Commission of the European Communities

Industrial health and safety

Biological indicators for the assessment of human exposure to industrial chemicals

Aldrin and Dieldrin N.J. van Sittert and W.F. Tordoir Arsenic V. Foà, A. Colombi, M. Maroni, M. Buratti Cobalt A. Fenoli, R. Roi, L. Alessio Endrin N.J. van Sittert and W.F. Tordoir Vanadium K.H. Schaller and G. Triebig

Edited by L Alessio, A. Berlin, M. Boni, R. Roi

Joint Research Centre Ispra Establishment

Directorate-General for Employment and Social Affairs Health and Safety Directorate Directorate-General for Science Research and Development

EUR 11135 EN 1987

Commission of the European Communities

Industrial health and safety

Biological indicators for the assessment of human exposure to industrial chemicals

Aldrin and Dieldrin N.J. van Sittert and W.F. Tordoir Arsenic V. Foà, A. Colombi, M. Maroni, M. Buratti Cobalt A Fenoli, R. Roi, L. Alessio Endrin N.J. van Sittert and W.F. Tordoir Vanadium K.H. Schaller and G. Triebig

Edited by L. Alessio, A. Berlin, M. Boni, R. Roi

Joint Research Centre Ispra Establishment

Directorate-General for Employment and Social Affairs Health and Safety Directorate Directorate-General for Science Research and Development EUR 11135 EN 1987 Published by the COMMISSION OF THE EUROPEAN COMMUNITIES Directorate-General Information Market and Innovation Bâtiment Jean Monnet LUXEMBOURG

LEGAL NOTICE

Neither the Commission of the European Communities nor any person acting on behalf of the Commission is responsible for the use which might be made of following information.

Cataloguing data can be found at the end of this publication.

Luxembourg: Office for Official Publications of the European Communities, 1987

© ECSC - EEC - EAEC, Brussels-Luxembourg, 1987

Printed in Italy Ill

Preface of the second volume

Last year we published a series of monographs in one volume under the title "Human biological monitoring of industrial chemicals series" in which Benzene, Cadmium Chlorinated Hydrocarbon Solvents, Lead, Manganese, Titanium and Toluene were discussed. When preparing these documents each author was asked to pay particular attention to the problem of the quantitative relationships between the levels of external and internal exposure and between exposure and effects. In a number of cases information on the levels of biological indicators which are indicative of current exposure without short term detectable health effects is available. However we were already confronted with the impossibility to determine at present if the levels for biological indicators without short term detectable health effects can also the considered as adequate with respect to longer term effects. In preparing the present series of monographs it became apparent that for a number of biological indicators corresponding to biochemical tests, it is not yet possible to establish if these are indicative of early reversible effects and would thus quality for terminology of "biological monitoring" as defined in the preface of the first volume. For many substances, extensily used in industry, biological indicators are being developed but still require exten• sive assessment before possible routine application. As the object of these monographs is to provide up to date scientific information not only for the chemical substances for which biological indicators could be rised routine, but also for the many more substances for which biological indicators are at the early stage of development it was considered advisable to change the title of the series to "Biological indicators for the assessment of human exposure to industrial chemicals". We hope that this new title will avoid giving the impression to the reader that for all the substances presented in this volume and the substances ones, biological indicators can be already routinely applied.

The Editors 1984 IV

Preface of the third volume

Following the established frequency, we are happy to present the third volume in the series of monographs on "Biological Indicators for the Assessment of Human Exposure to In• dustrial Chemicals", which are addressed to occupational health physicians, industrial hygienists and, in general, to all who are concerned with prevention in the workplace. The original title of the first volume of the series "Human Biological Monitoring of Industrial Chemicals Series" was changed in the second volume and this change is now further justified by the four monographs making up the third volume: alkyl lead compounds, dimethylfor- mamide, mercury and organophosphorus pesticides. As in the previous volumes, the scope of the publication has not been limited to the most widely known and used toxic industrial agents. It was felt that consideration should also be given to other substances, where recent scientific advances have suggested the need to verify how far assessment of exposure using biological indicators is reliable in real in• dustrial situations. One of the aims of this series is, in fact, to stimulate further research, especially applied research, that would have the task of validating, on large groups of workers, preliminary scientific observations that are usually obtained from studies on relative• ly small groups of subjects and often in controlled experimental exposure situations. Eighteen monographs have now appeared in this series published by the Commission of the European Communities. The previous two volumes covered fourteen monographs on Acrylonitrile, Aluminium, Benzene, Cadmium, Chlorinated Hydrocarbon Solvents, Chromium, Copper, Lead, Manganese, Styrene, Titanium, Toluene, Xylene, Zinc. Highly competent scientists from the following European scientific and research institutes have contributed in preparing the monographs: Cattedra di Medicina del Lavoro dell'Univer• sità di Parma (Italy), Clinica del Lavoro "L. Devoto" dell'Università di Milano (Italy), Coronel Laboratorium, Universiteit van Amsterdam (the Netherlands), Institut für Arbeits- und Sozialmedizin der Universität Erlangen-Nürnberg (F.R. Germany), Unité de Toxicologie In• dustrielle et Médicale, Université de Louvain, Bruxelles (Belgium). It is planned in the future to extend cooperation to other scientific institutes and thus in• volve a wider number of scientists and experts. The fourth volume, which is already under way, will include monographs on "Aldrin, Dieldrin and Endrin" by N.J. van Sittert, Shell Internationale Petroleum (the Netherlands); "Non- Substituted Aliphatic Hydrocarbons" by K.N. Cohz, Danish National Institute of Occupa• tional Health, Hellerup (Denmark); "Arsenic" by V. Foà, Cllnica del Lavoro, University of Milan, (Italy); "Vanadium " by K.-H. Schaller, Institut für Arbeits- und Sozialmedizin, Univer• sity of Erlangen-Nürnberg (F.R. Germany).

The Editors 1986 Preface of the fourth volume

The fourth volume of the series of Monographs on "Biological Indicators for the Assess• ment of Human Exposure to Industrial Chemicals" of the Commission of the European Com• munities includes monographs on Arsenic, Aldrin and Dieldrin, Endrin, Cobalt and Vanadium. For the first time some chemical substances widely used in agriculture, are also examin• ed; for these substances, in addition to the occupational exposure, particular considera• tion has to be given to exposure of family members and general population consuming agricultural products treated with pesticides. It is evident from these documents, that the information on biological indicators resulting from both occupational data relative to users and manufacturers, and from non-occupational data relative to consumers, contributes to interpret the biological tests used to evaluate the exposure and / or the early effects. Occupational and environmental exposure to chemicals represent two situations, with uncertain limits, that tend to get nearer and nearer, and require a common study methodology. A volume which will include monographs on: "Nickel", by P. Grandjean (Inst, of Communi• ty Health-Odense University, Denmark); "Aromatic Hydrocarbons Nitro and Amino Com• pounds", by R. Lauwerys (Dept. of Occupational Medicine and Hygiene-Catholic University of Louvain-Brussels, Belgium); "Carbamate Pesticides", by M. Maroni (Inst, of Occupational Health- "Clinica del Lavoro L. Devoto" - University of Milano, Italy), is under preparation and should be published in 1988.

The Editors 1987 VI

Editorial Board

Prof. Lorenzo Alessio Institute of Occupational Health Piazzale Spedali Civili 1 University of Brescia 25100 Brescia, Italy

Dr. Alexander Berlin Commission of the European Communities Directorate General Employment and Social Affairs Health and Safety Directorate 2985 Luxembourg

Dr. Renato Roi Dr. Mirto Boni Commission of the European Communities Joint Research Centre - ECDIN Group 21020 Ispra (VA) Italy

List of Authors

ALDRIN AND DIELDRIN N.J. van Sittert and W.F. Tordoir Health, Safety and Environment Division Shell Internationale Petroleum Maatschappij B.V. The Hague, Holland ARSENIC V. Foà, A. Colombi, M. Maroni, M. Buratti* Institute of Occupational Health "Clinica L. Devoto' University of Milan Vla S. Barnaba, 8 20122 Milan, Italy * Laboratory of Clinical and Pathological Research Istituti Clinici di Perfezionamento Via S. Barnaba, 8 20122 Milan, Italy COBALT A. Fenoli, R. Roj*, L Alessio * Commission of the European Communities Joint Research Centre 21020 Ispra, Italy Institute of Occupational Health University of Brescia Piazzale Spedali Civili 1 25100 Brescia, Italy ENDRIN N.J. van Sittert and W.F. Tordoir Health, Safety and Environment Division Shell Internationale Petroleum Maatschappij B.V. The Hague, Holland VANADIUM K.H. Schaller and G. Triebig Institute of Occupational and Social Medicine of the University Erlangen-Nürnberg Schillerstr. 25/29 8520 Erlangen - Fed. Rep. Germany Contents

Preface Ill

Aldrin and Dieldrin 3

Arsenic 23

Cobalt 47

Endrin 63

Vanadium 77

Biological indicators for the assessment of human exposure to industrial chemicals

Aldrin and Dieldrin N.J. van Sittert and W.F. Tordoir

Summary

Aldrin and dieldrin are of a high order of acute oral toxicity and of a moderate order of percutaneous toxicity to mammals. Studies in experimental species and experience in occupational^ exposed workers have shown that the central nervous system is the critical organ. Clinical symptoms, e.g., vomiting, muscular jerking, convulsions and EEG abnormalities have been reported in workers occupational^ exposed to aldrin and dieldrin for periods of 1 week up to 5 years. The dieldrin concentration in blood below which no signs and symptoms of intoxication have been reported is 200 ng/ml, provided that blood samples are taken immediately after the end of exposure. Short-term and long-term studies with aldrin and dieldrin in experimental species show that the liver is also a major target organ. Studies in occupational^ exposed persons (up to 22 years) did not show any indication of disturbed liver functions compared with control persons. Workers with an average dieldrin concentration in blood of 105 ng/ml, which is equivalent with a daily oral intake of 1221 ug aldrin-dieldrin, did not show an increased activity of liver microsomal enzymes. Metabolism takes place mainly in the liver where aldrin is readily biotransformed to dieldrin. At a slow rate dieldrin is biotransformed into hydrophilic metabolites which are excreted via the bile and the urine. Dieldrin accumulates in body tissues. The determination of the concentration of dieldrin in blood is the most relevant test for assessing aldrin-dieldrin exposure and dieldrin body burden. The proposed biological limit value for dieldrin in blood for the prevention of acute health effects is 200 ng/ml.

Abbreviations

ALAT aspartate aminotransferase

ASAT alanine aminotransferase

ESR erythrocyte sedimentation rate HEOD (1R, 4S, 5S, 8R) -1, 2, 3, 4,10,10-hexachloro-1, 4, 4a, 5, 6, 7, 8, 8a-octahydro-6, 7-epoxy-1, 4: 5, 8-dimethano naphtalene

HHND (1R, 4S, 5S, 8R) -1, 2, 3, 4, 10, 10-hexachloro-1, 4a, 5, 8, 8a - hexahydro - 1,4 : 5,8 diamethanonaphtalene

LDH lactate dehydrogenase

RBC red blood cell count WBC white blood cell count

Aldrin and Dieldrin

Introduction

Aldrin and dieldrin are the common names of insecticides containing 95% of HHDN and 85% of HEOD respectively. Aldrin is a highly effective broad-spectrum soil insecticide. It has good activity against ter• mites and has uses in industry wherever termite protection is required. Dieldrin is a well established and versatile insecticide used in agriculture, public health and industry. In agriculture dieldrin can be used for control of many soil pests as soil or seed treatment, in public health to control tsetse flies and other vectors of debilitating tropical diseases and in industry dieldrin is very active against termites, wood borers and textile pests. Human exposure to pesticides can take place firstly in occupational^ exposed persons and secondly in members of the general population. In the former, exposure may occur (a) in plant workers during manufacturing, formulation and repacking processes and (b) in the users during the application of pesticides in the form of a suitable formulation. In members of the general population exposure can take place via the food, water, air. Aldrin and dieldrin can be taken up in the body via the dermal, oral and inhalational route. In occupational^ exposed persons the dermal route of exposure is the most important (van Heemstra and van Sittert, 1986). Aldrin is readily metabolised to dieldrin (HEOD) in plants and animals and only rarely are residues of aldrin found as such in food or in the great majority of animals and then only in very small amounts. Therefore, national and international regulatory bodies have con• sidered these two closely related insecticides together. The practicality of considering them jointly is further emphasised by the lack of significant differences in their acute and chronic toxicity and by their common mode of action.

Chemical and Physical Properties Aldrin

Chemical name of HHDN (1R, 4S, 5S, 8R) - 1, 2, 3, 4, 10, 10-hexachloro-1, (IUPAC) 4, 4a, 5, 8, 8a-hexahydro-1, 4: 5, 8-dimethanonaphtalene. Chemical structure of HHDN CI

Molecular formula Relative molecular mass Physical form Vapour pressure 8.6 χ 10"3 N/m2 at 20° C (6.5 χ IO"5 mm Hg) Density 1.54 g/ml at 20° C Solubility moderately soluble in most paraffinic or aromatic hydrocarbon solvents and esters or ketones; spar­ ingly soluble in alcohols; pratically insoluble in water (0.027 mg/l at 25°C).

Dieldrin

Chemical name of HEOD (1R, 4S, 5S, 8R) -1, 2, 3, 4,10,10-hexachloro-1, 4, (IUPAC) 4a, 5, 6, 7, 8, 8a-octahydro-6,7-epoxy-1,4: 5,8-dime- thano naphtalene. Chemical structure of HEOD

Relative molecular mass 381

Molecular formula C-I2 H8 Cl6 O Physical form buff to light tan, solid, dry flakes. Vapour pressure 4x10-4 N/m2 at 20° C (3.1 χ 10-6 mm Hg at 20° C). Density 1.62 g/ml at 20° C Solubility Dieldrin is moderately soluble in most aromatic and halogenated hydrocarbon solvents and esters or ketones; less soluble in paraffinic or naphthenic solvents and only sparingly soluble in alcohols. Dieldrin is practically insoluble in water (0.186 mg/1 at 25° C).

Effects on humans

Aldrin and dieldrin are of a high order of acute oral toxicity and of a moderate order of percutaneous toxicity to mammals. The typical oral LD50 value for the adult rat ranges from 40-70 mg/kg for aldrin and from 40-90 mg/kg for dieldrin. The acute percutaneous LD50 for male rats was 98 mg/kg for aldrin and 90 mg/kg for dieldrin (Gaines, 1960). The acute effects involve the central nervous system. Cases of accidental poisoning have occured by ingestion of formulated material, mostly in children by mistake or by adults with suicidal intent. Estimates of dosages suggest that fatalities have occured with inges­ tion of about 10 mg dieldrin per kg body weight (Hayes, 1982). Acute intoxications occurr­ ing In case of single or a very few consecutive, massive exposures by ingestion or extensive skin contamination are manifested, usually without prodromi, by peracute epileptiform con­ vulsions. The onset of clinical intoxication is pratically always acute also in those cases where the accumulation of dieldrin in the target tissues has taken place during a much longer period. The latter cases are, therofore, usually clinically indistinguishaole from a true acute intoxication. Twelve cases of neurotoxicity resulting from the repeated consumption of wheat into aldrin and gamma BHC powder had been mixed accidently, occurred in India (Gupta, 1975). The patients consumed this wheat for 6-12 months before showing typical clinical symptoms, including convulsions. Electroencephalograph^ (EEG) tracings were consistent with a diagnosis of organochlorine insecticide poisoning. All patients recovered. Short-term and long-term toxicity studies with aldrin and dieldrin in experimental species show that the liver is also a major target organ. In several long term studies hepatocellular carcinoma were reported in the mouse but not in the rat or hamster. The overall no-effect level in long-term feeding studies in rats is 0.5 ppm aldrin or dieldrin. At feeding levels of 1 ppm and above an increasing, dose-related, hepatomegaly and histological changes in the liver characterized as "chlorinated hydrocarbon insecticide ro­ dent liver (CHIRL)" occured. Studies in long-term exposed persons, either human volunteers, manufacturing and for• mulation workers or agricultural workers, did not show any indication of disturbed liver func• tions compared with control groups (Hunter et al, 1969, Jager, 1970; Morgan and Roan, 1974; Sandifer et al, 1981). Duration of exposure and dieldrin concentration in blood or serum of occupational^ exposed workers are shown In Table 1. Experiments with animals have shown that the earliest physiological sign of exposure to dieldrin is induction of in• creased activity of certain liver microsomal enzymes. The dietary intake of dieldrin required for the induction of this enzyme system in monkeys was approximately 1 ppm correspon• ding to an intake of 25-30 microgram/kg body weight/day (Wright et al, 1978). No enzyme induction has ever been found in workmen exposed to aldrin and dieldrin (Jager, 1970; Morgen and Roan, 1974). Numerous studies In occupational^ exposed persons do not show any indication of a carcinogenic hazard (Jager, 1970; Ditraglia et al, 1981; Rib• bens, 1985). A rare case of dieldrin-induced ¡mmunohaemolytic anaemia in a non-occupationally expos• ed subject was reported by Hamilton et al, 1978). The patient had a haemolytic anaemia with a positive direct Coombs test. The serum contained antibodies selectively active against erythrocytes coated with dieldrin. HEOD concentrations in blood and fat were similar to those of the general population. According to Christophers (1969) there is no increase in incidence of blood dyscrasias among populations most exposed.

Metabolism

Absorption

Aldrin and dieldrin are readily absorbed into the blood from the gastro-intestinal tract, through the skin or by inhalation of the vapour. It has not been possible to determine the percentage of an ingested dose of dieldrin that is actually absorbed into the body because part of the dieldrin which is excreted unchang• ed in the faeces originates from dieldrin which is passed unabsorbed through the intestinal tract and eliminated from the body, part originates from dieldrin which is excreted unchanged from the liver into the bile. Work with human volunteers showed that 7.7% of dermally applied 14C-HEOD was ex• creted in the urine over a five-day period (Feldmann and Maibach, 1974). This proportion of the applied dose cannot be equated with that portion which is absorbed since a major route of elimination of HEOD and metabolites in man is via the faeces. Inhalational studies with human volunteers suggested that about 50% of the inhaled aldrin vapour is absorbed and retained in the human body (Beyermann and Eckrich, 1973, Bragt et al, 1985).

Distribution

After absorption, aldrin is rapidly converted to dieldrin (HEOD), mainly in the liver. In humans, aldrin is rarely if at all found in human blood or other tissues, except in acute poisoning by accidental or intentional ingestion of massive doses. Dieldrin is a very lipophilic compound which is rapidly distributed particulary to tissues with a high fat content, i.e., the adipose tissue and a continuous exchange between the blood and the tissues takes place. The biotransformation of dieldrin into more hydrophilic metabolites proceeds slowly, and after repeated exposure to aldrin/dieldrin, dieldrin accumulates in these tissues. At the steady state, the concentration of HEOD in blood and tissues approach a plateau, the concentration of which is related to the average daily intake. An amount of HEOD equal to the average daily intake is eliminated each day when the steady state is reached. In a human volunteer study, three groups of 3 men received daily oral doses of 10, 50 or 211 microgram (ug) HEOD per man for 18 months. A fourth group of 3 males acted as control. From the 18th month to the 24th month the six volunteers given the two highest doses continued to receive HEOD at those two levels; the daily dose of the 10 microgram group was increased to 211 ug whilst the control group was also given 211 ug HEOD daily for this sixth month period. At the time of the study the average daily diet in the United Kingdom contained 20 ug of dieldrin, corresponding to an average HEOD concentration in blood of members of the 10 general population of 1.8 nanogram (ng)/mL and in body fat of 0.21 mg/kg (Hunter and Robin­ son, 1967; Hunter et al, 1969). Fig. 1 shows the relationship between the time of oral HEOD exposure and its concentra­ tion in human blood and adipose tissue in the 211 microgram and in the control group.

