the Western blot technique (Rogiers et al., 1995). 

Furthermore, GST activity is increased significantly, in a dose dependent 

way by ethanol. Both GST protein and mRNA amounts (in particular, GST 

subunits 3 and 4) were increased by this compound (Coecke et al., 1995c). 

Results obtained using different substrates suggested that the GST isoenzyme 

profile changes as soon as hepatocytes are seeded in culture 

(Vandenberghe et al., 1988b). By a combination of GSH agarose affinity


chromatography and reversed phase HPLC, the GST subunits of cocultured 

rat hepatocytes were purified and separated (Vandenberghe et al, 

1988a). Alterations comparable to those observed for mono-cultures 

suggested changes towards a more ‘foetal-like’ state. Less variations, 

however, were noticed in the GST subunit pattern of co-cultured 

hepatocytes when various media conditions were compared. Incorporation 

of 35 S-methionine in the medium showed the ability of co-cultured rat 

hepatocytes to synthesize the different GST subunits and suggested that 

changes in GST subunit expression under various culture conditions were 

the result of in vitro ‘de novo’ synthesis (Vandenberghe et al., 1990a). 

Northern blot analysis, using specific cDNA probes showed that the 

mRNA levels encoding GST subunits 1/2, 3/4 and 7 were very dependent 

on the culture medium. 

Again in co-cultures, the changes observed were much less marked than 

was the case for mono-cultures (Morel et al., 1989; Vandenberghe et al., 

1990b). As already mentioned for conventionally cultured rat hepatocytes, 

phenobarbital had inducing effects on all the GST subunits, but to a 

different extent for each subunit (Vandenberghe, 1989). The increased 

steady-state mRNA levels observed in co-cultures after phenobarbital 

exposure were the result of an increased transcriptional activity of the GST 

genes together with a stabilizing effect of the compound (Vandenberghe et 

al., 1991). 

Also of interest is that the hormonal regulation of GST is maintained in 

co-cultures of male rat hepatocytes. 17β-Oestradiol, triiodothyronine and 

thyroxine cause a significant decrease in GST activity. Both the overall GST 

activity and in particular that of GST 3–3 and 3–4 are decreased (Coecke 

et al., 1995c). In contrast, male sex hormones and human growth hormone 

had little effect on the overall activity. The effects of triiodothyronine and 

thyroxine were particularly oriented towards GST subunits 3 and 4 and 

towards an as yet unidentified GST subunit, which was significantly 

increased (Coecke et al., 1995c). 17β-Oestradiol shifted the GST subunit 

pattern towards the one observed in freshly isolated cells whereas growth 

hormone had no specific effect on the individual protein classes (Coecke et 

al., 1995c). 

These results clearly show a hormonal regulation of GST in co-cultured 

rat hepatocytes, although previous work with mono-cultures failed to 

prove any direct effect (Gebhardt et al., 1990). The effects are most 

pronounced for the Mu-class GSTs. In man, Mu-class GST genes are 

structurally very similar to the rat genes and are of particular interest 

because 45 per cent of the European population fails to express a 

transferase at the GST M 1 locus (Zhong et al., 1993). It is this class of GSTs 

that is very effective in deactivating mutagenic and carcinogenic epoxides. 

UDP glucuronyltransferases (UDP-GT) have been much less studied in 

cocultures than GST. From the work of Niemann et al. (1991) it appears


that 1-naphthol UDP-GT activity is maintained well in co-cultures whereas 

this is not true for morphine UDP-GT, although for both a steady state 

situation was reached. These data point to a shift towards a more 

differentiated pattern since 1-naphthol is considered to be more specific for 

the late foetal form of UDP-GT and morphine for the neonatal form. The 

maintenance of 1-naphthol UDP-GT has been confirmed by Utesch and 

Oesch (1992). Other studies on preservation of phase 2 enzymatic activity 

in co-cultures have dealt with the identification of the metabolites formed 

when drugs are added to the culture medium. In human co-culture it could 

be shown that the glucuronide metabolite of ketotifen was still present 

after 3 weeks whereas it became undetectable after 6 days in mono-culture 

(Bégué et al., 1984b). Similar observations have been made for other drugs 

such as caffeine and theophylline (Ratanasavanh et al., 1990). 

Conclusions 

At present no ideal culture system for hepatocytes can be proposed. In all 

models reported in the literature, phenotypic changes occur, affecting the 

various components of phase 1 and phase 2 xenobiotic metabolism to a 

different extent. An interesting conclusion, however, remains from the 

observation that, when co-cultures and mono-cultures of hepatocytes are 

compared, cocultures exhibit higher biotransformation capacities which are 

better and preserved for longer than is the case for mono-cultures. The 

inducibility by common inducers is fairly well maintained and seems, to a 

certain extent, comparable with the in vivo situation. In addition, hormonal 

regulation of phase 1 and phase 2 key enzymes seems to be well maintained 

and comparable with the in vivo situation. Co-cultures of hepatocytes with 

rat liver epithelial cells are therefore already of importance as an alternative 

model for risk assessment. In particular, when long-term effects of a 

chemical are to be expected. 

Some experience already exists concerning the application of co-cultured 

hepatocytes for the study of pharmaceuticals. As far as chemical products 

other than pharmaceuticals are concerned, experience is lacking although 

interesting results are to be expected particularly in those cases where 

chemicals can interfere with the human organism via the food chain or by 

occupational exposure. In vitro exploration of this new field in toxicology 

is a challenge for the coming years. 

Notes 

1a. 88/379/EEC 

Directive du Conseil du 7 juin 1988 concernant le rapprochement des 

dispositions législatives, réglementaires et administratives des Etats membres

elatives à la classification, à l’emballage et à 1’étiquetage des préparations 

dangereuses. Journal Officiel des Communautés européennes no L187, 16 

juillet 1988, p. 14. 

1b. 93/18/EEC 

Directive 93/18/CEE de la Commission du 5 avril 1993 portant troisième 

adaptation au progrès technique de la directive 88/379/CEE du Conseil 

concernant le rapprochement des dispositions législatives, réglementaires et 

administratives des Etats membres relatives à la classification, à l’emballage 

et à 1'étiquetage des preparations dangereuses. Journal Officiel des 

Communautés européennes no L104, 29 Avril 1993, p. 46. 

2a. 91/414/EEC 

Richtlijn van de Raad van 15 juli 1991 betreffende het op de markt 

brengen van gewasbeschermingsmiddelen. Publikatieblad van de Europese 

Gemeenschappen no L230, 19 Augustus 1991, p. 1. 

2b. 93/71/EEC 

Directive 93/71/CEE de la Commission du 27 juillet 1993 modifiant la 

directive 91/414/CEE du Conseil concernant la mise sur le marché 

des produits phytopharmaceutiques. Journal Officiel des Communautés 

européennes no L221, 31 Août 1993, p. 27. 

3. 93/C239/03 

Voorstel voor een richtlijn van de Raad betreffende het op de markt 

brengen van biociden. Publicatieblad van de Europese Gemeenschappen no 

C239, 3 September 1993, p. 3. 

4a. 76/768/EEC 

Richtlijn van de Raad van 27 juli 1976 betreffende de onderlinge 

aanpassing van de wetgevingen der Lid-Staten inzake kosmetische produkten. 

Publikatieblad van de Europese Gemeenschappen no L262, 27 juli 1976, p. 

169. 

4b. 93/35/EEC 

Directive 93/35/CEE du Conseil du 14 juin 1993 modifiant pour la sixième 

fois la directive 76/768/CEE concernant le rapprochement des lé-gislations des 

Etats membres relatives aux produits cosmétiques. Journal Officiel des 

Communautés européennes no L151, 23 juin 1993 p. 32. 

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PART FIVE 

Mechanisms of toxicity of industrial 

chemicals

17 

Peroxisome Proliferation 

BRIAN G.LAKE and ROGER J.PRICE 

BIBRA International, Carshalton, Surrey 

Introduction 

Peroxisomes (sometimes referred to as ‘microbodies’) are single 

membranelimited cytoplasmic organelles which are characterised by their 

content of catalase and a number of hydrogen peroxide generating oxidase 

enzymes (Cohen and Grasso, 1981; Reddy and Lalwani, 1983). In rat liver, 

peroxisomes are normally spherical or oval in shape, approximately 0.5 µm 

in diameter and contain a finely granular matrix with a crystalline nucleoid 

core. A number of reviews have been published dealing with various 

aspects of hepatic peroxisome proliferation (Cohen and Grasso, 1981; 

Reddy and Lalwani, 1983; Hawkins et al., 1987; Stott, 1988; Lock et al., 

1989; Moody et al., 1991; Bentley et al., 1993; Lake, 1993). This chapter 

will focus on mechanisms of hepatocarcinogenesis, species differences in 

response and risk assessment of rodent peroxisome proliferators. 

Peroxisome proliferation in rodent liver 

Since the initial observations on the hepatic effects of the hypolipidaemic 

agent clofibrate (Paget, 1963; Hess et al., 1965) many compounds have 

been shown to produce hepatic peroxisome proliferation in rats and mice. 

Liver enlargement is due to both hyperplasia and hypertrophy and 

organelle proliferation is associated with a differential induction of 

peroxisomal enzyme activities. Peroxisomes, like mitochondria, contain a 

complete fatty acid β-oxidation cycle (Lazarow and DeDuve, 1976). While 

the enzymes of the β-oxidation cycle (normally assessed as cyanideinsensitive 

palmitoyl-CoA oxidation) are markedly induced, only small 

changes are observed in other peroxisomal enzyme activities such as 

catalase and D-amino acid oxidase. Apart from stimulating peroxisomal 

fatty acid metabolism, peroxisome proliferators also increase microsomal 

fatty acid ( -l)- and particularly -hydroxylase activities. This is due to 

induction of cytochrome P-450 isoenzymes in the CYP4A subfamily and is

224 PEROXISOME PROLIFERATION 

normally measured as lauric acid 12-hydroxylase (Sharma et al., 1988a, b; 

Gibson, 1989). Peroxisome proliferators also markedly increase carnitine 

acetyltransferase activity which is localised in peroxisomal, mitochondrial 

and microsomal fractions (Ishii et al., 1980; Bieber et al., 1981). In rat liver 

good correlations have been reported between the induction of 

peroxisomal fatty acid β-oxidation and organelle proliferation and between 

the induction of peroxisomal and microsomal fatty acid oxidising enzyme 

activities (Lake et al., 1984a; Lin, 1987; Sharma et al., 1988a, b; Dirven et 

al., 1992). 

Several laboratories have demonstrated that the characteristics of 

peroxisome proliferation in vivo may also be observed in vitro in primary 

rat and mouse hepatocyte cultures. Indeed, hepatocyte cultures have been 

employed for studying various aspects of peroxisome proliferation 

including structureactivity relationships and species differences in response 

(Gray et al., 1982; Elcombe, 1985; Bieri, 1993; Lake and Lewis, 1993; 

Foxworthy and Eacho, 1994). 

Rodent peroxisome proliferators 

Many different classes of chemicals have been found to produce 

peroxisome proliferation in the rat and mouse (Cohen and Grasso, 1981; 

Reddy and Lalwani, 1983; Stott, 1988; Moody et al., 1991; Bentley et al., 

1993; Lake and Lewis, 1993). Classes of industrial chemicals include 

plasticisers, chlorinated solvents (e.g. trichloroethylene, perchloroethylene), 

chlorinated paraffins and other chemicals (e.g. perfluoro-n-octanoic acid). 

Types of plasticisers known to produce peroxisome proliferation include 

phthalate esters (e.g. di-(2-ethyl-hexyl)phthalate (DEHP), di-(isodecyl) 

phthalate), adipate esters (e.g. di-(2-ethyl-hexyl)adipate (DEHA)) and other 

compounds (e.g. tri-(2-ethylhexyl) trimellitate). Apart from industrial 

chemicals other known rodent hepatic peroxisome proliferators include 

herbicides, hypolipidaemic and other categories of therapeutic agents, 

certain steroids, food flavours and natural products. 

While peroxisome proliferators appear to be structurally diverse, at least 

for some compounds, similarities in their three-dimensional structures have 

been reported (Lake et al., 1988; Lake and Lewis, 1993). Many studies 

have demonstrated structure-activity relationships for various classes of 

peroxisome proliferators including industrial chemicals (Lake and Lewis, 

1993). A characteristic feature of many, but not all, peroxisome 

proliferators is the presence of an acidic function (Lake et al., 1988; Lock 

et al., 1989). This acidic function is normally a carboxyl group, either 

present as a free carboxyl group in the parent structure or one that is 

unmasked by metabolism. Alternatively, the chemical may contain a 

chemical grouping which is a bioisostere of a carboxyl group (Thornber,

1979), such as tetrazole or a sulphonamide moiety (Eacho et al., 1986; 

Lock et al., 1989). 

It should be noted that rodent liver peroxisome proliferators exhibit 

marked compound potency differences. While potent peroxisome 

proliferators include compounds developed as hypolipidaemic agents (e.g. 

ciprofibrate, Wy-14,643), plasticisers such as DEHP are less potent and 

chemicals such as acetylsalicylic acid are even less potent (Reddy et al., 

1986; Barber et al., 1987; Lake and Lewis, 1993). For example, in a 30day 

feeding study, a similar magnitude of induction of palmitoyl-CoA 

oxidation was observed in rats fed 0.001 per cent ciprofibrate and 0.5 per 

cent DEHP diets, whereas for DEHA a dietary level of >1.0 per cent but


demonstrated to be effective in rat liver tumour promotion studies (Cattley 

and Popp, 1989; Bentley et al., 1993; Popp and Cattley, 1993). 

Mechanisms of hepatocarcinogenesis 

Several hypotheses have been proposed to account for why peroxisome 

proliferators can produce liver tumours in rodents. These mechanisms 

include: 

(a) Induction of sustained oxidative stress to hepatocytes (Reddy and 

Lalwani, 1983; Reddy and Rao, 1989). 

(b) A role of increased cell proliferation (Marsman et al., 1988; Popp and 

Marsman, 1991). 

(c) The promotion of spontaneously formed preneoplastic liver lesions 

(Schulte-Hermann et al., 1989; Cattley et al., 1991; Grasl-Kraupp et 

al., 1993). 

(d) A combination of two or all of the above factors. 

The oxidative stress hypothesis is based on the observation that the chronic 

administration of peroxisome proliferators produces a sustained oxidative 

stress in rodent hepatocytes due to an imbalance in the production and 

degradation of hydrogen peroxide (Reddy and Lalwani, 1983; Reddy and 

Rao, 1989). Peroxisome proliferators markedly induce the peroxisomal 

fatty acid β-oxidation cycle, but produce only a small increase in catalase 

activity. The first enzyme of the β-oxidation cycle, acyl-CoA oxidase, 

produces hydrogen peroxide and hence the cyclic oxidation of a single fatty 

acid molecule can result in the production of several molecules of hydrogen 

peroxide (Lazarow and DeDuve, 1976). Any excess hydrogen peroxide not 

destroyed by peroxisomal catalase can diffuse through the peroxisomal 

membrane into the cytosol where it will be a substrate for cytosolic 

selenium-dependent glutathione peroxidase. However, this enzyme activity 

and that of other enzymes including superoxide dismutase and glutathione 

S-transferases are often reduced by the administration of peroxisome 

proliferators to rodents (Reddy and Rao, 1989; Bentley et al., 1993; Lake, 

1993). These enzyme changes are postulated to result in increased 

intracellular levels of hydrogen peroxide which, either directly or via 

reactive oxygen species (e.g. hydroxyl radical), can attack membranes and 

DNA (Reddy and Lalwani, 1983; Reddy and Rao, 1989). 

A number of experimental observations have provided support for the 

involvement of oxidative stress in the hepatotoxicity of peroxisome 

proliferators (Reddy and Rao, 1989; Lake, 1993). For example, 

peroxisome proliferators have been reported in some studies to increase 

hepatic lipid peroxidation and lipofuscin deposition, to modulate levels of 

hepatic antioxidants and to increase levels of 8-hydroxydeoxyguanosine in

B.G.LAKE AND R.J.PRICE 227 

hepatic DNA (Reddy and Rao, 1989; Bentley et al., 1993; Lake, 1993). 

However, the available data suggest that sustained oxidative stress is 

unlikely to be solely responsible for per oxisome proliferator-induced 

hepatocarcinogenesis in rodents. Although evidence of oxidative damage to 

hepatocytes has been observed in some studies, the magnitude of such 

effects does not correlate with the potency of the compound to produce 

tumours. For example, oxygen radical attack on DNA is known to result in 

a variety of modified DNA bases including 8-hydroxydeoxyguanosine. 

However, at bioassay dose levels both DEHP and DEHA produce similar 

increases in hepatic 8-hydroxydeoxyguanosine levels, but only DEHP 

produced liver tumours in male F344 rats (NTP, 1982a, b; Takagi et al., 

1990; Lake, 1993). 

Many studies have demonstrated that cell proliferation is an important 

factor in the development of tumours by both genotoxic and nongenotoxic 

agents (Cohen and Ellwein, 1990, 1991). For example, an enhanced rate of 

cell replication can increase the frequency of spontaneous lesions and the 

probability of converting DNA adducts from both endogenous and 

exogenous sources into mutations before they can be repaired (Cohen and 

Ellwein, 1990, 1991; Popp and Marsman, 1991). Peroxisome proliferators 

are known to produce a burst of cell replication in rodent hepatocytes 

during the first few days of administration (Reddy and Lalwani, 1983; 

Eacho et al., 1991). In some studies peroxisome proliferators have also 

been shown to produce a sustained stimulation of replicative DNA 

synthesis (Lake, 1993). Apart from intrinsic compound potency, dose is an 

important factor in determining whether a particular compound can 

produce either a transient or a sustained stimulation of replicative DNA 

synthesis in rodent hepatocytes. For example, low doses of nafenopin and 

Wy-14,643 do not produce a sustained stimulation of cell replication, 

whereas higher doses do produce this effect (Eacho et al., 1991; Price et al., 

1992; Wada et al., 1992; Lake et al., 1993). 

Several studies have demonstrated the presence of numerous foci of 

putative preneoplastic cells in the livers of untreated old rats and mice 

(Schulte-Hermann et al., 1983; Grasl-Kraupp et al., 1993). These lesions 

are considered to represent spontaneously initiated cells as they have 

similar biological characteristics to those of cells initiated by genotoxic 

carcinogens (Grasl-Kraupp et al., 1993). The ability of peroxisome 

proliferators to produce tumours in young compared to old rats has been 

investigated in studies with nafenopin (Kraupp-Grasl et al., 1991) and 

Wy-14,643 (Cattley et al., 1991). In both studies more adenomas and 

carcinomas were produced in old as against young rats.


Species differences in response 

Many studies have examined species differences in hepatic peroxisome 

proliferation (Cohen and Grasso, 1981; Rodricks and Turnbull, 1987; 

Stott, 1988; Lock et al., 1989; Moody et al., 1991; Bentley et al., 1993). 

Based on both marker enzyme activities and ultrastructural examination 

the rat and mouse are clearly responsive species, the Syrian hamster 

appears to exhibit an intermediate response, whereas in most studies the 

guinea pig is either nonresponsive or refractory. For example, DEHP 

readily produces peroxisome proliferation in the rat and mouse, to a lesser 

extent in the Syrian hamster but not in the guinea pig (Osumi and 

Hashimoto, 1978; Lake et al., 1984b). Similar results have been obtained 

with more potent compounds including ciprofibrate, clobuzarit, LY 

171883 and nafenopin (Orton et al., 1984; Eacho et al., 1986; Lake et al., 

1989; Makowska et al., 1992). 

When assessing species differences in response a number of factors 

should be considered. These include the metabolism, disposition and dose 

of the test compound, sex differences, as well as intrahepatic differences in 

response. The importance of metabolism is illustrated by the industrial 

solvent trichloroethylene which produces peroxisome proliferation and 

liver tumours in the mouse but not in the rat (NCI, 1976; Elcombe, 1985). 

Metabolic studies demonstrated that the trichloroethylene was extensively 

metabolised to trichloroacetic acid in the mouse, whereas this was a minor 

saturable route of metabolism in the rat. That the difference in 

trichloroacetic acid formation was responsible for the observed species 

difference was demonstrated by the fact that this compound produced 

peroxisome proliferation in rat and mouse hepatocytes both in vivo and in 

vitro (Elcombe, 1985). An example of compound disposition is provided 

by DEHP which is known to be more extensively absorbed after oral 

administration in the rat than in the marmoset (Rhodes et al., 1986). 

However, the observed in vivo species differences in response are supported 

by the observation that metabolites of DEHP which produce peroxisome 

proliferation in rat hepatocytes in vitro have no significant effect in 

cultured marmoset hepatocytes (Elcombe and Mitchell, 1986). Generally, 

in vitro studies with primary hepatocyte cultures from the rat, mouse, 

Syrian hamster, guinea pig and marmoset have supported the results of in 

vivo studies in these species (Elcombe 1985; Elcombe and Mitchell, 1986; 

Lake et al., 1986; Bieri, 1993; Bentley et al., 1993; Foxworthy and Eacho, 

1994). 

Several studies have examined the ability of rodent peroxisome 

proliferators to produce effects in primates and humans. With respect to 

primates, studies with a number of compounds in both New (e.g. 

marmoset) and Old (e.g. Rhesus monkey) World monkeys have failed to 

provide any evidence of significant hepatic peroxisome proliferation

(Rodricks and Turnbull, 1987; Bentley et al., 1993). However, albeit at 

high doses two compounds, namely ciprofibrate (Reddy et al., 1984) and 

DL-040 (Lalwani et al., 1985), have been reported to produce hepatic 

peroxisome proliferation in Cynomolgus and/or Rhesus monkeys. In 

humans, studies have been conducted in patients treated with several 

hypolipidaemic agents (all being rodent peroxisome proliferators) including 

ciprofibrate, clofibrate, fenofibrate and gemfibrozil (Bentley et al., 1993). 

While most studies have failed to detect any significant changes, clofibrate 

was reported to produce a small increase in the number of peroxisomes 

(Hanefeld et al., 1983) and ciprofibrate to produce a small increase in the 

pro portion of the hepatocyte cytoplasm occupied by peroxisomes (cited in 

Bentley et al., 1993). However, owing to the large interindividual variation 

in peroxisome morphometrics observed in these studies, together with cell 

to cell variations and lobular variations, it is difficult to attach any clear 

biological significance to these findings (Bentley et al., 1993). Generally, 

peroxisome proliferators have not been reported to produce any significant 

effects on marker enzyme activities and/or peroxisomes in cultured primate 

and human hepatocytes (Bieri, 1993; Bentley et al., 1993; Foxworthy and 

Eacho, 1994). 

Some studies have also examined species differences in effects on cell 

replication. Both nafenopin and Wy-14,643 have been reported to 

stimulate replicative DNA synthesis in rat, but not in Syrian hamster, 

hepatocytes (Price et al., 1992; Lake et al., 1993). Although peroxisome 

proliferators can stimulate DNA synthesis in cultured rat hepatocytes, 

methylclofenapate was reported to be ineffective in guinea pig, marmoset 

and human hepatocytes (Elcombe and Styles, 1989). Similarly, nafenopin 

has also been reported not to induce replicative DNA synthesis in human 

hepatocytes (Parzefall et al., 1991). 

Risk assessment of rodent liver peroxisome proliferators 

The key issues concerning the risk assessment of rodent liver peroxisome 

proliferators include: 

(a) Genotoxicity. 

(b) Likely human exposure. 

(c) Compound potency and no effect levels. 

(d) Precise mechanism(s) of liver tumour formation. 

(e) Species differences in response. 


Generally, peroxisome proliferators are considered to be non-genotoxic 

agents (Bentley et al., 1993; Budroe and Williams, 1993) and hence should 

be assessed differently from genotoxic carcinogens (Weisburger, 1994). 

Human exposure to rodent peroxisome proliferators depends on the


intended usage of the particular compound. While hypolipidaemic agents 

are only administered to a restricted population of humans, exposure to 

industrial chemicals such as plasticisers is obviously far more widespread. 

For example, based on food surveillance surveys the daily human exposure 

to DEHA was reported to be 16 and 8.2 mg per person per day in 1987 

and 1990, respectively (MAFF 1987, 1990). In another study, where 

DEHA intake was assessed by measuring urinary levels of the major 

metabolite 2-ethylhexanoic acid, a median value of 2.7 mg per person per 

day was reported (Loftus et al., 1994). 

Apart from likely human exposure, consideration should be made of the 

relative potency of the particular compound to produce peroxisome 

proliferation and liver tumours in rodents. Plasticisers such as DEHP and 

DEHA are far less potent than certain therapeutic agents and 

experimentally used compounds (Reddy et al., 1986; Barber et al., 1987; 

Bentley et al., 1993; Lake and Lewis, 1993). Moreover rodent liver 

peroxisome proliferators exhibit clear no effect levels for both peroxisome 

proliferation and for tumour formation. For example, in the rat no effect 

levels for liver tumour formation have been observed in studies with 

several compounds including bezafibrate, clofibrate, DEHA and DEHP 

(Hartig et al., 1982; NTP 1982a, b). In addition, the threshold for tumour 

formation in rodents is appreciably greater than the threshold for 

peroxisome proliferation (Hartig et al., 1982; Reddy et al., 1986; Bentley 

et al., 1993). 

Several mechanisms have been proposed to account for why peroxisome 

proliferators produce tumours in rodent liver. If these various hypotheses 

are combined then a role for increased cell replication in peroxisome 

proliferatorinduced hepatocarcinogenesis may be readily identified. For 

example, if hepatocytes are transformed by either oxidative stress-induced 

damage or by alternative mechanisms, such initiated cells may be promoted 

to liver tumours by enhanced cell replication. Certainly peroxisome 

proliferators are effective promoters of certain populations of initiated cells 

and recent studies suggest that peroxisome proliferators can influence rates 

of both cell replication and cell death in particular populations of 

hepatocytes (Grasl-Kraupp et al., 1993; Popp and Cattley, 1993; Marsman 

and Popp, 1994). 

With respect to species differences, rats and mice are clearly responsive 

species, whereas the majority of both in vivo and in vitro studies suggest 

that primates including man are either essentially refractory or certainly 

much less responsive to rodent peroxisome proliferators. However while 

effects on peroxisome morphology and marker enzyme activities have been 

extensively studied, few investigations have examined species differences in 

peroxisome proliferator-induced cell replication and liver tumour 

formation. As enhanced cell replication appears to play a role in peroxisome 

proliferator-induced hepatocarcinogenesis in rats and mice, it would

appear to be an important biomarker for assessing species differences in 

response. Rodent peroxisome proliferators do not appear to stimulate 

replicative DNA synthesis in vivo in Syrian hamster hepatocytes and in 

vitro in human hepatocytes (Elcombe and Styles 1989; Parzefall et al., 

1991; Price et al., 1992; Lake et al., 1993). With respect to tumour 

formation, nafenopin and Wy-14,643 (two potent peroxisome 

proliferators) were reported not to produce liver lesions in the Syrian 

hamster although both compounds produced liver nodules and 

hepatocellular carcinoma after 60 weeks in the rat (Lake et al., 1993). 

Similarly, clofibrate was reported not to increase liver weight or produce 

liver tumours in marmosets after 6.5 years treatment (Tucker and Orton, 

1993) and in an ongoing study ciprofibrate was found not to produce any 

morphological changes in marmoset liver after 3 years administration 

(Graham et al., 1994). 

In conclusion, the present literature suggests that rodent peroxisome 

proliferators are non-genotoxic agents which should be assessed differently 

from genotoxic compounds for human hazard (Weisburger, 1994). 

Assessment of likely human exposure and compound potency are also 

important factors together with information on compound no effect levels 

and evidence of species differences in response. Rodent liver peroxisome 

proliferators as a class of chemicals thus do not appear to pose any serious 

hazard for man. However, it would be desirable to elucidate further the 

mechanism(s) of peroxisome proliferator-induced hepatocarcinogenesis in 

susceptible species (i.e. the rat and mouse). From such studies the most 

appropriate biomarkers of liver tumour formation could be identified and 

examined in studies of species differences possibly including in vitro studies 

with human hepatocytes. Finally, further carcinogenicity studies in partially 

responsive (e.g. Syrian hamster) and non-responsive (e.g. guinea pig) 

species would strengthen the conclusion that peroxisome proliferators do 

not constitute any significant hazard to man. 

Acknowledgement 

We thank the UK Ministry of Agriculture, Fisheries and Food for financial 

support of BIBRA studies on hepatic peroxisome proliferation. 

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18 

Neurotoxicity Testing of Industrial Compounds: 

in vivo Markers and Mechanisms of Action 

KORNELIS J.VAN DEN BERG, 1 JAN-BERT 

P.GRAMSBERGEN, 2 ELISABETH M.G.HOOGENDIJK, 1 

JAN H.C.M.LAMMERS, 1 WILLEM S.SLOOT 1 and 

BEVERLY M.KULIG 1 

1 TNO Nutrition and Food Research Institute, Rijswijk; 2 

Erasmus University, Rotterdam 

Introduction 

Neurotoxicity assessment is designed to provide an answer to the question 

of whether or not a particular chemical is able to evoke some form of 

adverse effect specifically associated with the nervous system. The 

development of risk assessment procedures is a long-term goal in view of 

the complexity of the target organ involved, e.g. the nervous system. As 

risk identification is a first step in the risk assessment paradigm of 

chemicals (NAS, 1983), for neuro-toxicity this translates at present into 

procedures and methods aimed at characterization of altered 

neurobehaviour of exposed experimental animals and abnormalities in the 

morphology of the nervous system. It has been suggested that risk 

assessment procedures may take more the approach of ‘exposuredoseresponse’ 

(Andersen, 1991) in which mechanisms play an important role at 

various levels, e.g. from disposition of the chemical through the body to 

target tissues, via biochemical interactions at the molecular level to a toxic 

response as altered behaviour. Understanding the mechanisms involved in 

this sequence of events is helpful to arrive in the future at quantitative risk 

assessment procedures, for instance biologically-based modelling (Borghoff 

et al., 1991). In addition, a better knowledge of the mechanistic principles 

of neuro-toxic agents may lead to the development of specific biomarkers 

that would further improve the efficiency of procedures for neurotoxicity 

screening, given the magnitude of the number of potentially neurotoxic 

compounds (NRC, 1992). 