. 25 5.0­

ε ε <> ­ 20 4. Ο­ o

­!■ ι "α o O o α ­ 15 3.0­ < Q Q O ι ii O ï c ­ 10 i I 2.0­ c o

o o (J ­ 5 ï 1,0. O

ï] 4_ ­ \ · I 1 I »s ά 6 12 18 24 ' 1 Time of exposure to HEOD (months)

• HEOD in blood: 211 μς per day group A HEOD in blood: control group oHEOD in adipose fat; 211 μ§ per day group i. HEOD in adipose fat; control group

• Mean value and range

Figure 1 The concentration of HEOD in samples of blood and adipose fat collected at different ti­ mes from the subjects given 211 μg HEOD per day and from control subjects (from Hun­ ter and Robinson, 1967).

The figure shows that dieldrin in blood and adipose tissue reached a steady state 8-18 months after the beginning of the study. The concentration of HEOD in blood in the 211 microgram group had increased approximately seven-fold (mean: 14.0 ng/mL) in comparison with the control group (mean: 2.0 ng/mL) and of HEOD in fat approximately nine-fold (mean: 2.2 mg/kg). From 21-24 months, the concentrations of HEOD in blood and adipose tissue fluc­ tuated but there was no indication of a significant continuing increase in either set of samples. For both the concentration of HEOD in blood and in fat the relationship with the time of oral exposure was of the form:

C = A - B.e -k-t

where • C is the HEOD concentration in blood (in ng/mL) attained at time t assuming a cons­ tant daily HEOD intake (in ug). is the HEOD concentration in blood (in ng/mL) attained at the steady state, by adven­ titious as well as by voluntary constant daily intake of HEOD (in ug). 11

Β Is the HEOD concentration In blood (in ng/mL) attained at the steady state, by volun­ tary constant daily intake of HEOD (in ug) only. Α-B is the background HEOD concentration in blood (in ng/mL) by adventitious cons­ tant daily intake of HEOD (ug) only, e.g., via the food. t is the period (in days) during which a subject has received his daily oral dose of HEOD. k is the first-order reaction rate constant of HEOD for tissues.

It appeared that the first-order reaction rate constants for blood and adipose tissue were similar (k = 0.0029), which suggests parallelism in respect to elimination of HEOD from the tissues which is consistent with a one-compartment model. The distribution ratio, con­ centration of HEOD in adipose tissue/concentration in blood, was 136. At the steady state the relationships between the daily intake of HEOD (D) and the HEOD concentrations in blood and adipose tissue were established. These relationships were of the form:

A in blood (ng/mL) D (ug/day) 0.086

A in adipose tissue (mg/kg) or D (ug/day) 0.0185

Numerous investigations of the concentration of HEOD in body fat, blood and other tissues from members of the general population showed, on the average, that the ratio for fat, brain and blood is 150 : 15 : 3 : 1. The concentration of HEOD in whole blood is less than in serum. Mick et al. (1972) found in aldrin formulators a ratio of HEOD concentrations in plasma and in erythrocytes of 4 : 1. Dale et al. (1965) found approximately the same ratio in members of the general population.

Biotransformation

Most of the currently available information on the biotransformation of dieldrin in mam­ mals Is based on studies In experimental animals. Biotransformation of HEOD occurs mainly in the liver but to a lesser extent also In some other tissues, e.g. the lung (Mehendale and El-Bassioumi, 1975). Although there appears to be species differences, the overall picture shows only quan­ titative variations between species. In the species studies, the major metabolite Is the syn-12 hydroxy derivate of HEOD (Fig. 2; VI). Three other minor metabolites have been detected in experimental animals: a. the trans 6, 7-dihydroxy derivative (Fig. 2; IV), b. the dicarboxyllc acid derived from the dihydroxy compound (Fig. 2; V) and c. the bridged pentachloroketone (Fig. 2; VII).

CI IV ALDRIN TRANSDIOL ALDRIN DICARBOXYLIC ACID

VI VII SYN-12 HYDROXY DIELDRIN DIELDRIN PENTACHLOROKETONE

Figure 2 Mammalian metabolites of dieldrin 12

The syn-12-hydroxy metabolite has been detected in the faeces of seven occupational^ exposed operators (mean value, 1.74 mg/kg; range, 0.95-2.80 mg/kg) and also in members of the general population (mean value, 0.058 mg/kg; range, 0.033-0.12 mg/kg). Unchanged HEOD was present in the faeces of the workmen (mean value, 0.18 mg/kg), whereas the concentrations in the samples of members of the general population were below the limit of detection. Neither the syn-12-hydroxy compound nor the other minor metabolites have been found in human blood or other tissues. Examination of the urine of the operators indicated that urinary excretion of HEOD metabolites is minor in respect of elimination of syn-12-hydroxy dieldrin via the faeces (Richardson, 1971).

Elimination

When the intake of aldrin/dieldrin ceases, HEOD Is eliminated from the tissues according to first-order kinetics. Jager (1970) determined the concentration of HEOD in the blood of 15 aldrin/dieldrin workers at intervals during 3 years after occupational exposure to aldrin/dieldrin was terminated. At the end of exposure the HEOD concentrations in blood and adipose tissue were at the state of equilibrium. The average biological half-life for dieldrin in blood was 267 days, cor­ responding with a reaction rate constant of 0.0026 days-1. This value is consistent with the k value for HEOD in blood in the previously described human volunteer study. (Hunter et al. 1969).

Biological Indicators

Indicators of exposure

Concentration of dieldrin in blood

The results obtained during medical surveillance of workers employed in the manufactur­ ing or formulation of aldrin/dieldrin, have shown that adverse effects induced by aldrin/dieldrin are related to the concentration of HEOD in blood (Brown et al. 1964; Jager, 1970). Therefore, the determination of the HEOD concentration in blood provides a powerful test for the assess­ ment of exposure and health risk. The test has shown to be valuable (a) as a specific differential diagnostic test in cases of aldrin/dieldrin intoxication (b) to assist in the decision when to suspend workers from further exposure and when workers may return to work. In cases of a repeated and regular exposure pattern, the personal daily Intake of aldrin + dieldrin can be calculated with a formula derived from the previously described equa­ tion (Hunter et al, 1969):

C = 0.086 χ D χ (i-e-°oo26.t)

For the determination of dieldrin in blood, two analytical procedures have been used (Dale et al, 1966; Richardson et al. 1967) which give significantly different results (Robinson et al. 1967): the acetone extraction procedure of Richardson et al., (method II in Table 1) gives results which are about 50% higher than the hexane extraction procedure of Dale et al (method I in Table 1). Using the same extraction method, the concentration of HEOD in whole blood is 0.66 χ the concentration of HEOD in serum (Dale et al, 1965; Mick et al, 1972) (see also section on analytical methods). Great differences exist between the average concentrations of HEOD in blood found in members of the general population, in obviousiy healthy occupational^ exposed workers without complaints or symptoms, and in persons with complaints and symptoms of adverse effects. 13

Table 1 - Concentrations of HEOD in blood or serum of 'healthy workers' occupationally exposed to aldrin and/or dieldrin. Blood samples were collected during manufacturing, formulation or spraying aldrin or dieldrin, unless stated otherwise.

Duration of HEOD in blood No of Year exposure or serum (ng/ml)D Method3 Reference subjects (years) mean min max 1964 63c <1-10 69 (B) 10 440 Jager, 1970 1965 63c <1-11 60 (B) 10 240 Jager, 1970 1966 64c <1-12 50 (B) 10 220 Jager, 1970 1967 72c <1-13 31 (B) 10 180 Jager, 1970 1968 70c <1-14 32 (B) 10 220 Jager, 1970 1969 62c <1-15 25 (B) 5 150 Jager, 1970 1967 28c 27-14066' 25 (S) 1.2 137 Hayes & Curley, 1968 1967 5C 1-5 100 (B) 24 250 Avar & Czegledi-Janko, 1970 1969 66c 4-5 4.2 (S)9 — — Warnick & Carter, 1972 1969 66d 4-5 21 (S)9 — — Warnick & Carter, 1972 1969 115e 4-5 1.5 (S)9 — — Warnick & Carter, 1972 ? 37c'd 5-22 84 (S) 40 350 Morgan and Roan, 1974 ? 7C 5 weeks 182 (S) 40 310 Mick et al, 1972 1974 7d n.s. 1.8 (B)9 1 5.5 Mac Cuaig, 1976 1975 7d n.s. 4.5 (B)9 1 8.8 Mac Cuaig, 1976 1978 27C >1 20 (B) 4.5 54 ? Sandifer et al, 1981 a. I = acetone extraction technique Il = hexane extraction technique b. S = samples of serum or plasma; B samples of whole blood. c. manufacturing and formulation workers. d. pest control operators. e. commercial spray-rig operators. f. hours. g. time of blood sampling not specified, n.s. not specified.

Occupationally exposed workers

Table 1 shows HEOD concentrations in blood of 'healthy workmen' occupationally expos• ed to aldrin/dieldrin during manufacturing, formulation or spraving activities. HEOD concentrations in the blood of these workers were considerably higher than in members of the general population. These 'healthy workmen' were men between 18 and 60 years of age who had no complaints or clinical or laboratory signs attributable to oc• cupational exposure.

Table 2 shows concentrations in blood of subjects exposed to aldrin/dieldrin with clinical symptoms. Concentrations of HEOD ranging from 40-530 ng/mL have been reported in the blood of persons who had more or less recently been poisoned and who recovered (Kazantzis et al, 1964; Brown et al, 1964,; Avar and Czegledi-Janko, 1970; Jager, 1970). It is unlikely that symptoms of persons with blood levels at the lower end of the range could be attributed to overexposure to aldrin/dieldrin. The low level of 40 ng/mL was found in a man with chronic nephritis, complaining of headache and nausea; the occupational cause of the symptoms is therefore doubtful (Kazantzis et al, 1964). In addition, the blood sample had been collected six weeks after the last exposure. In another worker a blood concentration of 130 ng/mL was found one month after a mild acute episode. (Brown et al, 1964). This suggests that in both subjects a much higher level must have occurred at the time of the complaints.

Members of the general population

HEOD is present in the blood at very low concentrations in the majority of members of the general population in the United States (EPA, 1983) and most other countries (Table 3). It is essential that the sensitivity of the analytical method is at least 0.1 ng/mL. 14

Table 2 ­ Concentration of HEOD in blood or serum in workers with clinical symptoms exposed to aldrin and/or dieldrin

No of HEOD in blood Duration of b Year sub­ or serum (ng/ml) Method3 Reference Clinical symptioms exposure jects0 mean min max

1963 4 1 week­3 years 40d (B) 530 I Kazantzis et al, 1964 vomiting: muscular jerking; convulsions; EEG abnormalities 1963 5 1 week­3 years 130e(B) 370e I Brown et al, 1964 EEG abnormalities 1964 1 4 months 230 I Jager, 1970 EEG abnormality 1967 10 2­5 years 203 (B) 50 280 II Avar & Czegledi­Janko, vomiting; muscular 1970 jerking; convulsions; EEG abnormalities. a. I = acetone extraction technique; II = hexane extraction technique b. Β = samples of whole blood c. manufacturing and formulation workers d. the sample had been collected six weeks after the last exposure e. determined some time after the acute episode

Table 3 ­ Concentrations of HEOD in blood or serum in subjects not occupationally exposed.

HEOD η bloodb Country Year No Of or serum (ng/ml) Method3 Reference subjects mean min. max.

USA 1967­68 1000 0.5 (S) — — I Watson et al, '70 USA 1967­71 970 0.9 (S) — — I Warnick, 1972 USA 1970 200 0.9 (S) — 10 I Wyllie et al, '72 USA 1977­81 Í4000 <1.0 (S) — 16.1 I EPA, 1983 U.K. 1966 55 1.8 (B) — 4.3 II Hunter et al, '67 U.K. 1968 18 0.9 (B) — 1.1 II Robinson & Roberts, '69 Switz 1972 100 1.1 (S) — — I Zimmerli and Marek 1973 The Neth. 1978 299 <0.4 (B) — — I Greve, 1985 a I = hexane extraction technique; Il = acetone extraction technique. b S = samples of serum or plasma; Β = samples of whole blood.

Concentration of dieldrin and metabolites in faeces and urine.

HEOD and its syn­12­hydroxy derivate have been detected in faeces in occupationally ex­ posed industrial workers. The average concentrations were 0.18 mg/kg and 1.74 mg/kg respectively (Richardson, 1971). The syn­12­hydroxy compound has also been demonstrated in the faeces of members of the general population (average 0.058 mg/kg), but the con­ centration of HEOD in faeces samples was below the limit of detection. Thus the determination of HEOD and its syn­12­hydroxy derivate in faeces can be used to measure aldrin/dieldrin absorption. However, these tests are not praticai in industrial conditions. Cueto and Biros (1967) reported that the concentration of HEOD in urine correlated with the exposure of workmen to aldrin/dieldrin. However, Richardson (1971) demonstrated that HEOD and its four known metabolites were not present in significant amounts in urine of workers with known high HEOD concentrations in blood.

Concentration of dieldrin in adipose tissue

As the concentration of HEOD in adipose tissue is much higher than that in blood, the deter­ mination of the concentration of HEOD in the body fat of members of the general popula­ tion who have had no occupational exposure to aldrin/dieldrin has been performed in numerous investigations. Surveys have been made in more than 20 countries, but in some surveys the number of samples of fat analysed is small. 15

In the United States and in the United Kingdom there have been several surveys during the period 1961-1977. The more recent results (from 1970 onwards) have been summaris­ ed in Table 4. Most of the mean values are in the range of 0.1 to 0.3 mg/kg fat. The concentration of HEOD in adipose tissue, in workers occupationally exposed to aldrin and dieldrin, has only been determined In one study. The group of 28 workers concerned had an average HEOD concentration in fat of 6.1 mg/kg (Hayes and Curley, 1968). The corresponding average HEOD concentration in plasma was 25 ng/mL (Table 1).

Table 4 - Concentration of HEOD in adipose tissue in subjects not occupationally exposed (1975 onwards)

HEOD concentration Country Year No of Tig/kg Reference samples3 mean maximum

USA 1970 40 (N) 0.15 — Warnick, 1972 USA 1970 68 (S) 0.29 0.73 Burns, 1974 USA 1970 202 (S) 0.2 1.0 Wyllie et al. 1972 USA 1970 1412 (NIS) 0.18 15.2 Kutz et al. 1979 USA 1971 88 (Β) 0.36 0.78 Burns, 1974 USA 1971 1615 (N/S) 0.22 2.91 Kutz et al. 1979 USA 1972 39 (S) 0.43 1.00 Burns, 1974 USA 1972 1913 (N/S) 0.18 2.91 Kutz et al. 1979 USA 1973 1094 (N/S) 0.18 5.64 Kutz et al. 1979 USA 1974 898 (N/S) 0.15 2.21 Kutz et al. 1979 U.K. 1969-71 201 (Ν) 0.16 0.68 Abbott et al. 1972 U.K. 1976-77 236 (Ν) 0.11 0.49 Abbott et al. 1981 Switz 1972 13 (S) 0.29 0.57 Zimmerli and Marck,. 1973 Denmark 1972-73 70 (Ν) 0.16 0.53 Kraul and Karlog, 1976 Belgium 1975 60 (Ν) 0.26 1.16 Dejonckheere et al. 1977 Belgium 1977 58 (Ν) 0.13 0.69 van Haver et al. 1978 The Neth. 1983 78 (Ν) 0.06 0.14b Greeve, 1985

3 Ν = samples taken at necropsy. S = samples taken during elective surgery, b 90 percentile.

Indicators of Effect Electroencephalography

As the central nervous system Is the target organ for aldrin/dieldrin, changes in the elec­ troencephalogram (EEG) may occur after absorption of high doses of these compounds. The EEG changes were first developed to be a praticai tool for monitoring effects on the central nervous system of workers exposed to aldrin/dieldrin and determining when they should discontinue exposure and when they could be allowed to resume work with aldrin/dieldrin. Characteristic changes - which, however are not specific for aldrin/dieldrin overexposure - include bilateral synchronous spikes, spike and wave complexes and slow theta waves (Spiotta, 1951; Winthrop and Felice, 1957; Hoogendam et al, 1964,1969; Kazant­ zis et al, 1964, Garretson and Curley, 1969, Avar and Czegledi-Janko, 1970; Jager, 1970). Nowadays the determination of the concentration of HEOD in blood has supplanted EEG examination as a method of exposure monitoring of operators exposured to aldrin/dieldrin.

Heod in Blood-Effect Relationships

Human intoxication

Since the introduction of the determination of the concentration of HEOD in blood, no com­ plaints of ill-health, or any objective clinical or laboratory tests of Ill-health, have ever been 16

noted In operators, occupationally exposed to aldrin/dieldrin for periods up to 15 years, whose HEOD concentration in whole blood was less than 200 ng/mL at the time of medical examination, which took place twice yearly before 1965 and once yearly thereafter. (Jager, 1970). Laboratory tests included liver function tests (ASAT, ALAT, alkaline phosphatase, LDH, total bilirubin, total protein, albumin and alfa, beta and gamma globulins), kidney function tests (serum creatinine, urea and uric acid), haemoglobin, RBC's, WBC's, ESR) and in ad­ dition serum calcium, inorganic phosphate, and cholesterol. The HEOD blood concentra­ tion of 200 ng/mL may, therefore, be considered to be a no­adverse­effect level. Concentrations exceeding 200 ng/mL HEOD in blood are not necessarily indicative of adverse effects, since frequently concentrations exceeding this level have been found in aldrin and dieldrin manufacturing and formulation workers without complaints of ill­health or laboratory test abnormalities. The maximum concentration found in the blood of an operator during aldrin or dieldrin manufacturing or formulation, not having clinical symptoms, was 440 ng/mL (Jager, 1970).

Effects on the central nervous system

The relationship between HEOD concentrations in blood and changes In the elec­ troencephalogram (EEG) has been reported by several authors (Kazantzis et al. 1964; Avar and Czegledi­Janko, 1970; Jager, 1970) (Table 2). The authors showed that EEG changes were reversible a few months after exposures to aldrin/dieldrin were discontinued. The interesting parallelism between the rate of diminuation of the EEG changes and the rate of decrease of the concentration of HEOD in the blood was reported in the case of an accidentally poisoned child (Garrettson and Curley, 1969).

Effects on the liver

Experiments with animals have shown that the earliest physiological sign of exposure to aldrin/dieldrin is induction of increased activity of certain liver microsomal enzymes (Wright et al, 1972). The dietary intake required for the induction of this enzyme system was 25­30 ug/kg body weight/day for monkeys fed dieldrin for six years (Wright et al, 1978). Three studies have been reported on the relationship between dieldrin concentrations in blood or serum and hepatic enzyme induction in workmen occupationally exposed to aldrin/dieldrin. The results of these studies are summarised in Table 5. Tests which were used as an indicator of enzyme induction were the concentration of pp'­ DDE in blood, the ratio of the concentrations of urinary 6­ß­hydroxycortisol and 17­hydroxycorticosteroids and the urinary excretion of D­glucaric acid. The table shows that pp'­DDE concentrations were not different in a group of 10 workmen with a mean HEOD concentration In blood of 105 ng/mL (range 80­150 ng/mL), corresponding with an average daily intake of 1221 ug dieldrin or 17.4 ug/kg body weight, compared with a con­ trol group (Jager, 1970). The same author showed that the ratio 6­ß­hydroxycortlsol and 17­hydroxycorticosteroids was not different in a group of 13 workmen with a mean HEOD

Table 5 ­ HEOD concentration in blood or serum and parameters of hepatic enzyme induction.