Only for a small select group of industrial chemicals are mechanisms of 

neurotoxic action more or less characterized. Perhaps the oldest examples 

are the class of organophosphate (OP) pesticides where the molecular 

targets have been identified as acetylcholinesterase (AChE) and neuropathy 

target esterase (NTE), the latter being associated with organophosphate-

induced delayed neuropathy (OPIDN) (Cherniack, 1988). This has led to 

investigations into structure-activity relationships of OPs with NTE in vitro 

that have proved to be valuable in predicting OPIDN (Davis et al. 1985). 

Furthermore, certain tissue culture systems, such as neuroblastoma cell 

lines, contain NTE and AChE activity that may be suited for identification 

of OPs causing OPIDN (Veronesi, 1992). Biochemical assays for NTE and 

AChE, in conjunction with neurobehavioural observations and 

neuropathology, are currently incorporated into US and Japanese 

neurotoxicity testing guidelines. 

A second group of industrial chemicals for which indications exist on 

their mechanism of action include certain compounds causing peripheral 

neuro-pathy such as acrylamide, carbon disulphide and n-hexane. It is 

generally assumed that the primary action of these chemicals is the 

crosslinking of axonal proteins (Graham et al., 1982), a process that blocks 

axonal transport (Sickles, 1991) and may lead to degeneration of distal 

axons (Spencer and Schaumberg, 1980). The basic information thus 

collected is yet to be developed into a useful biomarker. 

The pyrethroid insecticides represent a third group of chemicals for 

which the neurotoxic mechanism of action has been elucidated. The major 

symptoms of pyrethroid intoxication, e.g. convulsions, tremors, paralysis, 

are primarily the result of interaction of the pyrethroids with sodiumchannels 

on nerve membranes (Lund and Narahashi, 1982). Whereas 

under normal circumstances a sodium channel is only opened during a 

depolarization event, pyrethroids prolong opening of these channels 

thereby causing repetitive excitation of nerve and nerve terminals. In tissue 

culture systems, e.g. neuroblastoma cell lines, direct electrophysiological 

studies on cells have been done to investigate the effectiveness of different 

pyrethroids for opening of sodium channels (Oortgiesen et al., 1989). 

Biomarkers of neurotoxicity 

It is clear that a mechanistic understanding of the neurotoxicity of 

suspected chemicals is and will be proceeding at a slow pace. Eventual 

utilization of this knowledge in the form of biological markers for 

neurotoxicity risk assessment procedures is, necessarily, a long-term 

objective. In the mean time alternative approaches have been proposed 

recently that (a) may provide biomarkers that could aid to further define 

underlying mechanisms and (b) are directly linked to neurotoxic 

mechanisms of actions. 

Gliotypic and neurotypic proteins 

K.J.VAN DEN BERG ET AL. 239 

Insults to the brain by a large variety of agents or conditions, e.g. viral 

infection, auto-immune encephalitis, trauma and chemicals, evoke a fairly

240 NEUROTOXICITY TESTING OF INDUSTRIAL COMPOUNDS 

stereo typic response of astrocytes (Eng, 1988). The normally rather 

quiescent astrocytes change into reaction astrocytes with a characteristic 

morphological appearance, a process known as reactive gliosis or 

astrogliosis. It has been proposed that this fairly universal astrocytic 

reaction might be a useful parameter for risk identification purposes 

(O’Callaghan, 1992; Rosengren and Haglid, 1989). The hypertrophic 

response of astrocytes is biochemically characterized by a greatly enhanced 

synthesis of glial fibrillary acidic protein (GFAP), a major structural protein 

of intermediate filaments (Eng, 1988). Under a variety of experimentallyinduced 

insults to the central nervous system, increased GFAP 

immunoreactivity has consistently been found in association with neuronal 

damage (Eng, 1988; O’Callaghan and Miller, 1989). It must be kept in 

mind, however, that reactive astrogliosis represents an indirect indication 

of nerve cell damage or loss. A strategy has been proposed (Brock and 

O’Callaghan, 1987) to collect quantitative data on astrogliosis, for 

example for GFAP concentrations, in association with information on 

changes of neuron-specific proteins. e.g. synapsin I or synaptophysin. The 

general idea is that enhanced GFAP levels in combination with decreased 

synaptophysin concentrations in the brain would be strongly indicative of a 

neurotoxic event. Recent methodological developments have led to 

quantitative procedures for assessment of GFAP and synaptophysin in 

nerve tissues by dot-blot immunoassay and ELISA (Jahn et al. 1984; 

O’Callaghan, 1991). 

Cerebral calcium accumulation 

Nerve cells, like most animal cells, possess a complicated system 

comprising calcium gates, pumps and channels to maintain free cytosolic 

calcium ion levels within a low physiological range (Pounds, 1990). Under 

pathological conditions, including hypoxia-ischaemia and status 

epilepticus, calcium overload in vulnerable neurons has been observed that 

is thought to be associated with the process of cell death (both necrosis and 

apoptosis) (Boobis et al. 1989; Siesjo and Bengtsson, 1989). Cerebral 

calcium accumulation has, therefore, been proposed as a potential index of 

brain pathology (Korf et al., 1986). 

Free radical formation 

Free radicals may be involved in mechanisms of toxicity, including neurotoxicity 

of chemicals and are gaining more attention as is obvious from a 

number of recent reviews on this topic (LeBel and Bondy, 1991; Halliwell 

et al. 1992; Aust et al., 1993). Basically, free radicals are molecules that 

contain one or more unpaired electrons, whereas most molecules are 

nonradicals. Because of the unpaired electron(s) free radicals are highly

eactive chemical intermediates. Once a free radical is formed it can initiate 

a chain of reactions as the free electron is passed from one molecule to the 

other. In vivo, free radicals involving oxygen species are continuously 

produced and evolution has also provided the body with defence 

mechanisms such as superoxide dismutase (SOD), catalase, and scavengers 

such as glutathione, ascorbate, etc. 

Oxidative stress by free radicals may be an important neuropathological 

mediator following exposure to a number of neurotoxic agents (LeBel and 

Bondy, 1991). Examples of neurotoxic chemicals suspected of causing free 

radicals include chlordecone, ethanol, methamphetamine, methyl mercury, 

toluene, triethyl lead and trimethyltin. Of special interest is a ‘designer’ 

drug called MPTP 1 causing destruction of dopaminergic neurons of the 

basal ganglia with symptoms similar to Parkinson’s disease. It is the 

neurotoxic metabolite MPP +2 that has been found to induce cerebral 

oxygen radical formation in vitro. There is also suggestive evidence to 

indicate that free radical scavengers may provide protection of the basal 

ganglia against neuro-toxic effects of MPP + (LeBel and Bondy, 1991). The 

occurrence of these kinds of compounds has led, among other things, one 

to suspect possible involvement of environmental chemicals and factors 

associated with diet and/or lifestyle in Parkinson’s disease (Russell, 1992; 

Semchuk et al., 1993). 

Model neurotoxins 

In order to validate an approach based on changes in these proposed 

biomarkers experimental studies were performed using various model 

neuro-toxicants, including trimethyltin, kainic acid, heavy metals such as 

lead, methylmercury and manganese, and developmental neurotoxicants, 

e.g. polychlorinated biphenyls. 

Trimethyltin 


Trimethyltin (TMT) is known to cause in adult rats a rather selective 

neuronal degeneration in specific regions of the brain, notably in limbic 

structures such as the hippocampus where extensive loss of pyramidal cells 

in CA fields are observed by standard histopathological procedures 

(O’Callaghan, 1988). A single systemic dose of TMT (7.5 mg kg −1 ) given to 

adult rats caused, after a period of 3 weeks, an approximately three-fold 

enhanced level of GFAP in hippocampus (Figure 18.1, upper left panel). In 

a number of other brain regions, e.g. different parts of the cortex, 

thalamus, striatum, cerebellum and brain stem, no significant changes in 

GFAP were observed. Assessment of synaptophysin, a structural protein of 

synaptic vesicles of neurons (Jahn et al. 1985) in the same brain regions is 

given in Figure 18.1 (lower left panel). TMT induced a significant


Figure 18.1 Changes in cerebral glial fibrillary acidic protein (GFAP) and 

synaptophysin concentrations by trimethyltin (TMT) and kainic acid. Results are 

expressed as mean±SEM. Open bars, control; closed bars, 7.5 mg trimethyltin 

(TMT) per kg or 12 mg kainic acid per kg; HP, hippopcampus; AM, amygdala; *, 

p

Kainic acid 


Kainic acid (KA), a glutamate agonist, is a potent model neurotoxin that 

causes neurodegeneration in a number of limbic structures including 

hippocampus, amygdala, piriform cortex (Sperk et al. 1983; Ben-Ari, 

1985). It is generally assumed that the mechanism of action involves 

interaction of KA with a special class of glutamate receptors and release of 

endogenous excitatory amino acids in levels that are detrimental to 

neurons (Meldrum and Garthwaite, 1990). 

Kainic acid did induce, in a single systemic dose (12 mg kg −1 ), highly 

increased concentrations of GFAP in a number of target structures. For 

instance in hippocampus and amygdala GFAP levels did rise to 650 and 

960 per cent of control levels (Figure 18.1, upper right panel). Recent 

results from this laboratory have revealed that in a time-course study (Van 

den Berg and Gramsbergen, 1993) maximum levels are obtained after 4 

weeks that remained highly elevated in most target brain regions for at 

least a period of 6 months (Figure 18.2). The quantitative data on GFAP 

concentrations were supported by increased GFAP immunoreactivity in 

hippocampal sections visualized by immunohistochemical procedures. In 

addition to the damage in the hippocampus, permanently enhanced GFAP 

levels were found in other brain regions, e.g. piriform cortex, septum 

(Gramsbergen and Van den Berg, 1994) known to be targets of KA 

neurotoxicity. 

Synaptophysin levels were significantly reduced by KA in hippocampus 

and amygdala to a comparable degree (Figure 18.1, lower right panel). 

These data indicate a decrease in synaptophysin content encountered in 

brain regions where neuronal elements are known to be lost by these 

model neurotoxins. The magnitude of the changes in synaptophysin 

concentrations were much smaller than those of GFAP in the same brain 

structures. This may be explained by the fairly selective neuronal loss in 

specific layers of, for instance, the hippocampus, by TMT and KA. The 

effect is thus rather diluted in a biochemical procedure. Once an effect is 

scored, more detailed biochemical analysis is possible by using a punch 

technique after microdissection of brain nuclei (Palkovits and Brownstein, 

1988). Alternatively, a follow-up by histopathological procedures, for 

example by the cupric silver degeneration stain, would provide further 

details of neuronal damage (O’Callaghan and Jensen, 1992). In the 

experiments described above there was no clear correlation between the 

graded regional GFAP response and decrease of synaptophysin 

concentration. This may suggest a differential region-specific response of 

astrocytes towards neuronal injury. 

In our laboratory cerebral calcium accumulation was determined 

recently after a single systemic dose of KA (12 mg kg −1 , i.p.), given to adult 

rats. A rapid uptake of 45 Ca was observed in various regions of the brain. A


Figure 18.2 Time-course of GFAP concentration in hippocampus by kainic acid. Rats 

were treated with a single systemic dose of kainic acid (12 mg kg −1 , i.p.). GFAP was 

determined by ELISA at the times indicated. Results are expressed as mean±SEM 

(from Van den Berg and Gramsbergen, 1993, with permission). 

time-course study, covering a period of 6 months, has indicated that in the 

hippocampus a peak of 45 Ca uptake occurred already after 4 days and 

normal values were reached only after another 2–4 months (Van den Berg 

and Gramsbergen, 1993). More recent results indicate that other limbic 

areas display similar kinetics, while 45 Ca uptake in striatum and cortex 

return faster to normal values, e.g. within 2 weeks. Only in the thalamus a 

long-term sustained 45 Ca uptake was present for a period of 6 months. In 

general a good correlation was found between the regions showing 

sustained 45 Ca accumulation and those known to be targets of KA 

neurotoxicity. When these data on 45 Ca accumulation were related to 

effects of KA on GFAP concentrations, also a good agreement was 

observed concerning the brain regions involved (Gramsbergen and Van den 

Berg, 1994). Histopathological examination of hippocampal sections 

revealed extensive neurodegeneration by KA in CA1, CA3 and CA4 

regions (Van den Berg and Gramsbergen, 1993). 

Heavy metals 

Lead may produce in children symptoms of acute encephalopathy after 

exposure to high doses and at lower dose levels learning disorders and


hyperactive behaviour. In adults neurotoxicity is more often encountered 

as peripheral neuropathy after chronic occupational exposure to lead 

(Marsh, 1985). A number of mechanisms have been implicated in lead 

neurotoxicity (Bressler and Goldstein, 1991), but as yet no central 

hypothesis has emerged. 

In order to investigate the effect of lead on CNS structural proteins, a 

subchronic dosing experiment with adult rats was performed in which 

animals received daily doses of lead acetate (4, 8, 12.5 mg kg −1 i.p.) for 28 

days. GFAP concentrations were subsequently determined in different brain 

regions. Already at the lowest dose level GFAP levels were found to be 

significantly increased in several brain regions, notably in different parts of 

the cortex, hippocampus and striatum, while cerebellum and brain stem 

remained unaffected. Neurobehavioural assessment of animals, preceding 

the neurochemical analysis, also revealed significant alterations of 

neuromuscular function, excitability and spontaneous activity. 

Methylmercury has caused a number of poisonings in man (Marsh, 1985) 

where it appears to affect in particular both the central and peripheral 

nervous system. In an animal experiment, adult rats were subchronically 

dosed with methylmercury (0.75 or 2 mg kg −1 ) for 28 days. 

Neurobehavioural assessment indicated that grip strength was significantly 

impaired. Neurochemical analysis of GFAP in the central nervous system 

was performed in selected brain regions and, in addition, in various 

segments of the spinal cord. Increased GFAP levels were observed in the 

cerebrum only in the frontal cortex, also in brain stem and in spinal cord. A 

further detailed analysis of brain stem sub-structures showed significantly 

enhanced GFAP levels in pons and medulla oblongata but not in midbrain. 

In spinal cord GFAP concentration was increased in specific 

sections, e.g. in the cervical and lumbar segments but not in the thoracic 

segment. 

The results with these particular examples of heavy metals have indicated 

the unsuspected presence of regions in the central nervous system with 

astrogliosis, as determined in a biochemical GFAP assay. The neuronal 

damage involved has not yet been confirmed independently, e.g. by 

assessment of synaptophysin. The possibilities remain, therefore, that 

reactive astrocytosis by these heavy metals may be indirectly a result of 

breaching the integrity of the blood-brain barrier (Bressler and Goldstein, 

1991) or of a direct toxic action on astroglial cells (Selvin Testa et al., 

1990; Stark et al., 1992). 

Manganese is a well recognized industrial neurotoxin associated with 

neurologic effects after prolonged exposure in occupational settings (Katz, 

1985). The clinical manifestations of manganism bear a large similarity to 

those of Parkinson’s disease (PD). The neurodegenerative disease PD is 

characterized by a selective loss of neurons in the basal ganglia.


Experimental studies were done to investigate the vulnerability of the 

basal ganglia in manganism. Manganese (as Mn 2+ ) was applied 

intrastriatally to rats and region-specific brain damage was assessed by 

determining regional 45 Ca accumulation using quantitative 

autoradiographic procedures developed in this laboratory (Gramsbergen 

and Van der Sluijs-Gelling, 1993). It appeared that Mn 2+ induced a timedependent 

45 Ca accumulation in most of the regions constituting the basal 

ganglia, e.g. striatum and several other structures such as globus pallidus, 

entopeduncular nucleus, several thalamic nuclei and substantia nigra (Sloot 

et al., 1994). 

As monoaminergic neurotransmission plays a predominant role in the 

basal ganglia, recent studies have indicated that concentrations of biogenic 

amines such as dopamine and its major metabolites, serotonin and 

noradrenaline were also reduced in striatum by Mn 2+ . The kinetics of this 

process indicated that concentrations of most monoamines and metabolites 

were temporarily reduced except for dopamine and metabolites in striatum 

that remained at a permanently reduced level (>90 days) (Sloot et al., 

1994). In order to demonstrate the specificity of the manganese effects, 

several other compounds were studied, including ferrous ions (Fe 2+ ). An 

equimolar dose of Fe 2+ , applied intrastriatally, produced, however, a much 

more extensive and widespread 45 Ca accumulation throughout the basal 

ganglia and, in addition, in nucleus accumbens and cerebral cortex. 

Ferrous ions were also three times more potent than manganese ions in 

causing depletion of dopamine in striatum (Sloot et al., 1994). 

The results based on both 45 Ca accumulation and biogenic amine levels 

are in concordance with the hypothesis that the basal ganglia, which are 

enriched in iron and iron-binding proteins, represent a selective target for 

manganese. The role and fate of endogenous iron in the brain, and the 

basal ganglia in particular, under toxic conditions including chronic 

manganese exposure, merits further investigations. 

Oxidative stress by free radical formation may play a role in these toxic 

events. Transition metals are known as strong promoters of reactive 

oxygen species. Especially iron, as the ferrous ion (Fe 2+ ), has been found to 

react with hydrogen peroxide to form the hydroxyl radical in the so-called 

Fenton reaction (Halliwell et al., 1992; Aust et al., 1993). It is thought that 

this mechanism plays an important direct role in iron poisoning (Aust et 

al., 1993). In an indirect way oxidative stress by iron may be initiated by 

chemicals that are able to ‘liberate’ iron from stores such as ferritin, 

transferrin, haemoglobin, etc. Recently, evidence has been obtained that a 

number of chemicals, including the pesticides paraquat and diquat, may 

release iron from ferritin in vivo as well as in vitro (Aust et al., 1993), 

involving organic radical and superoxide formation. 

These observations are particularly relevant for the interpretation of the 

observed neurotoxic effects of iron described above. Dopamine is relatively

easily subjected to a process of auto-oxidation. The decrease in dopamine 

levels by iron may have been caused by oxygen radicals. Circumstantial 

evidence suggests that a similar mechanism of action may underly 

manganese neurotoxicity. The decrease in dopamine in the basal ganglia by 

manganese is also thought to occur through catalysis of dopamine 

oxidation (LeBel and Bondy, 1991), possibly involving radical oxygen 

species. An open question is whether manganese might participate in the 

‘iron release’ hypothesis (Sloot and Gramsbergen, 1994). Current efforts 

are being made to determine free radical formation by iron and manganese 

and to relate this to dopamine depletion. For this purpose rats are given 

salicylic acid (SA) and subsequently the SA hydroxylation products are 

measured in cerebrospinal fluid and brain tissues as indices of hydroxyl 

radical formation. 

Developmental neurotoxins (PCBs) 


Concern has been raised about the long term consequences of low level 

intake of polychlorinated biphenyls (PCBs) with respect to neurotoxicity as 

it relates to nervous tissue development and intellectual performance in the 

juvenile and adult stages. Several epidemiological studies with infants have 

shown a negative correlation between PCB levels in cord blood and 

cognitive functions and a positive correlation between PCB levels and 

altered neurological parameters such as hypotonia and hyporeflexia (Rogan 

et al., 1988; Jacobson et al., 1990). Experimental studies in various species 

including primates have also provided arguments for neurotoxic effects in 

offspring after perinatal exposure to PCBs (Tilson et al., 1990). In this 

laboratory evidence has recently been obtained to indicate dramatic 

reduction in sexual behaviour and reproduction in offspring that was 

perinatally exposed to PCBs (Smits-van Proojie et al., 1993). 

In order to investigate eventual structural alterations in the CNS, 

pregnant Wistar WU rats were exposed to Aroclor 1254 on days 10–16 of 

gestation. At a young age (3 weeks) and adult age (3 months) offspring 

were sacrificed and various brain regions were dissected for assessment of 

gliotypic and neurotypic proteins. In untreated control animals 

developmental aspects of astrocytes in the central nervous system were 

encountered. Both in hypothalamus and cerebellum of control animals 

GFAP levels were increased by 200–300 per cent between 3 weeks and 3 

months postnatally. A developmental GFAP increase was also found in 

brain stem, striatum and lateral olfactory tract, although to a lesser extent. 

In hippocampus and prefrontal cortex, GFAP levels remained virtually 

unchanged between 3 weeks and 3 months. These results, therefore, 

indicate that glial cell maturation and/or differentiation is not uniformly 

distributed over the whole brain. It appeared that glial cells at birth in rats 

were fully developed in brain regions dealing with cognitive functions, e.g.


cortex and hippocampus, while in regions such as hypothalamus and 

cerebellum this process occurred entirely neonatally. It should be kept in 

mind that in the rat also neuronal maturation and differentiation are not 

completed at birth and continue to a large extent postnatally. 

Exposure of pregnant rats to Aroclor 1254 did lead to a number of 

alterations in GFAP levels in various brain regions in offspring as compared 

to control animals. The most striking differences were observed in brain 

stem. While in control animals of both sexes GFAP levels increased 

postnatally, the normal developmental increase in perinatally exposed 

animals was absent (Figure 18.3, upper panel). At 3 months the relative 

deficit in GFAP levels was 41 per cent for male and 30 per cent for female 

progeny. Already at the lowest dose of Aroclor 1254 (5 mg/kg) a maximum 

decrease was observed (Figure 18.3, upper panel). In addition to brain stem, 

a similar effect on GFAP levels was found in striatum although the relative 

deficit at 3 months was somewhat less. Quite an opposite pattern emerged 

in brain regions such as cerebellum, lateral olfactory tract and prefrontal 

cortex, where GFAP levels were increased relative to unexposed progeny. 

In hippocampus no significant changes in GFAP levels were encountered. 

The question arose whether the observed neurodevelopmental toxicity in 

rats by PCBs was specific for astroglial cells or also involved neuronal 

maturation, differentiation and death. For this purpose the same nervous 

tissues were used for quantitative assessment of a neuronal marker in the 

form of synaptophysin. In brains of untreated adult control rats, regionspecific 

differences were observed in synaptophysin concentrations, being 

high in the prefrontal cortex, striatum and hippocampus and relatively low 

in lateral olfactory tract, cerebellum and brain stem. Synaptogenesis for the 

brain as a whole largely takes place in the rat from birth until postnatal 

day 70 (Knaus et al., 1986). Regional differences in the speed of postnatal 

synaptogenesis from 3 weeks to 3 months were found in a number of 

structures being most pronounced in cerebellum (190 per cent increase) and 

prefrontal cortex (170 per cent increase) while in brain stem little change was 

observed. 

As a result of perinatal exposure to Aroclor 1254, altered expression of 

synaptophysin was observed. In most brain structures examined, including 

brainstem (Figure 18.3, lower panel) significant decreases in synaptophysin 

concentrations were found. This suggests that during development of the 

central nervous system, PCBs may interfere with the formation of synaptic 

vesicles, synaptogenesis, or formation of nerve terminals. 

A straightforward interpretation of the present results is not possible at 

this stage. However, what is clear is that perinatal exposure to PCBs may 

cause changes in the structural composition of the central nervous system 

both in the neuronal and the glial cell compartment. Apparently there are 

different effects on nerve cells of the CNS depending on the brain region 

involved. The brain stem and striatum are regions with decreased


Figure 18.3 Developmental effects on glial fibrillary acidic protein (GFAP) and 

synaptophysin concentrations in brain stem after perinatal exposure to Aroclor 

1254. Results are expressed as mean±SEM with the values of the control animals at 

day 21 set at 100 per cent. Open bars, control; shaded bars, 5 mg Aroclor 1254 per 

kg; closed bars, 25 mg Aroclor 1254 per kg; D 21 and D 90, postnatal days; *, p


Table 18.1 Cerebral calcium accumulation 

a Systemic dose. 

b Intrastriatal dose. 

synaptophysin and increased GFAP in structures such as cerebellum, lateral 

olfactory tract, and prefrontal cortex. Such a situation could arise when 

neurons in these latter regions fail to receive proper input from developing 

neurons originating in brain stem and striatum. Further investigations into 

this interesting topic may provide further clues for the developmental 

toxicity of PCBs and related compounds that possibly could aid in the 

interpretation of neurobehavioural effects, for instance altered sexual 

behaviour and reproduction (Smits-van Prooije et al., 1993). 

The findings observed in cerebellum are consistent with a phase of 

hypothyroidism during development but it is clear that a conclusive role 

for thyroid hormone remains to be further established. The present results 

furthermore suggest that alterations in synaptophysin/GFAP levels may be 

useful and sensitive parameters to study compounds suspected of 

developmental neurotoxicity. 

Conclusion 

The results with various neurotoxins demonstrate that assessment of 

gliotypic proteins such as GFAP may be a useful tool to identify and 

quantify persistent toxic insults of the CNS, especially when this is backed 

up by indications for loss of neuronal elements, e.g. decreased 

synaptophysin concentration. Also in circumstances of developmental 

neurotoxicity, caused by chemicals such as PCBs, an approach based on 

changes of gliotypic and neurotypic proteins may provide a promising 

biomarker. Of course, further investigations with various other compounds 

are required to substantiate these interesting findings. 

Cerebral calcium accumulation may be useful as an early indicator of 

neurotoxicity on a more prospective basis as is suggested by various 

examples of neurotoxic compounds (summarized in Table 18.1). Because in 

unlesioned brain regions the background levels of calcium uptake remain

very low, areas where neurotoxic events take place are easily identified and 

quantified by autoradiographic or scintillation counting procedures. 

Finally, the role of oxidative stress by free radical formation in the 

nervous system merits further studies since it may open up possibilities for 

providing a biomarker that is linked to a mechanism of neurotoxic action. 


The author is grateful for unpublished information provided by Dr Didema 

de Groot on neuropathology, by Dennis C.Morse, MSc and the students 

Wendelien Wesseling and Annemiek Plug, Agricultural University 

Wageningen, on developmental effects of PCBs; and for expert technical 

assistance by Alita van der Sluijs-Gelling. 

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19 

Endocrine Toxicology of the Thyroid for 

Industrial Compounds 

CHRISTOPHER K.ATTERWILL and SAMUEL 

P.AYLWARD 

CellTox Centre, University of Hertfordshire, Hatfield, Herts 

General introduction 

Classification of endocrine toxicity 

Xenobiotic-induced endocrine dysfunction and toxicity is a common 

finding in safety studies and an increasingly important consideration in the 

riskassessment process. The effective identification of potential endocrine 

toxicological effects depends upon xenobiotic effect classification in 

relation to normal endocrine function and pathology. 

Classifications of different types of endocrine toxicity have been 

proposed in previous publications on this subject. Capen and Martin 

(1989) proposed a detailed classification based on clinical endocrine 

function and pathology. On the other hand, Baylis and Tunbridge (1985) 

proposed a simpler classification based on the adverse endocrine reactions 

of xenobiotics which are observed clinically. From a toxicological or 

preclinical safety testing point of view, Atterwill and Flack (1993) favoured 

classifying endocrine toxicology of xenobiotics in a manner which is 

similar in concept to that for classifying other toxicological phenomena 

(CIOMS, 1983) taking into account the unique nature of the endocrine 

system. This is as follows (see also Figure 19.1): 

Class 1 Effects which can be predicted from the endocrine pharmacology 

of compounds. An example would be the oestrogens and progestogens 

which have a plethora of effects on metabolic parameters in addition to their 

actions on oestrogen sensitive target sites when administered at 

pharmacological doses (the therapeutic dose levels). Further examples are 

certain dopamine and 5-HT antagonists acting on hypothalamic and 

pituitary receptors which may disrupt ‘downstream’ endocrine functions. 

Class 2 Effects which again can be predicted from the endocrine 

pharmacology of the compound when administered at doses well in excess 

of the therapeutic dose level. An example would be adrenal steroid 

suppression and general excessive catabolism observed with high dose and

256 ENDOCRINE TOXICOLOGY OF THE THYROID 

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

prolonged use of glucocorticoids. In addition, the use of thyromimetic agents 

which may suppress pituitary thyrotroph function. 

Class 3 Effects which could not have been predicted from the 

pharmacology of the compound. This group can be subclassified into: (a) 

Effects which are direct or primary actions on an endocrine gland. 

Examples of this might be the action of ketoconazoles on adrenal and 

testicular function and the action of alloxan and streptozocin on the β-cell 

of the pancreas; and (b) effects on endocrine glands which are indirect or 

secondary to changes in other organs or control mechanisms 

(homeostasis). Examples here would be the actions of phenobarbitone and 

PCBs on the rat thyroid and the effects of lactose and polyols on the rat 

adrenal medulla. 

Class 4 Effects which cannot be predicted from pre-clinical studies 

because of idiosyncratic effects on the endocrine system. 

As indicated by Baylis and Tunbridge (1985) the adverse effects of 

pharmaceutical agents on the endocrine system are generally due to normal

C.K.ATTERWILL AND S.P.AYLWARD 257 

or exaggerated pharmacological responses, that is Classes 1 and 2. 

Furthermore, it appears that endocrine toxicology can be detected reliably 

in pre-clinical studies. What is essential is that Class 3 toxic endocrine effects 

are classified appropriately. Class 3a effects would be expected to be seen 

across species, whereas there are numerous examples where Class 3b 

effects appear to be species-specific. There seems to be a paucity of clear-cut 

examples for Class 4 effects. There are several examples of compounds 

which have idiosyncratic immunotoxicological effects in susceptible humans 

which cannot be predicted from pre-clinical studies. 