Parameter of hepatic HEOD in blood enzyme induction (mean values) No of or serum (ng/mL) 6­ß­hydroxycortlsol Method0 Reference subjects pp'­DDE D­glucaric acid In blood 17­hydroxycorticosteriods in urine mean min. max. (ng/mL) (ug/mg) (uMol/g creat.) 130 5(B) 13.2 II Jager, 1970 10 105(B) 80 150 14.4 II Jager, 1970 20 5(B) 26.6 II Jager, 1970 13 51 (B) 18 110 25.3 II Jager, 1970 68 1 (S) 1 2 19 I Morgan & Roan, 1974 11 4(S) 2 10 19 I Morgan & Roan, 1974 11 29 (S) 10 40 19 I Morgan & Roan, 1974 37 84 (S) 40 350 22 I Morgan & Roan, 1974 27 20(B) 4.5 54 no changes ? Sandifer et al, 1971

S = samples of serum; Β = samples of whole blood I = hexane extraction technique II = acetone extraction technique 17 concentration in blood of 51 ng/mL (range 18-110 ng/mL) compared with a control group. Using the same parameter for hepatic enzyme induction, Sandifer et al, (1981) showed no differences compared with a control group in subjects with a mean HEOD concentra­ tion in blood of 20 ng/mL (range 4.5-54 ug/ml). D-glucaric acid concentrations in urine were not different compared with a control group in workmen with a mean HEOD concentration in serum of 84 ng/mL (range 40-350 ng/mL), which is equivalent with approximately the same levels in blood (see section on analytical methods), (Morgan and Roan, 1974). Jager (1970) examined liver enzymes of a group of 233 men with more than 4 years exposure to aldrin/dieldrin, endrin or telodrin. Results of liver enzyme activities did not show any in­ dication of an influence of the degree or the duration of exposure to these insecticides on the parameters. Morgan and Roan (1974) and Sandifer et al (1981) did not find differences in liver enzyme activities between workmen exposed to aldrin/dieldrin and a control group.

Analytical Methods

The most popular method for the determination of dieldrin in blood is the method of Dale et al. (1967) (see Tables I and II), even though the authors show that quantitative recovery is not obtained. This is probably due to binding of dieldrin to blood proteins. Richardson and Robinson (1967) describe a method in which the HEOD is quantitatively extracted from the blood with acetone and then concentrated by extraction with hexane (acetone extrac­ tion method). A comparison of the two methods showed that the concentration of HEOD in blood with the acetone extraction method is about 1.5 χ the concentration of HEOD in blood with the hexane extraction method. (Robinson et al, 1967). Using the same extrac­ tion method, the concentration of HEOD in blood is about 0.66 χ the concentration of HEOD in serum (Dale et al. 1965; Mick et al. 1972). An inter-laboratory comparison of the hexane extraction method showed large variations in results between the laboratories (Thompson, 1976).

By using the above conversion factors it is possible to compare the HEOD concentrations in blood or serum determined with either the hexane or the acetone extraction method (Tables 1-3). For example, a HEOD concentration in serum of 350 ng/mL determined with the hexane extraction method (Table 2: Morgan and Roan, 1974) is equivalent with the same HEOD concentration in blood, determined with the acetone extraction technique (350 χ 0.66 χ 1.5 ng/mL = 350 ng/mL).

It should be realised that the no-adverse effect HEOD concentration In blood of 200 ng/mL Is based on the acetone extraction technique and that this figure should be modified if the hexane extraction technique is used or If HEOD is determined in serum. This also applies for the calculation of the personal daily aldrin + dieldrin intake from dieldrin concentra­ tions in blood.

Conclusions

For the measurement of the absorption of aldrin + dieldrin in the body, the determination of the concentration of dieldrin in blood is the most relevant test. This test is also an In­ dicator of dieldrin body burden. After long term repeated exposures e.g., In industrial workers and pest control operators (Table 1), the dieldrin concentration in blood will be in equilibrium with the dieldrin concen­ tration in adipose tissue. In such situations the personal daily intake of aldrin + dieldrin can be calculated from the blood dieldrin concentration, using the equation of Hunter et al. (1969). The time of collection of the blood sample is not critical at equilibrium. However, after short term (over) exposures to aldrin/dieldrin, the dieldrin concentration In blood Initially will be higher than after distribution of dieldrin in the adipose tissue. In such cases, samples should be collected Immediately after the end of exposure to ascertain whether or not there is any risk for imminent intoxication. No cases of intoxication due to aldrin/dieldrin exposure have been reported at dieldrin con­ centrations in blood below 200 ng/mL, determined with the acetone extraction technique, provided that samples were collected at the end of exposure. Physical, neurological and laboratory examinations were within reference limits. Hepatic enzyme induction could not be demonstrated in workmen with dieldrin concentrations In blood (acetone extraction techni- 18 que) in the range of 80-150 ng/mL (mean: 105 ng/mL) and with dieldrin concentrations in serum (hexane extraction technique) in the range of 40-350 ng/mL (mean: 84 ng/mL). Electroencephalography is an obsolete method for the monitoring aldrin/dieldrin exposure. The test still can be used for confirming a diagnosis of aldrin/dieldrin intoxication. Guidelines for biological limit values of workmen exposed to aldrin/dieldrin were discussed in the workshop 'Epidemiological toxicology of pesticide exposure', organised by the Scientific Committee on Pesticides of the International Committee on Occupational Health in Amster• dam in 1971 (Zielhuis, 1972). It has been suggested that the following levels could be distinguished:

Level 1

Blood HEOD concentration: Below this value no harmful effect should be expected. 100 ng/mL

Level 2

Blood HEOD concentration: Values require action, e.g. repeat monitoring of dieldrin 100-200 ng/mL blood concentration after appropriate interval; guard against exceeding the upper value of this level, in• cluding appraisal of general and individual work situa• tion. Reduce exposure.

Level 3

Blood HEOD concentration: Values require remedial action as there is danger of >200 ng/mL imminent or overt intoxication. Workers have to be removed from further exposure until blood levels are within acceptable limits. The work situation has to be appraised and exposures have to be reduced.

These suggested biological limit values can be sustained by the absence of any reports on clinical symptoms in subjects with HEOD concentrations in the blood below 200 ng/mL after 1971. Because of the improved plant hygiene, presently dieldrin concentrations in blood of most manufacturing and formulation workers are below 10 ng/mL.

Research Needs

Several studies have been reported on the long term effects of aldrin and dieldrin in man, (Jager, 1970; Versteeg and Jager, 1973; van Raalte, 1977; Ditraglla et al., 1981; Ribbens, 1985). Further follow-up studies on these previously studied populations are in progress. These studies are characterised by:

a large number of exposed workers providing acceptabel statistical power; long term exposure (1953 onwards); high exposures, particularly in the first 10-15 years of manufacturing and formulation; personalised exposure data provided by dieldrin blood concentration data allowing the calculation of individual daily intake. 19

References

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HAYES, W.J. (1982) Chlorinated hydrocarbons insecticides. In: Pesticides studies in man, p. 172-283, Baltimore, Williams and Wilklns HAYES, W.J.; CURLEY, A. (1968) Storage and excretion of dieldrin and related compounds. Arch. Environ. Hlth. 16:155-162 HEEMSTRA, E.A.H. VAN; SITTERT, N.J. VAN (ed.) (1986) Biological monitoring of workers manufacturing, formulating and applying pesticides. Pro­ ceedings of the seventh International workshop of the scientific committee on pesticides of the international commission on occupational health, Szeged, Hungary, April 15-17,1986. Toxicology Letters 33. HOOGENDAM, I.; VERSTEEG, J.P.J.; VLIEGER, M. DE (1967) Aldrin and dieldrin - the safety of present exposures of the general populations of the United Kingdom and the United Stated. Fd. Cosmet. Toxicol, 5:781-787 HUNTER, CG.; ROBINSON, J.; ROBERTS, M. (1969) Pharmacodynamics of dieldrin (HEOD). II. Ingestion by human subjects for 18 to 24 months, and postexposure for eight months. Arch. Environ. Hlth. 18:12-21 JAGER, K.W. (1970) Aldrin, Dieldrin, Endrin and Telodrin: An epidemiological and toxicological study of long- term occupational exposure. New York, Amsterdam, Elsevier Pubi. Co. KAZANTZIS, G.; Mc LAUGHLIN, A.I.G.; PRIOR, P.F. (1964) Poisoning in industrial workers by the insecticide aldrin. Brit. J. Ind. Med. 21:46-51 KRAUL, I.; KARLOG, O. (1976) Persistent organochlorinated compounds in human organs collected in Denmark 1972-1973. Acta Pharmacol. Toxicol. 38:38-48 KUTZ, F.; STRASSMAN, S.; YOBS, A. (1979) Survey of pesticide residues and their metabolites in the general population of the United States. In: Berlin, Α.; Wolff, A.H.; Hasegawa, Y. (eds.). Use of biological specimens to assess human exposure to environmental pollutants, 267-274. The Hague, Martinus Nljhoff MacCUAIGH, R.D. (1976) The occurrence of insecticides in the blood of staff of a locust control organisation. Bull. Environ. Contamin. Toxicol. 15:162-170 MICK, D.L; LONG, K.R.; BONDERMAN, D.P. (1972) Aldrin and dieldrin in the blood of pesticide formulators. Am. Ind. Hyg. Ass. J. 33:94-99 MEHENDALE, Η.M.; SKRENTNY, R.L AND DOROUGH, H.W. (1975) Uptake and disposition of aldrin and dieldrin by isolated perfused rabbit. Drug Metab. Dispos. 3:543-556 MORGAN, D.P.; ROAN, C.C. (1974) Liver function in workers having high tissue stores of chlorinated hydrocarbon pesticides. Arch. Environ. Hlth., 29:14-17 RAALTE, H.G.S. VAN (1977) Human experience with dieldrin in perspective. Ecotoxicol. Environ. Saf. 1:203-210 RIBBENS, P.H. (1985) Mortality study of industrial workers exposed to aldrin, dieldrin and endrin. Int. Arch. Oc- cup. Environ. Health 56, 75-79 RICHARDSON, Α.; ROBINSON, J. (1971) The identification of a major metabolite of HEOD (dieldrin) in human faeces. Xenobiotica 1:213-219 RICHARDSON, Α.; ROBINSON, J.; BUSH, B.; DAVIES, J.M. (1967) Determination of dieldrin (HEOD) in blood. Arch. Environ. Hlth. 14:703-708 ROBINSON, J.; ROBERTS, M. (1969) Estimation of the exposure of the general population to dieldrin (HEOD). Fd. Cosmet. Tox­ icol. 7:501-514

ROBINSON, J.; RICHARDSON, Α.; DAVIES, J.M. (1967) Comparison of analytical methods for determination of dieldrin (HEOD) in blood. Arch. En­ viron. Hlth. 15:67-69 SANDIFER, S.H.; CUPP, CM.; WILKENS, R.T.; LOADHOLT, B.; SCHUMAN, S.H. (1981) A case-control study of persons with elevated blood levels of dieldrin. Arch. Environ. Con- tarn. Toxicol. 10(1):25-45 21

SPIOTTA, E.J. (1951) Aldrin poisoning in man. Report of a case. Arch. Ind. Hyg. Occup. Med. 4:560­566 THOMPSON, J.F. (1976) Manual of analytical quality control for pesticides in human and environmental media. U.S. Environmental Protection Agency VERSTEEG, J.P.J.; JAGER, K.W. (1973) Longterm occupational exposure to the insecticides aldrin, dieldrin, endrin and telodrln. Br. J. industr. Med. 30:201­202 WARNICK, S.L.; CARTER, J.E. (1972) Some findings in a study of workers occupationally exposed to pesticides. Arch. Environ. Hlth. 25:265­270 WATSON, S.L.; BENSON, W.W.; GABICA, J. (1970) Serum organochlorine pesticide levels in people in southern Idaho. Pestle. Monit. J. 4(2):47­50 WINTHROP, G.J.; FELICE, J.F. (1957) A clinical toxicological study of sprayman of a chlorinated hydrocarbon insecticide. Bol. of Sanit. Panam., 43:512­517 WRIGHT, A.S.; POTTER, D.; WOODER, M.F.; DONNINGER, C; GREENLAND, R.D. (1972) The effects of dieldrin on the subcellular structure and function of mammalian liver cells. Fd. Cosmet. Toxicol. 10:311­322 WRIGHT, A.S.; DONNINGER, C; GREENLAND, R.D.; STEMMER, K.L.; ZAVON, M.R. (1978). The effects of prolonged ingestion of dieldrin on the livers of male Rhesus monkeys. Ecotox. Environ. Saf. 1:477­502. WYLLIE, J.; GABICA, J.; BENSON, W.W. (1972) Comparative organochlorine pesticide residues in serum and biopsied lipoid tissue: a survey of 200 persons in southern Idaho ­ 1970. Pestle, Monit. J. 6(2):84­88 ZIELHUIS, R.L. (1972) Epidemiological toxicology of pesticides exposure ­ Report of an international workshop. Arch. Environm. Health 25, 399 ZIMMERLI, B.; MAREK, Β. (1973) Die Belastung der schweizerischen Bevölkerung mit Pestiziden. Mitt. Geb. Lebensmitt. Hyg. 64:459-479

Biological indicators for the assessment of human exposure to industrial chemicals

Arsenic V. Foà, A. Colombi, M. Maroni, M. Buratti

25

Summary

Human exposure to arsenic may occur from natural or occupational sources (e.g. smelteries, glass and pesticides industry). Inorganic As and its compounds may be absorbed Into the human organism through ingestion, inhalation and skin absorption. Metabolism of As is not completely known. Inorganic arsenic may be biotrasformed to and eliminated in various forms, namely arsenite, arsenate, monomethylarsonic acid, and dimethylarsinic acid. Seafood is extraordinarily rich in arsenobetaine a peculiar form of organic As. Arsenic from seafood does not mix with the inorganic As pool in the organism, and its toxicity is remarkably lower than that of inorganic As. Although very small amounts of certain As compounds may have beneficial health effects, As compounds are generally regarded as very potent poisons. Acute toxicity, however, varies greatly among organic and inorganic compounds, depen• ding on oxidation state and solubility in biological media. Chronic As poisoning may occur in occupational and environmental exposure. In occupationally exposed workers local effects in the mucous membranes of the respiratory tract and skin effects are prominent signs observed. Involvement of the nervous and cir• culatory system and the liver as well as cancer of the respiratory tract may also occur. With long term exposure to arsenic via ingestion in food, drinking water or medication, symp• toms are partly different from those observed after inhalation. Vague abdominal symptoms, skin lesions, anemia, leukopenia, vascular involvement as well as skin cancer dominate the clinical picture. The determination of Total As concentration in urine has been the dose test most widely used for monitoring human exposure to As, and in general there is a good correlation bet• ween level of exposure and urinary excretion. However, total arsenic concentration in urine may provide a misleading estimate of intake of inorganic As, since even a sigle meal of fish or marine food causes greatly elevated arsenic excretion for several days. Inoganic arsenic and its metabolites may be differentiated from other forms of organic As in urine and quantitated by ion-exchange chromatographic separation coupled with AAS. Total As in blood is an indicator mainly reflecting recent exposure and, as urinary excre• tion, is influenced by marine food consumption. There is sufficient information neither to establish a relationship between intensity of As exposure and its concentration in blood nor to assess the validity of measuring As metabolites in blood.

Abbreviations

InAs = inorganic arsenic MMA = monomethylarsonic acid DMA = dimethylarsinic acid AsF = arsenic in marine food TMA = total metabolites of inorganic arsenic (InAs + MMA + DMA) TAs = total arsenic (InAs + MMA + DMA + AsF)

27

Arsenic

Introduction

Chemical and Physical Properties

Arsenic is a silver-grey brittle, metallic-looking substance, symbol As, a metalloid, atomic number 33, atomic weight 74.92, specific gravity 5.73 (25°C), melting point 817°C at 28 atm, boiling point 613°C (sublimes), v.p. 1 mm Hg at 372°C. Naturally occurring arsenic is made up of one stable isotope As75. Artificially produced arsenic has a number of isotopes, the most important of which are As74, and As76. When arsenic is heated in air, it burns forming a white smoke of arsenic trioxide As203. From both the biological and toxicological points of view, there are three major groups of arsenic compounds: inorganic compounds, organic compounds, arsine gas (Table 1). Inorganic arsenic compounds may have different oxidation states with valence ranging from - 3 to +5. Of the many inorganic arsenic compounds, several are quite soluble in water, e.g. arsenic trichloride, sodium arsenite and sodium arsenate, and interchanges in valence state may occur in water solutions, depending on the pH of the solution as well as the presence of other substances which can be reduced or oxidized. Among organic compounds (e.g. monomethylarsonic acid, dimethylarsinic acid and arsenobetaine), some are stable in the environment and have chemical and toxicological properties different from those of the inorganic compounds.

Human Exposure to Arsenic

There are many natural and artificial sources of arsenic. Arsenic is abundant in the earth's crust and the number of arsenic-containing minerals exceeds 160. The most common of them are: arseno-sulpho-pyrites, FeAsS (As: 46%); arsenopyrite, FeAs (As: 73%); arsenosulphides As2S3,As4S4 (As: 60-70%). Average As content in soil ranges from 2 to 5 mg/kg but in regions of recent vulcanism can be up to 20 mg/kg. Arsenic is present in traces In every type of water: sea water ordinarily contains from 0.006 to 0.03 mg/kg of As, river water from undetectable levels to 0.23 mg/kg, and tap water from undetectable levels to 0.1 mg/kg (but In general less than 0.005 mg/kg) (WHO, 1981). Plants contain As in varying quantities: the richest In As are vegetables grown on arsenical- compound sprayed soils. Wine made from grapes sprayed with arsenic-containing insecticides may contain levels up to 0.5 mg/l of arsenic, mainly in trivalent from. Also tobacco may contain arsenic originating from insecticides, mainly lead arsenate. Values ranging between 12-42^g of arsenic per cigarette have been reported in various brands of American cigarettes, with values declining to 8 mg/kg of tobacco during the last twenty years due to decreased use of arsenical pesticides (Ishinishi et al., 1986). In most european countries the use of arsenical pesticides is forbidden by law since many years. Also food contains As in varying quantities: with the exception of some kind of fish, crustácea and shellfishes, most foods contain low levels of arsenic, normally below 0.24 mg/kg (NAS, 28

Table 1. Inorganic and organic arsenic compounds relevant to human health: environmental occurrence or industrial use.