Incidence of thyroid toxicity and tumours of the thyroid 

Most information on incidence is derived from pharmaceutical toxicity 

databases. Such data from Ribelin (1984) suggest that the endocrine system 

of the rats is particularly sensitive to toxicity from xenobiotics. This is also 

supported by Heywood (1984) in which he examined the target organ 

toxicity for 42 pharmaceutical compounds in the rat and dog. The 

endocrine system of the rat was only second to the liver as the most 

frequently affected target organ (38 per cent liver, 31 per cent endocrine). 

Ribelin (1984) reported that the most frequent endocrine lesion occurs in 

the adrenals followed by the testes but this analysis was conducted on 

chemicals and pharmaceuticals, and the data indicated that it was the 

cortical layers of the adrenal that were being predominantly effected, 

suggesting that the adrenal changes may be reflecting general stress 

responses rather than direct adrenal gland toxicity. 

In another analysis conducted in conjunction with the Centre of 

Medicines Research this area was further explored. Toxicology data on 

124 compounds (all pharmaceuticals) were analysed. Just under 50 per 

cent (61/124) of these compounds have effects on one or more endocrine 

glands. Similar to Ribelin (1984) the adrenals were the most frequently 

affected, followed by the testes and the thyroid. An extensive survey of the 

different types of thyroid toxicity for both pharmaceutical agents and 

industrial chemicals was presented by Atterwill et al. (1993). 

Perturbation of thyroid function 

Thyroid function can be perturbed by agents affecting a number of 

processes involved in the regulation of the hypothalamic-pituitary-thyroidliver 

(H-P-T-L) axis (Figure 19.2). These agents can affect function directly 

by interacting with thyroid cell receptors on their intracellular transduction 

mechanisms (see Figure 19.3). Alternatively thyroid function may be 

altered indirectly by agents affecting thyroid hormone metabolism and/or 

distribution—this event being followed by the release of thyrotrophic 

factors, or by xenobiotic-mediated alterations in the release of these factors


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

themselves from the hypothalamus or pituitary gland (see also review by 

Cavalieri and Pitt-Rivers (1981) and Atterwill et al. (1993)). For example, 

thyromimetics such as L-T 3 or D-T 3 can cause pituitary atrophy and 

thyroid ‘shutdown’ and toxicity due to suppressive effects on the 

thyrotrophs in the adenohypophysis (Atterwill, 1988).

Figure 19.3 Thyroid follicular epithelial cell function. 


Control of thyroid function and pathobiology of thyroid 

lesions 

Xenobiotic toxic effects on the hypothalamic-pituitarythyroid-liver 

(H-P-T-L) axis 

The control of mammalian thyroid follicular function is shown 

schematically in Figure 19.2 together with the points at which agents may 

perturb function and cause toxicity and thyroid lesions. The most common 

classes of agent affecting function are discussed at the different loci in 

terms of the categories (1–4) of endocrine toxicity (Atterwill and Flack, 

1993). 

The various control mechanisms and factors influencing hormone 

synthesis, distribution and metabolism can be summarised as follows: 

thyroid hormone (T 3 and T 4) synthesis and secretion from thyroid gland 

are controlled by thyroid stimulating hormone (TSH) released from the 

pituitary gland. This in turn is under control by hypothalamic thyrotrophin 

releasing hormone (TRH) and circulating levels of the thyroid hormones. 

Thyroid hormones exist in the circulation in the free (free T 4 (FT 4) and 

free T 3 (FT 3)) and protein-bound forms (approx 99 per cent of total T 4- 

TT 4 and total T 3-TT 3) and it is the FT 3 hormone produced by deiodination 

from T 4 which has both a physiological action on T 3 nuclear receptors in 

target tissues and influences pituitary TSH output.


Protein-binding of the hormone in the circulation takes place on several 

moieties, the profiles of which vary between different species (Dohler et al., 

1973). The major proteins are thyroglobulin (TBG), thyroid binding 

pre albumin (TBPA) and albumin. Free and bound T 3 and T 4 are in 

dynamic equilibrium in the circulation and because of differences in affinity 

for the binding proteins there is more T 3 in the unbound form (approx 0.4 

per cent of total) than T 4 (approx 0.04 per cent of total). 

Thyroid hormone synthesis takes place at the apical membrane of the 

polarised follicular epithelial cells and depends on TSH-stimulated active 

iodide uptake under the influence of many extracellular and intracellular 

factors and signals (see Figure 19.3). Iodide is oxidised by peroxidase 

enzymes which mediate the incorporation of iodine into the tyrosyl 

residues of the colloidal glycoprotein, thyroglobulin. Colloid-bound 

thyroid hormone is stored in the follicular lumen and is released into the 

circulation by lysosomal action on the colloidal complex following 

endocytosis of colloid droplets into the cells. The incorporation and 

organification of iodide into thyroid hormone within thyroid follicles and 

the toxicological effects of xenobiotics can be studied in vivo using the 

‘perchlorate-discharge test’ (Atterwill et al., 1987) or in vitro using cultured 

thyrocytes (see Figure 19.4) from different species (Atterwill and Fowler, 

1990). 

Catabolic metabolism of the hormonal products of the thyroid gland, FT 3 

and FT 4, is achieved via two major pathways, deiodination and 

conjugation (glucuronidation and sulphation yielding a more water-soluble 

product for biliary excretion) and one minor route, deamination 

(decarboxylation, see Figure 19.9 later). 

The metabolic fate of thyroxine relies predominantly on its deiodination 

to T 3, only 20 per cent of circulating T 3 (the thyromimetic) is secreted by 

the thyroid (Engler and Burger, 1984). The remaining 80 per cent is derived 

from the deiodinative conversion of thyroxine (FT 4) to T 3. Deiodination 

can occur at several sites—the ones of major importance from the clearance 

aspect being liver and kidney whilst pituitary deiodination is essential for 

controlling responsiveness to circulating FT 4 levels. The deiodinases exist 

as three iso-zymes: Type I (5′-D; localised in liver, kidney, thyroid and 

central nervous system (CNS) tissue; and is propylthiouracil (PTU) 

sensitive); Type II (5'-D; localised exclusively in the CNS, brown adipose 

tissue, and pituitary; PTU-insensitive); and Type III (5-D, CNS; PTU 

insensitive) and have different affinities for T 4, different maturational 

patterns and different compensatory responses to hypothyroidism (see 

Kohrle et al., 1987). 

Bastomsky (1973), using Gunn rats, congenitally jaundiced due to UDPglucuronyl 

transferase deficiency (including T 4-glucuronyltransferase) 

demonstrated that the rate limiting step in hepatic thyroxine clearance (via


the biliary excretion pathway) is the formation of the glucuronic acid 

conjugate, T 4-glucuronide. 

Hepatic conjugation, either sulphation (preferring FT 3) or 

glucuronidation (preferring FT 4) yields a more water soluble product, 

excreted in the bile/ biliary duct is a major pathway in T 4 excretion. 

Further deiodination via the deiodinase group of enzymes of T 4 T 3conjugates, 

T 4-amines and T 3 to thyromimetically inactive iodothyronines 

(rT 3 Triace, Tetrace, T 2 and T 1; see Figures 19.8 and 19.9) plays an 

important part in the T 4/T 3 biotransformation cascade and completes the 

thyroid hormone metabolic profile. 

The key factor in maintaining correct thyroid follicular capability is an 

appropriate TSH output to TRH stimulation alongside circulating levels of 

FT 4. Perturbation of this homeostatic control results in a classical thyroid 

response. The initial thyroid responses to increasing TSH levels are 

follicular cell hypertrophy, loss of colloid and vascular dilatation. In 

conventional animal toxicology studies performed for regulatory 

authorities one of the first indices of thyrotoxicity, therefore, is the 

observation of altered thyroid histopathology, primarily as follicular cell 

hypertrophy and/or diffuse hyperplasia, often leading to focal hyperplasia, 

thyroid adenomas and adenocarcinomas in longer term toxicity studies 

after longer term exposure. 

Pathobiology of thyroid follicular cell hyperplasia and 

neoplasia 

Thyroid neoplasia (see Figure 19.5) develops predictably in experimental 

species exposed to any procedure inducing prolonged and excessive TSH 

secretion (for example, the administration of chemical goitrogens, chronic 

iodine deficiency or subtotal thyroidectomy) although humans and mouse 

appear to be more resistant to TSH-induced thyroid neoplasia than rat. 

The number of cytogenetic abnormalities within the thyroid epithelium 

increases with duration of excess TSH exposure, with follicular cell 

hyperplasia potentially leading to neoplasia. The histopathological 

sequence of events is as follows (see Zbinden, 1987): following 

hypertrophy of the follicular epithelial cells focal hyperplasias appear in the 

gland which are distinct areas of papillary growth with enlarged epithelia. 

As these foci continue to grow they form nodules partly surrounded by 

collagenous fibres. These lesions are transition states between focal 

hyperplasias and adenomas. Adenomas are larger nodules that compress 

the surrounding tissue and have a distinct capsule. Follicular 

microcarcinomas (characterised by irregular gland-like structures, 

basophilia and nuclear crowding) may appear in some nodules. Larger 

carcinomas usually retain a follicular structure but sometimes consist of 

solid sheets of polymorphous cells (Zbinden, 1987).


Figure 19.4 Cultured porcine thyrocytes in vitro. These scanning electron 

micrographs show (a) individual cultured ‘inverted’ follicles, and (b) individual 

follicular epithelial cells at higher magnification (apical membrane facing upwards) 

displaying TSH-stimulated microvilli.


Figure 19.5 Factors influencing the development of thyroid neoplasia. 

Two consecutive processes are thought to occur in the development of 

thyroid tumours; first, initiation which occurs quickly and is irreversible 

and secondly, promotion which occurs slowly and is reversible (and for 

which cell proliferation may be a necessary but not sufficient condition). 

Initiators may be ionising radiation, chemical/biological agents or genetic 

factors, with TSH acting as a promotion agent. Spontaneous thyroid 

follicular cell tumours arise from unknown aetiologies and factors and ageinduced 

changes in the cell membrane, growth factor and signaltransduction 

mechanisms may be involved. Small subpopulations of hyperreactive 

epithelial cells retaining the high replication rate of the foetal stage 

have been identified (Peter et al., 1982; Smeds et al., 1987) which may 

enter clonal expansion following only slight elevations in TSH (such as 

those in handled or stressed animals) or other growth factors, leading to 

spontaneous nodular goitres (see Zbinden, 1987). 

Many experimental studies have confirmed the key role of TSH as a 

stimulator of thyroid growth. In a rat thyroid cell line (FRTL-5) TSHstimulated 

growth of the cells was found to be associated with a marked 

increase in c-fos and c-myc oncogene expression (Colletta et al., 1986). 

Another example of the tumour promoting capacity of TSH is given by


studies where rats were given carcinogens such as N-methyl-N-nitrosourea 

(MNU) and then phenobarbital, or put on an iodine deficient diet. These 

treatments cause an early and increased incidence of thyroid follicular 

lesions and tumour formation (Hiasa et al., 1982; Oshima and Ward, 1984). 

The duration of exposure to high circulating TSH concentrations is also 

important in that intermittent administration of chemical goitrogens with 

TSH ‘normalisation’ does not appear to lead to follicular neoplasias. 

An elaborate series of studies have shown that a sustained elevation of 

serum TSH in the rat leads to three phases of thyroid growth (Figure 19.6): 

(1) a phase of rapid growth lasting 1–2 months, followed by (2) a plateau 

phase of 3–6 months (growth desensitising mechanism (GDM) limiting 

epithelial cell mitotic response), followed eventually (3) by the appearance 

of multiple follicular cell tumours (loss of GDM; see Wynford-Thomas et 

al., 1982; Stringer et al., 1985; Smith et al., 1986). The reversibility of TSHinduced 

thyroid focal hyperplasia will evidently depend, therefore, on the 

stage during these ‘timed’ cellular changes in the first 6 months at which 

the TSH stimulus is withdrawn. Once the GDM is non-operative 

reversibility is not possible. 

Tumour progression seems to occur by a multi-stage process involving 

clonal ‘expansion’ and naturally occuring clones of cells have been 

demonstrated with high intrinsic proliferation potential in the mouse 

thyroid gland (Smeds et al., 1987), perhaps helping to explain the focal 

nature of hyperplastic and neoplastic lesions. The loss of a GDM within 

the follicular cells appears to be accompanied by an altered dependence or 

sensitivity to certain growth factors as well as the possible loss of an antioncogene 

which limits the follicular cells’s growth response to TSH. For 

example, the growth of normal cultured human thyroid cells requires TSH 

and insulin-like growth factor 1 (IGF1) in combination, whereas cells from 

adenomatous tissue in vitro proliferate in response to either TSH or IGF 

independently (Williams et al., 1987). This is due to the acquisition of 

autocrine production of IGF 1 by the tumour cells themselves (see Thomas 

and Williams, 1991). Since the differentiation and growth of thyrocytes 

under TSH is regulated by cyclic-AMP-dependent mechanisms 

(Figure 19.2), tissue hyperplasia and hyperthyroidism might be expected to 

result when activation of the adenyl cyclase-cAMP cascade becomes 

unregulated. This can occur, for example, when somatic mutations impair 

the GTPase activity of G-protein coupled reactors, which may thus behave 

as proto-oncogenes. Such a mechanism is probably responsible for the 

development of a minority of monoclonal hyperfunctioning thyroid 

adenomas (Parma et al., 1993) (these also result in a silencing of normal 

thyroid function in extra-adenomatous tissue). Other non-genotoxic 

factors, such as agents affecting patterns of DNA methylation when 

coupled with a growth stimulus, should also be given consideration when


Figure 19.6 Pathobiology of thyroid tumorigenesis.


attempting to define mechanisms in thyroid carcinogenesis (Thomas and 

Williams, 1992). 

In summary, there is, therefore, good evidence that sustained TSH drive 

to the thyroid gland can lead to a de-regulation of thyroid function. When 

investigating xenobiotic or drug-induced thyroid tumour formation, the 

mechanisms whereby TSH drive is increased can be understood by 

undertaking a series of experimental studies using in vitro and in vivo 

techniques. Having delineated the mechanism of the thyrotoxic effect it 

may then be possible to determine whether a particular drug or compound 

elicits a similar response in different species (including humans) and to 

investigate the dose-response relationship for this effect. 

Investigative toxicological studies and examples of 

xenobiotics causing thyroid toxicity via the H-P-T-L axis 

Introduction 

Atterwill et al., (1993) give extensive examples of both pharmaceutical and 

industrial compounds causing thyroid toxicity via the five main sites along 

the H-P-T-L axis as shown in Figure 19.7 and readers should refer to this 

for further and more detailed information. In this chapter, three of these 

five thyroid toxicity loci are described in relation to the endocrine effects 

produced, industrial xenobiotic examples, and investigative in vivo and in 

vitro tests to delineate mechanisms and species-specific effects. This 

information is further summarised in Figure 19.8. 

In terms of industrial compounds the most frequently cited examples 

causing thyroid toxicity appear to be in the categories of: (i) those 

potentially affecting the plasma protein binding of thyroid hormones—for 

example, the nitrile herbicide, ioxynil (Ogilvie and Ramsden, 1988); (ii) 

those acting directly on the thyroidal peroxidase enzyme as goitrogens, and 

blocking thyroid hormone synthesis and secretion—for example, the coal 

derived hydroxyphenol products (Lindsay et al., 1992); and (iii) those 

affecting the hepatic metabolism and elimination T 3 and T 4—for example, 

compounds such as β-naphthoflavone, PCBs and alachlor (Ogilvie and 

Ramsden, 1988). Tables 19.1–19.3 show examples of these three class 

effects, compounds producing the effects and some of the range of 

investigative tests currently available. 

in vivo and in vitro studies of xenobiotics acting on the 

hepatic metabolism and clearance of thyroxine 

There is a growing list of agents, both pharmaceutical and industrial 

xenobiotics, which act in rodents by interfering with thyroid hormone

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


metabolism, hepatic elimination and thus circulating TSH levels (see also 

Capen and Martin, 1989; McClain, 1989; Atterwill et al., 1993). The 

xenobiotics include phenobarbital (McClain, 1989), β-naphthoflavone 

(Johnson et al., 1993), the polychlorinated biphenyls (Bastomsky, 1974), 

diproteverine (a calcium antagonist; Flack et al., 1989), SC37211 (a Searle 

imidazole antimicrobial (Comer et al., 1985), L649923 (a leukotriene D 4 

antagonist; Saunders et al., 1988), a novel oxyacetamide-FOE 5043 

(Christenson et al., 1993), alachlor (Brewster et al., 1993), PCNB 

(pentachloronitrobenzene; Story et al., 1993), and hexachlorobenzene 

(Ogilvie and Ramsden, 1988). 

Most of these compounds have thus far been assumed to act in vivo via 

the induction of hepatic uridine diphosphate glucuronosyltransferase (UDP- 

GT) in the rat, with species-specific formation of thyroid tumours in 

carcinogenicity studies (see McClain, 1989). Indeed many of the 

compounds, including phenobarbital do lead to increased hepatic UDP-GT 

activity and appearance of glucuronidated T 4 in the bile, sometimes with 

elevated bile flow rates (see McClain, 1989). However, others such as the


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

pharmaceutical temelastine increase predominantly the clearance of free 

T 4, though the bile product is not in conjugate form (Poole et al., 1989, 

1990). Other compounds such as the food dye FD&C Red No 3 (Capen 

and Martin, 1989) are able to lower circulating triiodothyronine (T 3) by 

altered deiodination suggesting the further existence of alternative 

mechanisms. Furthermore, not all chemicals inducing hepatic neoplasia in 

rodents cause thyroid neoplasia (McClain, 1989). 

We and others have reported that two SK&F histamine antagonists, 

temelastine (SK&F 93944) and lupitidine (SK&F 93479) produce ratspecific 

thyroid lesions via perturbation of the hepatic locus (Atterwill et

Table 19.1 

Table 19.2 


al., 1989). Increased thyroxine clearance from the circulation, followed by 

elevated TSH ‘drive’ and increased thyroid follicular cell growth were 

observed. These com pounds act rapidly within minutes—hours of in vivo 

drug administration and are apparently able to increase the accumulation of 

T 4 directly in vitro by cultured rat hepatocytes (Atterwill et al., 1989; Poole 

et al., 1990). Phenobarbital appears to share this property in vitro


Table 19.3 

(Aylward et al., 1994) and increases the accumulation of thyroxine in 

treated rat liver in vivo (Oppenheimer et al., 1968). 

Thyroxine transport (Figure 19.9) is regulated by specific components 

located within the plasma membrane in various cell types including 

fibroblasts and hepatocytes and is an important prerequisite for both 

hormone metabo lism and nuclear hormonally-mediated events (Pliam and 

Goldfine, 1977; Krenning et al., 1981; Blondeau, 1986). Investigations 

indicate that there exist two distinct transport systems specific to 

thyroxine: a high-affinity, low capacity, energy-dependent ATP-ase linked 

transport system and a low-affinity, high capacity transport mechanism 

(Sorimachi and Robbins, 1978; Krenning et al., 1981, 1983; Blondeau, 

1986; Rao, 1991). 

Many of the compounds listed as indirect carcinogens in rat (due to an 

ability to induce liver microsomal enzymes and increase glucuronidated 

thyroxine elimination in the bile) have been shown to increase hepatic UDP- 

GT activity or cause liver hypertrophy indicative of following repeated 

dosing (Comer et al., 1985; McClain, 1989; Johnson et al., 1993). 

However, there have been no attempts to provide a definitive link between 

UDP-GT induction and thyroid pathology or to prove a primary 

endocrinological effect via UDP-GT. Our previous work with temelastine 

in the rat in vivo (Atterwill et al., 1989) was able to demonstrate that the 

increased clearance of T 4 from the rat circulation appeared within a few 

hours of a single compound dosing (Atterwill et al., 1989). Even

Figure 19.9 Hepatic events leading to hormone elimination. 


phenobarbital is able to increase the thyroxine clearance in a relatively 

short timespan (Atterwill, unpublished observations). New in vitro data are 

not inconsistent with the time course of the in vivo phenomena where 

enhanced T 4 hepatocytic accumulation by cultured rat hepatocytes 

following compound exposure occurred as early as 60–90 min after 

exposure (Aylward et al., 1994). There was no membrane cytotoxic effect 

of the compounds at the threshold concentrations producing these effects in 

vitro. This shows a potential rapid direct effect of the xenobiotics on


Table 19.4 Species and energy dependence of enhanced thyroxine accumulation in 

vitro 

Key: ↑T 4, increase; ↓T 4, decrease; ↔T 4, no change; Temperature? ATP?, 

temperature/ATP dependent; NA, not applicable. 

hepatocellular T 4 accumulation. The correlation between in vivo—in vitro 

species-specific toxicological effects is also evident (Figure 19.10 and 

Table 19.4). One of the features of temelastineinduced thyroid toxicology 

in vivo was the apparent species-specificity to the rat (Atterwill et al., 

1989; Poole et al., 1989). Temelastine-mediated thyroid follicular 

hypertrophy and hyperplasia was not observed in dog, mouse or monkey 

following temelastine treatment (Figure 19.10). In vitro, no enhanced 

thyroxine accumulation in response to temelastine or phenobarbital was 

observed in guinea pig or dog hepatocytes (Aylward et al., 1994). In 

support of these findings, it has been demonstrated that the guinea pig is 

insensitive to thyroid pathological changes after phenobarbital or βnaphthoflavone 

administration in vivo (Johnson et al., 1993; Wyatt et al., 

1993). In support of, and as an extension of these findings, we now present 

important new findings to demonstrate conclusively that some of the rapid 

‘effectors’ of thyroid toxicity via the liver, such as temelastine, do so 

independently of a primary action on UDP-GT, whereas other cytochrome 

P450 inducers such as phenobarbital may have a combined effect. This 

work was carried out using hepatocytes prepared from UDP-GT system 

deficient Gunn rats. 

Studies on Gunn rat hepatocytes in vitro 

Hepatocytes were prepared from the normal or Gunn rat (deficient in UDP- 

GT isozymes conjugating thyroxine) and exposed to either temelastine or 

phenobarbital (2 or 20 µM) for 3 h as before (Aylward et al., 1994). The 

results show (Figure 19.11) that whereas temelastine was able to enhance 

thyroxine accumulation in both types of hepatocytes, phenobarbital only 

produced alterations in hormone accumulation in normal cells, supporting


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

(1989)).


Figure 19.11 Effect of temelastine and phenobarbital on thyroxine accumulation in 

vitro by hepatocytes from control and Gunn rats. 

earlier in vivo findings where phenobarbital was toxicologically inactive in 

this strain of rat (Bastomsky, 1973).


Conclusions 

The observations now lend further and strong support to the hypothesis 

that indirect xenobiotic-induced thyroid toxicology can arise from direct 

effects on hepatic membrane-located thyroxine transport proteins. It also 

suggests that the species-specificity of this toxic effect in vivo of some 

xenobiotics may be attributed to actual species differences in the sensitivity 

of these hepatic carriers to the compounds and not simply or primarily to 

changes in T 4 glucuronidation via UDP-GT induction. For the first time we 

have demonstrated the usefulness of Gunn rat hepatocytes in vitro for 

discriminating between the two ‘hepatic subclasses’ of xenobiotics causing 

thyroid toxicity in rodents. 

A number of practical in vivo and in vitro investigative tests are now 

available for delineating mechanisms of thyroid toxicity along the H-P-T-L 

axis, and which also provide screening tools for examining chemical series 

of potentially toxic molecules: (i) Direct block of thyroid function via 

peroxidase inhibition can be measured in vivo by the perchlorate discharge 

test (Atterwill et al., 1987); (ii) it can also be measured in vitro using 

cultured thyrocytes (Atterwill and Fowler, 1990); (iii) indirect effects on 

hepatic thyroxine clearance can be assessed in vivo (Atterwill et al., 1989); 

or (iv) in vitro using cultured hepatocytes from different species or Gunn 

rat (Aylward et al., 1994). Effects on receptors at the hypothalamic and 

pituitary levels can also now be studied extensively using both in vivo and 

in vitro approaches (Buckingham and Gillies, 1993). This battery of 

technology now available will greatly advance the mechanistic 

understanding and screening of thyroid endocrine toxicants. 

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WILLIAMS, D.W., WYNFORD-THOMAS, D. and WILLIAMS, E.D., 1987, 

Human thyroid adenomas show escape from IGF-1 dependence for growth, 

Ann. Endocrinol., 48(2), 11. 

WYATT, I., GYTE, A., FOSTER, J.R., WILLIAMS, S.M. and ELCOMBE, C.R., 

1993, Differences in the response of the rat and guinea pig liver and thyroid to 

hepatic enzyme inducers, Hum. Exp. Toxicol., 12(6), 561. 

WYNFORD-THOMAS, D., STRINGER, B.M. J. and WILLIAMS, E.D., 1982, 

Dissociation of growth and function in the rat thyroid during prolonged 

goitrogen administration, Acta Endocrinol., 101, 210–16. 

ZBINDEN, G. 1987, Assessment of hyperplastic and neoplastic lesions of the thyroid 

gland, TIPS (Dec., 1987), 8, 511–14.

20 

Testing and Evaluation for Reproductive Toxicity 

ANTHONY K.PALMER 

Huntingdon Research Centre, Huntingdon 

Introduction 

Most of the presentations at this meeting refer to high level, scientific 

investigations of one or two, highly important, high production volume 

chemicals, for which an adverse effect has been demonstrated. They are 

studies of characterisation, because they elaborate on known effects using a 

wealth of available information. But, how were the adverse effects of these 

few substances first discovered, what were the initial clues? Sadly, for 

many, the observation of adverse effects in humans was the trigger to 

intensive investigations, which is akin to ‘shutting the stable door after the 

horse has bolted’. 

This presentation is concerned with detecting effects of substances for 

which little or no information is available and, preferably, before they 

cause harm to humans. This requires a different kind of science, for which 

the main asset is the ability to predict, with reasonable accuracy, possible 

activity from minimal information. It requires wide experience and a 

balance between imagination and pragmatism. These attributes are 

especially important for toxicity to reproduction, which triggers instinctive 

reactions in even the coolest and most objective scientist. 

Identifying the cause of adverse effects on human reproduction has long 

been surrounded by controversy and uncertainty. In respect of the 

evaluation of substances for reproductive toxicity this state of affairs seems 

likely to persist for years to come. The main obstacle to any attempt to 

rationalise the situation is that any discussion on evaluation almost 

inevitably gravitates to the black hole of regulatory guidelines. All 

guidelines are flawed because science fact is compromised by bureaucracy 

and science fiction. For many reasons, but especially the unwillingness of 

any establishment to change the status quo, guidelines provide the worst 

starting point for developing a strategy for evaluation. 

Most guidelines are concerned only with methods for gathering specified 

information. On its own this information (on hazard) is insufficient and 

needs to be supplemented with other information, from other sources, to

Table 20.1 Production volume triggers for industrial chemicals 

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

A.K.PALMER 281 

Notes: 

Tests for ecotoxicity are not included. 

Progression, by rote, from base set to level 2 would involve duplication of effort in 

several areas. 

The state of confusion regarding testing for toxicity to reproduction is portrayed by 

the failure to decide on requirements at base set level and the curious mixture of old 

and new terminology regarding tests. 

predict whether humans might be affected. Most guidelines are a watershed 

in a broader spectrum of testing and assessment. They represent a point at 

which it may be decided that the only way to gather more information is to 

take the final step of exposing humans. Exposure of humans provides the

282 TESTING AND EVALUATION FOR REPRODUCTIVE TOXICITY 

Figure 20.1 Overlap of toxicity. 

only certain way to determine whether reproduction would be affected, but 

we need to do the best we can before taking the chance. 

The paradox in this is that exposure of humans is facilitated by failure to 

demonstrate toxicity in animals. But, lack of activity, being negative, cannot 

be proven, only presumed. To make this presumption investigations must 

be extensive and comprehensive to convey reasonable assurance that failure 

was not due to deficiencies in methodology. 

There is an exception. For industrial chemicals, testing of all substances 

to these criteria would be a monumental task, therefore, less stringent 

testing is allowed according to production volume, which serves as an 

approximation for the extent of the population likely to be exposed 

(Table 20.1). It is a strategy based on population risk, the downside of 

which is an increased risk to the individual. The strategy takes advantage 

of the fact that, in a small population, the chance of identifying cause and 

effect is poor. At the base set level and level 1, as outlined by the EC, all 

tests are equivalent only to the voluntary preliminary studies conducted for

medicines, agrochemicals and food additives. They may be sufficient to 

detect potent toxicity but not comprehensive enough to allow presumption 

of the absence of hazard, especially as the base set does not include tests 

for reproductive toxicity (Table 20.2). 

As production volume increases, more extensive testing should be 

undertaken. Often, this has been neglected, prompting the development of 

the OECD guidelines 421 and 422. These tests were intended to recover a 

situation that never should have arisen. 

A better approach? 

There is no question that evaluation for reproductive toxicity could be 

improved considerably. The question is whether industry and agencies are 

willing to do so. It would require a change of attitude in industry and 

agencies alike. Industry’s ‘passive avoidance’ of testing would need to be 

replaced by ‘active participation’. For a new substance the first step should 

be an integrated assessment of commercial prospects and potential toxicity 

over a broad spectrum (Figure 20.1). Early identification of ‘serious bad 

actors’, which tend to effect many systems, can save time and effort. Given 

the prognosis of problems ahead, it may be better to devote resources to 

finding safer alternatives, or to risk management, rather than to endless 

testing. For materials with a high commercial potential, the aim should be 

to get to full scale tests by the quickest route. Following the EC levels by rote 

is very inefficient since there is duplication with successive steps. 

Methods 

With an active participation policy, a much broader scope of methodology 

can and should be considered, ranging from searches for structure-activity 

relationships, through various in vitro methods, whole animal tests, wild 

life surveys and human surveys (Table 20.3). Due to time constraints I will 

concentrate on whole animal test systems. 

Structure-activity databases 

Structure and activity relationships are an obvious place to start any 

evaluation, despite the fact that currently available databases are far from 

perfect. Their reliability could be improved dramatically by adding unused 

information currently hidden in industry and agency archives. 

Entire mammalian tests 


Tests in entire mammals provide the only way of assessing what effects a 

substance may evoke in the complex, integrated and dynamic process of


Table 20.3 Methods for detecting effects on human reproduction 

reproduction. To detect the wide range of possible effects it is necessary to 

expose mammals to a substance from conception through sexual maturity. 

It is necessary to look for consequences of this exposure through at least 

one life cycle (Figure 20.2). This long observation period is required for 

detection of latent manifestations of developmental toxicity, such as those 

induced by lead, alcohol, diethylstilboestrol and other hormonally active 

substances. The only means of covering all these aspects is a two generation 

study or the equivalent in a combination of tests. 

Restricted test systems 

The use of lesser tests for industrial chemicals is a concession. Examination 

for some effects is omitted because they are not perceived to be important 

or because they would be difficult to detect, or because they occur very 

rarely. For example, first detection of effects in offspring of second 

generations is rare so such activity does not have a high priority. 

With these restricted tests, emphasis should be on detecting effects and 

not on manipulating a no effect level. Tests that could be considered would 

include OECD 421 and 422, the OECD single generation study and the old 

FDA Segment I study for medicines. The latter two are restricted tests 

because they do not allow detection of latent manifestations of 

developmental toxicity. 

The best return for effort is afforded by OECD 422, which combines 

examination for general or systemic toxicity, as well as reproductive 

toxicity. However, realising its potential requires an experienced laboratory 

team, the courage to modify the test and the conceptual ability to know 

how to interpret the results.

Figure 20.2 Cycle of life/reproduction. 


OECD 421 involves treatment of both sexes from about 2 weeks prior to 

mating through to termination, a few days after birth of offspring 

(Figure 20.3). Assessment of male fertility is achieved in two parts. Males 

are paired with females for detection of effects unrelated to 

spermatogenesis, for example, effects on sexual behaviour, libido or 

ejaculation and functional maturation of sperm. 

For detecting effects on spermatogenesis, direct methods, particularly 

histopathological examinations of testes and epididymides are used. Sperm 

analysis (seminology) could be added, although it does not seem to be 

better than histopathology. These methods could be incorporated into 

systemic toxicity studies rather than in the reproduction study per se. 

In respect of fecundity, treatment and observations of females include 

most of those applied in full scale tests. An exception is the lack of 

observations for delayed, post-natal manifestations. Detailed examination 

of foetuses for skeletal and soft tissue abnormalities is not included. The 

potential for prenatal effects is deduced by observation of post-natal 

differences in numbers pregnant, litter size, litter and mean pup weight at 

birth and to day 4 post partum. 

As with any guideline, OECD 421 should be used with commonsense 

and flexibility. If pretesting prognosis suggests that prenatal effects are 

unlikely, extension of the study to weaning of the offspring (Figure 20.3) 

provides added safeguards at little extra cost. Increase group size and it


Figure 20.3 OECD 421 priority selection test. 

Figure 20.4 Fertility and embryotoxicity. 

provides the equivalent of the OECD single generation study. Conversely, 

if pretesting prognosis suggests a high probability of prenatal effects, 

including induction of malformation, then females could be killed just 

before delivery and foetuses examined for structural defects (Figure 20.4). 

This provides the equivalent of a fertility and embryotoxicity study we will 

see again later. 

OECD guideline 422 simply adds to OECD 421, elements for assessment 

of systemic and neurotoxicity. For those who have never conducted such a 

test it seems impossibly complex, but it is neither as difficult to perform, 

nor to interpret, as is feared. Its rejection by EC Officialdom makes it an 

even better proposition, since there need be no inhibitions about modifying 

the design according to circumstances. 

Some brief examples of results that may be encountered with positive 

materials are illustrated by the examples of Carbendazim (metabolite of 

Benomyl), DEHP, Cyclophosphamide and ethylene glycol methyl ether 

(EGME, 2-methoxyethanol). With Carbendazim (Table 20.4) macroscopic 

and microscopic examinations show unequivocal effects on testes and 

epididymides indicating an effect on spermatogenesis. An effect on females 

and offspring is indicated by an increased duration of pregnancy, reduction 

in the number of females with live young and lower values for litter size,

Table 20.4 OECD 422: carbendazim, tabular summary 

Notes: 

a Malformations included hydrocephaly and misaligned tails. 

pp=post partum 

Bold type indicates treatment effects including macroscopic and microscopic 

changes in testes and epididymides (an effect on spermatogenesis), an increased 

duration of pregnancy, reduction in the number of females with live young and lower 

values for litter size, litter weight and mean pup weight. The dosage related pattern 

of response provides added emphasis, as does the observation of malformed 

foetuses. 


litter weight and mean pup weight. The dosage related pattern of response 

provides added emphasis, as does the observation of malformed foetuses. 

With DEHP (Table 20.5) an effect on spermatogenesis is evident at 2000 

mg kg −1 . Treatment at this dosage had to be withdrawn shortly after 

mating to avoid further mortalities of the more susceptible females. There 

is a marked reduction in the number of pregnancies. At lower dosages an 

increased duration of pregnancy, reduced litter size and litter weight, 

indicates an effect on the female and/or offspring. The higher mean pup 

weights are consequent to the longer duration of pregnancy. 

With cyclophosphamide (Table 20.6) female deaths at 3 and 4.5 mg kg −1 

are attributable to treatment as, at 6.7 mg kg −1 , all females died. There was 

a reduction in the number of females with young, litter size, litter weight 

and mean pup weight, providing clear evidence of effects on the female and


Table 20.5 OECD 422: DEHP, tabular summary 

Notes: 

[] Treatment at 2000 mg kg −1 was withdrawn after mating (4 weeks of treatment) 

due to loss of condition and mortality of females. 

pp=post partum. 

Bold type indicates treatment effects on spermatogenesis at 2000 mg kg −1 . There is 

a marked reduction in the number of pregnancies. At lower dosages an increased 

duration of pregnancy, reduced litter size and litter weight, indicates an effect on 

the female and or offspring. The higher mean pup weights are consequent to the 

longer duration of pregnancy. 

offspring. No effects on spermatogenesis were reported but, if a dosage 

inducing effects on the male had been selected, all the females would have 

died. 

With EGME (Table 20.7), dosages were based on acute toxicity and 

limited repeat dose toxicity studies only. This provided a more realistic 

representation of the testing of a new substance. The first consequence was 

the occurrence of systemic toxicity at 500 and 1000 mg kg −1 . Treatment at 

the high dosages had to be withdrawn prior to mating, investigating 

recovery became a new objective. At 100 mg kg −1 profound effects on 

spermatogenesis were evident. Some pregnancies were obtained, but no live 

young were born. For the high dosage groups, effects on males remained 

evident several weeks after withdrawal of treatment. The duration of 

pregnancy was increased. A consequence of this is the higher mean pup 

weight. Values for numbers of implantations, live young and litter weight 

were lower than control values. So, not only did the test show the 

anticipated effects on reproduction, it also demonstrated that recovery of

Table 20.6 OECD 422: cyclophosphamide, tabular summary 

Notes: 

pp=post partum. 

Bold type indicates treatment effect. Female deaths at 3 and 4.5 mg kg −1 are 

attributable to treatment as, at 6.7 mg kg −1 , all females died. There was a reduction 

in the number of females with young, litter size, litter weight and mean pup weight. 

No effects on spermatogenesis were reported but, if a dosage inducing effect on the 

male had been selected, all the females would have died. 

females was slow. As far as I know this has not been mentioned in the 

extensive literature on EGME. 

These examples and others show that the OECD tests are capable of 

detecting substances with marked effects on reproduction. With such 

results it would be foolhardy to consider higher level guideline tests. If the 

substance is not abandoned any further testing would require case by case 

designs to characterise the detected effects more completely. 

Full scale testing 

If pretesting prognosis suggests that a substance is unlikely to present 

problems of toxicity, fast track progression to level 2 testing should be 

considered (Table 20.2) to avoid unnecessary duplication of step by step 

testing. Expected tests include a two generation study in rats and an 

embryotoxicity study in rats and rabbits. In the harmonised guideline for 

medicines this same strategy is just one of several options and the same or 

greater flexibility should be made available for testing industrial 

chemicals. 

Tests for embryotoxicity 


An anomaly in this strategy is the specification for embryotoxicity studies 

in two species, when only one species is required for all other aspects of


Table 20.7 OECD 422: EGME, tabular summary 

Notes: 

[] Treatment at 500 and 1000 mg kg −1 withdrawn prior to mating due to loss of 

condition and mortality. Animals in withdrawal phase. 

NE not examined 

pp=post partum 

Bold type indicates treatment effect. At 100 mg kg −1 profound effects on 

spermatogenesis were evident. Some pregnancies were obtained, but no live young 

were born indicating effects on females and the conceptus. 

For the high dosage groups, effects on testes and epididymides remained evident 

several weeks after withdrawal of treatment. The duration of pregnancy was 

increased with a consequent increase in mean pup weight. Values for numbers of 

implantations, live young and litter weight were reduced indicating slow recovery. 

This does not appear to have been mentioned in the extensive literature on EGME. 

reproductive toxicity. A more sensible strategy would be to identify a 

relevant species before testing. It is pointless to conduct a test in an 

unsuitable species and doubly pointless to conduct tests in two irrelevant 

species. 

The requirement for detailed examination of foetuses for abnormalities 

is based on an exaggerated perception of risk prompted by fear. For many 

reasons the risk of inducing abnormalities is extremely low. Those same 

reasons make detection by direct observation of malformations unreliable.

Table 20.8 Post-natal detection of prenatal effects 


Notes: 

Prenatal effects include any soft tissue or skeletal changes (variants, anomalies or 

malformations) or altered foetal weight. Post-natal effects include reduced litter size 

at birth, reduced mean foetal weight or increased post natal mortality. 

In the few studies in which post-natal effects were not detected, prenatal changes 

were of a ‘minor’ nature (e.g. reduced ossification) and sometimes not conclusive. 

The lower number of litters reared in the Japanese Experiment 2 design would 

contribute to the slightly higher rate of ‘failures’. 

Where more serious prenatal effects such as the observation of malformations or 

prenatal death were observed a post-natal effect was always observed. 

The dimensions of the tests we conduct are too small and dosage regimes 

are contradictory to the basic principle that malformations are induced by 

application of a precise dosage at a precise time. Whilst direct observation 

of malformations is unreliable the saving grace is that induced 

malformations always occur within a wider spectrum of embryotoxicity. 

This wider spectrum provides a more reliable, if indirect, means of 

detecting substances that might cause malformations. Also, these effects are 

important in their own right. 

Major effects such as altered offspring weight and prenatal death can be 

observed postnatally. For example both EC Segment I and Japanese 

Experiment 2 studies include examinations for foetal abnormalities as well 

as postnatal observations. A survey of such studies shows that, when 

malformations were observed, a post-natal effect was always observed 

(Table 20.8). Surveys such as this should have been conducted or 

sponsored by agencies and industry before formulating guidelines. Why 

have they not done so? 

The obsession with abnormalities is based on the fear of another 

thalidomide. The great contradiction is that a considerable number of rat 

embryotoxicity studies with thalidomide failed to provided convincing 

demonstration of teratogenicity. Conversely, reproduction studies in rats 

provided unequivocal, post-natal evidence of effect in the form of a marked 

reduction in the number of females with young and a marked reduction in


Table 20.9 Thalidomide—rat reproduction studies 

Notes: 

Thalidomide was administered in the diet to provide a daily intake of 200 mg kg −1 

bodyweight from 60 days prior to mating (US FDA two litter test). Later studies 

demonstrated that the reduced percentage of females with young and the lower live 

litter size in females with young was associated with embryolethality. 

litter size of the few that had young. Two studies, each containing several 

matings showed the reproducibility of the results (Table 20.9). 

Two generation studies 

The emphasis on structural abnormalities detracts from examination for 

other, important and more likely manifestations of reproductive toxicity 

(Figure 20.5). For example, current and proposed test guidelines for nonmedicines 

lack procedures for detection of developmental neurotoxicity (or 

behaviour). This is a curious contradiction given the current fashion for 

investigation of adult neurotoxicity. 

For detecting other effects on reproduction, all regulatory versions of the 

two generation study leave something to be desired. Even more 

disappointing is that newer versions proposed by the US FDA, the US EPA, 

the EC and OECD have recycled many of the old flaws. Truly, there has 

been a great deal of activity but very little progress. 

All current and proposed guidelines continue to require a prolonged 

premating treatment period for the F0 or parent generation. The claim that 

it is necessary to treat males for a full spermatogenic cycle is pure science 

fiction. Spermatogenesis is not a cycle but a sequence of overlapping batch 

processes. At any one time, all stages of spermatogenesis are present and a 

short dosing period is sufficient to cause effects. 

To detect these effects, direct histopathological methods can be applied 

shortly after treatment. Direct methods are quicker and more certain than 

mating to females. Mating trials are inefficient and lack sensitivity due to

Figure 20.5 Manifestations of developmental toxicity 


the high sperm production capacity of animals compared with humans. 

Interim results from an ongoing survey show that, where data are 

available, direct methods are effective and that a combination of direct 

methods with a premating dosing time of 2 weeks or less is even more 

effective (Table 20.10). The combination compares favourably with 

prolonged premating treatment and mating. Why have agencies and 

industry failed to conduct or sponsor such surveys? 

For non-medicines, use of a prolonged premating treatment period for the 

parent generation is an unnecessary duplication as treatment is continued 

into the F1 generation; this cannot be mated until animals have reached 

sexual maturity. Using science facts, a more efficient design for a two 

generation study (Figure 20.6) would include the following features: 

– One control and 2–4 test groups with dosages set at 2–5 fold descending 

intervals from a high dosage. 

– The high dosage should be a limit dose (1000 mg kg −1 ) or one inducing 

a minimal systemic effect on adults. 

– A short 2–4 week premating treatment period for both sexes of the 

parent (F0) generation. 

– A greater group size for the F0 generation to allow balanced selection of 

the F1 generation.


Table 20.10 Detecting effects on males 

Notes: 

a Unknown=data not available or not examined 

A, organ weights, histopathology, serum chemistry. 

B, sperm analysis, count, motility, morphology. 

Data derived from an ongoing survey of 150 substances for which an effect on 

males has been claimed, mostly from human studies. To date 80 of the substances 

have been evaluated. Results indicate that use of a prolonged premating treatment 

period (>2 wks) has not been helpful for indicating effects on humans and is no 

better than use of a short premating period alone (


For example, a simple division of a two generation study will provide a 

developmental toxicity study in which pregnant females are treated from 

implantation through lactation or beyond and an F1 generation reared 

through to sexual maturity (Figure 20.8). The counterpart to this is a study 

of fertility in which both sexes are treated from about 2 weeks prior to 

Figure 20.6 A two generation study.


Figure 20.7 Reproductive toxicity—selecting studies. 

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

mating through to termination of males after a minimum of 4 weeks of 

treatment overall (Figure 20.9). Treatment of females continues to 

implantation and they may be killed and examined at about day 13–15 of 

pregnancy. 

Alternatively, treatment of females can be continued beyond closure of 

the palate or even through pregnancy (Figure 20.4). Foetuses can be

Figure 20.9 Fertility 

delivered and examined for abnormalities according to procedures used in 

embryotoxicity studies. (Note that this study is almost identical to a 

modified OECD 421 study.) Combination with the ICH developmental 

toxicity study provides the equivalent of a two generation and 

embryotoxicity study. 

Interpretation of studies 


For regulatory agencies, one of the purposes of these tests is to gather 

information for labelling. By common consensus, a substance is labelled as 

a reproductive toxicant only if it induces effects at dosages below those 

causing systemic toxicity. Such labelling, however, can be very misleading 

especially with industrial chemicals. Even if the animal species can be 

shown to be a good surrogate for humans by means of kinetic and other 

studies, it is necessary to take into account the relationship between 

exposures causing effects and those likely to be encountered by humans. 

For example, a material can be labelled as a reproductive toxicant but 

present little or no real risk to humans because the effects are induced at 

exposures well in excess of those encountered by humans (Figure 20.10). 

Conversely, a substance not labelled as a reproductive toxicant could cause 

reproductive effects in humans if these are induced at exposures 

encountered by humans. In other words toxicity is relative, a matter of 

dosage and situation. Labelling without reference to exposures is 

incomplete.


Figure 20.10 Interpretation/extrapolation of reproductive toxicology. 

Conclusions 

In conclusion I should mention that time constraints enforce superficial 

mention of many important aspects. I would like to emphasise that 

improved testing and evaluation for toxicity to reproduction can be 

achieved with methodology that exists within and without regulatory 

guidelines. This has been illustrated by the special cases presented at this 

meeting. 

There is no necessity to be restricted to specific guidelines, there never 

was. We should make use of any and all test methods available as 

appropriate for the substance being investigated. Whatever the type of 

substance, testing involves looking for the same hazards. Having identified 

a hazard, methods of assessing risk, essentially, are the same (although 

PBPK models for reproductive toxicity would be more complex than those 

used for systemic toxicity). 

The methodology is available, what is required is the willingness and 

wisdom to use it effectively and efficiently (Palmer, 1993a, b). The goal 

should be to investigate a specific substance to the extent necessary, no 

more and no less. For this there needs to be a change in attitude by 

industry, agencies and academia. Therein is the greatest problem since 

‘Change is not made without inconvenience, even from worse to better’ and 

people are very unwilling to change their prejudices and habits.

References 


PALMER, A.K., 1993a, Identifying environmental factors harmful to reproduction, 

Environm. Hlth Perspect. Supplements, 101(2), 19–25. 

PALMER, A.K., 1993b, Introduction to (pre)screening methods, Reproduct. 

Toxicol., 7, 95–8.

PART SIX 

Toxicity of selected classes of industrial 

chemicals

21 

Special Points in the Toxicity Assessment of 

Colorants (Dyes and Pigments) 

HERMANN M.BOLT 

Institut für Arbeitsphysiologie an der Universität Dortmund, 

Dortmund 

Introduction 

Colorants (dyes and pigments) are very important industrial chemicals. A 

special point in the toxicological assessment of such compounds is their 

bioavailability upon inhalation. From the technological point of view 

pigments are colorants which are insoluble whereas dyes are soluble in the 

application mixture. 

Biologically, the most relevant route of potential exposure of humans to 

colorants is by inhalation. If a pigment is biologically insoluble, it may 

finally be removed from the airways by clearance mechanisms. However, in 

practice the situation is much more complicated. For instance, chromates 

are technically important pigments which are well investigated. Biochemical 

and toxicological research has shown that the common toxicological 

principle of chromates which penetrate the cell membrane and, after 

intracellular transformation, exert genotoxic effects, is the chromate anion 

(CrO ). In terms of inhalatory carcinogenicity, the very water-soluble 

alkali chromates and the practically insoluble lead chromate have the 

lowest potency. Pigments of an intermediate solubility, e.g. calcium 

chromate, zinc chromate, strontium chromate, have a high carcinogenic 

potency on the respiratory tract. Local storage of chromate particles in the 

airways with a slow but continuous local release of CrO seems therefore 

to be an important factor in respiratory carcinogenesis induced by 

chromates. But also lead chromate, which is technically regarded as 

insoluble, is bioavailable to some extent; this is visualized by practical cases 

of occupational lead chromate exposure which display markedly elevated 

blood lead levels. 

The question of systemic bioavailability, upon inhalation, became of 

recent regulatory importance for azo colorants based on carcinogenic 

aromatic amines. This problem has already been addressed in detail 

(Myslak and Bolt, 1988).

302 SPECIAL POINTS IN THE TOXICITY ASSESSMENT OF COLORANTS 

Table 21.1 Number of azo colorants based on cancerogenic aromatic amines (2napthylamine, 

benzidine and its derivatives) listed in Colour Index (3rd edn, 3rd 

Rev., 1987) 

The problem of carcinogenic azo colorants 

In the past, azo colorants based on benzidene, 3,3′-dichlorobenzidine, 3,3′dimethylbenzidine 

(o-tolidine), and 3,3'-dimethoxybenzidine (odianisidine) 

have been synthesized in large amounts and numbers, 

especially in the German chemical industry. The Colour Index (1987) lists 

a total number of more than 2000 azodyes, 452 of them being based on 2naphthylamine, 

benzidine, or benzidine derivatives (Table 21.1). 

Azo colorants have a number of properties that have made them 

invaluable for dyeing a wide variety of materials, including natural and 

artificial fibres, plastics, resins, textiles, leather, paper, glass, ceramics, 

cement, inks, printing inks, chalks, crayons and carbon papers, as well as 

cosmetics, food and beverages. Interesting with respect to potential 

exposure of painters is the use of azo colorants in the coloring of oil-, 

resins-, emulsion-, lime-, and other aqueousbased paints, distempers, 

transparent laquers, spirit and oil wood stains, and varnishes (Colour 

Index, 1987). In all these fields, particularly benzidine-based azo colorants 

have found widespread use (Gregory, 1984). 

In the UK, the Carcinogenic Substances Regulation led in 1967 to 

discontinuation of the use of benzidine in the production of azo colorants 

(Martin and Kennelly, 1985). The US government in 1974 promulgated 

regulations to control benzidine at the workplace (Gregory, 1984). 

Nevertheless, in the period of 1972–4, more than 150000 persons in the 

USA were potentially occupationally exposed to benzidine-based colorants 

(Gregory, 1984); in 1978, approximately 1.7 million US pounds of 

benzidine-based azo colorants were manufactured, and a further 1.6 

million pounds were imported into the USA (Lynn et al., 1980). 

In Germany, over 30 different benzidine-based azo colorants were 

manufactured in the early 1960s. The manufacture of these colorants was 

stopped in 1971, with the exception of one dye (Direct Black 4; C.I. No. 

30245); the manufacture of the latter was continued until 1973. Azo

H.M.BOLT 303 

colorants based on carcinogenic congeners of benzidine (e.g. 3,3′dimethoxybenzidine; 

3,3′-dimethylbenzidine) are most likely still being 

manufactured in some countries. The case of pigments based on 3,3′dichlorobenzidine 

is discussed below. 

Azo colorants are biologically active through their metabolites. 

Azoreduction of these compounds occurs in vivo (Radomski and 

Mellinger, 1962; Rinde and Troll, 1975; Robens et al., 1980) by an 

enzyme-mediated reaction. Azoreductases are found in mammalian tissues, 

particularly in liver (Fouts et al., 1957; Walker, 1970; Martin and Kennelly, 

1981; Kennelly et al., 1982) and also in gut bacteria (Yoshida and 

Miyakawa, 1973; Chung et al., 1978; Hartmann et al., 1978; Cerniglia et 

al., 1982; Bos et al., 1986). The result of this azoreduction is the release of 

the (carcinogenic) aromatic amine from the colorant (Martin and Kennelly, 

1985). Studies performed on exposed workers have demonstrated that the 

azoreduction of benzidine-based colorants occurs in man (Genin, 1977; 

Boeninger, 1978; Lowry et al., 1980; Meal et al., 1981; Dewan et al., 

1988). Studies of Lynn et al., (1980) and Bowman et al. (1983) have 

demonstrated that the metabolic conversion of benzidine-, 3,3′dimethylbenzidine- 

and 3,3′-dimethoxybenzidine-based colorants to their 

(carcinogenic) amine precursors in vivo is a general phenomenon that must 

be expected for each member of this class of chemicals. 

However, in contrast to water-soluble dyes, the question of biological 

azoreduction of (practically insoluble) pigments was a matter of discussion 

in the recent years. One study has claimed the presence of 3,3'dichlorobenzidine 

in the urine both of experimental animals fed with 

Pigment Yellow 12 and of exposed workers (Akiyama, 1970). However, 

other experimental studies, using more modern analytical tools, did not 

confirm these results (DHEW, 1978; Leuschner, 1978; Mondino et al., 

1978; Nony et al., 1980). 

Several epidemiological studies have demonstrated that the use of the 

benzidine-based dyes has caused bladder cancer in humans. In a Japanese 

study, the risk of bladder cancer among dye applicators (kimono painters) 

was 6.8 times the expected rate (Yoshida et al., 1971). In a British study, 

workers performing the dyeing of textiles (and not exposed to benzidine 

itself) had a higher risk of bladder cancer (RR=3.4) than expected 

(Anthony, 1974). In a USSR study, an increased incidence of bladder 

cancer was found in workers who dried or ground benzidine-based dyes 

(Genin, 1977). 

In our own study on bladder cancer in painters (Myslak et al., 1991), the 

time of first exposure (painters with bladder tumors) dated mostly back to 

the first half of the century. Two factors may have been relevant: (1) at 

that time, a large number of benzidine-based azodyes was in manufacture, 

especially in Germany; (2) during that time it was usual for painters in 

Germany to prepare their paints themselves. This work included grinding


and mixing of the dyes and preparation of the coloring mixture by addition 

of solvents. All painters we had interviewed reported that this type of work 

had been regularly associated with considerable occurrence of dye dust in 

the atmosphere, up to the end of the 1950s. 

The very long latency period may explain why an enhanced risk of bladder 

cancer in German painters (due to previous exposure to azo dyes) is 

observed even today. Similar arguments have also been put forward for 

other occupational groups associated with an increased risk of bladder 

cancer, and where a causal connection with benzidine-based azo dyes had 

been proven or suggested, e.g. for textile dyers (Jenkins, 1978), leather 

dyers and shoeworkers (Decouflé, 1979), hairdressers (Guberan et al., 

1985), and tailors (Anthony and Thomas, 1970). The results of our own 

survey of painters are very probably not relevant for the present working 

conditions in Germany and other highly industrialized countries, because 

of different materials, working methods, and hygienic standards introduced 

in recent years. They are, however, quite relevant for matters of 

compensation of persons who are now diseased. 

Regulatory aspects (FRG) 

The arguments described above have led the German Commission for 

Investigation of Health Hazards of Chemical Compounds in the Work 

Area (MAK-Commission) to include the following chapter in the MAKlist, 

since 1988 (DFG, 1988): 

Azo colorants are characterized by the azo group -N=N-. They are 

made by the coupling of singly and multiply diazotized aryl amines. 

Of particular toxicological importance are colorants from double 

diazotized benzidine and from benzidine derivatives (3,3′dimethylbenzidine, 

3,3′-dimethoxybenzidine, 3,3′-dichlorobenzidine). 

In addition, aminoazo-benzene, aminonaphthalene and monocyclic 

aromatic amines are encountered. Reductive fission of the azo group, 

either by intestinal bacteria or by azo reductases of the liver and 

extrahepatic tissues, can cause these compounds to be released. Such 

breakdown products have been detected in animal experiments as 

well as in man (urine). Mutagenicity, which has been observed with 

numerous azo colorants in in-vitro test systems, and the 

carcinogenicity in animal experiments are attributed to the release of 

amines and their subsequent metabolic activation. There are now 

epidemiological indications that occupational exposure to benzidinebased 

azo colorants can increase the incidence of bladder carcinomas. 

Thus, all azo colorants whose metabolism can liberate a carcinogenic 

aryl amine are suspected of having carcinogenic potential. Due to the 

large number of such dyes (several hundred) it seems neither possible

nor justifiable to substantiate this suspicion in each individual case by 

means of animal experimentation according to customary 

classification criteria. Instead, scientifically justifiable models have to 

be relied on. Therefore, as a preventive measure to avoid putting 

exposed persons at risk, it is recommended that the substances be dealt 

with as if they were classified in the same categories as the 

corresponding carcinogenic or suspected carcinogenic amines (A1, A2, 

B) 

If there are indications that the colorant itself (e.g. a pigment) or 

any carcinogenic breakdown products are not biologically available, 

the exclusion of risk should be experimentally proven or 

substantiated by biomonitoring. Suitable animal experiments can also 

rule out suspicion of carcinogenic potential. 

On the basis of this general view, which had been endorsed by the German 

Ministry of Labor (TGS 900, Bundesarbeitsblatt 1/1990, p. 63), the 

identification of the aromatic amine component of azo colorants is of key 

importance. A suitable compilation of azo colorants, according to their 

aromatic amine components, has been published by Myslak (1990). 

Azo pigments 

H.M.BOLT 305 

The postulate of further research on the bioavailability and/or 

carcinogenicity of azo pigments, especially those based on 3,3′dichlorobenzidine 

(v.s.), has focused interest on this particular problem. 

Azo pigments based on 3,3′-dichlorobenzidine (e.g. Pigment Yellow 12, 

Pigment Yellow 13, Pigment Yellow 14) have been orally administered to 

rats, hamsters, rabbits and monkeys, at doses up to 400 mg pigment kg −1 

b.w. (Leuschner, 1978; Mondino et al., 1978; DHEW, 1978; Nony et al., 

1980; Decad et al., 1983; Sagelsdorff et al., 1990; Hoechst AG, 

unpublished data). These authors could not find 3,3'-dichlorobenzidine in 

the urine of animals treated with 3,3'-dichlorobenzidine pigments. Decad 

et al. (1983) demonstrated that 14 C-labelled Pigment Yellow 12, orally 

administered to rats, was completely excreted in the faeces. On the basis of 

these investigations of disposition of 3,3′-dichlorobenzidine-based 

pigments, it is clear why none of the long-term animal carcinogenicity 

studies performed so far (Leuschner, 1978; DHWE 1978; ICI, unpublished 

data) has demonstrated a carcinogenic effect of a diaryl pigment. 