Chemical abstract Valence ( ), Physical Industrial use or biological Chemical Name Synonyms and formula Reg. Serial No. properties occurrence

Arsine 7784-42-1 Arsine. Arsenic trihydride, (-3); colourless neutral gas, organic synthesis, solid state hydrogen arsenide. AsH3 slightly w.s. electronic components, by product of metal smelting

Arsenic (elemental) 7440-38-2 Grey arsenic, metallic (0); gray, shiny, brittle alloys in order to increase arsenic. As metallic-looking rhom- the hardness and heat resi• bohedra; w.i. stance

Arsenic trichloride 7784-34-1 Butter of arsenic. (+3); a yellowish oily liquid; pottery industry, manufac• AsCI3 w.s., decomposes turing of clorine-containing arsenicals

Arsenic trioxide 1327-53-3 Arsenic Oxide, As203. (+3); white amorphous or manufacture of glass, White arsenic; arsenic crystalline powder. Heated insecticide, rodenticide sesquioxide, arsenious sublimes. Siowly w.s. anhydride As203 Combines with water to form arsenious acid

Sodium arsenite 7784-46-5 Arsenenous acid, sodium (+3); white or grayish veterinary use, insecticide, salt. Arsenious acid sodium hygroscopic powder, very wood preservative. Used meta-arsenite. NaAs02 w.s. also in combination with other salts such as: calcium arsenite, lead arsenite, cu- pric acetoarsenite

Arsenic pentoxide 1303-28-2 Arsenic oxide, As205. (+5); white amorphous manufacture of colored Arsenic acid, arsenic anhy• deliquescent powder. Freely glass, insecticide, wood dride. As205 w.s. Combines with water to preservative arsenic acid

Lead arsenate 7784-40-9 Arsenic acid, H3ASO4, (+5); heavy white powder, constituent of various lead (+2) salt, (1:1). Acid w.i. On heating emits toxic insecticides also with other lead arsenate; arsenate of fumes salts: calcium arsenate, lead, approx. PbHAs04 sodium arsenate, potassium arsenate

Methylarsonic acid 2163-80-6 Arsenic acid, methyl-mono- white powder, w.s. constituent of various pesti• sodium salt. Methanearsonic cides also as mixture with acid monosodium salt; methylarsonic acid disodium sodium methylarsonate, salt. Excretion product of sodium methanearsonate. mammalian metabolism CH3AsO(OH)ONa

Dimethylarsinic acid 124-65-2 Arsenic acid, dimethyl- colourless cristais hygro• constituent of various sodium salt sodium salt. Cacodilic acid scopic, w.s., slightly pesticides. Excretion sodium salt, sodium cacodi- soluble in ethanol product of mammalian late, sodium dimethylar- metabolism sinate, arsine oxide, hydro- xymethylsodium salt. (CH3)2AsO(ONa)

Trimethylarsine 593-88-4 Gosio gas. As(CH3)3 colourless neutral gas, produced after metabolic slightly w.s., decomposes transformation of As compounds by bacteria and fungi especially in sewage

Arsanilic acid 98-50-0 4-aminobenzenearsonic white christalline powder, stimulator of growth of food- acid, aminophenylarsine w.s. producing animals NH2C6H4-AsO(OH)2

+ Arsenobetaine (CH3)3As CH2COOH w.s. organic arsenical compound in marine organisms

Arsenocholine (CH3)3As + CH2CH2OH w.s. organic arsenical compound in marine organisms

Legenda: Chemical Abstracts services names are heavy typed w.s. = water soluble w.i. = water insoluble 29

1977). Arsenic concentrations ¡n marine organisms may range between 1 and 100 mg/kg depending on different species. Both lipld-soluble and water-soluble organic compounds have been Isolated from marine organism; arsenobetaine, a betaine compound in which As replaces N, seems to be the most commonly occurring water-soluble arsenical in fish and crustácea (Vahter et al., 1983; Yamauchl et al., 1985). The amount of arsenic Ingested daily by humans via food Is greatly influenced by the amount of seafood in the diet. An average daily intake of arsenic from food ranging from 0.01 to 0.02 mg has been estimated in U.S.A. Higher values, ranging between 0.07 and 0.37 mg, have been reported for the Japan population (WHO, 1981; Ishinishi et al., 1986). Arsenic is normally present ¡n trace amount in the human organism with an estimated body burden of about 10-20 mg. The biological role, if any, of this "normal" amount in man is unknown. Burning of coal and smelting of metals are major souces of arsenic environmental pollu• tion (NIOSH, 1975; Sabbioni et al., 1983; Sabbioni et al., 1984). In special circumstances, environmental background exposure to arsenic may be con• siderable and toxic effects to humans from this source have sometimes been reported (Ben- cko et al., 1977). World annual production of arsenic and its compounds during the past few decades was around 60.000 tons, with some decrease in the 1970s (WHO, 1981). Arsenic is used in a number of industrial processes: in the chemical industry, the manufac• ture of special glasses or crystals, as a pesticide and in drugs. It is also present in the processing of copper, gold and lead ores. Occupational exposure to arsenic may occur in any process where it is used. The most serious possibility of exposure to fumes and dust occurs in connection with the smelting of metal-ores and in the manufacture and use of arsenic-containing pesticides. Airborne arsenic concentrations in smelting industries have been found to vary between 0.005 and 20 mg/m3 (NIOSH, 1975). The levels reported ¡n the past were very high in relation to the currently recommended exposure limits of 0.01 mg/m3 (TWA-OSHA for inorganic As and its compounds) and 0.2 mg/m3 (TLV-ACGIH for As and its soluble compounds) (Nordberg, 1983). Since inorganic arsenic is recognized as a human carcinogen, it is still largely debated if acceptable regulatory decisions about permissible limits of long term exposure can' be setted.

Metabolism

The metabolism of arsenic in man is rather complex since its fate in human body is depen• dent on the type of compound and its chemical form.

Absorption

Human intake of inorganic arsenic and its compounds occurs mainly through ingestion or through inhalation of arsenic containing airborne particulates. Absorption of arsenic from gastrointestinal tract can occur following ingestion of food, water or drugs containing arsenic or as result of Inhalation and subsequent mucocilary clearance. Gastrointestinal absorption depends on whether the arsenic compounds is given In solu• tion or as undissolved particles. Both human and animal data indicate that over 90% of an Ingested dose of dissolved inorganic arsenic (trivalent or pentavalent) is absorbed within a few hours (Ishinishi et al., 1986; Thome et al., 1986). In the case of arsenic trioxide, which is only slightly soluble in water, gastrointestinal absorption is slower and depends on parti• cle size and pH of the gastric juice. Organic arsenic compounds in seafoods are readily absorbed after ingestion. Absorption of arsenic following inhalation has been much less investigated. Retention, deposi• tion, and absorption in the respiratory system again depends on size and solubility of inhal• ed particles. In animals, all the commonly occurring compounds of arsenic have been found to be cleared from the lungs with a half-life of a few days (Smith and Thome, 1986). Human absorption of airborne arsenic has been investigated ¡n copper smelter workers (Pinto et al., 1976; Smith et al., 1977; Vahter et al., 1986) and in volunteers exposed to smoke of arsenic impregnated tobacco (Holland et al., 1959; Arnold et al., 1970). A biphasic model of pulmonary clearance, in which 75% of deposited materials Is cleared with a half-life of 4 days and 25% is cleared with a half-life of 10 days, has been proposed 30

(Thorne et al., 1986). However, high concentrations of arsenic found in autoptlc lung samples of smelter-refiner workers who had retired 2-19 years before death, indicate the possibility of a long retention time of inhaled arsenic on certain exposures (Brune et al., 1980).

Distribution

After absorption by the lungs or the gastrointestinal tract, arsenic is rapidly distributed through the blood. As a consequence of the efficient interchange of valence in vivo for inorganic arsenic (i.e. trivalent into pentavalent and viceversa), tissue partitioning ¡s somewhat similar for the different chemical forms. In man as well as in most animal species, excluding hair and nails where a long-term retention is seen, arsenic is uniformly distributed throughout all organs and tissues, though after ex• posure it may be Initially concentrated to some extent in kidney, liver and lungs (Bertolero et al., 1981; Vahter and Marafante, 1985). In rats, however, a substantial part of the absorbed arsenic is accumulated into the red blood cells where It is bound to haemoglobin, mainly in the form of dimethylarsinic acid (Vahter et al., 1984). Studies on the tissue distribution of arsenobetaine in mice, rats and rabbits have shown that this arsenical is rapidly cleared from most tissues (Vahter et al., 1983).

Biotransformation

The biotransformation of inorganic arsenic depends on the type of compound and its chemical form. Trivalent inorganic arsenic is oxidized in vivo as indicated by the findings of pentavalent arsenic in urine of both animals and humans exposed to arsenite (Mealey et al., 1959; Ben- cko et al., 1976; Vahter and Envall, 1983; Lovell and Farmer, 1985). In animals, and recent• ly in man, the opposite reaction has also been observed (Vahter and Envall, 1983; Lovell and Farmer, 1985). Both trivalent as well as pentavalent arsenic, the latter probably after reduction to trivalent, are methylated in the body to methylarsonic acid (MMA) and dimethylarsinic acid (DMA), which may be considered to be detoxification products since they have a lower affinity for tissue constituents than their precursors (Vahter & Marafante 1985). In man after environmental background exposure, the urinary excretion of As consists of about 5 to 10% unmethylated arsenic, 5 to 10% MMA and 20 to 30% DMA and 30 to 60% AsF, (Buchet et al., 1981a; Foà et al., 1984; Vahter and Lind, 1986). Experimental findings suggest that two different enzymatic activities are involved in the methylation of inorganic arsenic in mammals. Although dimethylarsinic acid seems to result from methylation of monomethylarsonic acid, the possibility that these metabolites are pra duced by two completely indipendent pathways cannot yet be conclusively excluded (Buchet and Lauwerys, 1985). Observations in man support the latter hypothesis, in volunteers given increasing amounts of As3+ the urinary excretion of dimethylarsinic acid increases only at the highest dose and faster than that of MMA which increases linearly over the whole dose range (Buchet et al., 1981b). Moreover an excretory pattern suggesting different rate limits for the two methylation steps was observed in patients acutely intoxicated by As3 + , where several days elapsed before DMA became the preponderant urinary metabolite (see Figure 1)(Mahieu et al., 1981; Foà et al., 1984; Lowell and Farmer, 1985). The metabolic fate of organoarsenicals in humans is uncertain. Elevated levels of inorganic arsenic, MMA or DMA have been never found in human blood or urine after ingestion of seafood, suggesting that organoarsenicals from marine organisms are excreted without mixing with the inorganic As pool in the body (see Figure 2) (Foà et al., 1984; Yamauchi and Yamamura, 1984). Also in animals, arsenobetaine is apparently not biotrasformed and excreted as such, mainly in urine (Vahter et al., 1983). Arsenocholine is to a great extent, oxidized to arsenobetaine (Marafante et al., 1984). Arsanilic acid and dimethylarsinic acid are not converted to in• organic arsenic in vivo (Overby and Frost, 1962; Vahter et al., 1984).

Excretion

Excretion of absorbed arsenic Is mainly via urine. Only a small amount s excreted in feces. In primates, excretion after a single injected dose of InAs is essentially complete in six days (Smith and Thorne, 1986). 31

10.000 100

5.000

D α tot. As o o in. As A *. DMA • « MMA

Figure 1. Urinary excretion of As and its metabolites in a suicidal acute poisoning after ingestion of about 3 g As203. No measure of As metabolites was performed till the 8th day (from Foà et al., 1984).

• tot. As

■ in. As A MMA * DMA

200 ­

100 _

40 hours

Figure 2. Urinary excretion of As and its metabolites after ingestion of 100 g shrimps (containing 1.1 mg arsenic as organoarsenical compounds) (from Foà et al., 1984). 32

Recent investigations using radioanalytical techniques have shown that retention of InAs In different animals is strictly correlated with the ability of different species to methylate inorganic arsenic, and also dependent on the degree of Interaction and binding with in­ tracellular components (Bertolero et al., 1981). In man, the biological half-life of arsenic has been measured after Intravenous injection of radiolabelled arsenite to patients with terminal cancer (Mealey et al., 1959). Excretion proceeds in three phases, with half-times of about 24 hr, 84 hr, and 8 days respectively. After oral intake of radiolabelled pentavalent arsenic (about 0.01 μς of arsenic), 66% was excreted with a half-time of 2.1 days, 30% with a half-time of 9.5 days and 3.7% with a half-time of 38 days in the three phases (Tarn et al., 1979; Pomroy et al., 1980). Experimental studles In volunteers with higher doses have shown that 50% of the orally administered dose (500μg of inorganic trivalent arsenic) is excreted in urine within 4 days (Buchet et al., 1981a). In the case of repeated oral intake of arsenite (500­800/¿g of As/day, over few days), 60­70% of the dally dose is excreted in the urine daily (Mappes 1977; Buchet et al., 1981b). Following inhalation of arsenic dust, there is a long retention of As in the lungs. In smelter workers, a sevenfold increase of arsenic in the lungs and a threefold Increase of arsenic in the liver has been reported (Brune et al., 1980; Wester et al., 1981). The arsenic in the liver decreased markedly with the length of retirement, but that in the lungs did not, in­ dicating a very long biological half life of As particulate probably related to its low solubility (Brune et al., 1980). Experimental studies with mice, rats, hamsters and rabbits have shown that pure arsenobe­ taine, when orally administered, is rapidly absorbed through the digestive tract and is mostly excreted into urine in a very short time, being less retained in organs and tissues (Vahter et al., 1983; Yamauchi et al., 1986). The urinary excretion pattern of arsenic in man follow­ ing a single oral ingestion of trimethylarsenic compounds contained in fish (750μg of total As intake mainly as arsenobetaine and arsenochollne), indicates that a portion correspon­ ding to 50% of the ingested amount of these compounds is excreted In urine during the first 6 hrs after the ingestion (Yamauchi and Yamamura, 1984).

Effects on Humans

Very small amounts of certain arsenic compounds may play a physiological role and act as growth promoting agents in animals (Anke et al., 1977). Arsenic compounds are otherwise generally regarded as very potent poisons. Their acute toxicity, however, varies greatly among organic and Inorganic compounds, depending on oxidation state, solubility in biological media and routes of entry. Moreover arsenic is biotransformed in the body and, as is probably true for other metals, it is becoming In­ creasingly apparent that the chemical form in which As is present In the organism may be a determinant of toxicity to the same extent as is its concentration. Effects of arsenic compounds on humans have been reviewed in details by: NIOSH (1975), NAS (1977), Nordberg et al., (1979), I.A.R.C. (1980), WHO (1981), Fowler (1983), Ishinishi et al. (1986). Only the main aspects are summarized in the following.

Acute poisoning

Acute arsenic poisoning is uncommon at the workplace but may result from improper use of chemicals, unadvertent ingestion of contaminated materials, or accidents. The severity of poisoning depends on the dose, the route of absorption and, when ingested as a particulate, on the particle size and solubility in water of the compound. Accidental or voluntary Ingestion of inorganic arsenicals, mainly arsenic trloxide, have been often described. Profound gastrointestinal damage resulting in severe vomiting and diar­ rhoea is the major lesion. Other acute symptoms and signs Include muscular cramps, facial oedema and cardiac abnormalities. Shock can develop rapidly as a result of the dehydra­ tion followed by systemic collapse. Symptoms may ensue within a few minutes after ex­ posure if the arsenic compound is in solution form, but may be delayed several hours ¡f it is a solid, the rate of absorption largely depending on its solubility. Subacute effects mainly involve the respiratory, gastrointestinal, cardiovascular and hematopoietic systems (with anemia end leukopenia, especially granulocytopenia). In sur­ vivors, a peripheral sensory neuropathy frequently develops a few weeks after ingestion. The fatal dose of ingested arsenic trioxide has been reported to He between 70 and 180 mg (WHO, 1981). Human data on differences in toxicity between trivalent and pentavalent arsenic are limited. A certain tolerance against acute poisoning is believed to develop upon repeated long term 33 exposure with low doses. This phenomenon, however, is not well documented in the scien• tific literature. Exposure to irritant and vescicant arsenic compounds in air, such as arsenic trioxide, arsenic trichloride and the arsenical war gases, have been known to cause acute damage to the mucous membranes of the respiratory system and acute symptoms on exposed skin. Severe irritation of the nasal mucosae, larinx and bronchi as well as conjunctivitis and dermatitis occur in such cases. Arsine gas is a powerful hemolytic poison. Acute arsine exposure ¡s characterized by nausea, vomiting, headache, shortness of breath, haemoglobinuria, and shock within 2-24 hours after exposure. Oliguria or anuria occurs commonly after about 24 hours, due to blockage of the renal tubules by hemoglobin casts (Foà and Bertolero, 1983; Ishinishi et al., 1986).

Chronic poisoning Systemic and local effects

Chronic effects after long-term exposure to arsenic do not constitute a well-defined clinical syndrome. With long-term exposure to arsenic via ingestion from food, drinking water or medication, symptoms are partly different from those observed after inhalation exposure.

Vague abdominal symptoms, flushing of the skin, pigmentation and hyperkeratosis dominate the clinical picture. Anemia, leukopenia and liver involvement often occur. In addition there may be vascular involvement. Intake of arsenic with drinking-water in Taiwan has been reported to have given rise to peripheral gangrene, the so called "blackfoot disease', (Tzeng, 1977). Such vascular changes have not been reported in occupational arsenic exposure; however peripheral vascular disorders, in the form of endoangitis obliterans and acroder• matitis atrophicans, have been found in vineyard workers and in workers employed in the manufacture of insecticides containing inorganic arsenic (Grobe, 1976). Effects on peripheral circulation in the extremities, manifested as white fingers (Raynaud's phenomenon) and increased prevalence of vasospastic reactivity (indicated by a low finger blood pressure after local cooling), have been reported among smelter workers exposed to arsenic dust (Lagerkvist et al., 1986). Liver involvement has been more commonly seen In persons exposed for a long time via oral ingestion than in those exposed via inhalation, particularly in vineyard workers con• sidered to have been exposed also through drinking contaminated wine. Liver damage, evolving to cirrhosis, has been also attributed to inorganic arsenic ingestion in people who had eaten contaminated food or had been treated with As-containing antiasthenlc drugs such as Fowler's Solution (liquor arsenicals B.P. 1983, arsenite solution containing 7.6 g As/I) (WHO, 1981).

A number of skin lesions have been attributed to chronic exposure to inorganic arsenic compounds and they are somewhat different depending on modality of exposure. Sym• metric verrucous hyperkeratosis of the palms an soles is a characteristic finding after long term ingestion of inorganic arsenic via drinking-water or drugs. These lesions have been reported from regions in Argentina, Chile, Taiwan, Japan and Mexico where the contents of arsenic in drinking-water were elevated (concentrations ranging between 0.4 and 1.8 mg/l). In two surveys performed respectively on a population of 40,121 in the endemic area of Taiwan (Tseng, 1977) and on 318 individuals in the North of Mexico (Cebrian et al., 1983) the prevalence of keratosis and hyperpigmentation was found to be significantly associated with age, arsenic level in drinking-water and estimated total dose of arsenic ingested. Similar findings have also been observed after long-term Ingestion of arsenic-containing drugs. In a study by Fierz on 262 patients treated 6-26 years earlier for chronic dermatosis with diluted (1:1) Fowler's Solution, a dose-response relationship was found between the amount of arsenic ingested and the incidence of palmoplantar hyperkeratosis. In the patients who had received the equivalent of more than 400 ml of Fowler's Solution (3 g of arsenic), the observed hyperkeratosis prevalence was of more than 50% (Fierz, 1965).

In occupational exposure, following skin contact, dermatological disorders characterized by eczematoid symptoms of varying degrees of severity can develop. Dermatitis is primarily localized on the most heavily exposed areas, In dependence on the concentration and dura• tion of exposure. Chronic dermal lesions such as hyperkeratosis, warts and melanosis of the skin lesions may follow. These chronic lesions may occur after many years of occupa• tional or environmental exposure and particularly hyperkeratosis may evolve Into precancerous and cancerous lesions (IARC, 1980). 34

In exposure to airborne arsenic, skin lesions may result from local irritation. Mucous mem• brane lesions are most classically reported as perforations of the nasal septum after chronic arsenic Inhalation exposure. The irritation of mucous membranes also extends to the larinx, trachea and bronchi, and damage to the respiratory tract has been found particularly in smeltery workers who were exposed to levels of airborne inorganic arsenic as high as 7 mg/m3, as well as to fumes and other inorganic dust (NIOSH, 1975; Pinto et al., 1976).

Peripheral nervous diseases are frequently encountered in individuals surviving acute and subacute oral poisoning by inorganic arsenic compounds or exposed through arsenic drugs; electromyographic abnormalities have sometimes been recorded ¡n people living ¡n areas highly contaminated by arsenic (NAS, 1977). Similar disorders have occasionally been observ• ed also after heavy occupational exposure to inorganic arsenic, but very few studies have dealt with possible neurological effects following occupational long-term low-level exposure to arsenic. No effects on the nervous system have been reported in the investigations per• formed in Argentina, Chile or Taiwan where the subjects studied used drinking-water con• taining remarkable amounts of arsenic (WHO, 1981). Arsenic polyneuropathy generally involves the long nerves of the legs and is characterized by both motor dysfunction and paresthaesia; in less severe cases only sensory or unilateral neuropathy may occur. An involvement of cranial nerves (loss of hearing) has also been claimed as an effect of arsenic over-exposure in children living in an arsenic polluted area (Bencko et al., 1977). Learning impairment and loss of hearing were also described among children who surviv• ed to a subacute arsenic poisoning occurred in Japan (Yamushita e al., 1972). The victims ingested dried milk contaminated with arsenic over a period of 33 days with an estimated amount of 1.3-3.6 mg of arsenic per day, and symptoms of arsenic poisoning developed when a total of about 80 mg of arsenic had been ingested (Hamomoto, 1955). All the above reported effects are attributable to inorganic arsenic. Far less is known about long-term toxicity of organic arsenic compounds. Medication with some organic arsenic compounds, such as (2-amino-2-oxyethyl)-amino-phenylarsonic acid (Tryparsamide), has in• duced side effects mainly to the central nervous system (WHO, 1981). Other organic compounds have limited industrial use, and information on long term toxici• ty on humans is not available. Toxicity of organic arsenical compounds present In sea food has not been extensively tested in animals and man. In male mice a LD50 higher than 10 g/kg has been observed with oral administration of pure arsenobetaine (Kaise et al., 1985). In human limited experience from single observations indicated that an oral dose in the order of 1-2 mg is not associated with observable effects.

Carcinogenicity

Existing studies consistently point to a causal relationship between skin cancer and heavy exposure to inorganic arsenic via contaminated drinking-water or use of medications con• taining arsenic. Skin cancer as a result of arsenical poisoning is characterized by multifocal lesions over the entire body; among the several types of skin cancer epitelioma, usually of the squamous type and developing on the site of keratotic lesions, Is the most common. The total number of observed cases is well over 1000, with exposure occurring most fre• quently via the oral route, either through contaminated drinking-water or medication (IARC, 1980). Ingestion has usually taken place over several decades with a daily dose of several mg of arsenic. In medication, the arsenic was mainly in the form of arsenite, while the form and the oxidation state of arsenic In contaminated water, has been poorly investigated. Preliminary results, obtained on health surveys of chronic arsenicism, indicate a propor• tion of trivalent and pentavalent arsenic of 30 and 70% of total arsenic content in well- waters both from Alaska and from the North of Mexico (Harrington et al., 1978; Cebrian et al., 1983). The relationship between arsenic exposure and risk of skin cancer has been investigated in different epidemiological studies. A positive relationship between the prevalence rate of skin cancer and total ingested dose of arsenic has been established in patients treated with Fowler's Solution (Fierz, 1965) and In chronic arsenicism by contaminated well-water (Tzeng et al., 1968; Tzeng, 1977). From these results after extrapolation to low doses with a linear non-threshold model, a 5% of skin tumor prevalence per 10 grams of total ingested arsenic may be estimated. Assuming a 2 litre/day average consumption of drinking water, and a life span of 70 years, the life-time risk from arsenic in drinking-water has been set at about 5% per 0.2 mg As/litre (25% per 1 mg As/litre) (WHO, 1981).

Arsenic exposure has been also associated with lung cancer. Lung cancer has been found to account for excess mortality in several groups of workers exposed to inorganic arsenic 35 in gold mining (Osburn, 1969) and In smelting of non-ferrous metals, expecially copper (Pinto et al., 1977). In the mortality study of Pinto, the excess of deaths was found to be highly correlated to the duration and intensity of exposure (Pinto et al., 1977; Pinto et al., 1978). Moreover, among smelter workers, a multiplicative effect of arsenic exposure and cigarette smoking has also been observed, and the latency time between onset of exposure and appearance of cancer resulted inversely related to exposure and to co-factors such as other smelter dusts, sulfut dioxide and cigarette smoked (Pershagen et al., 1981). The studles available do not solve the question of whether arsenic alone is a lung car• cinogen, or concomitant exposure to sulfur dioxide or other dust, such as occurs in smelter operations, ¡s necessary for the development of lung cancer. Excess mortality from respiratory cancer has been also reported among pesticide manufac• turing workers exposed to lead or calcium arsenate (Mabuchi et al., 1979), and in vineyard sprayers exposed to arsenic compounds (Roth, 1957). Revising the available literature the International Agency for Research on Cancer has classified inorganic arsenic compounds as lung and skin carcinogens for humans (I.A.R.C, 1980). There is also some evidence pointing to the possibility that persons exposed to inorganic arsenic compounds via medication, insecticides, contaminated wine or drinking-water may suffer a higher incidence of other types of cancer, such as angiosarcoma of the liver and possibly stomach cancer. However, the relationship between arsenic exposure and a higher risk of cancer in organs other than skin and lungs has not been confirmed. In spite of the epidemiological evidence of association between arsenic exposure and skin and lung cancer in humans, a reliable and consistent induction of cancer in animals with arsenic has not yet been achieved. According to I.A.R.C, so far available evidence of arsenic carcinogenicity to animals has to be considered inadequate. Evidence of mutagenicity and activity ¡n short-term tests is limited as well (I.A.R.C, 1980; I.A.R.C, 1982).

Biological Indicators

Indicators of internal dose

Human exposure or body burden of arsenic has been assessed through the measurement of the total arsenic concentration in blood, hair and urine. Depending on the source (am• bient or industrial air, drinking-water, or food) and the length of exposure, these biological indicators assume a different meaning.

Arsenic in blood

Total arsenic In blood has been widely used to monitor the environmental and the occupa• tional exposure to arsenic. The figures available on blood arsenic content are reported in Table 2. Values observed in the general population greatly vary from one study to another, depending also on diet and on different analytical techniques used. Due to analytical limitations on available methods, arsenic has generally been evaluated as total content and only few investiga• tions have dealt with blood measure of inorganic arsenic and its metabolites (Yamamura and Yamauchi, 1980; Yamauchi and Yamamura, 1984). Because of the short half-life of both Inorganic and organic arsenic ¡n blood, in episodical or short term exposure blood content reflects the extent of exposure for only a short period after absorption and the time elapsed between the onset of exposure and sampling is of relevance in the interpretation of the results. When exposure is continuous and steady, as it would be the case of chronic exposure through contaminated drinking-water, blood arsenic should reach a steady state, quantitative• ly dependent on intensity of exposure, but no data are available for this relationship. Limited data are only available in occupationally exposed workers. On chronic low level exposure in glass industry blood As concentration was found higher in exposed subjects than in controls and, on a group basis, proportional to the levels of As exposure (Foà et al., 1984). Seafood content in diet may greatly influence blood arsenic levels as observed in reference values of the general population (Foà et al., 1984). Rapid changes in the blood concentra- 36

Table 2. Blood Arsenic concentrations in reference populations and in exposed subjects

General population or No Mean Value ± SD Unit of Analytical Reference Source of exposure Measure (*) Technique

General population 8 4.0 μθ/kg NAA Brune et al., 1966

Danish general population 7­1 e 2.5 μ­g/l NAA Heydron, 1969

Taiwan general population 6­17 22 /ug'i NAA Heydron, 1969 Taiwan blackfood disease patients 60 μφ NAA Heydron, 1969

U.S.A. general population 5 10­130§ μg/kg COL Wagner and Weswing, 1974

Forestry workers (dimethylarsinic 5 10­270§ μg/kg COL Wagner and Weswing, acid exposed subjects) 1974

10­yrs old boys exposed to airborne 10 4.5 t 2 μ9^ NAA Bencko and Symon, arsenic 1977 10­yrs old boys as reference 10 1.9 i 1 μθ/kg NAA Bancko and Symon, 1977

U.S.A. general population 49 1.5 /"g/i AAS Kagey, 1977 (Female non smokers) U.S.A. general population 50 2.3 /"g/i AAS Kagey, 1977 (Female smokers)

Japan general population 10 9 /ug'kg AAS Yamamura and Yamauchi, 1980

Plant workers (arsenic acid 6 17­47§ /"g/kg AAS Yamamura and manufacture) Yamauchi, 1980

India general population 8 5.9 μg/kg** NAA Dang et al., 1983

Mexico general population 100 2 í 1 μ­g/i AAS Olguin et al., 1983

Drinking water exposed subjects 50 81 5 μ­g'i AAS Olguin et al., 1983 (arsenic content = 0.41 mg/l)

Italy general population 148 5.1 £7 μ­g/i AAS Foà et al., 1984

Glass industry workers 38 15.7*13 μ­g/i AAS Foà et al., 1984 (As203 exposed)

Workers applying monosodium Abdelghani et al., 1986 methanearsonate as herbicide*** pre­exposure values 19 131 6 μ­g/l AAS post­exposure higher values 10 23 i 6

•whole blood **post mortem samples '"arithmetic mean t st. error § minimum an maximum values Abbreviations: NAA = neutron activation analisis; AAS = atomic absorption spectrometry COL = colorimetrie method

tion of total arsenic have been observed after a single seafood meal as little as a portion of prawns (750 μg of As as total arsenic intake) (Yamauchi and Yamamura, 1984). A model for the different metabolic forms of arsenic that accounts for both the short half­ time of arsenic in blood and the biological half­time in the whole body, has yet to be ade­ quately constructed (Thorne et al., 1986). For these reasons, blood total arsenic has to be regarded as an indicator of exposure of limited value.

Arsenic in hair

In the general population and in exposed subjects, arsenic has been usually found ¡n higher concentrations in hair and nails than in other tissues or parts of the body. This has been explained by the high affinity of keratin­SH groups for trivalent arsenic. Arsenobetaine, unlike Inorganic arsenic, does not accumulate ¡n hair (Vahter et al., 1983). The raising of arsenic content in hair after exposure to inorganic arsenic via drugs, drink­ ing water or airborne arsenic in industrial as well as ambient air, may be due both to incor­ 37 poration of absorbed arsenic into the growing portion of the hair root, and to external contamination. An average level of less than 1 mg/kg has been observed in subjects from different coun­ tries (Smith, 1964; Leibscher and Smith, 1968) and values 10­20 times higher have measured in populations chronically exposed to arsenic­rich drinking water (Terada et al., 1960; Hind­ marsh et al., 1977, Olguin et al., 1983) and in patients treated with arsenic­containing drugs (Shapiro, 1967; Pearson and Pounds, 1971; Lelslie and Smith, 1978; Ribard et al., 1986). The mean concentration of arsenic In hair has been found to reflect the level of air­pollution, since hair arsenic concentrations in the general population were generally found to increase as the concentrations of arsenic in air increased (Heydron, 1969; Hammer et al., 1971). Values of arsenic in hair ranging from 0.6 to 10 mg/kg (colorimetrie method, reference value 0.15 mg/kg) were observed in boys living in the proximity of a power plant burning local coal with a high arsenic content (Bencko et al., 1977). Increased levels in the range of 1­17 times of reference values, were also reported among children living near an open cast metal mine in Ireland (Corridan, 1974) and ¡n various copper smelter towns in U.S.A. (Baker et al., 1977). The arsenic content in hair of occupationally exposed workers can reach values very high. Levels of several hundreds mg/kg have been observed in arsenic mining or refining workers (Yamamura and Yamauchi, 1980), but values within a normal range have been found In hair of retaired workers extensively exposed to arsenic in the past even in the presence of symptoms of chronic arsenic poisoning, such as dermatological disorders, perforation of nasal septum, peripheral nervous disturbance or sequelae of earlier effects of poisoning (Ishinishi et al., 1977). Many attempts have been made to correlate the content of arsenic in hair and the level of exposure. A high degree of correlation has been observed between hair arsenic content and the dose ingested after short term treatment whith trivalent arsenic as Fowler's Solu­ tion (Pearson and Pound, 1971). Nevertheless, no correlation was found between arsenic hair content and either observed signs of chronic arsenicism (Terada et al., 1960) or water arsenic content in subjects long term exposed to contaminated well­water (Hindmarsh et al., 1977). In occupational exposure, no quantitative evaluation has been reported on the relationship between arsenic hair content and ambient levels of exposure. From the available results on the whole, arsenic content in hair seems to be a relevant indicator of inorganic arsenic exposure through ingestion, mainly on a group basis and pro­ viding that external contamination can be excluded or ¡s only slight. The use of hair arsenic as indicator of exposure through inhalation is limited as no reliable methods are available to distinguish between hair arsenic from airborne external contamina­ tion and arsenic incorporated into hair after absorption and metabolism in the body (Bos et al., 1985). However it can be of value on a group basis when external contamination can be controlled in some way.

Arsenic in urine

The levels of total arsenic in urine of the general population without anomalous sources of exposure, are reported ¡n Table 3: values are generally in the range of 5­50 μg As/I. The urine arsenic concentration appears to be the best indicator of current or recent past (few days) exposure to organic and inorganic arsenic and total arsenic in urine has been used traditionally to assess occupational exposure to Inorganic arsenic (see Table 3). However total arsenic content in urine, as previously mentioned, is greatly Influenced by dietary intake of organic arsenic compounds, mainly arsenobetaine, which is present in very high concentrations in certain fish species and crustaceans and is rapidly excreted in the urine after fish intake. One single meal of this food may Induce an urinary excretion of more than 1000μg As/I, the normal excretion in persons without fish consumption being 5 to 30 μς As/I (Foà et al., 1984; Yamauchi and Yamamura, 1984) (see also Fig. 2). In biological monitoring of exposure to Inorganic arsenic, the separate measure of inorganic arsenic and its methylated compounds is required (Schrenk and Schreibeis, 1958; Bertazzi et al., 1982). The concentration of total metabolites of inorganic arsenic, which include in• organic arsenic itself, monomethylarsinic acid, and dimethylarsonic acid, ¡s not influenced by diet and is always lower than total urinary arsenic concentration, which also includes dietary arsenic (as example see Fig. 3). In environmental exposure, a high correlation coefficient was found between arsenic con• centration in drinking water and in the urine (Harrington et al., 1978). In occupational exposure, significant but rather low correlation coefficients between total 38

Table 3. Urine arsenic contents in reference populations and in exposed subjects.

No Urine Arsenic tpgl\) Subjects Reference subjects In As MMA DMA TMA TAs Pilot plant workers 29 — — — — 20-2000 Schrenk and Schreibis, 1958

Unexposed subjects 4 5.8 1.6 15.0 22.6 22.5 Braman and Foreback, 1973

Smelting ores workers 24 — — — — 38-539 Pinto et al., 1976

Exposed 10 yrs old 25 25125 Bencko and Symon, boys (****) 1977 Reference 10 yrs old 44 — — — — 96 1 15 boys

General populat. (**) 41 2.6 ±2 3.4 12 11.5 ï 2 17.5 21.2 ±2 Smith et al. 1977 Copper smelter 83 3.8-11.7 4.9-20.8 17.0-64.1 25.7-96.6 24.7-66.1 workers (***)

Copper smelter 1000 — — — — 220-400 Pinto et al., 1978 workers (*****)

Belgium general pop. 10 5.3-35 3.1-12.5 7.1-74.6 15.5-113.1 28-170 Buchet et al. 1980 Mediterraneum gen. 10 0-3.8 0.6-2.2 6.2-22.4 6.9-25.6 36-1180 population

Ingestion of Asili 5 31-165 13-170 30-596 74-934 91-953 Buchet et al., 1980 (3 mg) in volunteers

Arsenic trioxide 18 16-57 8-15 77-87 — — Yamamura and exposed workers Yamauchi 1980

Glass industry 25 — — — — 304 Roels et al., 1982 exposed workers

General population 924 270-1200 Bertazzi et al., 1982 after AS2O3 environ, contam. (*****) " " " "

Worker exposed to lead arsenate pre-exposure values 5 1.4 i 1 1.8 ±2 11.4 Í 11 14.6 14.2 ί 11 Wojeck et al., 1982 post-exposure values 5 26.6 í 26 49.8 ± 63 189.8 1111 266.2 266.1 ï 157

Reference italian 148 1.9 + 1 1.9 ±1 2.1 ; 1 5.9 17.2 1 11 Foà et al., 1984 population

Glass industry 38 7.1 ï 6 4.7 ±8 18.9 i 19 30.7 37.7 130 Foà et al., 1984 exposed workers

Copper smelter 526 — — — — 14-29§ Lilis et al., 1984 workers (***)

Mine and refinery 200 — — — — 160-270 Javelaud. 1986 As203 workers (*****)

General popul. from 49 12.4 î 11 72.9190 Vahter and Lind, 1986 Stockholm (*) General popul. from 50 — — — 9.7 ί 7 41.1 146 Västeras (*)

Workers exposed to methanearsonate pre-expos. values** 13 2-83 Abdelghani et al. 1986 post-expos, values** 14 2-1637

Glass industry 16 22.8113 18.1 ±12 67.7 ±45 103 ±63 — Foà et al.. 1987 exposed workers

Values expressed as mean ±SD, or as observed range *) values adjusted to a specific gravity of 1.016 **) values as geometric mean ± standard error ***) values as range of geometric mean observed in low and high exposée groups ****) values expressed as mg/l and measured with colorimetrie method (SDDC) *****) lower and higher mean values in different surveys § values expressed asfjglg creatinine Abbreviations: InAs = Inorganic Arsenic; MMA = Monomethylarsonic acid; DMA = dimethylarsinic acid; TMA = Total metabolites of arsenic (InAs + MMA + DMA); T As = Total Arsenic (InAs + MMA + DMA + AsF); AsF = Arsenic ¡n marine food 39

As μ g/' '500. ] tot. As

I in. As

400­ | MMA

U DMA

300.