It therefore appears that the aromatic amine components from azo 

pigments are practically not bioavailable, as demonstrated for several 

pigments on the basis of 3,3′-dichlorobenzidine (ETAD, 1990; see also Bolt 

and Golka, 1993). Hence, it is now very unlikely that occupational 

exposure to insoluble azo pigments would be associated with a substantial 

risk of (bladder) cancer in man.


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KENNELLY, J.C., HERTZOG, P.J. and MARTIN, C.N., 1982, The release of 4,4′diaminobiphenyls 

from azo dyes in the rat, Carcinogenesis, 3, 947–51. 

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pigments in mice and rats, Toxicol. Lett., 2, 253–60. 

LOWRY, L.K., TOLOS, W.P., BOENINGER, M.F., NONY, C.R. and 

BOWMAN, M.C., 1980, Chemical monitoring of urine from workers 

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MATHEWS, H.B., 1980, Metabolism of bisazobiphenyl dyes derived from 

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biding of biphenyl-based azo dyes, Drug Metab. Rev., 16, 89–117. 

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and ZANOLO, G., 1978, Absence of dichlorobenzidine in the urine of rats, 

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7. 

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1980, Metabolism studies of an azo dye and pigment in the hamster based on


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University Park Press.

22 

Toxicology of Textile Chemicals 

DIETER SEDLAK 

EnviroTex GmbH, Augsburg 

The former main aspects of textiles like fashion or usefulness seem to be 

pushed back by a new phenomenon, the ecological and toxicological aspect 

of textiles. Although everybody is talking about textiles in this context, 

everybody means the textile chemicals (and dye-stuffs) on the textile. With 

respect to this situation we have to consider new developments in Germany 

(Figure 22.1). 

For two years now textile finishing plants have to be approved within 

the German federal immission control act. This means that all sorts of 

immissions to the working place or the surroundings of the plant must be 

defined in quantity and quality. In the meantime so-called emission factors 

have been developed. But this project is not yet finished. Even for insiders 

it was surprising how many textile chemicals show unexpected behaviour 

during their application. The reasons are: 

– unknown byproducts or impurities, 

– unknown reactivities between components, 

– unknown interactions with substrate, 

– unknown dependencies of process parameters. 

Today producers and users realise that the inherent toxicological properties 

of many textile chemicals evaluated within a typical safety data sheet 

represent only a part of the whole knowledge you need for safe handling 

and processing. 

A further interesting development undoubtedly is the discussion around 

the ecolabelling of textiles. There definitely is some danger so that different 

labels based on different commercial interests should be evaluated. A 

positive step could be the fact that MST and Ecotex have joined. But other 

societies like TÜV or GSF create their own labels including statements that 

they use the better (right) label criteria. This leads quite clearly to total 

confusion for the consumer. However, what the textile industry needs 

today is confidence. The developments discussed above will not help. From 

a technical point of view we have the same problem as discussed with 

respect to the immissions. Too little is known about the real properties and

310 TOXICOLOGY OF TEXTILE CHEMICALS 

Figure 22.1 New developments in Germany regarding toxicology of textile 

chemicals. 

behaviour of textile chemicals to define the absolute label criteria. These 

uncertainties can be solved only by cooperation of the different groups and 

not by aggressive competition. How can we handle both developments in a 

proper way? 

At first we have to be clear about the interactions of both problem fields. 

Let us start with the textile chemical delivered to a finishing plant 

(Figures 22.2 and 22.3). Following the directives for dangerous substances 

or safety data sheets the product is exactly labelled and its toxicological

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

D.SEDLAK 311 

properties are well described in the SDS. This is only as good as the 

information given to the product safety manager responsible for the 

possible ways of handling and processing. However, in most cases textile 

chemicals can be described as harmless. 

Nearly all show an acute oral toxicity greater than 2000 mg kg −1 

although they may contain toxic substances in diluted form. Only a few of 

them possess an irritant character or even CMT properties. Nevertheless, 

many of them have impurities with these properties which are given more 

and more, even in low concentrations. This is a very important process


Figure 22.3 Toxicological profile of textile chemicals. 

because these low concentrations may also lead to high vapour 

concentrations at the workplace if these substances are volatile. This is 

mostly the case. 

For example, the carcinogenic substance acrylonitrile has a product label 

only giving a concentration of 0.2 percent and more in the corresponding 

polymer dispersion. Even 0.02 percent may lead to workplace 

concentrations a hundred fold higher than the TLV allows. Product data 

sheets with these ‘properties’ still exist with no additional information 

because it is not needed! 

Another problem is that nearly all textile chemicals are exposed to 

temperatures between 100°C and 230°C during application which leads to 

many only poorly defined substances which may all be set free in the 

workplace and the surroundings. Here we are also confronted with a 

typical juristic phenomenon. Many suppliers do not take any responsibility 

regarding these questions if the user defines his process parameters himself, 

especially when using recipe components from different suppliers. But 

everybody knows that the user has not the means to evaluate the 

toxicological profile of his chemicals mixture and process. This lack of 

responsibility should be clarified. 

Now let us observe the result of such chemical treatment of textiles. 

Normally you will only discuss the summarised toxicological profiles of the 

used chemicals in an additive way. This means that the combination of 

different products—all with no acute toxicity—will also show no acute

Figure 22.4 Typical composition of a flame retardant recipe. 

D.SEDLAK 313 

toxicity. Furthermore the concentration of the active ingredients in the new 

matrix textile is much lower than in the matrix water most textile 

chemicals are based on. This strategy seems to be appropriate. On the 

other hand take the irritant property of, for example, fatty amine based 

emulsifier. This property is lost during formulation of the emulsifier in the 

textile chemical by homogeneous dilution. The average concentration after 

application to the textile is surely lower. But who has information about 

the actual form of this fatty amine in the textile. Is it distributed 

homogeneously, is it more located at the surface, does it interact with other 

substances and even lose its irritant character? Many questions seem to be 

unanswered. 

Both complex questions—the toxicology of the handling and processing 

of textile chemicals and the toxicology of the result of the processing on 

textiles— will be discussed by means of an admittedly drastic example, a 

flame retardant process. This example is rather representative because an


Figure 22.5 Total composition of a flame retardant recipe. 

EEC-directive is ‘under construction’ just now which demands the finishing 

of all upholstery covers with flame retardant chemicals. 

The composition of this recipe, shown in Figure 22.4, doesn’t seem to be 

too complex. No special toxicological properties are apparent. The 

corrosive character of the phosporic acid vanishes during the application

D.SEDLAK 315 

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

by neutralizing the finished textile with alkalis. Complex reactions of the 

different components are expected which are supposed to form not well 

defined polycondensates fixed to the textile fibre. On this basis there is no 

apparent reason to think of any toxicological side effects. 

In Figure 22.5, however, you will get a good idea about the real 

situation. But this detailed composition should not be deceptive about the 

fact that this is the ‘wanted’ technical composition verified by a few 

analytical data. In this case about 500 actual substances can be expected. 

Even if we should have available all the necessary toxicological data for the 

components it would not help, because there is poor information about the 

result of the finishing process itself. This is the only reality. 

A first step in the right direction would be the analysis of the process 

offgas. The result reflects approximately the activities on the textile during


Figure 22.7 Needs for the textile toxicology assessment. 

processing. Figure 22.6 shows in which way this knowledge can be used to 

perform a proper risk assessment based on defined substances. However, in 

the case of evaluation of the skin toxicity we are confronted with a lack of 

data concerning the bio-availability of the named substances. Not before 

having cleared these additional questions can the toxicologist seriously 

start his routine activities, the risk assessment. Anyway this seems to be 

much easier than to discover the chemical basis for this assessment. 

With respect to the toxicology of textile chemicals there is a lot to do 

mostly in the field of understanding the chemicals and processes used. 

After this we recommend the development of models for evaluating the 

bioavailability of substances and test kits to characterize the toxicological 

properties of textiles as a whole. Last but not least we hope for much 

better communication between the industries involved (Figure 22.7).

23 

Antioxidants and Light Stabilisers : Toxic Effects 

of 3,5-Di-alkyl-4hydroxyphenyl Propionic Acid 

Derivatives in the Rat and their Relevance for 

Human Safety Evaluation 

HELMUT THOMAS, PETER DOLLENMEIER, ELKE 

PERSOHN, HANSJÖRG WEIDELI and FELIX WAECHTER 

Ciba-Geigy Limited, Basle 

Introduction 

Antioxidants and light stabilisers are important additives for a wide range 

of plastic materials used for industrial as well as for medicinal and food 

packaging purposes. Technical efficiency requires that these compounds are 

mobile within the polymer network. This implies that humans may be 

exposed to such compounds not only during the manufacturing process but 

also in the course of the decay of polymers, by direct contact with 

substances that have migrated to the surface of plastic materials or by 

ingestion of diffusion-contaminated food. It is of considerable interest, 

therefore, to be aware of the toxicological properties of these compounds 

and the relevance of these properties for the safety assessment in humans. 

Sterically hindered phenolic antioxidants 

Phenolic antioxidants have been widely used as food preservatives and 

their almost classical representatives, 2,6-di-tert-butyl-4-methyl phenol 

(BHT) and 2-tert-butyl-4-methoxyphenol (BHA) have been characterised in 

extenso with respect to their biochemical and toxicological properties 

(Søndergaard and Olsen, 1982; Conning and Phillips, 1986; Ito et al., 

1986a; Williams et al., 1990a,b; Verhagen et al., 1991). Being nongenotoxic, 

these compounds were found upon oral administration to 

rodents to cause slightly increased liver weight, to induce mainly epoxide 

hydrolase and phase II drug metabolising enzymes and to exert some anticarcinogenic 

effect (Benson et al., 1979; Choe et al., 1984; Gregus and 

Klaassen, 1988; Prochaska and Talalay, 1988; Perchellet and Perchellet, 

1989; Rodrigues et al., 1991). Pulmonary toxicity and carcinogenicity 

particularly of BHT in the mouse and formation of forestomach carcinoma 

and papilloma in the rat and Syrian hamster, respectively, have been 

reported and are considered to be largely species-specific effects (Abraham 

et al., 1986; Anon, 1986; Ito et al., 1986a, b; Verhagen et al., 1989). 

Industrial phenolic antioxidants as used in the polymer technology, by

318 ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT 

contrast, although related to BHT, require higher molecular weights and 

hydrophobicity in order to retain them in the polymer matrix. This can be 

achieved, for example, by introducing alkyl chains or (esterified) 

carboxyalkyl moieties in the para-position to the phenolic hydroxyl group 

with variations in the aliphatic substitution pattern of the 2- and 6positions. 

With these modifications the questions arose, whether or not the 

toxicology of the chemically modified offspring would still be related to 

that of the phenolic core of the BHT ancestor or completely different, 

although a common metabolic degradation pathway to structural entities 

resembling BHT could be anticipated. Clarification of this question was 

expected to contribute to the understanding of the toxicology of an entire 

class of phenolic antioxidants. A number of differentially esterified 4hydroxy-3, 

5-dialkyl-phenylpropionic acid derivatives were subsequently 

subjected to subchronic toxicity testing in rats with the result that the 

effects encountered were largely dependent on the alcohol moiety used for 

esterification and the size of the alkyl-substituent in the 3- and 5-positions: 

most frequently, increased liver weights were encountered in compounds 

with bulky 3,5-substituents such as tert-butyl. In some instances, however, 

initial hepatomegaly was followed after several weeks or months of 

treatment by increases in thyroid weights and a proliferation of the thyroid 

follicular epithelium. The mechanism leading to the latter effect has been 

investigated in detail using the di-ester of 3-tert-buty1–4-hydroxy-5-methylphenylpropionic 

acid with ethylene glycol, Compound B, as a model 

compound (Table 23.1). 

Blood kinetics and blood metabolites 

Compound B is a symmetrical di-ester compound (Table 23.1). When 

administered at a single oral dose of about 10 mg kg −1 body weight to male 

rats, 14 C-phenyl-labelled Compound B was readily adsorbed, and maximal 

blood radioactivities were reached after 1 h. Thereafter, blood radioactivity 

declined rapidly and only minute amounts were detected 48 h after 

treatment. At any time point investigated Compound A, the free carboxylic 

acid of Compound B, was the dominating blood metabolite, whereas the 

parent compound constituted a minor component only (Table 23.2). These 

findings are indicative of an efficient first pass hydrolysis and suggest that 

the carboxylic acid metabolite, Compound A, might be responsible for the 

toxicological profile of this antioxidant in the rat. 

Liver enzyme induction 

Compound B was administered in the feed to male rats at dose levels of 50, 

150, 500 and 1000 ppm. After treatment for 14 days, absolute liver 

weights were dose-dependently increased. Biochemically, this

Table 23.1 3.5-Substituted 4-hydroxyphenyl propionic acid esters 

H.THOMAS ET AL. 319 

hepatomegaly was accompanied by an induction of total microsomal 

cytochrome P450, microsomal epoxide hydrolase and UDPglucuronosyltransferase, 

cytosolic glutathione S-transferase and 

peroxisomal β-oxidation. The observed induction of the cytochrome P450 

content was reflected in increased activities of ethoxycoumarin O-deethylase 

as well as lauric acid 11- and 12-hydroxylases (Table 23.3). This 

suggests an induction of cytochrome P450 isoenzymes of the subfamilies 

CYP2B and CYP4A. Indeed, increased amounts of CYP2B and CYP4A 

proteins were found in liver microsomes from treated animals by means of 

Western-blotting with monoclonal antibodies specific for these iso-enzymes 

(data not shown). Within 28 days after cessation of a 14-day treatment at


Table 23.2 Parent equivalents and blood metabolites of [ 14 C]-labelled Compound B 

after single oral administration of 9.5 mg kg −1 body weight to male rats 

Note: 

bld: below the limit of detection. 

Blood was taken at the indicated time intervals and extracted with ethyl acetate for 

analysis. Compound B and its free carboxylic acid metabolite (Compound A, 

Table 23.1) were identified by thin-layer co-chromatography with authentic 

reference samples. Quantification of Compounds A and B was accomplished by 

radiometric scanning of the plates following thin-layer chromatography of the 

respective blood extracts. 

1000 ppm, liver weights as well as the investigated biochemical parameters 

returned to control levels. Therefore, Compound B may be addressed as a 

reversible barbiturate- and peroxisome proliferator-type inducer in the rat, 

as characterised by its liver enzyme induction profile. 

Effects on serum thyrotropin and thyroid hormones 

When male rats were fed Compound B admixed in the diet for 14 days at 

dose levels of 50, 150, 500 and 1000 ppm, liver and thyroid weights were 

increased in a dose-dependent manner. Histopathological examination of 

the thyroid gland revealed hypertrophy of the follicular epithelium and 

thinning of colloid at 150, 500 and 1000 ppm. Morphological changes in 

the pituitary gland comprised enlarged thyrotropin (TSH)-producing cells 

with foamy or vacuolated cytoplasm. In addition, treatment resulted in 

markedly increased serum TSH and reverse triiodothyronine (rT 3) 

concentrations, whereas serum thyroxine (T 4) and triiodothyronine (T 3) 

levels were found slightly decreased. The effects observed at 1000 ppm 

were found to be reversible after a 28-day recovery period (Muakkassah- 

Kelly et al., 1991). 

In additional experiments, male rats were rendered hypothyroid, fed 

Compound B at 1000 ppm for 21 days and infused for the last 7 days with 

slightly supraphysiological concentrations of T 4. The observed changes in


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

male rat liver 

Notes: 

Values are means of 9 (14 days treatment) or 4 (14 days treatment/28-days 

recovery) animals. 

Standard deviations are given in parentheses. 

0/0: Control recovery. 1000/0: High-dose recovery. 

* p


metabolism has been observed with various deiodinase inhibitors (Hill et 

al., Liang et al., 1993) and hepatic enzyme inducers (McClain, 1989; 

Curran and DeGroot, 1991; Barter and Klaassen, 1992; Visser et al., 

1993). 

Induction of thyroid neoplasia 

In a long-term feeding study, administration of Compound B to rats at a 

dose level of 1000 ppm was associated with an increased incidence of 

thyroid gland follicular adenoma and carcinoma (Muakkassah-Kelly et al., 

1991). Compound B was shown to be devoid of mutagenic and clastogenic 

activity (Muakkassah-Kelly et al., 1991). Therefore, it is likely that thyroid 

tumour induction by Compound B was not the result of a direct, genotoxic 

effect on this organ, but rather a consequence of the hormonal imbalance 

induced by this antioxidant in the rat. 

An intact hypothalamic-pituitary-thyroid axis is able to respond to a 

chemically induced alteration in peripheral hormone metabolism with 

increased hormone production by the hypertrophic gland. However, it is 

known that chronic, excessive stimulation of the gland can lead to 

follicular hyperplasia and ultimately progress to thyroid neoplasia (Paynter 

et al., 1988; McClain, 1989; Curran and DeGroot, 1991; Johnson et al., 

1993). For Compound B, this hypothesis is in agreement with the doseresponse 

characteristics obtained in the long-term study, where thyroid 

tumours were induced exclusively at a dose-level sufficiently high to cause 

hormonal imbalance (Muakkassah-Kelly et al., 1991). 

Implications for human risk assessment 

For human risk assessment, it is of critical importance to identify the 

mechanism by which Compound B caused thyroid neoplasia in the rat, a 

species most often used for carcinogenic hazard identification. Hyperplastic 

changes in the thyroid are frequently observed in rat carcinogenicity 

studies, and this species appears to be very sensitive to compounds which 

interfere with thyroid hormone synthesis and/or catabolism. They evoke an 

immediate stimulation of the gland upon short-term treatment as a 

consequence of an increased pituitary TSH secretion (Zbinden, 1987; 

Paynter et al., 1988; McClain, 1989). 

However, the effects of xenobiotics on the pituitary-thyroid axis in 

rodents cannot necessarily be extrapolated to man since rodents and man 

are distinguished by many important physiological and biochemical 

differences (Gopinath, 1991): e.g. the amount of thyroxin-binding 

globulin, the half-life of T 4 and its biliary excretion as well as plasma THSlevels 

and their response to thyrotropin releasing hormone. These 

differences render the rat very sensitive to small changes in the plasma T 4

level, whilst humans are essentially insensitive (Hill et al., 1989; Grasso et 

al., 1991). 

Therefore, in contrast to rats, there is no conclusive evidence for a critical 

role of TSH in thyroid stimulation and carcinogenesis in humans (Hill et 

al., 1989). In cultured human thyroid cells, for example, THS was unable 

to induce proliferation whereas a stimulation of growth was observed in 

rat thyroid cells (Westermark et al., 1985). Clinical data are available for 

some compounds such as the anticonvulsants phenobarbitone, 

diphenylhydantoin and carbamazepine as well as the antibiotic rifampicin, 

which are known liver microsomal enzyme inducers in man. They increase 

thyroid hormone metabolism and excretion and eventually decrease serum 

thyroid hormone levels. There is also evidence, that administration of these 

drugs leads to thyroid stimulation, however, largely in the absence of 

increased TSH levels (Curran and DeGroot, 1991). In addition, 

epidemiological data are not in favour of a link between human use of such 

compounds with an increased incidence of thyroid tumours (Curran and 

DeGroot, 1991), nor have increased rates of thyroid cancers been reported 

in areas of endemic iodine deficiency (McClain, 1989). Therefore, the 

currently available data do not support the idea, that thyroid stimulation 

as a response to chemically induced increases in circulating TSH 

concentrations significantly contributes to thyroid tumour formation in 

man. 

Compound B has been shown to cause thyroid tumours in the rat. In a 

series of speciality studies, the compound was identified as an enzyme 

inducer and a 5′-deiodinase inhibitor in the rat liver. These findings argue 

for a rodentspecific, indirect mechanism leading to the formation of thyroid 

tumours. Moreover, the observed dose-response characteristics are 

indicative of a threshold process, e.g. liver enzyme induction and inhibition 

of 5′-deiodination are irrevocable prerequisities for thyroid tumour 

formation in this species. 

Benzotriazole-based light stabilisers 


Ester derivatives of the 3-[3-(2H-benzotriazole-2-yl)-5-tert-butyl-4hydroxy-phenyl] 

propionic acid represent potent UV-light absorbers and 

constitute an important class of industrial plastic additives and light 

stabilisers (Compounds C-F, Table 23.1). Toxicologically, this class of 

chemicals is characterised by generally low acute oral or dermal toxicity 

and the lack of genotoxicity in the commonly employed battery of 

bacterial and cellular mutagenicity tests. Irrespective of the alcohol moiety, 

however, all compounds, when administered subchronically to rats, 

displayed very similar predominantly hepatotrophic effects, with spleen and 

kidney weights in addition being only slightly affected: pronounced 

hepatomegaly, hepatocyte hypertrophy, and concomitantly increased


plasma transaminase activities. Upon electron microscopical examination, 

a striking peroxisome proliferation was the major finding. 

Compound F, a di-ester ‘product by process’ obtained upon esterification 

of 3-[3-(2H-benzotriazole-2-yl)-5-tert-buty1–4-hydroxyphenyl] propionic 

acid with polyethyleneglycol 300, when tested for its toxicity to rat 

reproduction in a Segment I study, gave rise to increased numbers of stillborn 

pups, decreased pup survival, decreased weight gain of surviving pups, and 

dark discoloured abdominal skin regions in a number of pups at higher 

dose levels. 

The common nature of general toxicology findings with all investigated 

derivatives of the addressed benzotriazole-based light stabilisers suggested a 

common basis of action and, depending upon this action, perhaps a very 

similar behaviour and extent of potency as foetotoxic agents. In order to 

investigate these interrelationships a series of mechanistic studies were 

conducted focusing on the kinetics, primary metabolism of the parent 

compounds in vitro and in vivo and their effect on selected biochemical 

liver parameters in rats. Compound F was selected as a model compound 

to investigate the mechanism of toxicity in pregnant female rats and 

foetuses. 

In vitro hydrolysis 

Compound D, the methyl ester of 3-[3-(2H-benzotriazole-2-yl)-5-tertbutyl-4-hydroxyphenyl] 

propionic acid was readily hydrolysed in vitro by 

rat serum as well as rat liver homogenate while a homogenate of rat small 

intestine when compared on a gram tissue basis, appeared to be less 

efficient by three orders of magnitude than the liver. Increasing the sterical 

hindrance around the ester bond by formation of the di-ester with a short 

chain alcohol reduced the rate of in vitro hydrolysis considerably as 

demonstrated for Compound F, the diester of hexane-l,6-diol with 

Compound C. When offered at a test concentration of 0.2 mM, essentially 

no hydrolysis of this compound was observed with rat serum, and the 

hydrolysis by liver and small intestine homogenate was estimated to 

proceed at least two and one orders of magnitude slower, respectively, than 

calculated for Compound D (Table 23.1 and 23.4). 

Blood kinetics and blood metabolites 

Assuming comparable extents of intestinal absorption in vivo, the observed 

differences in the in vitro hydrolysis rates might as well suggest 

significantly different in vivo hydrolysis rates and consequently quite 

different residence times for both parent compounds in the rat in vivo. 

Different residence times, on the other hand, may eventually allow not only 

for additional routes of metabolism but also for an intensification of toxic

Table 23.4 Kinetic parameters for the in vitro hydrolysis of Compound D and E by 

rat serum and organ homogenates 

Notes: 

a The apparent Vmax value for serum is given in µmol min −1 ml −1. 


b Initial velocity of hydrolysis at 0.2 mM ester concentration. 

Hydrolysis of Compound D was determined in 50 mM Tris/phosphate buffer, pH 7. 

5, containing either 1 per cent (v/v) rat serum, or 1.25 per cent (w/v) rat liver 

homogenate or 10 per cent (w/v) small intestine homogenate. Similarly, hydrolysis 

of Compound E was assessed using 98 per cent (v/v) rat serum or in the presence of 

10 mM Tris/HCl buffer, pH 7.5, containing 250 mM sucrose and either 24.5 per 

cent (w/v) rat liver homogenate or 19.6 per cent (w/v) small intestine homogenate. 

effects. This question was addressed in a pharmacokinetic study under 

conditions of single oral administration of Compounds D and E at a dose 

level of 10 mg kg −1 each (Table 23.5). 

14 C-Phenyl-labelled Compound D was readily absorbed from the 

gastrointestinal tract. Maximal blood radioactivity was reached between 1 

and 2 h and subsequently eliminated with an apparent half-life of 10.0–11. 

8 h. After 48 h only minute amounts of radioactivity equalling about 3 per 

cent of the blood levels in T max were detectable. Analysis of the resulting 

blood metabolite pattern largely confirmed the findings of the preceding in 

vitro investigation of enzymatic Compound D hydrolysis: particularly 

during periods of high parent equivalent concentrations in blood as 

recorded between 30 min and 6 h after dosing, hydrolysis appeared to be 

the major metabolic pathway as evidenced by the high concentrations of 

the carboxylic acid, Compound C (34–77 per cent), and an unidentified 

metabolite (17–36 per cent) which was regarded to have evolved from 

Compound C by an additional metabolic step (Table 23.5). 

Quite surprisingly, Compound E was found to be absorbed to a much 

lower extent than Compound D, with a C max after 1 h of less than one 

tenth of the value seen with the latter. Elimination with an apparent halflife 

of 12.0 h and slightly higher residual radioactivity of approximately 4.8 

per cent of the blood levels recorded at T max after 48 h indicated only a 

slightly reduced elimination rate as compared to Compound D. Also, quite 

different from the initially anticipated result, hydrolysis contributed 

substantially to the rapid metabolism of Compound E. The 24 h AUC 

values revealed a 23 per cent and 36 per cent contribution of the carboxylic


Table 23.5 Summary of pharmacokinetic parameters obtained after single oral 

administration of 10 mg kg −1 Compound D and E to two male rats each 

Notes: 

a Below the limit of reliable quantification. 

pe: Parent equivalent. 

acid, Compound C, and unidentified metabolites, respectively, to the total 

AUC. It is assumed that the majority of unidentified metabolites have 

arisen from further biotransformation of the carboxylic acid. The 

significant contribution of hydrolysis products to the total AUC is still 

evident after 168 h although low blood radioactivity levels 24 h after 

dosing generally prevented accurate quantitation of the metabolites and 

thus slightly diminished their apparent overall share (Table 23.5). 

Liver enzyme induction 

Subchronic oral (gavage) administration of single daily doses of Compound 

C for 14 days, Compound D for 14 days, Compound E for 13 weeks, and 

Compound F for 114 days to male rats (Tables 23.6 and 23.7) was 

correlated with a dose-dependent massive increase in absolute liver weight 

up to about 190 per cent of control at the highest dose level irrespective of 

the treatment period and the test compound. This pronounced 

hepatomegaly was paralleled by a comparably small two-fold elevation of 

the microsomal cytochrome P450 contents and an about 50 per cent 

decrease in total UDP- glucuronosyltransferase activity. Essentially no 

changes were recorded for the cytochrome P450 dependent 

ethoxycoumarin O-de-ethylase activity while microsomal epoxide 

hydrolase activities appeared to vary, with slight increases in Compound C 

and Compound D treated animals, no changes in Compound E treated 

animals and even a dose-dependent reduction to 46 per cent of control in 

rats treated with Compound F at 100 mg kg −1 (Table 23.6). 

Strongly induced peroxisomal fatty acid β-oxidation activities for all 

model compounds as well as lauric acid 12-hydroxylase activities tested for 

Compounds E and F were accompanied by significant dose-dependent 

decreases in glutathione S-transferase activities to 32, 51, 40 and 15 per 

cent of control at the highest dose level tested for Compound C, D, E and

Table 23.6 The effect of various benzotriazole-based light stabilisers on selected biochemical parameters related to and indicative 

for a barbiturate and/or polycyclic aromatic hydrocarbon type enzyme induction in male rat liver 

H.THOMAS ET AL. 

327 

Notes: 

dnt: dose level not tested, 

nd: not determined. 

Microsomal epoxide hydrolase and total microsomal UDP-glucuronosyltransferase activities were determined with styrene oxide 

and 3-methyl-2-nitrophenol as substrate, respectively. 

Values are means±standard deviation of 10 control and 5 treated animals (a, b) or 8 animals (c) or 5 animals per group (d). 

Asterisks indicate results significantly different (two-sided Dunnett’s test) from control: * p


F, respectively. For Compound F, which appeared to be the most potent 

inducer of peroxisomal β-oxidation and lauric acid 12-hydroxylase 

activities, a concomitant strong 90 per cent reduction of morphine UDPglucuronosyltrasferase 

and marked 2.5-fold increase in bilirubin UDPglucuronosyltransferase 

activity was recorded (Table 23.7). Electron 

microscopy confirmed what had already been indicated by changes in the 

investigated enzyme levels, a striking proliferation of peroxisomes with the 

same appearance of these organelles regardless of the compound tested: 

vigorous increase in number, a striking number of markedly enlarged 

peroxisomes frequently containing matrical inclusions (matrical plates) and 

peroxisomes forming arrays or clusters (polyperoxisomes) in the virtual 

absence of any significant proliferation of smooth endoplasmic reticulum 

(data not shown). 

Consequently, the hepatotrophic effects of the tested benzotriazole-based 

light stabilisers were clearly assigned to their action as peroxisome 

proliferators in rat liver. Mindful of the different durations of treatment 

their potency was found to rank in the order Compound E

Table 23.7 The effect of various benzotriazole-based light stabilizers on selected biochemical parameters related to and indicative 

for a peroxisome proliferator type enzyme induction in male rat liver 


Notes: 

dnt: dose level not tested. 

nd: not determined. 

Cyanide-insensitive peroxisomal fatty acid -oxidation and cytosolic glutathione S-transferase activities were determined with 

[l-14C]-palmitoyl-CoA and1-chloro2,4-dinitrobenzene as substrate, respectively. 