200­

100­

mM- m¿ u ■ m m jfai ι Subjects

Figure 3. Patlern of urinary excretion of As and its metabolites in 9 glass workers exposed to As203. Subjects A, B, C and I with low exposure and consumption of marine food. Sub­ jects D, E, F, G and H, with high occupational exposure and no consumption of marine food (from Foà et al., 1984).

arsenic ¡η urine and the concentration of inorganic arsenic in the air were found (Pinto et al., 1976; Smith et al., 1977; Vahter et al., 1986). As recently pointed out by many authors (Landrigan, 1981 ; Foà et al., 1984; Vahter et al., 1986), in monitoring occupational exposure the time intercurring between onset of exposure and sampling is very important in order to assess the meaning of the arsenic concentra­ tion in urine. In smelter workers, at the beginning of exposure or on the first working day following an interruption of exposure (at least for 48h), the urinary arsenic concentration (Total Metabolites) has been observed to increase gradually during the first 24­hours of exposure and to level off at 48­hours (mean urinary levels of 220μg/l of total arsenic metabolites), indicating an equilibrium between exposure and excretion. The concentration in the late evening after a day of exposure was very similar to that in the early morning the day after, and both values appeared well correlated to the total daily excretion (Wahter et al., 1986). Similar behaviour of urinary arsenic has been observed in the biological monitoring of glass

Industry workers, exposed to As203. Exposure to As was assessed by measuring urine As concentration in three spot urine samples collected in the morning, at the beginning of the workshift and at the end of the workshift after four days of exposure. The results in­ dicated a significant increase in urinary As concentration in exposed workers when com­ pared with a reference population (urinary levels between ΘΟ­ΙΟΟμρ/Ι of total arsenic metabolites), but no significant differences were seen among the three samples collected at the different times of the day (Foà et al., 1987). Changes in urinary arsenic values between the beginning and the end of a five­day ex­ posure period, suggesting accumulation during the week, have been observed in subjects

heavily exposed to airborne As203 (Pinto et al., 1976; Javelaud, 1986). In these studies the reported values of total arsenic ¡n urine ranged between 200 and 400μ9/Ι. An increase in geometric means arsenic concentration in urine has been also observed during the work­ week in subjects occupationally exposed to methanearsonate (Abdelghani et al., 1986) and to lead arsenate (Wojeck et al., 1982). From these results it may be concluded that at high levels of exposure an incomplete ex­ cretion of arsenic and consequent body accumulation may occur, as reflected by the in­ crease of urinary arsenic concentration during the work­week. Depending on levels and length of previous exposure, the difference between urinary arsenic concentration in samples collected before and after the work­shift, may reflect with different accuracy the daily levels of exposure. Since urinary concentration gives a good estimate of exposure on an individual basis, to carry out analyses of arsenic in urine on a routine basis is the best approach to monitor occupational exposure to arsenic. 40

Indicators of effects

Chronic effects after long­term exposure to arsenic do not constitute a well­defined clinical syndrome. Besides carcinogenicity, a great variety of effects involving peripheral nervous system, skin, peripheral circulation, liver and blood forming system have been reported after occupational exposure to arsenic. Their attribution to arsenic toxicity ¡s in some cases uncertain and not sufficiently demonstrated and the dose­response relationship for all these effects is not known. No information is available on specific indicators of early functional or biochemical effects related to As exposure. However attempts have been made to study the effects related to chronic low dose exposure and to Identify indicators usable for biological monitoring of early toxic response. In glass workers chronically exposed to low levels of airborne arsenic (range of methylated arsenic in urine between 5­130μg/l) neurological and neurophysiological effects were In­ vestigated and the results showed the absence of any clinical or subclinical modification pertinent to the peripheral nervous system (Foà et al., 1983). In a study of an analogous group of glass workers (range of methylated arsenic in urine between 20 and 240μg/l), kidney toxicity has been investigated by the use of very sen­ sitive and specific tests, namely urinary excretion of brush­border antigens of the kidney proximal tubule (measured by mouse monoclonal antibodies), ß2­microglobulin, retinol bin­ ding protein and albumin. No significant difference has been observed, for any renal func­ tion parameter with the exclusion of a borderline significance of retinol binding protein, after comparing the exposed subjects with a matched reference group (Foà et al., 1987).

Methods for As determination in biological materials

The chemical forms of arsenic present in urine may be differentiated and quantitated by ¡on­exchange chromatographic separation coupled with arsine generation and atomic ab­ sorption spectrometry. The most recent methods were developed from those originally pro­ posed by Tarn et al. (1979) and Uthe et al. (1974) with some modifications (Buratti et al., 1984). Due to interference of the organic matrix, these methods are scarcely applicable to blood samples, where only the total­As concentration after dry ashing may be accurate­ ly measured (Stringer & Attrep, 1979). The ashing procedure may be adopted also to deter­ mine total­As concentration ¡n urine samples and total As organic forms in marine food. There are other methods to determine the concentration of arsenic in urine without any influence from organic arsenic compounds deriving from food. The most simple one is the determination of arsenic in urine carried out by arsine generation and atomic absorption spectrophotometry on the urine sample in toto (Norin and Vahter, 1981 ; Vahter et al., 1986). This method, however, only provides a value which represents the total sum of Inorganic arsenic and its methylated matabolites. When using spot urine samples to monitor As exposure, correction of arsenic excretion by creatinine concentration or specific gravity of urine is frequently performed. There is uncertainty about the validity of these adjustments and no specific studies are available concerning this practice with urinary arsenic (Alessio et al., 1985).

Conclusions

Recent studies suggest that in human organism inorganic As is in part converted to methylated from that are eliminated in the urine. Organic As from seafood Is unable to mix with the inorganic As pool in the organism, and this may be one of the reasons why its toxicity differs remarkably from that of inorganic As. The determination of As concentration In urine has been the dose test most widely used for biological monitoring of human exposure to As. There is evidence that total urinary excretion of As may not be a reliable indicator of ex­ posure to inorganic As and the chemical speciation of the different forms of As in urine is required. This fact has Implications not only for the surveillance of people with occupa­ tional or environmental exposure to As, but also for the development of a correct dose­ response relationship for As toxicity. When a differentiated measure of urinary arsenic is not feasible, monitoring of exposure to inorganic arsenic should be accompanied with a detailed record of dietary habit, refer­ red to marine food consumption for at least the two days preceding the biological sampling. 41

Blood As concentration has to be regarded as an indicator of exposure of limited validity, being scarcely correlable to urinary excretion of total As. However, in workers exposed to As203, blood As average concentrations on a group basis, are proportional to the Inten• sity of exposure. Time elapsed between onset of exposure and sampling is critical for blood and to some extent for urine, whereas hair could be a better Indicator for past exposures if external contamination of the hair can be sufficiently controlled.

Need for Further Research

Further research is required on the following aspects: 1. relationship between external dose and internal dose for both low and high water soluble As compounds; 2. relationship among the indicators of internal dose (blood As and urinary As forms); 3. experimental studies to determine the mechamism of transformation from inorganic to organic arsenic and from As5+ to As3+ and viceversa. 4. identification of mechanisms of As toxicity and of biochemical receptors of As in the tissues in order to Identify possible early indicators of effect; 5. relationship between Indicators of internal dose and early toxic effects; 42 References

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SABBIONI E., GOETZ L, BRIGNOLI G.: Health and environmental implications of trace metal released from coal­fired power plants: an assessment study of the situation in the Euro­ pean Community. Sci. Tot. Environ. 40, 141­154, 1984. SABBIONI E., GOETZ L, SPRINGER Α., PIETRA R.: Trace metals from coal­fired power plants: derivation of an average data base for assessment studies of the situation in the European Communities. Sci. Tot. Environ. 29, 213­227, 1983. SCHRENK H.H., SCHREIBEIS L. Jr.: Urinary arsenic levels as an index of industrial ex­ posure. Am. Ind. Hyg. Ass. J. 19, 225­228, 1958. SHAPIRO H.: Arsenic content of human hair and nails. Its interpretation. J. Forensic Med. 14. 65­71, 1967. SMITH A.D., THORNE MC (eds): Review of recent literature: arsenic. In: Pharmacodynamic models of selected toxic chemical in man. MTP Press Limited, Lancaster, for the Commis­ sion of the European Communities, vol. 2, pp. 1­15, 1986. SMITH H.: The interpretations of the arsenic content of human hair. J. Forensic Sci. Soc. 4, 192­199, 1964. SMITH T.J., CRECELIUS E.A., READING J.C Airborne arsenic exposure and excretion of methylated arsenic compounds. Environ. Health Perspect., 19, 89­93, 1977. STRINGER CE., ATTREP M.: Comparison of digestion methods for determination of organoarsenicals in waste water. Anal. Chem. 51, 731­734, 1979. TAM G.K.H., CHARBONNEAU S.M., BRYCE F., POMROY C, SANDI E.: Metabolism of in­ organic arsenic (74As) In humans following oral ingestion. Toxicol. Appi. Pharmacol. 50, 319­322, 1979. TERADA H., KATSUTA K., SASAGAWA T., SAITO H., SHIRATA H., FUKUCHI K., SEKIYTA T., HIROKAWA S., WATANABE Y., HASEGAWA K., OSHINA T.: Clinical observations of chronic toxicosis by arsenic. Nihon Rlnsho, 118, 2394­2403, 1960 (EPA Translation n. TR 106­74). THORNE MC, JACKSON D., SMITH A.D. (eds): Appendix 1 : arsenic. In: Pharmacodynamic models of selected toxic chemicals In man. MTP Press Limited, Lancaster, for the Com­ mission of the European Communities, vol. 1, pp. 30­102, 1986. TSENG W.P.: Effects and dose­response relationships of skin cancer and blackfoot disease with arsenic. Environ. Health Perspect. 19, 109­119, 1977. TSENG W.P., CHU H.M., HOW S.W., FONG J.M., LIN CS., YEH S.: Prevalence of skin cancer ¡n an endemic area of chronic arsenicism in Taiwan. J. Natl. Cancer Inst. 40, 453­463, 1968. UTHE J.F., FREEMAN HC, JOHNSTON J.R., MICHALIK P.: Comparison of wet ashing and dry ashing for the determination of arsenic in marine organisms, using methylated arsenicals for standards. J. Assoc, of Anal. Chem. 57, 1363­1365, 1974. VAHTER M., ENVALL J.: In vivo reduction of arsenate in mice and rabbits. Environ. Res. 32, 14­24, 1983. WAHTER M., FRIEBERG L, RAHN3TEN B., NYGREN Α., NOLINDER P.: Airborne arsenic and urinary excretion of metabolites of Inorganic arsenic among smelter workers. Int. Arch. Occup. Environ. Health 57, 79­91, 1986. VAHTER M., LIND Β.: Concentrations of arsenic in urine of the general population in Sweden. Sci. Tot. Environ. 54, 1­12, 1986. VAHTER M., MARAFANTE E.: Reduction and binding of arsenate ¡n marmoset monkeys. Arch. Toxicol. 57, 119­124, 1985. VAHTER M., MARAFANTE E., DENCKER L: Metabolism of arsenobetaine ¡n mice, rats and rabbits. Sci. Tot. Environ. 30, 197­211, 1983. VAHTER M„ MARAFANTE E., DENCKER L: Tissue distribution and retention of 74As­ dimethylarsinic acid in mice and rats. Arch. Environ. Contam. Toxicol. 13, 259­264, 1984. WAGNER S.L., WESWIG P.: Arsenic in blood and urine of forest workers. Arch. Environ. Health, 28, 77­79, 1974. WESTER PC, BRUNE D., NORDBERG G.F.: Arsenic and selenium in lung, liver and kidney tissue from lead smelter workers, Br. J. Ind. Med. 38, 179­184, 1981. W.H.O.: Environmental health criteria. 18 ­ Arsenic. World Health Organization, Geneva, 1981. WOJECK G.A., NIGG H.N., BRAMAN R.S., STAMPER J.H., ROUSEFF R.I.: Worker exposure to arsenic In Florida grapefuit spray operations. Arch. Environ. Contam. Toxicol., 11, 661­667, 1982. 46

YAMAMURA Y., YAMAUCHI H.: Arsenic metabolites in hair, blood and urine in workers exposed to arsenic trioxide. Ind. Health, 18, 203-210, 1980. YAMASHITA N., DOI M., NISHIO M., HOJO H., TANAKA M.: Current sta:e of Kioto children by arsenic tainted Morinaha dry milk. Jpn. J. Hyg. 27, 364-399, 1972. YAMAUCHI H., YAMAMURA Y.: Metabolism and excretion of orally ingested trimethylarsenic in man. Bull. Environ. Contam. Toxicol. 32, 682-687, 1984. YAMAUCHI H., KAISE T., YAMAMURA Y.: Metabolism and excretion of orally administered arsenobetaine ¡n the hamster. Bull. Environ. Contam. Toxicol. 36, 35C-355, 1986. Biological indicators for the assessment of human exposure to industrial chemicals

Cobalt A. Ferioli, R. Roi, L. Alessio

49

Summary

Cobalt ¡s an essential element for man. It is widely used in the metallurgical industry main• ly for the production of hard metals. The main sources of human absorption are diet and occupational exposure. The most Im• portant route of absorption in occupational situations is the respiratory apparatus, by the inhalation of dusts or fumes. On workers, long term exposure to cobalt and its compounds mainly involves effects on respiratory apparatus, skin and myocardium. Prolonged inhalation of metallic cobalt may cause a chronic airway obstruction or, for very high exposures, an interstitial pulmunary fibrosis. In subjects occupationally exposed, this metal and its compounds were demonstrated to produce an allergic contact dermatitis. Ingestion of cobalt chloride has been related to a serious epidemic outbreak of car• diomyopathy observed in heavy beer drinkers. In occupational exposure evidences of cardiomyopathy have been occasionally described in workers after long term exposure in metallurgical industry. Occupational exposure to cobalt can be monitored by measuring its concentrations in am• bient air and, for soluble cobalt compounds, also by measuring the concentration of the metal in urine and blood. The data so far available on the use of blood and urinary cobalt for the biological monitor• ing, indicate that these indicators are suitable for assessing exposure on a group basis. Of them, the determination of urinary cobalt level seems to offer more informations on the degree of exposure in the course of each working day and over the whole working week. It also appears to be sufficiently sensitive since it permits the evaluation of exposure in workplaces where the environmental cobalt concentrations are around 0.1 mg/m3. Actually no Indicators of early biological effects are available for the biological monitoring of occupational exposure to cobalt. It is not possible to fix health based limit values for these biological indicators since studies are lacking on which concentrations of cobalt in urine and blood may be considered "safe" both in short and long term exposure.

Abbreviations

CoB cobalt in blood CoU cobalt in urine Co Air cobalt in air AAS Atomic Absorption Spectrometry NAA Neutron Activation Analysis

51

Cobalt

Introduction Chemical and Physical Properties

Cobalt (chemical symbol Co) is a gray, hard, magnetic, ductile metal which exists in two allotropie form. In Table 1 are summarized the physochemical properties of Cobalt and its compounds. (Stockinger 1981; Izmerov, 1986). It is a relatively rare element. In soil the concentration of cobalt is in the order of 1 to 40 mg/Kg. In water a few/zg/l or less, and in ambient air about 1 ,ug/m3 or less. Levels ex­ ceeding 10 ,ί/g/m3 have occasionally been reported in heavily industrialized towns (Schroeder et al., 1987; Elinder and Friberg, 1986).

The main sources of human absorption are diet and occupational exposure.

Table I - Physochemical properties of cobalt and its compounds

Molecular Molecular Melting Boiling Solubility in Solubility in biological Substance formula weight point point chemicals fluids

Cobalt (ll-lll)* Co 58.94 1493°C 3100°C HCI,H2S04,HN03 0.03mg/100ml(37°C) physiological solution 15.25mg/100ml (37°C) biood plasma 20 mg/100 ml (37°C) blood serum

Cobaltous oxide (II)* CoO 74.94 1935°C H20 (0.313mg/100g) blood serum: 27.3mg/100g (37°C) in 10 days 49mg/100g (37°C) in 30 days

Cobalt oxides (ll-lll)* Co30„ 240.80 900°C NaOH (molten); Na2C03 (solution)

Cobaltic oxide (III)* Co203H20 183.88 H20 (84/g/100g, 37°C 5.39mg/100g (37°C) in 30 days) blood serum

Cobaltous sulfate CoS04 155 H20 (39.3g/100g, 25°C) CoS047H20 96.8°C

Cobaltous chloride CoClj 129.84 724°C 1049°C H20 (52,9g/100g, 20°C)

Cobaltous carbonate C0CO3 118.94 H20 (0,11g/100g, 15°C, under C02 pressure)

Cobaltous acetate Co (C2H302) 249.08 H20 4H20

'valency states 52

Industrial uses

About 60% of the world's production of cobalt comes from the mines in Zaire, and to a lesser extent, from Zambia, Canada and United States (Payne 1977). It is obtained by sulfurizing roasting of cobalt-containing material, along with other methods of pyrometallurgy. This includes converting cobalt into solution and separating it from associated metals, chiefly chromium and nickel. In industry, 80% of the cobalt produced is used in metallic state: with chromium In certain heat resisting steels (Co: 25-65%); with chromium, nickel and molybdenum ¡n vitallium alloy (Co: 65%) used for biomedical protheses; with nickel and aluminium in the production of magnets; as Co60 in radiotherapy to replace radium. Cobalt oxides, are used as catalysts in after burners of Internal combustion engines to reduce environmental pollution; as a pigment cobalt Is used in paints (Stocklnger, 1981). It is a component of vitamin B12and it ¡s used as CoEDTA for the treatment of severe cyanide poisoning (Nagler et al., 1978). The most important use as far as industrial hygiene is concerned, is in the production of hard metals. This covers the powder metallurgy sector where, with a ceramic type pro• cess at controlled temperature, pressure and atmosphere, special metallic carbides are bonded by means of a metallic binding agent. In practice, the most widely used carbide is tungsten: the percentage of cobalt Is between 6% and 9% in three-quarters of produc• tion, and occasionally reaches 30% (Reinl et al., 1979).