Values are means±standard deviation of 10 control and 5 treated animals (a, b) or 8 animals (c) or 5 animals per group (d). 


330 

ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT 

Table 23.8 Morphological changes of dam and foetal hepatocyte organelles after treatment with Compound F from day 6 through 

days 14, 17 and 20 of gestation

iochemical investigations which revealed a moderate four-fold induction 

of peroxisomal β-oxidation in dams and an up to 15-fold induction of this 

activity at day 21 of gestation in foetal livers, respectively, with the final 

foetal activity exceeding dam activity by 40 per cent. Strong increases in 

lauric acid 11- and 12-hydroxylation, which are known to be associated 

with isoenzymes of the cytochrome P450 CYP4A gene family and the 

phenomenon of peroxisome proliferation in rodents, of up to 2.6- and 10. 

5-fold in dams and 11.5- and 23.2-fold in foetuses, respectively, at day 21 

of gestation were recorded as well, indicating again a slightly higher final 

activity in foetal than in dam liver. By contrast, catalase, known as a 

detoxifying enzyme for hydrogen peroxide generated particularly in the 

course of increased peroxisomal activity, was shown to be induced up to 5. 

6-fold in dam and 8.2-fold in foetal liver at day 21 of gestation, leaving 

foetal liver, however, with only about 30 per cent of the final activity seen 

in dams (Table 23.9). This discrepancy between the strong induction of 

hydrogen peroxide generating peroxisomal β-oxidation and the 

considerably less potent induction of the hydrogen peroxide destroying 

catalase activity appears to be reflected in the 1.7-fold increase in the lipid 

peroxidation product malondialdehyde in foetal livers as well as in the 

decrease of total and reduced hepatic glutathione levels at day 21 of 

gestation (Tables 23.9 and 23.10). Surprisingly, foetal defense systems 

against oxidative stress such as selenium-dependent and 

seleniumindependent glutathione peroxidase were found poorly developed 

throughout the investigated periods of gestation and barely inducible by 

the test article leaving the pups with little protection against any kind of 

oxidative insult (Table 23.10). Consequently, Compound F was clearly 

identified as a peroxisome proliferator in pregnant rat as well as foetal liver 

with high potential for the initiation of oxidative damage in foetal tissues. 

Implications for human safety assessment 


The understanding of the mechanisms by which benzotriazole-based UV 

light stabilisers exert their hepatotrophic effects in rodents is of crucial 

importance for the assessment of safety aspects in humans. The presented 

rat studies have shown that the liver effects exerted by Compound C and 

its ester derivatives Compounds D, E and F are clearly related to the 

induction of peroxisome proliferation. The identical nature of the observed 

effects regardless of the alcohol component in the investigated esters 

suggests that the toxic potential resides solely with the 3-[3-(2Hbenzotriazole-2-yl)-5-tert-butyl-4-hydroxy-phenyl] 

propionic acid whereby 

the individual potency appears to be mainly determined by the different 

bioavailablity of the respective ester compound and the extent and velocity 

of its hydrolysis in vivo. Also, it appears, that in vitro hydrolysis studies do


not actually reflect the hydrolytic capacity of the in vivo system for a given 

ester. 

Liver enlargement and the induction of diagnostic enzyme activities is a 

characteristic response of rodents to treatment with peroxisome 

proliferators and results from a combination of both hypertrophy and 

hyperplasia. According to current opinion, peroxisome proliferation and 

liver growth are closely associated with the formation of hepatocellular 

tumours in rats and mice (Hawkins et al., 1987; Lock et al., 1989; Bentley 

et al., 1993). However, a number of feeding studies have demonstrated 

that there may actually be two types of threshold with respect to dose 

relationships: at very low doses, administration of peroxisome proliferators 

will not result in any liver response at all (Bentley et al., 1993). With 

increasing doses the first threshold will be exceeded with subsequent 

stimulation of peroxisome proliferation and DNA synthesis. 

As a result of several studies it is obvious that a limited extent of liver 

growth does not automatically lead to tumour formation as has been 

demonstrated, for example, with fenofibrate and diethylhexylphthalate in 

carcinogenicity assays (Mitchell et al., 1985; Price et al., 1986; Keith et al., 

1991). Thus, a second threshold has to be exceeded at which the 

magnitude of effects is sufficient to cause tumour development in rodents. 

In addition, extended administration of the peroxisome proliferator 

appears a necessary prerequisite to exceed this tumourigenic threshold. 

Also, a large number of in vitro and in vivo studies have provided ample 

evidence for a marked species difference in susceptibility to the effects of 

peroxisome proliferators. Rats and mice are extremely sensitive while 

hamsters show a markedly smaller response and non-human primates and 

humans appear to be insensitive or non-responsive (Lake et al., 1989; 

Bentley et al., 1993; Graham et al., 1994). The latter finding is supported 

by epidemiological evidence from long-term treatment of patients with 

hypolipidaemic agents (Bentley et al., 1993). Therefore the available 

evidence strongly supports the conclusion that the effects of benzotriazolebased 

light stabilisers in rodents are of no relevance to human safety 

assessment. 

The action of Compound F as a strong peroxisome proliferator in foetal 

livers starting as early as day 15 of gestation suggests that treatment related 

initiation of high level oxidative stress under conditions of poorly 

developed foetal protection and detoxification systems. This view is 

supported by substantially elevated hepatic malondialdehyde levels and 

essentially depleted glycogen stores in hepatocytes of foetuses from treated 

dams on day 21 of gestation. The latter indicates an extensive glucose 

consumption, presumably via the pentose phosphate pathway, to supply 

the hydrogen peroxide and lipid peroxide detoxifying glutathione 

peroxidase system with the necessary reduction equivalents (NADPH). 

Under conditions of limited degradation of hydrogen peroxide, this

Table 23.9 The effect of Compound F on dam and foetal absolute liver weight and selected biochemical liver parameters related to 

and indicative for peroxisome proliferation after treatment from day 6 through days 14, 17 and 20 of gestation 


Notes: 

bld: below the limit of detection. 

Cyanide-insensitive peroxisomal fatty acid -oxidation and catalase activities were determined with [lhydrogen 

peroxide as substrate, respectively. 

C]-palmitoyl-CoA and 

Values are means±standard deviation from 6 dams per group and 6 pools of 2–9 foetuses per dam depending on the litter size. 


334 

ANTIOXIDANTS AND LIGHT STABILISERS: TOXIC EFFECT 

Table 23.10 The effect of Compound F on selected biochemical dam and foetal liver parameters related to defence mechanisms 

against oxidative stress after treatment from day 6 through days 14, 17 and 20 of gestation 

Notes: 

Selenium-dependent and selenium-independent glutathione peroxidase activities were determined with cumene hydroperoxide and 

hydrogen per-oxide as substrate, respectively. 

Values are means±standard deviation from 6 dams per group and 6 pools of 2–9 foetuses per dam depending on the litter size. 


compound is known to pass the liver and to be systemically distributed 

throughout the foetal body. Thus the endothelial cell injuries observed in 

the course of the Segment II study are regarded secondary to the effect of 

peroxisome proliferation and to have arisen as a consequence of oxidative 

damage which has been demonstrated to occur in various endothelial cell 

systems, for example as a result of iron-mediated oxygen free radical attack 

(Brieland et al., 1992; Barchowsky et al., 1994; Krautschick et al., 1995). 

Therefore, the observed foetotoxicity of Compound F in rats is regarded as 

occurring solely as a consequence of and secondary to peroxisome 

proliferation, and there is no evidence to assume that this effect as a result 

of exposure to benzotriazole based light stabilisers may occur in species 

that are non-responsive to the action of peroxisome proliferators as stated 

above, including humans. 

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24 

Toxicology of Surfactants: Molecular, 

Mechanistic and Regulatory Aspects 

WALTER STERZEL 

Henkel KGaA, Düsseldorf 

Introduction 

The vast distribution of surfactants in various products in everyday use 

requires that the unwanted effects as well as the desired properties are 

known in order to recognize possible risks and prevent any damage to the 

health of humans. In order to understand the effects of surfactants on the 

organism, their most important biochemical effects, which depend on the 

interaction of surface active agents with basic biological structures like 

membranes, proteins and enzymes, are discussed. Following this discussion 

the local effects of surfactants are described. Local in this sense are all the 

effects encountered directly at the point of contact with the outer surfaces 

of the body, such as skin and mucous membrane irritation as well as 

allergies arising from skin contact. After this section the toxicokinetic 

properties of surfactants, providing information about type and extent of 

absorption by organisms, metabolic pathways and their elimination are 

discussed. The section on systemic effects deals, in contrast to local effects, 

with reactions arising after the substance has entered the organism by 

swallowing, skin penetration or inhalation. 

Due to their technical and economic importance, surfactants have been 

used extensively for decades. This resulted in an abundance of scientific 

publications concerning their effects on organisms. As a complete review 

would exceed the scope of this contribution, the focus will centre on a 

description of exemplary data which are important for the evaluation of 

the safety of surfactants. 

Biochemical properties of surfactants 

Surfactants come into immediate contact with the body during cleaning of 

the skin and act on the skin cells directly. When surfactants are swallowed 

unin tentionally, tissue damage is also possible. The question of the effects 

on the cells and cell components like membranes, proteins and enzymes is 

therefore also important from a toxicological point of view.

340 TOXICOLOGY OF SURFACTANTS 

Interactions with membranes 

Due to their ability to absorb at interfaces, surfactants can interact with 

biological membranes. This interaction depends on the concentration of the 

surfactant and can be described in the following sequence (Helenius and 

Simons, 1975). In the first instance the monomeric surfactant molecule 

adsorbs onto the membrane. For a low surfactant/membrane ratio this 

changes the permeability of the membrane and leads to cell lysis at higher 

concentrations. At even higher surfactant concentrations, the lamellar 

structure of the membrane is lost and it is solubilized. A further increase in 

surfactant concentration results in the separation of the phospholipids from 

the protein. This allows surfactant molecules to adsorb on previously 

hidden regions of the protein molecule. For the solubilization of integral 

membrane proteins the formation of micelle/protein complexes seems to be 

a prerequisite. A significant solubilization of these proteins is possible only 

if the critical micelle concentration c M is exceeded. This is indicated by the 

fact that the microsomal membrane bound enzyme arylsulphatase-C could 

only be extracted from the membrane with retention of the biological 

activity after micelles were formed (Chang et al., 1985). 

As a consequence of these interactions, surfactants are able to influence 

the metabolism of membrane components (DeLeo, 1989). This has been 

demonstrated by studies on the pathophysiology of surfactant-mediated 

skin irritation. In vitro cultured corneocytes showed an increased release of 

cholin metabolites after incubation with anionic surfactants. This effect 

was less pronounced after treatment with nonionic surfactants. In 

conclusion, these investigations demonstrated that the release of 

metabolites is correlated with the irritation potential of surfactants. 

Interactions with proteins 

Depending on the structure of the surfactant the interactions with proteins 

are based on polar or hydrophobic interactions. The binding of surfactants 

to protein molecules is a function of the concentration of free surfactant in 

equilibrium with the protein. The binding is affected by the pH, 

temperature and ionic strength of the solution. These factors can lead to 

conformational changes of proteins and thereby increase or decrease the 

number of available binding sites. Natural bovine albumin, for example, 

has 10 binding sites for decyl glucoside at 10°C and 13 at 25°C 

(Wasylewski and Kozik, 1979). According to a theory developed by Jones 

(1975), surfactants adsorb onto proteins in multiple equilibria. Only a few 

surfactant molecules (

W.STERZEL 341 

pyruvate oxidase can form this type of binding (Schwuger and Bartnik, 

1980). Binding of more surfactant molecules leads to conformational 

changes in the protein. It is obvious that conformational changes allow 

binding of further surfactant molecules on hydrophobic regions which were 

previously not exposed. 

According to their different chemical structure (e.g. anionic, cationic, 

amphoteric or nonionic) surfactants differ significantly in their ability to 

carry out cooperative binding and therefore they differ in their biological 

activity. Anionic surfactants form adsorption complexes with proteins due 

to polar and hydrophobic interactions. Polar interactions between the 

negatively charged hydrophilic group of the surfactant and the positively 

charged groups of the protein molecule are the precondition for the 

formation of hydrophobic associations between surfactant molecule and 

protein molecule (Garcia-Dominguez, 1977; Schwuger and Bartnik, 1980). 

In the case of dodecylsulphate and tetradecylsulphate the binding results in 

denaturation of the proteins (Makino et al., 1973). Cationic surfactants can 

interact by polar and hydrophobic binding as well. Polar interactions result 

in electrostatic bonds between the negatively charged groups of the protein 

molecule and the positively charged surfactant molecule. For example, the 

enzyme, glucose oxidase, is deactivated by hexadecyl trimethyl ammonium 

bromide through formation of an ion pair between the cationic surfactant 

and the anionic amino acid side chain of the enzyme molecule (Tsuge, 

1984). Nonionic or amphoteric surfactants and proteins show either no 

interaction at all or interactions that are extremely weak and normally 

close to the limits of sensitivity of the analytical methods used. For this 

reason, nonionic surfactants will not dissolve sparingly soluble proteins, 

denature proteins, or contribute to a swelling of the epidermis. Figure 24.1 

shows the solubility of the protein zein, which is almost insoluble in water, 

and is more or less solubilized by sodium dodecyl sulphate and alkyl 

ethyleneglycol ether sulphates, while the nonionic ethoxylated nonylphenol 

is ineffective (Schwuger and Bartnik, 1980). A further reason for the poor 

interactions between nonionic surfactants and proteins could be that the 

concentration necessary for cooperative binding with the protein is not 

attained with nonionic surfactants due to their low critical micelle 

concentration c M (Makino et al., 1973). 

An important consequence of the interactions between anionic 

surfactants and proteins is the swelling of the stratum corneum of the skin. 

Hydrophobic interactions between surfactant chains and the protein result 

in pendant ionic head groups and subsequently in swelling because of 

electrostatic repulsion between them. As the substrate matrix expands and 

the tertiary structure is disrupted, hydration occurs which leads to swelling 

(Blake-Haskins, 1986).


Figure 24.1 Zein solubility c z of saturated solutions as a function of surfactant 

concentration. Zein concentration, 50 g lit −1 , mixing period, 2 h, temperature, 40°C 

( , sodium dodecyl sulphate; , alkyl ether sulphate (2EO); , nonyl phenol 

ethoxylate (9EO)). 

Interactions with enzymes 

Surfactants which are capable of massive cooperative binding, such as 

many anionic and cationic surfactants, induce conformational changes in 

the protein molecule which in general lead to loss of biological activity. 

The following mechanisms of enzyme inactivation by surfactants have to 

be considered (Ne’eman et al., 1971): 

1. Disruption of the quaternary structure of the enzyme when the enzyme 

protein consists of several subunits.

2. Induction of conformational changes in the tertiary or secondary 

structure of the enzyme protein. 

3. In the case of membrane-bound enzymes, separation of the enzyme 

protein from essential membrane lipids. 

4. Binding at active sites of the enzyme. 

While the effect of cationic surfactants on membranes is comparable to 

that of anionic surfactants, many proteins are obviously more resistant 

towards the denaturing activity of .cationic surfactants (Nozaki et al., 

1974). Binding of tetradecyl trimethyl ammonium chloride onto bovine 

serum albumin and other proteins is comparable to that of sodium dodecyl 

sulphate. However, the cooperative binding with subsequent denaturation 

requires a ten-fold higher concentration of cationic surfactant. The 

saturation of the surfactant/protein complex is prevented by the competing 

formation of surfactant micelles. Contrary to the irreversibly denaturing 

effect of sodium dodecyl sulphate, the effect of some cationic surfactants on 

proteins is reversible (Nakaya et al., 1971). 

Local effects 

Skin compatibility 

W.STERZEL 343 

The damaging effects of surfactants on skin manifest themselves in dryness, 

roughness and scaling. In addition, symptoms of inflammation (reddening, 

swelling) can develop, which can result, in severe cases, in complete 

destruction of the tissue. All these symptoms are a result of the described 

biochemical properties of surfactants. The skin is defatted by the more or 

less pronounced property of the surfactants to emulsify lipids and thus 

partially or completely removing the surface film of lipids. This leads to a 

disturbance of the barrier function of the skin resulting in increased 

permeability for chemical substances and a loss of water. Anionic 

surfactants can cause swelling of the skin. As a result, they facilitate the 

transport of substances to lower layers where inflammation reactions can be 

induced (Scholz, 1967). The reaction of surfactants with proteins dissolves 

proteins out of the skin and leads to their denaturation. These changes in 

the matrix material have an effect on the resistance of the skin (Götte, 

1967) and, along with degreasing and drying, are an additional cause of an 

increase in skin roughness (Imokawa, 1975). 

The majority of the knowledge about skin compatibility of surfactants 

originates from studies with experimental animals, preferably rabbits. 

Furthermore, it is possible to evaluate new substances directly on human 

skin after careful exclusion of unreasonable risks. A critical overview of 

different test methods is given by Kästner (1980). In this context, the


problem of labelling chemical products with respect to their toxicological 

properties has to be addressed. Different national or international 

regulations, e.g. the EG directive 67/548 within the European Community, 

dictate that these products are labelled ‘irritating’ or ‘corrosive’ whenever 

exactly defined effects are observed in appropriate tests. With consumer 

protection in mind, exceedingly stringent test procedures have been 

established. These conditions frequently result in an unfavourable 

classification especially for surfactants. When interpreting data from such 

studies, it is important to consider that unrealistic conditions of exposure 

were involved. 

Since anionic surfactants are the class with the greatest economic 

importance, they are the best studied. No general statement is possible with 

regard to a classification of the various groups of anionic surfactants in 

order of their skin compatibility, since within each class of substances 

significant differences exist in their effect on skin depending on the 

respective structure. Opdyke et al. (1965), for example, found a decrease in 

the skin irritation potential of different alkyl ether sulphates with 

increasing level of ethoxylation. The effect of the alkyl chain length of 

anionic surfactants was examined in different test models for soaps, alkyl 

sulphates, alkyl sulphonates, alkylbenzene sulphonates as well as alphaolefin 

sulphonates (Kästner, 1980). As shown in Table 24.1, it could be 

established in all cases that compounds with a saturated side chain of 10– 

12 C atoms exert the largest effect, or rather, have the highest potential for 

damage. When the results of skin compatibility tests for the most 

important classes of anionic surfactants are summarized, it becomes 

evident that the undiluted products have to be regarded as strongly 

irritating substances. Even at concentrations of 10 per cent moderate to 

strong effects have to be expected. However, at concentrations less than 1 

per cent, which is the range corresponding to typical use levels in 

detergents, only minimal irritation is observed. 

Nonionic surfactants have a good skin compatibility at normal use 

levels. Although studies with alcohol ethoxylates were reported in which a 

strong irritant effect was observed (Grupp et al., 1960), these studies used 

concentrations far above the usual exposure levels of consumers. 

Independent of their structure, cationic surfactants cause severe skin 

damage in high concentrations, while typical application levels are 

generally tolerated well. 

Mucous membrane compatibility 

When talking about mucous membrane compatibility one has to consider 

not only the mucous membranes of the eye. In addition, the mucous 

membranes in the mouth, upper and lower gastrointestinal tract as well as 

the urogenital tract have to be considered. In general, the effects of

Table 24.1 Structure/activity relationships of anionic surfactants 

W.STERZEL 345 

a Test model: A=epicutaneous, mouse; B=intracutaneous, mouse; C=epicutaneous, 

man; D=roughness of skin; E=swelling of collagen in vitro; F=denaturation of 

protein in vitro. 

b Number of carbon atoms in the alkyl chain. 

surfactants on mucous membranes are based on the same biochemical 

mechanisms that are described in the chapter on skin compatibility. Special 

characteristics in the fine structure of mucous membranes, like the absence 

of keratin, result in a significantly higher sensitivity of these tissues towards 

chemical substances. Irritating materials affecting the eye cause reddening 

through increased blood flow in the conjunctivae with enlargement of the 

blood vessels. This can finally lead to the destruction of the cell walls 

accompanied by bleeding. Depending on the severity of the effects, a more 

or less pronounced swelling or reflex-induced closure of the eyelid will 

occur, followed by tearing and secretion. If the degree of irritation is low, 

epithelium damage develops on the cornea which can be visualized only 

with special techniques (staining, slit lamp microscope) and which is 

generally reversible. In severe cases the effects result in irreversible clouding 

of the cornea and therefore lead to an impairment of the eyesight. 

The classical method for the evaluation of mucous membrane 

compatibility of chemicals is the so-called Draize test on the rabbit eye 

(Draize et al., 1944). A structure/activity relationship with respect to the 

length of the respective alkyl chains of anionic surfactants can, as for the 

skin compatibility, also be observed for the mucous membrane 

compatibility (Kästner, 1980). According to this, the maximum irritation 

occurs at chain lengths of C 10−14. for n-alkyl sulphates as well as for nalkyl 

sulphonates. Although the irritation potential of the different 

surfactant classes extends over a large range, it can be concluded that the 

mucous membrane compatibility decreases in the following order:


nonionic>anionic>cationic surfactants (Draize and Kelley, 1952; Hazleton, 

1952; Grant, 1962). 

Sensitization 

Aside from acute irritation, chemical substances can cause allergies after 

contact with the skin or a mucous membrane. The development of an 

allergy is dependent on certain preconditions. An essential factor is the 

individual disposition which is predominantly genetically determined. An 

additional important point is the extent of damage to the tissue at the point 

of contact of the chemical substance (inflammation), which promotes 

sensitization. In addition, the sensitization potential of a substance is of 

decisive importance. For products with low molecular weights, this 

potential is dependent on their chemical properties. Small molecules are by 

themselves not able to trigger a reaction of the immune system. They 

become immunologically active only after binding to endogeneous 

proteins. Since the majority of the surfactants can only form weak and 

reversible bindings via hydrophobic and electrostatic interactions, this 

prerequisite is not fulfilled. 

Once the organism is sensitized towards a certain chemical, renewed 

contact with trace amounts of this material can provoke allergic reactions, 

which especially affect the skin and respiratory tract. Typical symptoms are 

itching, eczema, exanthema, rhinitis and bronchial asthma. 

Anionic surfactants and surfactant containing products were tested for 

sensitizing properties by numerous laboratories (Götte, 1967; Kästner, 

1980; Siwak et al., 1982) without detecting any significant increase in risk. 

The same holds true for nonionic surfactants (Siwak et al., 1982). Some 

cationic surfactants, which are able to form stable complexes by the 

formation of ion pairs with anionic groups of proteins, proved to be 

allergenic (Schallreuter and Wood, 1986). 

Toxicokinetics 

Percutaneous absorption 

The most important exposure of humans occurs through the skin with the 

use of cosmetics and toiletries. The skin comes in contact with surfactants 

also during dishwashing or when washing hands. Since these products are 

used over a long period of time, possible long-term effects must be 

evaluated. Measurement of percutaneous absorption of surfactants is 

important because it provides data for the toxicologist concerning the 

amount of surfactants which could enter the body through the skin in the 

most unfavourable case. Together with other toxicological information,

this allows a realistic evaluation of the risk when these compounds are 

used. 

Due to their economic importance, most studies have been carried out 

with anionic surfactants. Fewer studies exist for the other classes of 

surfactants. In vitro measurements of the percutaneous absorption of 

sodium dodecyl sulphate indicated a low absorption value for rat skin as 

well as for human skin (Blank and Gould, 1961; Embery and Dugard, 

1969; Howes, 1975). The low cutaneous absorption of sodium dodecyl 

sulphate could also be confirmed in experiments with rats (Greb, 1980). 

After application of a 0.7 per cent aqueous solution of sodium dodecyl 

sulphate (contact time 15 min), a cutaneous absorption of 0.26 µg cm −2 

within 24 h was measured (Howes,1975). 

In summarizing the results of the available studies, one can conclude that 

only small amounts of surfactants are resorbed through the intact skin. 

Since human skin in general is less permeable to chemicals (Rice, 1977; 

Wester and Maibach, 1982), the amounts of surfactants absorbed 

cutaneously in everyday use are probably even smaller. If the epidermis is 

removed completely or partially, e.g. damaged skin, the degree of 

absorption can increase substantially (Scala et al., 1968). In vitro studies 

demonstrated that cationic surfactants are absorbed by the skin to a much 

lesser extent than anionic surfactants (Scala et al., 1968; Geisler, 1976; 

Faucher et al., 1979). 

The degree of percutaneous absorption is generally larger for nonionic 

surfactants than for anionic or cationic surfactants. Studies on the 

percutaneous absorption of alkyl polyethyleneglycol ethers of the structure 

C 12-(CH 2-CH 2-O) 3H, C 12-(CH 2-CH 2-O) 6H , C 12-(CH 2-CH 2-O) 10H, and 

C 15-(CH 2-CH 2-O) 3H were performed under conditions of use (Black and 

Howes, 1979). The aqueous solutions of the applied surfactants were 

washed off after a contact time with the skin of 15 min. Under these 

conditions, the penetration of the alkyl polyethyleneglycol ethers was 

greater than the penetration of the analogous alcohol sulphates or alcohol 

ether sulphates. The penetration increased with increasing length of the 

carbon chain. Percutaneous absorption decreases for an ethylene oxide 

content of 6 moles or more in the ethoxylate moiety. 

Intestinal absorption, metabolism and excretion 

W.STERZEL 347 

The ingestion of surfactants is possible e.g. through the use of 

surfactantcontaining toothpaste, through residues from dishwashing 

detergents and through traces of surfactants in potable water. Anionic 

surfactants are resorbed well in the intestine (Michael, 1968; Black and 

Howes, 1980; Bartnik and Künstler, 1987). After absorption, a part of this 

is excreted together with bile in the faeces and is subject to a enterohepatic 

cycle. The majority of the absorbed surfactant is metabolized in the liver


and the respective metabolites are eliminated in the urine. The metabolic 

degradation of the linear alkyl chain is performed by -oxidation followed 

by β-oxidation. The ether-linkage in the ethoxylate portion of sulphated 

alcohol ethoxylates seems to be resistant to metabolism. 

Linear alkylbenzene sulphonates and branched alkylbenzene sulphonates 

are metabolized to short chain sulphophenyl carboxylic acids. N-alkyl 

sulphates are metabolized by -oxidation of the hydrophobic end followed 

by β-oxidation. Butyric acid-4-sulphate and acetic acid-2-sulphate are the 

end products, which are then further converted in small amounts nonenzymatically 

to sulphate and -butyrolactone (Ottery et al., 1970). Studies 

by Taylor et al. (1978) demonstrated that alkyl sulphonates are degraded 

via the same pathway as alkyl sulphates. 

Cationic surfactants can be assumed to be resorbed in the intestine only 

to a small extent. This was confirmed in a study with trimethyl cetyl 

ammonium bromide (Isomaa, 1975; Isomaa et al., 1976). Due to the low 

level of resorbed surfactant, an unquestionable identification of the 

metabolites was not possible. Parts of absorbed cationic surfactants were, 

as found for anionic surfactants, excreted together with bile in the faeces 

and to a lesser degree with the urine. 

Nonionic surfactants are resorbed to a large degree in the intestine 

(Drotman, 1980). A significant part of the material is eliminated with the 

bile. Cleavage of the ether linkage is obviously possible. Homologous 

ethyleneglycol ethers are probably generated as metabolites along with the 

corresponding carboxylic acids which are formed through oxidation of the 

terminal hydroxymethyl group (Drotman, 1980). The sorbitan fatty esters, 

which are often used as emulsifiers, and the ethoxylated fatty acid esters 

are hydrolysed in the gastrointestinal tract after oral administration 

through cleavage of the ester bond. While the resulting fatty acid is treated 

metabolically like a natural fatty acid, the polyol component of the 

sorbitan fatty acid is absorbed in the intestine, but is not further oxidized 

and is eliminated predominantly with the urine (Elworthy and Treon, 1967). 

Systemic effects 

Talking about systemic effects means, in contrast to local effects, the 

description of reactions arising after the substance has entered the organism 

after swallowing, skin penetration or inhalation. For surfactants, 

resorption through the skin has to be considered in particular. As described 

in the previous section, it is relatively small. But for products that 

frequently come into close contact with the skin, either unintentionally or 

due to their intended use, the resorption of very small amounts over a long 

period of time cannot be prevented.

Acute toxicity 

In general, the acute oral toxicity of surfactants is low. The LD 50 values 

typically range between several hundred and several thousand mg kg −1 of 

bodyweight. This is of the same order of magnitude as for table salt 

(Swisher, 1968). The most important effects are damage to the mucous 

membranes of the gastrointestinal tract. High doses induce vomiting and 

diarrhea (Weaver and Griffith, 1969). Surfactants exhibit significantly 

higher toxicity when the gastrointestinal tract is by-passed through 

intravenous injections. Even at very low concentrations, the interaction 

with the membrane of erythrocytes leads to their destruction. Inhalation of 

surfactant-containing dusts or aerosols in higher concentrations leads to 

disturbances of the lung function (Coate et al., 1978). This effect can be 

attributed to interactions with the surface active film that lines the vesicles 

of the lung (Kissler et al., 1981). As with local compatibility, there are also 

pronounced structure/activity relationships for acute toxicity. Gale (1953) 

has investigated the acute toxicity of sodium alkyl sulphates from C 8 to C 18 

and found the strongest effect for C 12 sulphate. 

The anaesthetic properties of certain alcohol ethoxylates which can be 

observed after intravenous application as well as after application to the 

skin or the mucous membranes are remarkable. Ethoxylates of unbranched 

primary alcohols with 9 ethylene oxide units were found to exhibit local 

anaesthetic properties starting with an alkyl chain of C 8. The activity 

increases with increasing chain length (Zipf and Dittmann, 1964). 