Effects on humans

Cobalt Is an essential metal for the human organism as a component of cyanocobalamin or vitamin B-|2. No function for cobalt in human nutrition other than in vitamin B12 has been established. The target organ for cobalt is dependent on the exposure route. In occupational exposure the effects of cobalt and its compounds mainly Involve the respiratory tract and the skin. Industrial exposure to high levels of metallic cobalt in the hard metal industry may result in a severe type of pulmonary fibrosis, named "hard metal disease" (Hartung and Valen• tin, 1983). Hard metal lung disease was first described in 1940 in German workers and since then many other observations have been made in many industrialized countries. Two different pictures of this disease have been described. The first is initially characterized by an Irritative respiratory syndrome with a dry, non pro• ductive cough which can be associated to a restrictive respiratory impairment. The disease progresses to interstitial fibrosis where the most significant symptom is an increasingly severe dispnoea. The onset of the fibrotic phase usually occurs after 10 years of exposure: if the subject is removed from exposure early enough, there may be a subjective improvement and regres• sion of the Irritative syndrome (Forrester et al., 1978; Sjogren et al., 1980; Meyer et al., 1981; Davison et al., 1983; Démets et al., 1984; Lahaye et al., 1984). In those few reports in which Co dust concentrations were measured, exposure levels exceeded the TLV-TWA (ACGIH) of 0.1 mg/m3 by at least ten folds (Stokinger, 1981). According to NIOSH (1981) mild fibrotic lung changes have been observed in workers exposed to Co Air concentra• tions of 0.1 - 0.2 mg/m3. It has been suggested that the onset of the fibrotic alterations is the result of an im• munological mechanism (Sjogren et al., 1980). The second form is initially characterized by work - related asmatic crises which can be induced also by broncoprovocation test with cobalt chloride. The disease progresses in this case to a chronic airway obstruction. According to conclusion reached by NIOSH (1981) respiratory irritation and lung function tests revealing signs of obstructive lung disease can be observed for Co Air concentrations of 0.06 mg/m3. This type of disease was suggested to have an allergic background (Elinder and Friberg, 1986). On the basis of experimental (Kerfoot et al., 1975) and human observations it has intended to change (already proposed from 1976) the ACGIH TLV - TWA from 0.1 mg/m3 to 0.05 mg/m3 (1984). An allergic contact dermatitis, closely resembling nickel dermatitis and characterized by erytematous papoules, has been observed in subjects exposed to cobalt. In these cases patch tests with 1 % cobalt chloride usually show positive reactions. Clinical observations seem to indicate that hard metal workers with simultaneous nickel and cobalt sensitivity 53

have more severe hand eczema than those with isolated cobalt or nickel sensitivity. Ce• ment dermatitis can be due to ¡persensitivlty to cobalt (Fisher and Rystedt, 1985; Garcia and Armisen, 1985). Long term peroral exposure to cobalt may produce lesions in the myocardium; this effect was also seen in animal esperiments after parenteral administration. The studies on these effects of cobalt were made after the reports of endemic outbreaks of fatal degenerative cardiomyopathy observed among heavy beer - drinkers, shortly after the Introduction of cobalt chloride and cobalt sulfate as foam stabilizers in beer (Bonenfant et al., 1967; Elinder and Friberg, 1986). Different cases of myocardiopathy have been associated with Industrial exposure, but unlikely cobalt exposure levels were not given. The autopsies performed on two subjects revealed higher cobalt concentrations in the myocardic tissues than in controls; the microscopic picture of myocardic tissue was found resembling that seen among cobalt beer drinkers (Barborik and Dusek, 1972; Kennedy et al., 1981). Moreover, in a review of Sovietic Literature (Izmerov 1986), 19 cases of cobalt cardiomyopathy have been reported, mainly among smelters. Their age was below 40 years and the average length of exposure was 8 years. Studies have also suggested that cobalt may play a role in the genesis of uremic heart disease (Lins and Pehrsson, 1976), since cobalt Is mainly excreted with the urine and in uremic patients there appears to be a marked retention of absorbed metals, with much higher serum Co levels in these patients compared to controls. It Is to note that in two recent investigations, workers exposed to Co Air concentrations often much more higher than the current ACGIH TLV-TWA (1986), did not show any biological or physiological alteration (Morgan, 1983; Angerer et al., 1985). There is still today no consensus on the possible carcinogenicity of cobalt. In rats, sub• cutaneous injection of cobalt induces sarcomas at the site of injection (Heath, 1980). Cobalt also possesses certain mutagenic properties demonstrated in "in vitro" experimental models of animals and plants (Elinder and Friberg, 1986). There are, however, no data available which suggest or Indicate that cobalt costltutes a cancer risk for occupationally exposed humans (Doll et al., 1980).

Metabolism The main routes of absorption of cobalt are via digestive and respiratory apparatuses. Dai• ly intake of cobalt from food is considered to be around 5 - 45//g Co/day (Elinder and Friberg, 1986). The gastrointestinal absorption (in the jejum) of soluble cobalt compounds is around 25% of the ingested amount, with large inter-individual variations. The most important route of absorption in occupational exposure is via the respiratory ap• paratus. Studies on hamster indicate that about one third of the inhaled Co-oxide is ab• sorbed (Wehner et al., 1977). For humans no quantitative data are yet available but from measurements of cobalt ¡n blood and urine samples obtained from exposed workers, (Co Air 0.09 mg/m3) it is evident that in humans inhaled cobalt is taken up from lung to a great extent (Alexanderson and Lidums, 1979). An almost constant feature in experimental observations and in studies on subjects expos• ed to hard metals is the presence in the lung tissue of a quantity of cobalt that is rather low (Kuhne, 1962; Tozawa et al., 1981). This seems to depend on the high solubility of cobalt ¡n fluids with high protein content, such as lung surfactant (Payne, 1977). One must however underline that, according to Hartung et al. (1982), the already quoted observations in humans had been obtained in studies performed with low reliable methods. These Authors on the contrary found concentrations of Cobalt of 1010//g/Kg wet weight in a lung biopsy of an exposed subject. This 36 year old man had been exposed to the grinding dust from sintered hard metal parts for more than ten years and presented a marked fibrosis at histology. As controls 21 autopsy lungs were used without exposure to cobalt. The values obtained were 3.0 - 33.0//g/kg wet weight. Furthermore in a study carried out on rats, after a single Intratracheal instillation of cobalt, it was demonstrated that the Co-oxide clearance took quite a long time: the time to clear 50% of the initial lung burden was 15 days (Rhoads and Sanders, 1985). In humans the highest concentrations are found in the liver: ¡n autopsy studies they have been shown individual values ranging from 0.01 to 0.07 mg Co/kg wet weight mainly in the form of vitamin B12. It has been estimated the liver to hold about one-fifth of the total body burden of cobalt. There are no indications of any accumulation of cobalt with age (Elinder and Friberg, 1986). 54

Once adsorbed from the gastrointestinal tract or the lung, cobalt will mainly eliminated via urine (about 80%) and via faeces (about 15%) (Elinder and Friberg, 1986). The result of a study made by Alexandersson and Lidums (1979) in cobalt exposed workers showed that the metal is usually excreted with the urine in two phases (Figure 1): a rapid phase that occours in the first 24­35 hours, and a subsequent slower phase. Certain data indicate that retained cobalt has a biological half­time in the order of years (Elinder and Friberg, 1986).

Biological Indicators

The concentration of Cobalt in biological fluids and tissues is very low. The colorimetrie techniques have been demonstrated to be not reliable to measure this metal. Since the introduction of sensitive and specific AAS­methods, the measure of this metal in urine and blood has been demonstrated to be of use in biological monitoring of workers exposed to its soluble compounds. The cobalt concentrations in urine obtained from not occupationally exposed subjects are in the order of 0.1­4.0/ig/l, while cobalt blood concentrations are In the order of 0.005­2.0 μgl\ (Table 2). These values have been obtained with AAS or NAA methods: CoU evaluated with a colorimetrie determination was found to be higher, with values between 1.5 and 7 /¿g/l and a mean of 3.6 μgl\ (Hubbard et al., 1966).

Table 2 ­ Urinary and blood cobalt concentrations of non­occupationally exposed subjects.

No. of Authors Method Mean Range S.D. Subjects

Coleman et al., 1973 MAA 10 0.5 figli Lidums, 1979 AAS 25 0.5 μgl\ 0.1 ­ 1.2 Kennedy and Dornan, 1981 AAS 0.18/(g/g cr. Schumaker and Angerer, 1981 AAS 10 0.38 /¿g/l 0.1 ­ 0.75 Hartung et al., 1982 AAS 1.3 μ g/l Scansetti et al., 1983 AAS 8 0.41 /¿g/l 0.22 CoU 0.28/ig/g cr. 0.18 Angerer et al., 1985 AAS 79 0.08/yg/l Ichlkawa et al., 1985 AAS 20 2.0/¿g/l 1.0­ 4.0 1.0 Molln­Christensen and Mikkelsen, 1985 AAS 46 0.8/¿g/l 1.59 Posma and Dijstelberger, 1985 AAS 0.0 ­ 2.0/¿g/g cr.

Coleman et al., 1973 MAA 5 0.7/¿g/l 20 0.6/¿g/l Wester, 1973 (χ) MAA 7 0.52/¿g/l 0.43 CoB Ichikawa et al., 1985 AAS 20 1.9/íg/i 1.1 Molin­Christensen and Mikkelsen, 1985 AAS 46 0.24/¿g/l 0.005 ­ 0.6 0.12 Posma and Dijstelberger, 1985 (χ) .AAS 0.0­ 0.3 ¿¿g/l

(x) cobalt in serum cr. = creatinine

In subjects occupationally exposed to cobalt the urinary excretion of cobalt and the blood or serum concentrations of the metal are significantly higher than in controls. The concen­ trations of soluble cobalt compounds in urine and blood increase in proportion to the level of exposure. Both these tests have been proposed as indicators of exposure in biological monitoring of workers. One of the first evidence of a relationship between CoB and cobalt exposure levels has 55 been reported by Alexandersson and Lidums (1979). They found in workers with heavy ex­ posure to cobalt (0.09 mg/m3) CoB levels of 10.5 ±10.9//g/l, while in another group ex­ posed to Co concentration of 0.01 mg/m3, they found CoB levels of 0.7 ±0.2 «g/l. Studing 10 groups of workers, Ichikawa et al. (1985) found a good correlation between CoB and Co Air, determined with personal samplers. It's to note that the Authors did not study the relationship between CoB and Co Air using the individual data of the workers (n. 147) expored to cobalt, but they used the mean values of the two parameters as observed ¡n the different groups (Figure 2). The blood and urinary cobalt concentrations prove to be usually well correlated (Figure 3) (Angerer et al., 1985; Ichikawa et al., 1985; Molin Christensen and Mikkelsen, 1985; Posma and Dljstelberg, 1985). However the evaluation of urinary cobalt excretion seems to offer more advantages than determination of blood cobalt concentration. Infact its two phases elimination curve can furnish useful information both on current and cumulature exposure (Figure 1). Moreover the CoU concentrations increase considerably with exposure, while the Increase of serum and blood cobalt levels are smaller. As studied on workers exposed to hard metal dust, the initial phase of cobalt elimination is rapid and due to this fact there Is an increase of urinary cobalt excretion at the end of the work­shift. Moreover there is a progressive increase In the urinary levels of cobalt as the work­week proceeds, with the highest levels of the metal being detected at the end of the work­shift (Scansetti et al, 1983 ­ Pellet et al, 1984) (Figure 4). Recently Scansetti et al. (1985) studing a group of 26 workers exposed to environmental Co concentration around 0.1 mg/m3 examined the relationship between environmental Co concentration and urinary cobalt levels of spot samples collected in different periods of the week. The results obtained in this work, seem to confirm the previous data. A good correlation (r = 0.83) was found between CoU values at the end of the work­shift on Mon­ day and environmental Co concentrations measured on the same day (Figure 5) and bet­ ween CoU values at the end of the work shift on Friday and the mean exposure values for that week (r = 0.75). Moreover they found that the differences of CoU concentrations between the end shift of Friday and the before shift on Monday, good reflect the mean cobalt exposure levels of the preceding weeks (Figure 6).

5000

4000

rr 3000 D

2000 O υ

1000

^^'cocP—0~σ­

T r 35 70 hours TIME AFTER EXPOSURE

Figure 1 Urinary cobalt concentrations ai different times after cessation of exposure ¡η 4 workers with different degrees of exposure (From Alexandersson and Lidums, 1979). 56

μ g/d I

2.50 y = 0.0044 χ +0.23

= r 0.96 χ ' ^ 2.00 Ρ 0.001 *>' D **" ^ O O s" ^^^^ _i m 1.50 ^ ^^^^ Σ *"" ^^^^^ «■■* "­­ ^^ ^^^ ^^^ < 1.00 O ''^s^^'"" υ s" ^¿^ ****

0.50 • -Ό<^--" —' ÆZ^^^

*s%-<Λ^s · ** Γ . ι ι 100 200 300 400 ug/m3 COBALT IN AIR

Figure 2 Relation between mean cobalt exposures and mean cobalt blood concentrations from ten groups of exposed workers (From Ichikawa et al., 1985).

//g/l y = ­11.2 + 7.52 r = 0.862

600

rr 400

< m O u 200

~ι 1 «g/l 10 20 30 40 COBALT IN BLOOD

Figure 3 Urinary and blood cobalt levels of single workers with different degrees of exposure (From Angerer et al., 1985).

m =morning e= evening

rr

cumulative exposure '5 days) < O O

basal MON. TUE. WED. THU. FRI. SAT. SUN. exposure

Figure 4 Theoretical behaviour of urinary cobalt levels before and after the working shift during a working week (From Pellet et al., 1984). 57

«g/i

60 y = 0.287 χ + 0.828 r =0.831

45­

Lil Ζ rx Ζ) Ζ Η _Ι 30­ < < Ο

15­

60 80 100 120 /¿g/mJ COBALT IN AIR Figure 5 End­shift urinary cobalt versus exposure on Monday (From Scansetti et al., 1985).

I 1 1 3 20 40 60 80 100 120 «g/m COBALT IN AIR

Figure 6 Δ urinary cobalt (after­shift on Friday/before­shift on Monday) versus mean exposure of preceding weeks (From Scansetti et al., 1985). 58

In the Authors' view adjustment for creatinine of urinary cobalt values does not seem to give any particular advantage as the relationships between CoU and environmental ex• posures were more significant when CoU values expressed in «g/l were considered. A similar Investigation was made by Ichikawa et al. (1983) who studled the relationship between Co Air and CoU of spot samples collected at the end of the shift on Thursday. They found an almost perfect relationship (r = 0.99) using Co Air and CoU mean values of 10 different groups of workers with different degrees of exposure. Following interruption of exposure, the urinary CoU levels In subjects exposed to Co Air concentrations around 0.1 mg/m3, return to "normal" rather slowly, reaching values com• parable to those of control subjects after about 4 weeks (Scansetti et al., 1985). Subjects with higher exposure (ranging from 0.07 to 8.61 mg/m3) after six weeks of cessa• tion of work showed not only higher levels of CoU (5 - 7 times) than the controls, but also of CoB (5 - 9 times). The values of both tests, on the same workers, did not drop significantly after two years, despite the fact the average concentration of Co in the environment were, after Improvement of workplace, around 0.05 mg/m3 (Molin Christensen and Mikkelsen, 1985) (Table 3).

Table 3 Urine cobalt concentrations of pottery plate painters measured before and after improvement of workplace (From Molin Christensen and Mikkelsen, 1985).

Pottery plate painters Pottery plate painters Variable 1982 1984

Urine Cobalt N 34 38 /íg/mmol créât. Mean 4.2 2.6 Median 2.3 2.0 SD 5.8 2.7 Range 0.24-29.1 0.16-16.1

It should be pointed out that these data are reported for exposure to soluble compounds, while following exposure to compounds with low solubility the levels of CoU and CoB rose only slightly (Molin Christensen and Mikkelsen, 1985).

Conclusions

Occupational exposure to cobalt can be monitored by measuring its concentrations in am• bient air and, for soluble cobalt compounds, also by measuring the concentration of the metal in urine and blood. Of the tests proposed for the biological monitoring of exposure, urinary cobalt determina• tion seems to offer more informations on the degree of exposure in the course of each working day and over the whole working week. Moreover the evaluation of CoU excretion appears to be a sufficiently sensitive indicator since it permits the evaluation of exposures in work places where the Co concentrations In air are around 0.1 mg/m3, which corres• pond to the current TLV - TWA proposed by ACGIH. The data so far available on blood and urinary cobalt seem to indicate that, due to the presence of individual differences in biological disposition of metal, these tests are suitable for assessing exposure on a group basis. Since the relationships between CoB, CoU and effects have been not investigated both in the short and long term exposure, it is not possible to identify at present values for these tests at which effects may occur. It should also be remembered that most of the pathological alterations caused by cobalt depend on an immunological mechanism and are therefore not dose - dependent: for this reason sensitized subjects probably would not be able to tolerate inahalation of even small amounts of Co. CoU and CoB can be therefore used only to evaluate the degree of exposure. 59

Research Needs

Further research is required dealing with the following aspects: critical evaluation and standardization of analytical methods for determination of CoB and CoU; evaluation of the values of CoB and CoU in non-exposed subjects and evaluation of the influence on these values of various factors, particularly smoking and drinking habits, with a view to setting up reference groups; standardization of the proper time (in the working day and in the working week) for collec• ting the biological samples for Co determination; evaluation of external/internal dose relationship at certain work places with various Co compound; identification of possible early metabolic effects and their correlations with CoU and CoB levels. 60 References

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LIDUMS V.V., 1979, Determination of cobalt in blood and urine by electrothermal atomic absorption spectrometry. Atom. Absorption Newsletter 18:71­72. LINS L.E., PEHRSSON K., 1976, Cobalt intoxication in uraemic myocardiopathy, Lancet 29:1191­1192. MEYER P.D., STOECKELC, GEIST T., LE BOUFFANT L„ ROEGEL E., 1981, A propos de trois nouveau cas de fibrose pulmonaire chez des affûteurs d'outils renforcés au carbure de tungstène, Poumon­Coeur, 37:165­170. MOLIN CHRISTENSEN J„ MIKKELSEN S., 1985, Cobalt concentration in whole blood and urine from pottery plate painters exposed to cobalt paint, in: International Conference "Heavy metals in the Environment", Athens­September 1985, Vol. 2, pp. 86­89, T.D. Lekkas, Ed. MORGAN L.G., 1983, A study into the health and mortality of men exposed to cobalt and oxides, J. Soc. Occup. Med. 33:181­186. NAGLER J., PROVOOST R.A., PARIZEL G., 1978, Hydrogen cyanide poisoning: treatment with cobalt EDTA, J.O.M. 20:414­416. NATIONAL INSTITUTE FOR OCCUPATIONAL SAFETY AND HEALTH, 1981, Criteria for Controlling Exposure to Cobalt, DHHS (NIOSH) Publication No. 82­107 p. 95. PAYNE LR., 1977, The hazards of cobalt, J. Soc. Occup. Med. 27:20­25. PELLET F., PEDRIX Α., VINCENT M., MALLION J.M., 1984, Dosage biologique du cobalt urinaire, Arch. Mai. Prof. 45:81­85. POSMA F.D., DIJSTELBERGER S.K., 1985, Serum and urinary cobalt levels as indicators of cobalt exposure in hard metal workers, in: International Conference "Heavy metals in the Environment", Athens­September 1985, Vol. 2, pp. 89­91, T.D. Lekkas, Ed. REINL W„ SCHNELLBACHER F., RAHM G., 1979, Lungenfibrose und entzündlichen Lungenerkraunkungen nach Einwirkung von Kobaltkontaktmasse, Zbl. Arsitsmed, 29:318­325. RHOADS K., SANDERS C.L, 1985, Lung clearance, translocation and acute toxicity of arsenic, beryllium, cadmium, cobalt, lead, selenium and ytterbium oxides following deposi­ tion in rat lung. Environ. Res. 36:359­378. SCANSETTI G., LAMON S., BOTTA G.C., TALARICO S., PIOLATTO G., 1983, Valutazione dell'esposizione a cobalto nella produzione di metalli duri con misure ambientali e biologiche. Med. Lav. 74:323­332. SCANSETTI G., LAMON S., TALARICO S., BOTTA G.C., SPINELLI P., SULOTTO F., FAN­ TONI F., 1985. Urinary cobalt as a measure of exposure in the hard metal Industry, Int. Arch. Occup. Environ. Health. 57:19­26. SCHROEDER H.A., NASON ΑΡ., TIPTON I.H., 1967, Essential trace elements In man: Cobalt, J. Chron. Dis., 20:869­890. SCHUMAKER­WITTKOPF E., ANGERER J., 1981, Praxlgerechte Methode zur kobaltbestlm­ mung in Harn, Int. Arch. Occup. Environ. Health 49:77­81. SJOGREN I., HILLERDALG., ANDERSSON Α., ZETTERSTROM O., 1980, Hard metal lung disease: importance of cobalt in coolants, Thorax, 35:653­659. STOKINGER H.E., 1981, The metals, in: Patty's Industrial Hygiene and Toxicology­Third Revis­ ed Edition ­ G. Glayton, F.E. Clayton Eds., pp. 1065­1619. A Wiley Interscience Publication ­ John Wiley and Sons, New York, Chichester, Brisbane, Toronto. TOZAWAT., KITAMURA H., KOSHI K., IKEMI Y., AMBE K„ KITAMURA H„ 1981, Experimen­ tal pneumoconiosis induced by cemented tungsten and sequential concentrations of cobalt and tungsten in the lungs of the rat. Jpn. J. Ind. Health, 23:216­228. WEHNER A.P., BUSCH R.H., OLSEN R.J., CRAIG D.K., 1977, Chronic Inhalation of cobalt oxide and cigarette smoke by hamsters, Am. Ind. Hyg. Assoc. J. 38:338­346. WESTER, P.O., 1973, Trace elements in serum and urine from hypertensive patients before and during treatment with chlorthalidone, Acta Med. Scand. 194:505­512.