Chronic toxicity 

W.STERZEL 349 

In order to exclude any adverse effects arising from the repeated exposure 

against small amounts of surfactants over a prolonged period of time, 

representatives of all important classes of surfactants were examined for 

chronic toxic effects. In these tests, dosages of several thousands ppm were 

administered over a period of up to 2 years. No observable effects were 

detected with linear alkylbenzene sulphonates in 2 year studies with rats 

using concentrations up to 0.5 per cent (feed) or 0.1 per cent (drinking 

water) (Buehler et al., 1971). A sodium alkyl sulphate with an average 

chain length of C12 was tolerated by rats up to 1 per cent in the feed for 1 

year without any remarkable side effects (Fitzhugh and Nelson, 1948). C14 −16α-olefin sulphonates were applied over 2 years in a feeding study in 

dosages up to 0.5 per cent without causing any remarkable effect (Hunter 

and Benson, 1976). Analogous studies were reported for alcohol 

ethoxylates and alkylphenol sulphates, which revealed no toxic symptoms 

at doses up to 0.1 per cent and 1.4 per cent, respectively (Larson et al., 

1963; Siwak et al., 1982). Studies on cationic surfactants reported a noobservable-effect-level 

of 0.25 per cent (Coulston et al., 1961). In all these


long-term studies, the dosages that were tolerated without damage were in 

the range of several thousand ppm, indicating large margins of safety. This 

was confirmed by Hunter and Benson (1976), who calculated for a 

relevant example that the respective dosage lies at least by a factor of 1000 

over the estimated maximum daily exposure level of humans. Besides these 

data from animal experiments, a series of studies exists in which volunteers 

ingested considerable amounts of anionic or nonionic surfactants over 

several weeks, without any noticeable severe adverse effects (Swisher, 

1968). 

Mutagenicity 

Mutagenicity is the induction of irreversible changes in genetic material. If 

normal cells (somatic cells) are the target, malformation results in 

the developing organism. In case of the mature organism, it can lead to 

tumour formation. If germ cells are the target, the danger exists that the 

genetic defect will be passed on to the offspring. All classes of surfactants 

have been evaluated in numerous test systems. The collected data allow the 

conclusion that surfactants pose no considerable risk of genetic damage 

(Yam et al., 1984; Fowler, 1988; Oba and Takei, 1992). 

Carcinogenicity 

Due to the widespread use and contact with surfactants the question of 

irreversible damage has to be raised in addition to the problem of other 

chronic effects. The following compounds were evaluated for 

carcinogenicity after administration in the drinking water or feed: 

alkylbenzene sulphonate (Buehler et al., 1971), alkyl sulphates (Fitzhugh 

and Nelson, 1948), α-olefin 

sulphonates (Hunter and Benson, 1976), secalkane 

sulphonate (Quack and Rend, 1976), alcohol ether sulphates 

(Tusing et al., 1962; Siwak et al., 1982), alcohol ethoxylates (Siwak et al., 

1982) and alkylphenol ethoxylates (Larson et al., 1963; Smyth and 

Calandra, 1969). None of these experiments provided any indication of 

increased risk of cancer after oral ingestion of surfactants. The question of 

possible carcinogenic effects of surfactants on the skin has also been 

studied extensively. Summaries exist by Oba and Takei (1992) and Siwak et 

al. (1982). 

Embryotoxicity 

The effects of substances on the organism during pregnancy can lead to 

delayed development or death of the embryo or malformation. Studies with 

the following surfactants revealed no indications of embryotoxic activity: 

alcohol ethoxylates (Nomura et al., 1980), α-olefin 

sulphonates (Palmer et

al., 1975), alcohol ether sulphates and linear alkylbenzene sulphonates 

(Nolen et al., 1975). Concerns which started with the publication in 1969 

(Mikami et al., 1969) that surfactants had caused malformations in animal 

studies could not be reproduced (Oba and Takei, 1980). The findings of 

Mikami et al. (1969) were interpreted to be a result of methodical 

inadequacies and misinterpretations (Charlesworth, 1976). 

Summary 

Due to their physico-chemical properties, surfactants are capable of 

reacting with biological membranes, proteins and enzymes. Most of their 

toxicological properties can be traced back to these interactions. During 

the application of surfactant-containing products, the most important 

aspect of consumer safety is local compatibility. No indications of 

systemic, chronic or irreversible damage could be found. Estimates of the 

amounts of orally ingested surfactants typically encountered were reviewed 

by several authors. Based on these estimates, a total daily intake of 

surfactants in the range of 0.3–3 mg per person was calculated by Swisher 

(1968). Due to the low rate of percutaneous absorption exposure through 

the skin can be neglected. If the above mentioned highest conceivable daily 

intake is compared with the dosage that was tolerated without adverse 

effects in studies concerning systemic effects, it becomes quite clear that 

these amounts can be regarded as harmless. In conclusion, it can be stated 

that the use of surfactants does not pose a health risk for humans. 

References 

W.STERZEL 351 

BARTNIK, F.G. and KÜNSTLER, K., 1987, Biological effects, toxicology and 

human safety, in Falbe, J. (Ed.). Surfactants in Consumer Products, Theory, 

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354

PART SEVEN 

Controversial mechanistic and regulatory 

issues in the safety assessment of 


25 

Low Dose of a Genotoxic Carcinogen does not 

‘Cause’ Cancer; it Accelerates Spontaneous 

Carcinogenesis 

WERNER K.LUTZ 

University of Würzburg, Würzburg 

Definitions of cancer risk 

The risk of cancer is normally expressed as a fraction of a population 

diagnosed with cancer within a specified period of time. In an animal 

bioassay for carcinogenicity, this period usually is 2 years; in cancer 

epidemiology, a life span of 65 years (0–64) is often used. The two periods 

can be considered equivalent with respect to the process of carcinogenesis: 

its rate in different species is inversely correlated with the natural life span 

and basal metabolism and the background cancer incidence in 2-year old 

rats or mice is very similar to the one seen in 65-year-old humans 

(Anisimov, 1989; Raabe, 1989; Tennant, 1993). 

For an individual, a cancer risk can only be 0 or 1, depending on 

whether the situation is analysed before or after the diagnosis of the 

tumour. The population-based expression of a cancer risk therefore is not 

easily visualized and does not take into account interindividual differences 

in susceptibility. 

In the following discussion, the dose-response relationship in chemical 

carcinogenesis is analysed in terms of an effect of a carcinogen on the 

individual tumour latency time (Kodell et al., 1980; Littlefield et al., 1980; 

Day, 1983; Gaylor, 1992). Together with the idea of background DNA 

damage responsible for what is considered ‘spontaneous’ tumour formation 

and including individual variability for the rate of this process, it will be 

shown that low doses of genotoxic carcinogens might accelerate the 

spontaneous process of carcinogenesis but are not expected to induce 

cancer ‘out of the blue’. 

Linear dose response for a DNA-reactive carcinogen 

The effect of a carcinogen at low dose is normally extrapolated from data 

obtained in 2-year bioassays. At the end of the 2-year treatment period, the 

surviving animals are killed and analysed for the presence of tumours. The

fraction of tumour-bearing animals is then plotted against the dose, and 

lowdose effects are estimated from some model curve fitted to the data 

points. 

The shape of the dose-response curve at the low-dose end is strongly 

debated, especially for nongenotoxic carcinogens. For DNA-reactive 

carcinogens, it is widely accepted that there is no dose without effect, and a 

linear extrapolation is used (Lutz, 1990b). This is based on the idea that 1 

molecule of a DNA-reactive carcinogen could form a dangerous DNA 

adduct in a critical gene and activate an oncogene or inactivate a tumour 

suppressor gene by mutation, if the adduct is not repaired before DNA 

replication. 

Table 25.1 shows the consequences of linear interpolation between the 

tumour incidence in the controls and in a dosed group of a bioassay for 

carcinogenicity. An organ-specific tumour incidence of 4 per cent in the 

control group (2/50) and 14 per cent (7/50) after treatment for 2 years at 

10 mg kg −1 per day is assumed. With linear interpolation, a treatmentrelated 

increment in tumour incidence of 1 per cent would be calculated 

per mg kg −1 per day so that at 1 mg kg −1 per day, a 5 per cent total tumour 

incidence would be expected. 

In humans, increments in cancer risk in the per cent range would not be 

acceptable. A risk of 1 in 1 million lives might be considered negligible and 

the respective exposure could be regarded as ‘virtually safe.’ With the 

example given in Table 25.1 and upon linear interpolation, this ‘virtually 

safe’ dose would be calculated as 0.0001 mg kg −1 per day. 

What does it mean: ‘1 additional tumour in 1000 000 

lives’? 

The fact that a cancer risk is only 10 −6 cannot be a consolation for the 

affected individual. For this person, the cancer risk was 1. In the public 

opinion, therefore, an increase by one tumour case per one million lives is 

often interpreted to mean that one additional individual has got cancer 

who could otherwise have lived a much longer tumour-free life. For 

reasons explained below, this fear appears unfounded. 

Endogenous DNA damage; individual susceptibility 

W.K.LUTZ 357 

Carcinogenesis is a multi-stage process based on the accumulation of a 

number of critical DNA-related changes, due to, for example, DNAcarcinogen 

adducts. Evidence of background DNA damage from 

endogenous and unavoidable substances is accumulating (Ames, 1989; 

Loeb, 1989; Lutz, 1990a). It is due to, for instance, electrophiles such as 

S-adenosylmethionine, epoxides, quinones, or aldehydes, or to reactive 

oxygen species. In addition, DNA is not a chemically stable molecule, it

358 CARCINOGENESIS AT LOW DOSE 

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

bioassay for carcinogenicity 

Notes: 

Data in boldface; extrapolations in italics. 

depurinates and deaminates spontaneously. Finally, DNA replication is not 

100 per cent correct so that mutations cannot be avoided completely. The 

resulting background DNA damage is responsible for spontaneous 

mutations and for what is called spontaneous tumour formation. The 

process of carcinogenesis therefore has a non-zero rate even if exposure to 

exogenous DNA-reactive carcinogens could be avoided. 

The level of the background mutation rate is expected to show 

interindividual variability. It depends both upon genetic and life-style 

factors which govern, for instance, enzyme activities responsible for 

carcinogen metabolism or DNA repair (Harris, 1989). The rate of 

spontaneous carcinogenesis is further governed by the inherited and 

acquired presence of activated oncogenes or absence of tumour suppressor 

genes (Scrable et al., 1990). Therefore, each individual in a heterogeneous 

population is expected to have its own endogenous cancer risk expressed as 

an individual time-to-tumour or tumourfree lifetime. 

Exogenous DNA damage; acceleration of spontaneous 

carcinogenesis 

Exposure to an additional, exogenous DNA-reactive molecule adds to the 

background DNA damage, increases the probability of a mutation and 

accelerates the multi-stage process of carcinogenesis. At low doses of the 

exogenous carcinogen, the rate of the process is expected to be dominated 

by the background damage so that the exogenous factor cannot constitute 

a cancer risk independent of the spontaneous process. The acceleration 

must be dose dependent and might be related to the background rate 

operating in each individual. 

It is true, therefore, that even a few molecules of a DNA-reactive 

carcinogen can have an effect. However, this effect cannot be a tumour

‘out of the blue’, in an individual that would otherwise have a low cancer 

risk. It ‘only’ reduces the individual’s tumour-free lifetime. 

No cancer ‘out of the blue’ 

W.K.LUTZ 359 

Figure 25.1 Schematic representation of the time course of tumour appearance in a 

group of individuals with large differences in susceptibility. Solid line: background 

process of spontaneous carcinogenesis; arrows: acceleration of the spontaneous 

process by exposure to an additional carcinogen. 

This interpretation does not contradict the understanding that a low dose of 

a carcinogen could increase the tumour incidence from 40000 to 40001 per 

1000000 lives, in the example shown in Table 25.1. The connection 

between the two approaches is shown in Figure 25.1. The solid line shows 

the appearance of a spontaneous tumour in individuals of a group of 20 

people. At the age of 65 years, 4 individuals have a tumour diagnosed. This 

is equivalent to a cumulative tumour incidence of 20 per cent. Exposure of 

this group to an additional exogenous carcinogen would result in some 

reduction in the tumour-free lifespan in all individuals. In the example 

shown in Figure 25.1, this shift would move one additional individual to 

an age of diagnosis

360 CARCINOGENESIS AT LOW DOSE 

years. This is equivalent to an increase from 40000 to 40001 as a 

cumulative incidence 0–64, but it has a completely different meaning. It can 

now be excluded that the additional individual would have lived tumour 

free for 80 years in the absence of the exogenous carcinogen. 

Final remarks 

With the ideas presented, fear of cancer from low dose or from rare 

exposures can possibly be reduced. This does not mean that small cancer 

risks should be tolerated. Carcinogens in the environment, for instance, 

affect all of us; the tumour-free life span is reduced in the entire 

population. 

The model should be valid for tissues with a high spontaneous tumour 

incidence and an exponentially steep age dependence indicative of a 

multistage requirement of 5–6 steps. For cells that can be transformed in 2 

or 3 steps or that are specifically sensitive in certain phases of the 

development (in utero or during childhood), the model has to be 

reconsidered. This might be necessary for tumours with incidence peaks at 

a young age (leukaemia, tumours of the lymphatic tissues, brain, testis). 

Nevertheless, the latter tumour types are rare in comparison with cancer of 

the old age so that the concept should hold for the majority of the tumours 

in humans. 

References 

AMES, B.N., 1989, Endogenous DNA damage as related to cancer and aging, 

Mutat. Res., 214, 41–6. 

ANISIMOV, V.N., 1989, Dependence of susceptibility to carcinogenesis on species 

life span, Arch. Geschwulstforsch., 59, 205–13. 

DAY, N.E., 1983, Time as a determinant of risk in cancer epidemiology: the role of 

multi-stage models, Cancer Surv., 2, 577–93. 

GAYLOR, D.W., 1992, Relationship between the shape of dose-response curves 

and background tumour rates, Regul. Toxicol. Pharmacol., 16, 2–9. 

HARRIS, C.C., 1989, Interindividual variation among humans in carcinogen 

metabolism, DNA adduct formation and DNA repair, Carcinogenesis, 10, 

1563–6. 

KODELL, R.L., FARMER, J.H., LITTLEFIELD, N.A., FRITH, C.H., 1980, Analysis 

of life-shortening effects in female Balb/c mice fed 2-acetylaminofluorene, J. 

Environ. Pathol. Toxicol, 3 69–88. 

LITTLEFIELD, N.A., FARMER, J.H. and GAYLOR, D.W., 1980, Effects of dose 

and time in a long-term, low-dose carcinogenic study, J. Environ. Pathol. 

Toxicol., 3, 17–34. 

LOEB, L.A., 1989, Endogenous carcinogenesis: molecular oncology into the 

twentyfirst century—Presidential address, Cancer Res., 49, 5489–96.

W.K.LUTZ 361 

LUTZ, W.K., 1990a, Endogenous genotoxic agents and processes as a basis of 

spontaneous carcinogenesis, Mutat. Res., 238, 287–95. 

LUTZ, W.K., 1990b, Dose response relationship and low dose extrapolation in 

chemical carcinogenesis, Carcinogenesis, 11, 1243–7. 

RAABE, O.G., 1989, Scaling of fatal cancer risks from laboratory animals to man, 

Hlth Phys., 57 suppl. 1, 419–32. 

SCRABLE, H.J., SAPIENZA, C. and CAVENEE, W.K., 1990, Genetic and 

epigenetic losses of heterozygosity in cancer predisposition and progression, 

Adv. Cancer Res., 54, 25–62. 

TENNANT, R.W., 1993, Stratification of rodent carcinogenicity bioassay results to 

reflect relative human hazard, Mutat. Res., 286, 111–18.

26 

Controversial Mechanistic and Regulatory Issues 

in Safety Assessment of Industrial Chemicals —an 

Industry Point of View 

HEINZ-PETER GELBKE 

BASF AG, Ludwigshafen 

Introduction 

In the appropriate classification and risk assessment of industrial chemicals 

three main players are involved: the scientific community, regulatory 

authorities and the chemical industry. The rules of the game are given by 

the definitions of the classification criteria and by guidelines for the risk 

assessment process. These definitions and guidelines are sometimes very 

flexible and open to different interpretations but in other cases precisely 

defined. Mostly these rules have been set up by regulatory bodies, e.g. the 

EU or national authorities, but sometimes also by scientific committees like 

the MAK commission in Germany or by IARC. 

Looking now at these three main players, the scientific community will 

provide the data as the starting point for each individual chemical. Possible 

controversies centre around the question, how data gaps may be bridged by 

scientifically valid assumptions. This indeed may often be necessary, if a 

complete toxicological and mechanistic data base is not available. 

Nevertheless, a scientifically based consensus can often be achieved for this 

bridging process. 

On the other hand, the other two players—industry and regulatory 

authorities—often do not reach mutually agreed decisions, although both of 

them finally strive for the same target: protection of human health and the 

environment in an industrialized community. Controversial issues may 

sometimes stem from different approaches for bridging the scientific data 

gaps, but mostly they arise from political, economic or social aspects, 

which often are not outspoken. Just to give some examples: anticipated 

reaction of the society, possible emotions of the consumer, possible 

influences on the next election, availability of technical alternatives, 

different perceptions of the risk-benefit balance, overall economic situation 

of the community, different evaluations in other countries, impact on 

worldwide competitiveness, etc. 

In the following an industry point of view will be presented for some 

specific problems of today in the area of classification which is a more

qualitative approach and for risk assessment which in addition has to take 

into account quantitative aspects. This will be discussed separately for 

toxicological effects with thresholds (‘classical’ organ toxicity, reproductive 

and developmental toxicity) and without thresholds (mutagenicity, 

carcinogenicity). Apart from this subdivision there is one general concern 

of industry, that is to appropriately take into account exposure. 

Exposure 

Every classification and especially risk assessment decision should not only 

be based qualitatively on the toxicological profile, but it should also take 

into account quantitatively the toxicological dose-response relationship as 

compared to human exposure. 

As a first rough approximation the different exposure profiles may be 

grouped into four main categories: 

1. 

Exposure during chemical production 

Many high production volume chemicals are used mainly or even 

exclusively as intermediates within the chemical industry. Although large 

amounts of these materials may be produced or processed within only a few 

facilities, exposure is often quite low and can be controlled or reduced by 

technical means. In addition there are many specific features enabling an 

efficient exposure control, such as: a trained workforce, site-specific and 

personal protection devices, stringent surveillance of workforce and work 

procedures, medical programmes tailored to the specificities of the work 

place, well defined exposures at the specific work sites, the exposed 

population is well known and limited, specified exposure duration, 

relatively homogeneous age and better health status of the workforce as 

compared to the general population, etc. 

2. 

Exposure of the downstream user during industrial/ 

manufacturing applications 

H.-P.GELBKE 363 

In principle, the exposure scenarios may be quite similar to those of 

chemical production, since many of the features described above also relate 

to or may be implemented at smaller workshops of the downstream user. 

Unfortunately, in reality often quite high exposures prevail in small 

workshops, possibly due to limited expenditures into exposure reduction 

measures or to a workforce not specifically trained for handling of 

dangerous chemicals.

364 CONTROVERSIAL ISSUES IN SAFETY ASSESSMENT 

3. 

Exposure of the consumer 

When looking at the vast array of chemicals in use today, most of them 

will have industrial applications and only relatively few will directly be 

used by consumers. Especially highly reactive chemicals with their inherent 

potential for health hazards will generally not enter into consumer 

application fields simply due to their limited stability. But if a chemical gets 

to the consumer, efficient control measures can hardly be implemented, be 

it for the exposure per se, the exposed population, appropriate handling or 

prevention of misuse. 

4. 

Exposure of the general population via the environment 

Apart from highly reactive substances, all chemicals will to some extent 

enter the environment depending on their application fields and their 

processing to other end products. The environmental concentrations are 

determined by the amount released, by the distribution media, local 

situations and the efficacy of the different degradation processes. In 

general, the exposure of the population via the environment will be very 

low as compared to the workplace and health hazards are not to be 

expected. Of course highly persistent and highly toxic substances can be an 

exemption to this rule and should be monitored carefully. 

These exposure scenarios can be exemplified by a textile dyestuff: 

manufacturing within the chemical industry makes use of various starting 

materials and intermediates, which will not end up—apart from minute 

impurities—in the final product, and downstream exposure to these 

materials will be negligible. The manufacture of the dyestuff, its further 

handling, processing and formulation within the chemical industry can 

easily be controlled. But when this material is used downstream in textiledyeing 

workshops an efficient exposure control and appropriate handling 

may not always be guaranteed and higher exposures are conceivable. 

During the dyeing process, parts of the material enter into the 

environment, for example into aqueous media. This might lead to an 

exposure of the general population, albeit at very low concentrations. 

Consumer exposure can occur via migration of the dyestuff from the 

textile, sweat being the carrier medium or for small children the saliva, but 

exposure will most often be so low that a health hazard is not to be 

expected. 

It is one of the main concerns of industry within the processes of 

classification and risk assessment that sufficient consideration is often not 

given to the different exposure profiles of each chemical. This results in 

simplified black and white decisions, which are not very helpful for an

appropriate and cost effective health protection within our industrialized 

world. 

Threshold effects 

Classification 


The classification system of the EU resides in the designation of 

appropriate R-phrases. In most cases they are rather meant for hazard 

identification (e.g. irritation, sensitization, carcinogenicity) and not so 

much for risk characterization. Sometimes, risk aspects are also involved 

when specific dose levels are decisive for a specific R-phrase (e.g. acute 

toxicity). Thereby not only the effect per se is taken into account but also 

the strength of the effect. 

This latter principle also applies to the R 48-phrase (‘danger of serious 

damage to health by prolonged exposure'), if severe (irreversible) toxic 

effects are observed at dose levels of ≤50 mg kg −1 body weight per day in a 

90-day test after oral administration. For other routes and durations of 

exposure similar provisions exist. The strategy to use dose limits is 

certainly appropriate for threshold effects. 

Although for reproductive and developmental effects thresholds are also 

accepted in most cases, no such dose limits are given apart from 1000 mg 

kg −1 body weight per day. This dose limit is not equivalent to that for the R 

48-phrase, because it only stems from the limit dose levels of the test 

guideline and is higher by more than one magnitude. Thus, for 

reproductive/ developmental toxicity classification, exposure and risk 

considerations are not influential in contrast to the R 48. Such a simplified 

‘yes or no’ approach for a threshold effect is not justifiable neither from a 

scientific point of view nor for an appropriate health protection. This is 

even more so, if the proposal for the ‘Restrictions on Marketing and Use 

Directive’ (13th Amendment to Directive 76/769/EEC) is implemented 

calling for a general prohibition of category 1 and 2 reproductive/ 

developmental toxicants in consumer products in concentrations of ≥0.5% 

—if no specific concentration limit has been accepted according to the 

preparation guideline (Commission Directive 93/18/EEC; Council Directive 

88/379/EEC). 

The inconsistency of the approaches for ‘classical’ organ toxicity and 

developmental/reproductive toxicity can easily be demonstrated by the 

following theoretical example: 

For neurotoxicity a LOEL of 40 mg kg −1 day −1 in a 90-day oral test will 

result in a R 48-phrase, but a LOEL of 80 mg kg −1 day −1 would not 

lead to classification. On the other hand, slight foetal weight


reduction or impairment of fertility observed at 800 mg kg −1 day −1 in 

anappropriate developmental or reproductive test without parental 

toxicity would lead to classification into the respective category 3 or 

possibly category 2, even if a NOEL was found at 400 mg kg −1 day −1 . 

This certainly is not appropriate when considering the reversibility 

and severity of the effects and the differences in the NOELs and 

LOELs of one order of magnitude. 


The general strategy in risk assessment of chemicals with threshold effects 

is to use ‘assessment’ factors (‘uncertainty’ or ‘safety’ factors) (AF) for 

setting appropriate exposure limits. This concept was originally introduced 

by the WHO to establish ADI values (acceptable daily intake) for pesticide 

residues in food. Here generally a ‘safety’ factor of 100 is applied to the 

NOEL of a chronic experiment; higher or lower factors might be used for 

specific effects or experimental conditions. 

This basic approach can be generalized from consumer exposure to 

pesticide residues in food to other chemicals and exposure scenarios. The 

main problem then will be the selection of an appropriate AF. First of all, 

there are some general considerations to be taken into account: 

– Should AFs for the workforce and the general population differ because 

of the age characteristics and the general health status of workers? 

– Should different AFs be selected for chemicals and pesticides, taking into 

account that pesticides are specifically tailored for biological activity? 

– What are suitable AFs for route to route extrapolations if the 

experimental exposure does not correspond to that of humans? Thereby 

metabolic firstpass effects in the liver and different efficacies of the 

adsorption barriers of skin, lung and the intestines have to be taken into 

account. 

– What is the appropriate dose parameter, mg kg −1 body weight, mg m −2 

surface area or concentration? 

– What are suitable AFs for developmental effects which may occur in 

principle after a single exposure and may lead to irreversible lifetime 

impairment? 

– Are specific AFs necessary for toxic effects on the reproductive organs as 

compared to toxic effects on other organ systems? 

Apart from these general considerations there are also specific criteria 

decisive for the selection of AFs depending on each single chemical, its 

total data base and the experimental details. Very often the final AF is 

obtained by additional default factors which are to account for


experimental insufficiencies or to bridge data gaps. Just to give some 

examples for such specific considerations: 

– reversible versus irreversible effects, 

– duration of the study, 

– NOAEL versus NOEL, 

– LOAEL versus NOAEL, 

– local versus systemic effects, 

– species-specific effects, 

– species differences in anatomy or physiology, 

– similar versus different results observed in experiments with various 

species, 

– biokinetics and metabolism (e.g. metabolic pathways are species specific 

or occur only at high doses), 

– structure-activity considerations. 

This listing certainly not being complete clearly demonstrates that 

appropriate AFs cannot be arrived at by a simple cook-book procedure, 

but a flexible case-by-case approach is required for each individual 

chemical and data set. This is extremely important in order to avoid overconservative 

AFs; and in the long run over-conservative risk assessments 

are just as prohibitive for an appropriate health protection in an 

industrialized world as an underestimation of risk may lead to a more 

immediate danger to health. Over-conservative risk evaluations will result 

in a wrong allocation of resources, an unjustified prohibition of valuable 

chemicals, a wrong or unnecessary selection of alternative materials. etc. 

What might be an indication for an over-conservative AF? In principle, 

AFs for threshold effects should then be questioned to be over-conservative 

if they lead to acceptable human exposures which would also be 

appropriate for non-threshold effects (e.g. carcinogenicity) This can be 

exemplified by the following consideration: 

For a carcinogenicity experiment a ‘LOAEL’ in classical terms would 

be equivalent roughly to a dose just leading to a statistically increased 

tumour incidence of about 5 per cent. A ‘virtual NOAEL’ in the same 

classical sense without a statistically significant increase could then be 

at a dose with an actual tumour incidence of 1 per cent, which will 

not show up as a substance related effect under usual experimental 

conditions. At such a dose level the extra tumour risk would be 1/100. 

Applying an AF of 1000 to this ‘virtual NOAEL' would result in an 

exposure level with a risk of 1/10 5 , and an AF of 10000 in one with a 

risk of 1/10 6 using a simple linear extrapolation without further 

default considerations. Exposure levels with a risk of 1/10 5 or 1/10 6 are 

under discussion as ‘virtually safe doses’ for the workforce or the


general population. Thus, AFs of >1000 should always be questioned 

as possibly over-conservative for threshold effects, since exposure 

levels thereby obtained could also be acceptable for non-threshold 

effects like carcinogenicity. The assumptions underlying such high AFs 

should be re-examined critically. 

In the light of these considerations AFs for a developmental toxicity of 

1000– 5000, as they are under discussion by some groups today, should 

also be questioned as possibly being over-conservative. It should not only be 

taken into account that developmental effects most often will have 

thresholds but also that their severities span a wide range from slight foetal 

weight impairment up to disabling malformations. 

In addition an unreflected selection of default factors in ‘classical’ organ 

toxicity can easily lead to over-conservative AFs: starting with a factor of 

10 each for inter- and intra-species variability yields the AF of 100 used for 

ADI-calculations. In addition the following default factors are sometimes 

proposed: 

– Extrapolation from subacute/subchronic exposure to chronic exposure: 

a default factor of 10. 

– Extrapolation to the NOEL, if only a LOEL was obtained: a default 

factor of 2–5. 

– Taking into account an inappropriate experimental design: a default 

factor of 2–5–10. 

Thereby, for a multiple dose study with a marginal effect at the lowest dose 

level and an experimental design not fully in accordance with today’s 

standards, these default factors would result in a final assessment factor of 

4000– 50000. It is highly questionable whether such an AF is really 

appropriate for threshold effects in comparison to the example given above 

for carcinogenicity. 

Non-threshold effects 

Other principles and approaches for classification and risk assessment have 

to be applied for non-threshold effects since safe exposure levels cannot be 

defined at which an adverse health effect will definitely not occur. Thus, 

for these compounds carcinogenic and mutagenic effects cannot be 

excluded even at very low dose levels albeit with extremely low 

probability.

Classification 


The present classification systems of scientific organizations (e.g. IARC, 

German MAK-commission) or regulatory agencies (e.g. EPA, EU 

commission) reside in quite a simplistic ‘strength of evidence’ approach: 

how valid are the experimental or epidemiological data? 

In future we should strive for a ‘weight of evidence’ approach which 

appears much more appropriate for a realistic human health protection, 

since it takes into consideration both risks and benefits of man-made and 

natural chemicals. Such a classification system basically asks the question: 

what do the experimental data really mean to humans at specific exposure 

levels? Thereby both qualitative and quantitative aspects are considered. 

Qualitatively whether and to what extent the mechanisms leading to an 

adverse effect in animals will also act in humans, and quantitatively to put 

the experimental dose-response relationship into context with human 

exposure. Of course, worst case exposure scenarios have to be taken into 

account. If under these considerations the experimental carcinogenic effect 

would not be relevant for humans a classification would not be 

appropriate. 

This latter quantitative aspect has led to discussions in several groups, as 

to whether a separate category for ‘weak carcinogens’ should be 

established, since more and more compounds turn out to be experimental 

carcinogens with a very weak or questionable effect. Such a category could 

be used for example for compounds which: 

– would only have insignificant effects even under worst case exposure 

scenarios, 

– did not give a carcinogenic response in appropriate animal experiments 

but are metabolized to carcinogenic intermediates or exert genotoxic 

effects in vivo, 

– show metabolic toxification to genotoxic metabolites only at high doses 

where the ‘normal’ metabolic detoxification pathway is overwhelmed. 

It is doubtful whether such a new category would really mean a step 

forward and be helpful. First of all why should compounds be classified if 

the carcinogenic effect is not to be expected in humans even under worst 

case exposure scenarios? And secondly how can the message of ‘weak 

carcinogenicity’ be brought over to the public without raising an emotional 

over-reaction to these compounds. 

Apart from these considerations on the general approach (‘strength’ or 

‘weight of evidence’) there is one specific major problem: presently, only 

criteria for classification are well defined, but those for non-classification 

are either not or only very vaguely described. It is also an important 

challenge to set up practicable and clear-cut non-classification criteria


which can be used by industry to tailor experiments in order to refute the 

classification of a questionable animal carcinogen. In the long run there 

will be no benefit if about 50 per cent of all chemicals are classified as 

carcinogens, neither for the public nor for industry nor for an adequate 

protection of human health. 


Most scientific committees and regulatory agencies refrain from 

scientifically based risk assessments for carcinogens but rather propagate 

an exposure as low as possible. And if a risk assessment is really carried 

out, it usually just applies a simplistic mathematical extrapolation using the 

linearized multistage model and highly conservative default assumptions to 

bridge data gaps. These mathematical procedures arrive at a scientifically 

unjustified numerical precision of the risk estimate. One of the problems is 

to explain to the public the real meaning and the uncertainties of such a 

risk assessment. A possible alternative could be to substitute the 

mathematical extrapolation by an appropriate assessment factor which of 

course has to take into account the severity and irreversibility of the 

carcinogenic effect. The simplistic mathematical modelling might be used 

only for selection of priority chemicals for further in-depth investigations. 

On the other hand, a mathematical risk assessment can be an 

appropriate procedure for chemicals with a broad experimental data base, 

when the most relevant default assumptions are substituted by real data. 

This would be of primary importance for: 

– the selection of the mathematical model: the simplistic linearized 

multistage model presently in use could be substituted by biologically 

driven models, like that proposed by Sielken (1989) or the MVK-model 

(Moolgavkar and Knudson, 1981; Moolgavkar et al., 1988). 

– dose scaling from animals to humans: presently the experimental dose in 

mg kg −1 body weight is often extrapolated to humans by transforming 

the dose to mg m −2 body surface. This is used both for compounds 

which are metabolically toxified and detoxified, the scientific basis for 

such an undifferentiated procedure is at best highly doubtful. In the 

future this default assumption could be substituted by physiologically 

based pharmacokinetic (PBPK) modelling. 

– estimation of the target dose: presently the external dose to which the 

animals are exposed is considered to be proportional to the dose 

reaching the target tissue or the target chemical entities—generally the 

DNA. Again in future this could be substituted by adequate PBPK 

modelling.

For the time being, there are only few compounds with an experimental 

data base broad enough to substitute these default assumptions in every 

respect. But there are important industrial chemicals—and their number 

will increase eventually—for which at least some data are available for a 

justified substitution of part of the default assumptions. And this should be 

done as far as possible when a mathematical risk assessment is carried out. 

Conclusions 

Problems of classification and risk assessment have been discussed 

separately for toxicological effects with thresholds (‘classical’ organ 

toxicity, reproductive and developmental toxicity) and without thresholds 

(mutagenicity, carcinogenicity). With regard to the general procedures of 

today, for industry there are two main points of concern: 

1. Not only for risk assessment but also for classification, exposure 

considerations should be taken into account. In principle, there are 

four different exposure scenarios: (a) for chemical production or within 

chemical industry, (b) for industrial application by downstream users, 

(c) for the consumer and (d) for the general population via the 

environment. For a given chemical the exposure may vary widely for 

the different scenarios, and there are many chemicals which are only 

used within chemical industry, which will never reach the consumer or 

which will enter into the environment only in minute quantities. 

Exposure estimates, including worst case scenarios, should be included 

in the processes of risk assessment and classification in order to avoid 

over-conservative results, which are not in the interest of adequate 

health protection. 

2. To get away from risk assessment procedures based on default 

assumptions which will often lead to over-conservative results; a case 

by case approach making use of all available data is scientifically far 

more appropriate. 

These problems have been elaborated for the: 


– classification of chemicals for reproductive or developmental toxicity 

within the regulatory framework of the EU, 

– selection of appropriate ‘assessment factors’ for chemicals with 

threshold effects, 

– classification and risk assessment of chemicals with non-threshold 

effects (especially carcinogens).


References 

MOOLGAVKAR, S.H. and KNUDSON, A.G., 1981, Mutation and cancer: a 

model for human carcinogenesis, Journal of the National Cancer Institute, 66, 

1037–52. 

MOOLGAVKAR, S.H., DEWANJI, A. and VENZON, D.J., 1988, A stochastic 

two-stage model for cancer risk assessment, I. The hazard function and the 

probability of tumour. Risk Analysis, 8, 383–92. 

SIELKEN, R.L.JR, 1989, Useful tools for evaluating and presenting more science in 

quantitative cancer risk assessments, Toxic Substances Journal, 9, 353–404.

Absorption of organic solvents 3–7 

Acceptable daily intake (ADI) 43, 168 

Acetylcholinesterase (AChE) 238 

N-acetyl-S-(5-ethoxy-l,2,4-thiadiazol-3yl-methyl) 

-L-cysteine (ET-MA) 26 

N-acetyl-β-glucosaminidase (NAG) 117 

Acid anhydrides 152–7 

Acrylamide 238 

Acrylonitrile 23, 313 

Adduct determination 81–7 

limitations of methods for 84 

Adduct formation of reactive 

compounds 74 

Adipate esters 223 

Airway hyperreactivity 128, 131 

Airway inflammation, animal models 

129 

Alachlor 208 

Alcohol ethoxylates, anaesthetic 

properties 349 

Alkaline phosphatase (ALP) 117 

Alkylating agents 76–9 

Alkyldeoxyguanosine adducts 186 

Allergic contact dermatitis 137 

Allometric equation 44, 46 

Allometric scaling 44–55 

Allometry 44 

Antioxidants 316–39 

Aromatic amines 74, 77 

Ascorbate 241 

Assessment factors (AFs) 366–71, 370 

Astrocytes 240 

ATP 78, 79, 159 

Axonal proteins 238 

Azo colorants 301 

carcinogenic 301–6 

Index 

Benzene 39 

Benzidene 301 

Benzo(a)pyrene (BP) 159, 186 

1,4-benzothiazines 64 

Benzotriazole-based light stabilisers 

322–29 

blood kinetics 323–29 

blood metabolites 323–29 

effects on rat dam and foetal liver 

329–4 

in vitro hydrolysis 323 

liver enzyme induction 327–31 

safety assessment 331, 334–7 

BGA, collaborative study 199 

Bioactivation mechanisms 11, 35–41 

Biocides 208 

Bioinactivation mechanisms 11 

Biological effect monitoring (BEM) 19 

Biological effects, determination of 189– 

6 

Biological monitoring (BM) 19–1 

definition 2 

organic solvents 2–3 

Biomarkers 

glutathione conjugation products as 

20–2 

of neurotoxicity 238–4 

Biotransformation 12 

Bisphenol A diglycidylether (BPADGE) 

83 

Body metabolic potential 48 

Body surface area (BSA) 46–8 

BP-DNA adducts 187 

Bromobenzene 63 

Bromo-diglutathionyl hydroquinones 

64

374 INDEX 

2-bromo-glutathionyl 64 

Bromohydroquinone 63–5 

o-bromophenol 63 

Bronchial challenge tests 149 

Bronchial hyperreactivity 151 

Bronchoalveolar lavage 117, 123 

1,3-butadiene 

PBPK/PBTK model 16 

physiologically based toxicokinetic 

modeling 31–3 

2-butoxyacetic acid (BAA) 172–80 

2-butoxyethanol, PBPK/PBTK model 

166–4, 172–80 

Calcium 245–9, 250, 267 

Cancer risk 

‘absolute’ 191 

assessment 188–6 

definitions 356 

management of 181 

Carbamazepine 322 

Carbendazim 285–8 

Carbon disulphide 238 

Carcinogen(esis) 179–6, 356 

azo colorants 301–6 

DNA-reactive 356 

dose-response relationship in 356 

epidemiological approaches to 

detect and identify 183–9 

genotoxic 181–8 

identification of 183–93 

peroxisome proliferation 224–30 

role of CYPs in activation and 

detoxication of 206 

spontaneous process 356, 357–3 

surfactants 350 

Carcinogenic potency 181 

Carcinogenic Substances Regulation 

301 

Catabolic metabolism 260 

Catalase 241 

Cell-mediated immune responses in 

chemical respiratory allergy 142–6 

Cellular nucleophiles 73 

Cerebral calcium accumulation 240 

Chemical pesticides, immunotoxicity of 

203 

Chemical respiratory allergy, cellmediated 

immune responses in 142–6 

Chlordecone 241 

Chlorinated solvents 223 

Chloroform 38–1 

Chromates 301 

Chrysotile asbestos fibres 

pulmonary toxic effects 116–30 

size-separation methods 118–3 

Classification systems 368–3 

Color Index 301 

Colorants (dyes and pigments) 301–10 

regulatory aspects (FRG) 304–7 

Cosmetics 208 

Cultured porcine thyrocytes 261 

Cyclophosphamide 286–90 

Cytochrome P-450 (CYP) 15, 60, 74, 

206–15, 318, 327, 332 

Cytokine products of murine Th 1 and 

Th 2 cells 141 

Cytokines 140 

Dangerous compounds 208 

1,2-DCV-Cys 29, 31 

DCV-G 29 

1,2-DCV-G 29, 31 

1,2-DCV-Nac 29 

2,2-DCV-Nac 29 

Dermal uptake of organic solvents 5–7 

Detoxication, enzymes involved in 60–4 

Developmental neurotoxins 247–2 

Diagonal radioactive zones (DRZ) 158 

Diarylide pigments 81–3 

1,2-dibromoethane 40, 65–8 

3,3′-dichlorobenzidine (DCB) 81–3, 

301, 303 

1,2-dichloroethane 65–8 

2,5-dichloro-3-(glutathion-S-y1) 

hydroquinone 64 

Dichloromethane 39 

1,3-dichloropropene (DCP) 21–3 

S-(l,2-dichlorovinyl)glutathione. See 1, 

2-DCV-G 

di-(2-ethylhexyl)adipate (DEHA) 223, 

224, 227, 229–4 

di-(2-ethylhexyl)phthalate (DEHP) 223, 

224, 227–4, 286, 287

5α-dihydrotestosterone 212 

Diisocyanate asthma 151 

di-(isodecyl)phthalate 223 

3,3′-dimethoxybenzine 301 

3,3′-dimethylbenzidine 301 

Dioctyl phthalate 152 

Diphenylhydantoin 322 

2,6-di-tert-butyl-4-methyl phenol 

(BHT) 316–20 

DNA adducts 74, 77–81, 83, 84, 158, 

159–9, 183–91, 189, 224 

immunoenrichment of 185–2 

DNA binding 60 

DNA damage 183, 192, 208, 356 

endogenous 356–1 

exogenous 357–3 

DNA reactions 73 

DNA-reactive carcinogens 356 

dose-response curve 356 

DNA repair 192 

DNA replication 180, 192, 357 

DNA synthesis 227, 230 

Dopamine 246 

Dose determination 188 

Dose-response relationship 

and exposure profiles 362–68 

DNA-reactive carcinogens 356 

in carcinogenesis 356 

low-dose range 190 

DTH reactions 196 

Dyes. See Colorants (dyes and 

pigments) 

EC annex VII and VIII toxicity tests 

280 

ECETOC 202 

EDB, conjugation in rats and man 67 

Electrophilic agents 60, 71 

Electrophilic centres 72 

Electrophilic compounds, examples of 

72 

Electrophilic metabolites 60 

Embryotoxicity tests 289–3 

Emulsifiers 312 

Endocrine dysfunction, xenobioticinduced 

254 

INDEX 375 

Endocrine toxicity, classification of 254– 

59 

Endocrine toxicology, thyroid 254–80 

Entire mammalian tests 283 

Environmental exposure 363 

Environmental monitoring (EM) 19–1 

Epoxide hydrolases (EH) 60 

Equivalent radiation dose concept 192– 

8 

Ethanol 241 

5-ethoxy-1,2,4-thiadiazole-3-carboxylic 

acid (ET-CA) 25–7 

Ethylene glycol methyl ether (EGME) 

286–91 

Etridiazol, disposition of 25–7 

European Union (EU) 167 

Exposure profiles 362–68 

Fatty amines 312 

FDA Segment I study for medicines 283 

Fecundity tests 284 

Fertility 

and embryotoxicity 285 

toxicity 281 

Fibre aerosols 97 

Fibre glass 91, 99–4 

airborne levels in workplace 109 

industrial hygiene studies 105–10 

lung fibre levels in workers 110 

Fibre recovery from lung tissue 118 

Fischer 344 study (Kimber-White) 198– 

4 

Flame retardant 312–18 

Flavin-containing monooxygenase 

(FMO) 211–16 

Fluoranthene-DNA adducts 187 

Foetal abnormalities 291 

Food additives 208 

Food and Agriculture Organization 208 

Fotemustine 46, 47 

Free radical formation 240–4, 246 

Full scale testing 288–1 

Furazolidone 68 

Gas chromatography (GC) 74, 76, 81

376 INDEX 

Gas chromatography/mass 

spectrometry (GC/MS) 74–9, 81, 84, 

185 

Genotoxic carcinogens 168 

Genotoxic hazards 181–8 

Genotoxicity 184, 281 

in vitro assays 184 

Germ cell mutagens 168 

Glial fibrillary acidic protein (GFAP) 

240–7, 247–2 

Gliotypic proteins 238–3 

Glutathione (GSH) 15, 20, 24, 184, 

241 

Glutathione conjugation, reversible 67– 

9 

Glutathione S-transferase (GST) 15, 20, 

24, 60–4, 66, 161, 212, 329 

Glycophorin A 160 

GM-CSF 143 

Growth desensitising mechanism 

(GDM) 263, 265 

Gunn rat hepatocytes in vitro, studies 

on 274 

Haemoglobin 76 

Haemoglobin adducts 189 

Health surveillance (HS) 19 

Heavy metals 244–9 

Hepatic metabolism, xenobiotics acting 

on 267–76 

‘Hepatic pharmacokinetic stuff’ 49 

Hepatocarcinogenesis, mechanisms of 

226–1 

Hepatocytes 

co-cultures 

phase 1 reactions 210–16 

phase 2 reactions 212–17 

long-term cultures of 207–14 

Herbicides 208 

Hexamethylene diisocyanate (HDI) 151 

n-hexane 3, 7, 238 

2,5-hexanedione 3 

HGPRT mutation assay 192 

HHPA 152 

Himic anhydride (HA) 152 

Hormone elimination 269 

Hormone synthesis 258 

Host resistance (HR) studies 200 

HPLC 74–83, 84 

Human allergic disease 142 

Human clearance prediction 49, 53 

Human exposure monitoring 188 

Human unbound clearances 51–3 

Hydroxymethylethenodeoxyadenosine 

(HMEdA) 83 

1-hydroxypyrene 159 

Hypersensitivity reactions, Type I-IV 

196 

Hypothalamic-pituitary-thyroid-liver 

(H-P-T-L) axis 256–60 

investigative tests on 267 

thyroid toxicity via 265–76 

toxicological 266 

xenobiotic toxic effects on 258 

Hypoxanthin guanine phosphoribosyl 

transferase (HPRT) 160 

ICICIS collaborative study 197–3 

IFN- 140–6 

IgA 199, 200 

IgE 153 

IgE antibody 138 

IgE antibody responses, induction and 

regulation of 140–5 

IgG 153, 200 

IgG2a antibody 140 

IgM 200 

Immune system, evaluation of toxicity 

to 196–10 

Immunoenrichment of DNA adducts 

185–2 

Immunological methods 77, 78 

Immunotoxic side effects, screening of 

197 

Immunotoxicity of chemical pesticides 

203 

Immunotoxicity testing, direct food 

additives 202 

Immunotoxicology 196 

collaborative studies 197, 198 

screening tests 196 

Insulin-like growth factor 1 (IGF,) 265 

Interferon (IFN- ) 140–6 

Interleukin 3 (IL-3) 141,143

Interleukin 4 (IL-4) 140, 141, 143 

Interleukin 5 (IL-5) 141–6 

Interleukin 10 (IL-10) 141 

Interleukin 12 (IL-12) 142 

International Agency for Research on 

Cancer (IARC) 91, 93–7, 163 

International Programme on Chemical 

Safety (IPCS) 91 

Iron 246 

Isocyanates 151–6 

Isophorone diisocyanate (IPDI) 151 

Joint Expert Committee on Food 

Additives (JECFA) 208 

Kainic acid (KA) 242–7 

Labelling 296–9 

Lactate dehydrogenase (LDH) 117 

Late respiratory systemic syndrome 

(LRRS) 152–7 

Lead 244–8 

Leaving group 65–8 

Life cycle exposure 283 

Life span correction 47–49 

Light microscopic histopathology 123–6 

Light stabilisers 316–39 

benzotriazole-based 322–9 

Linearized multistage cancer model 

(LMS) 170 

Lipophilic compounds 206 

Liver enzyme induction 318–2 

benzotriazole-based light stabilisers 

327–31 

LOAEL 168, 367 

Local lymph node assay (LLNA) 196 

LOEL 43 

Low molecular weight (MW) organic 

chemicals 187 

Lung burden analysis 103–8, 121–5 

Lung digestion/biodurability studies 

125 

Lung dissection 118 

Lung fibre burden 98–1 

Lung tissue, fibre recovery from 118 

MAK-list 304 

INDEX 377 

Maleic anhydride (MA) 152 

Malformations 291 

Malignancy, critical mutations leading 

to 190 

Manganese 245, 246 

Man-made vitreous fibres (MMVFs) 

animal inhalation studies 95–106 

carcinogenic potential 91–115 

cell culture studies 94 

comparison of human exposures 

used in rat chronic inhalation studies 

108–13 

epidemiological studies 91, 93–7 

implantation studies 95 

potential biological effects 91 

previous inhalation studies 106–9 

toxicologic studies 94 

Maximum life potential (MLP) 47–54 

Maximum tolerated dose (MTD) 170 

Meehs Formula 47 

Mercaptans 21 

Mercapturic acids 21–4, 184–90, 187–3 

toxicokinetics of 23 

urinary excretion 15, 25 

Metabolic activation 60–4 

Metabolism and toxicity 206–12 

Methamphetamine 241 

Methylene chloride 65 

Methylene diphenyldiisocyanate (MDI) 

151 

Methylmercury 241, 245 

N-methyl-N-nitrosourea (MNU) 263 

Michaelis-Menten kinetics 173 

Mitogenic stiraulation (ConA.LPS) 199 

Model neurotoxins 241–52 

Monitoring 

environmental (EM) 19–1 

human exposure 188 

in occupational toxicology 19–1 

polycyclic aromatic hydrocarbon 

(PAH) exposure 158–6 

see also Biological effect monitoring 

(BEM); 

Biological monitoring (BM) 

Monoclonal antibodies (Mabs) 185–2, 

318 

MPP + 241 

MPTP 241

378 INDEX 

mRNA analysis 211 

Mutagenic potency 190–6 

Mutagenicity, surfactants 349–2 

N 7 -deoxyguanosine (N 7 -dG) 186 

NADPH-cytochrome P450 reductase 

15 

Naphthalene diisocyanate (NDI) 151 

β-naphthoflavone 266, 267 

2-naphthylamine 40–3, 301 

Neuropathy target esterase (NTE) 238 

Neurotoxicity, biomarkers of 238–4 

Neurotoxicity assessment 285 

Neurotoxicity testing 237–56 

Neurotypic proteins 238–3 

Nitroarenes 74, 77 

NK activity 199 

NK test 200 

NOAEL 168, 174, 367 

NOEL 43, 366, 367 

Nucleophilic centres 73 

Occupational asthma 148–9 

chemical agents causing 148–5 

incidence 148, 151 

initial diagnosis 149 

Occupational toxicology, monitoring in 

19–1 

OECD guidelines 421 282–7 

OECD guidelines 422 282–91 

OECD single generation study 283 

17β-oestradiol 212 

Organic solvents 

absorption of 3–7 

biological monitoring of 2–3 

dermal uptake of 5–7 

pulmonary uptake of 3–5 

Organophosphate-induced delayed 

neuropathy (OPIDN) 238 

Organophosphate pesticides 237–2 

Parkinson’s disease 241, 245 

PBPK/PBTK models 

1,3-butadiene 16 

2-butoxyethanol 166–4, 172–80 

development of 31–3 

in risk assessment 170–80 

PCA-DNA adducts 187 

PCNB 267 

Pentafluorophenyl isothiocyanate 

(PFPITC) 76 

Pentafluorophenyl thiohydantoine 

(PFPTH) 77 

Perchlorate-discharge test 260 

Peroxisome proliferation 223–40 

carcinogenicity 224–30 

in rodent liver 223–8 

mechanisms of 226–1 

risk assessment 229–5 

rodent 223–9 

species differences in response 227–3 

Pesticides 203, 202 

human exposure to 208 

immunotoxicity of 203 

organophosphate 237–2 

Pharmaceuticals 202 

Phencyclidine 48 

Phenobarbital 267, 273 

Phenobarbitone 322 

Phenolic antioxidants 316–25 

blood kinetics 317–1 

blood metabolites 317–1 

effects on serum thyrotropin and 

thyroid hormones 318–4 

liver enzyme induction 318–2 

model compound 317 


Phthalate esters 223, 224 

Phthalic anhydride 152 

Physiologically-based pharmaco 

(toxico-) kinetic models. See PBPK/ 

PBTK models 

Phytopharmaceuticals 208 

Pigments. See Colorants (dyes and 

pigments) 

Plaque assay (PFCA) 199, 200 

Plasticisers 223, 229–4 

Polychlorinated biphenyls (PCBs) 247– 

2, 267 


(PAHs) 74–7, 78, 158, 186 

biomonitoring exposure 158–6 

Polyisocyanates 151–6 

Postlabelling 78–79, 81 

Post-natal manifestations 284

Post-radiolabelling technology 184 

Prenatal effects 284, 285, 289 

Production volume triggers 280, 281 

Propylthiouracil (PTU) 260 

Protein adducts 73, 74, 79, 81, 184–90 

Protein-binding 258 

Pulmonary cell proliferation 118, 124–7 

Pulmonary hyperreactivity to industrial 

pollutants 128–9 

guinea-pig model 131–6 

rat model 134–7 

Pulmonary sensitization, mechanisms of 

137–51 

Pulmonary toxic effects 

chrysotile asbestos fibres 116–30 

para-aramid fibrils 116–30 

Pulmonary uptake of organic solvents 

3–5 

Pyrethroid insecticides 238 

Pyromellitic dianhydride (PMDA) 152 

Quinone-thioethers 63 

Quinones 63–6 

Rad-equivalence values 192 

Radioallergosorbent tests (RAST) 138 

Radioimmunoassays 77 

Reactive airways dysfunction syndrome 

(RADS) 128, 148, 151, 153 

Reactive chemicals, metabolism 60–71 

Reactive compounds 

adduct formation of 74 

determination in unknown mixtures 

83–6 

determination of 71–88 

interaction with cellular constituents 

71–5 

Reactive metabolites 71 

Refractory ceramic fibres (RCFs) 91, 

99–2 

industrial hygiene studies 105 

Repeated dose toxicity 281 

Reproductive toxicity 279–298 

detecting effects on males 293 

evaluation for 282 

interpretation/extrapolation of 296– 

9 

INDEX 379 

labelling 296–9 

manifestations of 291–4 

methods for detecting effects on 282 

overlap of 281 

Respiratory allergic hypersensitivity 

137 

Restricted test systems 283 

Rifampicin 322 

Risk assessment 162–9, 166–84 

and exposure profiles 362–68 

approaches to 167–6 

cancer 188–6 

guidelines for 361 

in vitro approach 207–13 

mathematical procedures 369–3 

non-threshold effects 368 

peroxisome proliferation 229–5 

phenolic antioxidants 321–5 

physiologically based 

pharmacokinetic (PBPK) models in. 

See PBPK/PBTK models 

textile chemicals 315 

threshold effects 366–71 

see also Safety assessment 

RNA probes 211 

Rock (stone) wool 91, 102–5 


R-phrases 364 

Safety assessment 361–74 

benzotriazole-based light stabilisers 

331–7 

see also Risk assessment 

Salmonella typhimurium 66 

SHIELD system 149 

Slag wool 91, 102–5 


Solid phase assays 77 

Sperm analysis 284 

Spermatogenesis 284, 286, 288 

Structure-activity databases 282 

Structure-activity relationships, 

surfactants 344 

Styrene 38–38 

Styrene oxide 38–38 

Superoxide dismutase (SOD) 241 

Surface markers 200

380 INDEX 

Surfactants 338–55 

acute toxicity 348–1 

biochemical properties 338–5 

carcinogenicity 350 

chronic toxicity 349 

embryotoxicity 350 

excretion 347–50 

interactions with enzymes 341–5 

interactions with membranes 339 

interactions with proteins 339–3 

intestinal absorption 347–50 

local effects 342–48 

metabolism 347–50 

mucous membrane compatibility 

345 

mutagenicity 349–2 

oral toxicity 348 

percutaneous absorptions 346–9 

sensitization 345–8 

skin compatibility 342–7 

structure/activity relationships 344 

systemic effects 348 

toxicokinetics 346–50 

Surveillance of Work-related and 

Occupational Respiratory Disease 

Project (SWORD) 148 

Synaptophysin 242 

Systemic bioavailability of colorants 

301 

Systemic toxicity assessment 285 

Target dose determination 188–4 

Temelastine 271 272, 273 

2-tert-buty1–4-methoxyphenol (BHA) 

316 

Testosterone 212 

Tetrachloroethane 7 

Tetrachloroethene 8, 7 

Tetrachlorophthalic anhydride (TCPA) 

152, 153 

Textile chemicals 308–18 

finishing plant 309–13 

handling and processing 312–18 

irritant properties 311, 312 

new developments regarding 

toxicology 309 

oral toxicity 310 

process off-gas 315 

risk assessment 315 

temperature effects 311 

toxicological profile 311 

toxicology assessment needs 315 

Thalidomide 291 

T helper (Th) cells 140–5 

Thioethers 21–3 

Threshold effects 

classification 364–9 


Thyroglobulin (TBG) 258, 260 

Thyroid 

endocrine toxicology 254–80 

tumours of 256 

Thyroid binding pre-albumin (TBPA) 

258–2 

Thyroid follicles 260 

Thyroid follicular capability 262 

Thyroid follicular cell hyperplasia and 

neoplasia, pathobiology of 262–7 

Thyroid function 

control of 258 

perturbation of 256–60 

Thyroid hormones 258, 260, 318–4 

Thyroid lesions, pathobiology of 258 

Thyroid neoplasia 262, 321 

Thyroid stimulating hormone (TSH) 

258, 260, 262, 263, 265, 320, 321, 

322 

Thyroid toxicity 256 

via H-P-T-L axis 265–76 

Thyroid tumorigenesis 264 

Thyrotrophin releasing hormone (TRH) 

258, 262 

Thyrotropin 318–4 

Thyroxine 260, 269 

accumulation 271, 273 

clearance of 267–76 

T lymphocytes 142 

Toluene 241 

2,4- and 2,6-toluene diisocyanate (TDI) 

151 

Toxicity, and metabolism 206–12 

Toxicity tests, EC annex VII and VIII 

280 

Toxicodynamic interactions 20 

Toxicokinetic interactions 20

Toxicokinetic parameters 16 

Toxicokinetics, principles of 15 

Transition metals 246 

Trichloroacetic acid (TCA) 28, 29 

Trichloroethane 7, 8 

Trichloroethylene (TRI) 26–31 

effects related to 26–31 

hepatocarcinogenicity induced by 

28 

oxidative metabolism 28, 29 

2,5,6-trichloro-3-glutathion-S-yl) 

hydroquinone 64 

Triethyl lead 241 

tri-(2-ethylhexyl)trimellitate) 223 

Trimellitic anhydride (TMA) 152, 153 

Trimethyltin (TMT) 241–6 

Tumour incidence 189–5 

Two generation studies 291–5 

alternatives to 295–8 

UDP-glucuronosyltransferase 321, 327, 

329 

UDP-glucuronyltransferase (UDP-GT) 

213, 260, 267, 271, 274 

University of Pittsburgh study 91, 92 

Urinary excretion 

mercapturic acids 15, 25 

xenobiotics 16–18 

US Environmental Protection Agency 

91 

US-NTP study 200 

Veterinary medicinal chemicals 202 

Vinyl chloride 36–9 

Western blotting 211 

World Health Organization (WHO) 91 

Xenobiotic-induced endocrine 

dysfunction 254 

Xenobiotics 11 

assessment of long-term toxicity 

207 

biological effects 14 

disposition of 12–15 

overall exposure to 20 

urinary excretion of 16–18 

INDEX 381

Toxicology of Industrial Compounds

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