Biological indicators for the assessment of human exposure to industrial chemicals

Endrin N.J. van Sittert and W.F. Tordoir

65

Endrin

Summary

Endrin is of a high order of acute oral and percutaneous toxicity to mammals. Studies in experimental species and experience in occupationally exposed workers have shown that the central nervous system is the critical organ. Convulsive intoxications due to short-term overexposure to endrin have been reported in endrin manufacturing and formulation workers. No cases of intoxication after short-term overexposures have been reported at endrin con• centrations in blood below 50-100 ng/ml, provided that blood was collected immediately at the end of exposure. As endrin is rapidly metabolised in humans, the determination of the urinary concentration of the major endrin metabolite, anti-12-hydroxyendrin, is more useful for measuring absorption of endrin in the body of occupationally exposed workers than the measurement of the con• centration of endrin in blood or serum. Hepatic enzyme induction has been demonstrated both in manufacturing and formulation workers and in spraymen repeatedly exposed to endrin. Following an exposure-free period this effect was shown to be reversible. It has been demonstrated that hepatic enzyme Induction does not occur below an urinary concentration of 130 mg anti-12-hydroxyendrin per gram creatinine.

Abbreviations

ALAT aspartate aminotransferase ASAT alanine aminotransferase LDH lactate dehydrogenase

67

Introduction

Endrin is the common name for an Insecticide belonging to the group of chlorinated cyclo­ dlene compounds, which is mainly used on cotton to control lepidopterous larvae. Endrin is the endo­endo stereo­isomer of dieldrin, but unlike dieldrin, is rapidly metabolised and excreted in mammalian species including man. Application of endrin to a variety of crops does not lead to residues in foodstuffs indicating little opportunity for residual build up in mammals or man (Donoso et al., 1979). Human exposure to endrin is therefore restricted to those occupationally exposed during manufacturing, formulation and repacking processes and during application on crops; the dermal route of exposure is the most important (Kummer and Van Sittert, 1986).

Chemical name (1R, 4S, 4aS, 5S, 7R, 8R, 8aR)­1, 2, 3, 4, 10, 10­hexachloro­1, 4, 4a, 5, 6, 7, 8, 8a­octahydrc~6, 7­epoxy­1, 4:5, 8­dimethanonaphthalene. Chemical structure

Molecular formula C12H8CI60 Relative molecular mass 381 Physical form light tan solid Purity 92% Vapour pressure 3.6 χ 10­4 N/m2 at 25°C (2.7 χ 10­7 mm Hg at 25°C) Density 1.64 g/ml at 20°C Solubility Endrin is moderately soluble ¡n acetone and benzene; sparingly soluble In alcohols and petroleum hydrocarbons; practically insoluble in water.

Effects on humans

Endrin is of a high order of acute oral percutaneous toxicity to mammals. Both the oral and percutaneous LD50 value in the male rat were 18 mg/kg (Gaines, 1960). The acute effects involve the central nervous system. Cases of accidental poisoning, sometimes lethal, have occurred by ingestion of endrin­contaminated flour or sugar (Davies 68 and Lewis, 1956; Coble et al., 1967; Weeks, 1967; Curley et al., 1970; Anon., 1985) and by accidental or intentional ingestion of formulated material (Jacobzlner and Raybin, 1959; Reddy et al., 1966; Karplus, 1971). Karplus (1971) estimated the lethal dose in man to be approximately 10 mg/kg. In cases of mild intoxication those affected have complained about dizziness, weakness of the legs, abdominal discomfort and nausea, usually without vomiting. Severe acute intoxications are manifested by peracute epileptiform convulsions usually without prodromal signs or symptoms. No neurological after-effects, except temporary elec- troencephalographical abnormalities, have been observed in survivors (Weeks, 1967). In seven cases of convulsive intoxication reported in endrin manufacturing and formulation workers, the return of the abnormal electroencephalographic pattern to the original pat• tern occurred within a few months. All seven intoxications were due to short-term overex• posure to endrin and took place in the first two years of employment of the workers. (Jager, 1970). Workers in this plant, exposed to endrin and related Insecticides, have been medically supervised since the start of endrin manufacture (Hoogendam et al., 1962 and 1965; Jager, 1970; Versteeg and Jager, 1973). Extensive pnyslcal, neurological, haematological and clinical determinations, including ALAT, ASAT, LDH, alkaline phosphatase, total serum protein and serum protein spectrum, did not reveal adverse effects attributable to endrin exposure. The rate of absenteism due to disease or accidents and the pattern of diseases were not different from operators working in other chemical plants.

Hepatic microsomal enzyme induction has been demonstrated both in manufacturing and formulation workers and in spraymen repeatedly exposed to endrin (Jager, 1970; Hunter et al., 1972; Ottevanger and Van Sittert, 1979; Kummer and Van Sittert, 1986). Ottevanger and Van Sittert (1979) showed that this effect is reversible. The relation between duration of exposure and the concentration of the urinary metabolite of endrin, antl-12-hydroxyendrin, In occupationally exposed workers in shown in Table 1.

Table 1. Concentrations of urinary anti-12-hydroxyendrin in occupationally exposed 'healthy workers'

Urinary Endrin in No Of anti-12-OH endrin Year blood Reference subjects (mg/l or mg/g creatinine) (ng/mL) mean min max

1973 71) 0.0623) 0.013 0.14 <2 Baldwin and Hufson, 1976 121a> 0.064") 0.016 0.118 <5 Ottevanger and Van Sittert, 1979 4 121b) 0.164 ) 0.025 0.325 <5 Ottevanger and Van Sittert, 1979 121c) 0.0634) 0.018 0.181 <5 Ottevanger and Van Sittert, 1979 1983 623) 0.06") 0.001 0.12 — Kummer and Van Sitten, 1986 222b) 0.25") 0.03 0.59 — Kummer and Van Sittert, 1986 102c> 0.55") 0.21 1.4 Kummer and Van Sittert, 1986

D Endrin manufacturing and formulations workers 1a) Urine samples collected after 6 weeks shut-down and maintenance of the plant 1b) Urine samples collected after 7 days production/formulation 1c) Urine samples after 3 days without exposure to endrin 2) Spraymen who sprayed a methylparathion/endrin/DDT ULV formulation 2a) Supervisors of spraymen urine samples were collected 20 hours 2b) Spraymen, after their 1st and 2nd application after the application. 2c) Spraymen, after their 3rd application 3) Concentrations expressed in mg/L The duration of one application 4) Concentrations expressed in mg/g creatinine was about 5 hours.

Numerous studies ¡n occupationally exposed persons do not show any indication of a car• cinogenic hazard (Jager, 1970; Wang and MacMahon, 1979; Ditraglia et al., 1981 and Rib- bens, 1985). However, in these studies workers mainly exposed to endrin were not separated from workers mainly exposed to other chlorinated hydrocarbon insecticides such as aldrin, dieldrin, heptachlor. 69 Metabolism

No studies under controlled conditions have been performed in human volunteers, relating defined oral or dermal doses to the parameter(s) used in biological monitoring. Some infor• mation on the metabolism of endrin in man has been obtained from occupationally expos• ed workmen (Jager, 1970; Ottevanger and Van Sittert, 1979; Baldwin and Hutson, 1980; Kummer and Van Sittert, 1986). The metabolism of endrin has extensively been studied in several animal species, e.g., rats, rabbits, mice, dogs and cows.

Absorption

Studies in experimental animals showed that endrin is readily absorbed into the blood from the gastrointestinal tract, through the skin or by inhalation of the vapour (Hine et al., 1954; Treon et al., 1955).

Distribution

Dogs were fed a diet containing endrin at a concentration of 0.1 mg/kg body weight for 128 days. The concentration of endrin in the blood was determined at weekly intervals and, although fluctuations occurred, no significant increase was found after the first week. The steady state endrin concentration in blood was 4 ng/mL. The concentration of endrin was highest ¡n adipose tissue (Richardson et al., 1967). In lactatlng cows, receiving endrin in the diet at a concentration of 0.1 ppm for 21 days, a steady state endrin in blood concentration of 0.7 ng/mL was reached in approximately one week after commencement of treatment. As in dogs, the concentration of endrin was highest in adipose tissue (Baldwin et al., 1976). In manufacturing and formulation workers, exposed to endrin for up to 19 years, the con• centration of endrin in blood was always below the limit of detection of 3 ng/mL (Hayes and Curley, 1968) or 10 or 5 ng/mL (Jager, 1970). Detectable concentrations of endrin in blood have only been found in individuals with recent signs of intoxication (Jager, 1970). These results indicate that endrin is rapidly biotransformated ¡n the human body.

Biotransformation

Biotransformation of endrin into polar metabolites occurs in the liver. The metabolism of endrin between animal species is very similar. In general, there are no major qualitative or quantitative differences in the excreted metabolites. The routes of excretion show varia• tions between the species, metabolites being excreted mainly via bile and faeces in the rat, while in the rabbit and the cow the urinary excretion is more Important. The major metabolite was anti-12-hydroxyendrin (Fig. 1 ; III) which is excreted partly as its sulphate and glucuronide conjugate. Four other metabolites have been reported, but the concentrations of these metabolites are generally small relative to that of anti-12-hydroxyendrin and its conjugates. These four metabolites are: syn-12-hydroxyendrin (Fig. 1; II), 3-hydroxyendrin (Fig. 1; IV), 12-ketoendrin (Fig. 1; V) and trans-4.5-dihydroisodrin-4.5-diol (Fig. 1 ; VI). Each of the hydroxy componds ¡s also excreted partly as its sulphate or glucuronide in the urine of animals (Bedford et al., 1975; Hutson et al., 1975). In occupationally exposed manufacturing and formulation workers, the glucuronide of anti-12-hydroxyendrin has been detected up to a concentration of 0.14 mg/L in the urine of workmen with no detectable concentration of endrin or its major metabolite in blood (detection limit 2 ng/mL). Both endrin and anti-12-hydroxyendrln were found in the faeces (Baldwin and Hutson, 1980).

Elimination

When the intake of endrin ceases, endrin is eliminated in an exponential way from the adipose tissue and the blood of rats (Klein et al., 1968; Korte et al., 1970). The half-life of endrin in blood of humans has been estimated to be 24 hours (Jager, 1970). Assuming first-order kinetics of elimination, the half-life of urinary anti-12-hydroxyendrin ex• cretion has been calculated to be 55-75 hours (Van Sittert, 1985). 70

Figure 1. Mechanisms of Biotransformation of Endrin (I) (Bedford and Hutson, 1976).

Biological Indicators

Indicators of exposure

As endrin is rapidly metabolised in humans, measurement of the concentration of endrin (or its metabolites) in the blood is only useful ¡n cases of short-term overexposures as this enables assessment of whether or not there is any risk of intoxication. In other cases the determination of the concentration of the urinary metabolite anti-12-hydroxyendrin is more relevant for measuring absorption of endrin in the body. (Ot- tevanger and Van Sittert, 1979; Baldwin and Hutson, 1980). As the analytical method for the determination of the urinary anti-12-hydroxyendrin con• centration became only recently available, most of the biological monitoring data on en• drin in occupationally exposed workers, however, have been obtained by the determination of endrin in blood. In members of the general population endrin concentrations have also been determined in the adipose tissue.

Concentration of urinary anti-12-hydroxyendrin Occupationally exposed workers

Table 1 shows concentrations of anti-12-hydroxyendrin in the urine of 'healthy workmen' occupationally exposed to endrin during manufacturing, formulation or spraying activities. These 'healthy workmen' were men between 18 and 60 years of age who had no com• plaints or clinical signs attributable to occupational exposure. The highest urinary concentrations were found in the urine samples of spraymen after repeated applications of a formulation containing methyl parathion, endrin and DDT by hand• held ultra low volume (ULV) spraying to cotton. Samples were collected 20 hours after the end of spraying. The formulation had to be applied three times per season with fortnightly intervals. The time needed for one application was approximately 5 hours. Antl-12-hydroxyendrin concentrations were highest In 10 spraymen after their third applica• tions, suggesting that accumulation of endrin was taken place (Kummer and Van Sittert, 1986). Workmen in an endrin manufacturing plant showed lower urinary concentrations. In a study carried out in 1973, a maximum concentration of 0.14 mg/L was reported (Baldwin and Hutsen, 1980). In another study carried out in 1976, concentrations of up to 0.325 mg/g creatinine were reported In urine samples collected at the end of seven days production 71 and formulation. After a three days without exposure, anti-12-hydroxyendrin was eliminated from the body with a half-life between 55 and 75 hours (Ottevanger and Van Sittert, 1979; Van Sittert, 1985).

Concentration of endrin in blood or serum Human intoxications

Table 2 shows endrin concentrations in blood or serum of 18 persons exposed to endrin by accidental contamination of food and of 1 person exposed to endrin in an endrin for• mulation plant. All subjects showed clinical symptoms, but recovered within 24-48 hours after exposure. The only study in which endrin in blood was determined in an occupationally exposed for• mulation worker, dealt with a formulator handling technical endrin powder without wearing a dustmask. He developed convulsions 4 hours after starting work and recovered fully with medical treatment the next day (Jager, 1970). The endrin concentrations in serum as reported in Table 2 ¡n subjects after having eaten contaminated food are generally lower than the concentration in blood of the formulation worker. This may be partly due to the fact that blood samples were taken at longer periods after ingestion. In addition, endrin concentrations in blood and serum cannot be compared as their relationship is unknown.

Table 2. Concentration of endrin in blood or serum in subjects with clinical symptoms

Time of Number of Endrin in blood or serum sampling Reference persons (ng/mL) and clinical symptoms

1* 30 min after 53 (S) Coble et al., 1967 convulsion 20 hours after convulsion 38 (S) 30 hours after convulsion 21 (S) 1* 8 1/2 hours after Coble et al., 1967 convulsion 4(S) 1* 19 hours after Coble et al., 1967 convulsion 3(B) 3* On the day of onset Curley et al., 1970 of symptoms*** 8,11 and 27 (S) 12* time of sampling mean 17 (S) Anon., 1985 after occurrence median 1 of convulsions was minimum 1 not specified maximum 254 1** immediately Jager, 1970 after convulsion 80 (B) 24 hours later 20 5 days later <5

S = samples of serum; B = samples of whole blood Accidental contamination of food Endrin formulation worker Symptoms were not specified

Occupationally exposed 'healthy workmen'

No endrin could be detected in the plasma of 49 men who were employed for a period of 1-19 years (average 12 years) in a plant manufacturing amongst other pesticides aldrin, dieldrin and endrin. The limit of detection for endrin in plasma was 3 ng/mL (Hayes and Curley, 1968). Workers in a Dutch plant, manufacturing aldrin, dieldrin and endrin always showed endrin concentrations In blood below the limit of detection of 10 ng/mL (before 1969) or 5 ng/mL (Jager, 1970; Ottevanger and Van Sittert, 1979), except in two cases of short-term overex• posures. 72

A formulator (case 1) and an operator (case 2) were accidently splashed with a 20% en• drin emulsifiable concentrate which was washed off within 10 minutes. Neither developed signs or symptoms of Intoxication (Jager, 1970). The endrin concentrations in blood of these two persons are shown in Table 3.

Table 3 Concentration of endrin in blood in 'healthy workmen' after a short-term overexposure to endrin (Jager, 1970)

Endrin in blood (ng/mL) Time of first Case sample 1st 12h 24h 36h 5 days sample later later later later

1 one hour after 90 <5 exposure 2 40 min. after 27 25 <5 exposure

Members of the general population

No endrin could be detected in the blood of 4000 individuals, examined between 1976 and 1980 (EPA, 1983). The detection limit was not specified but was probably below 1 ng/mL.

Indicators of effect Electroencephalography

As the central nervous system is the critical organ for endrin, changes in the elec• troencephalogram (EEG) may occur after absorption of high doses of these compounds. The EEG changes were first developed to be a practical tool for monitoring effects ¡n the central nervous system of workers exposed to endrin and to determine when they should discontinue exposure and when they could be allowed to resume werk with endrin. Characteristic changes - which, however are not specific for endrin overexposure - include bilateral synchronous spikes, spike and wave complexes and slow theta waves (Hoogen- dam et al., 1965). In all seven cases of convulsive intoxication in an endrin manufacturing and formulation plant, the return of the abnormal electroencephalographic pattern to the pattern recorded before exposure occurred within a few months (Jager, 1970). No EEG measurements have been reported in subjects whose endrin concentration in blood or serum had been measured following endrin intoxication. Therefore the relationship bet• ween the rate of dlminuation of EEG changes and the rate of decrease of endrin in blood ¡s unknown.

Enzyme induction

Several studies have been reported which showed increased liver microsomal enzyme ac• tivity in subjects occupationally exposed to endrin (Jager, 1970; Hunter et al., 1972; Ot• tevanger and Van Sittert, 1979; Kummer and Van Sittert, 1986). The studies have been summarised in Table 4. Enzyme induction was demonstrated by significantly lower p, p'- DDE concentrations in the blood and a significant increase of the ratio of the concentra• tions of urinary 6-beta hydroxycortisol and 17-hydroxycorticosteroids compared with con• trol groups (Jager, 1970). Enzyme induction in endrin manufacturing and formulation workers was further demonstrated by examining the urinary excretion of D-glucaric acid (Hunter et al., 1972; Ottevanger and Van Sittert, 1979). In the latter study it was shown that the effect of enzyme induction is reversible, as D- glucarlc acid concentrations were within normal limits after a six weeks period of shut• down and maintenance of the plant (Table 4; I). In addition urinary D-glucaric acid concen• trations were diminished after a three-days leave (Table 4; III), following 7 days production 73

and formulation of endrin (Table 4; II). Urinary D-glucaric acid concentrations appeared to decrease at a slower rate after a period of no exposure than urinary anti-12-hydroxyendrln concentrations. There was a statistically significant correlation between the urinary concentrations of anti-12-hydroxyendrin and of D-glucaric acid in samples collected after 7 days production and formulation (Table 4; II). From these data a concentration of anti-12-hydroxyendrin in urine of 0.130 mg/g creatinine was suggested as a threshold value below which enzyme induction (i.e. D-glucaric acid concentrations above 20 mg/g creatinine) does not occur (Ottevanger and Van Sittert, 1979). In spraymen who applied a formulation containing methyl parathion, endrin and DDT by hand-held ULV equipment to cotton, the urinary excretion of the methyl parathion metabolite p-nitrophenol (PNP) was measured ¡n samples collected 20 hours after the end of spraying. Urinary PNP concentrations were significantly lower in spraymen after their third applica• tion than in spraymen after their first and second application of the formulation. Assuming that exposures were similar, this difference in urinary PNP concentrations was ¡nterpretated as being caused by enzyme induction by endrin. This resulted in an increas• ed rate of metabolism of methyl parathion and subsequently, as a result of the more rapid excretion of p-nitrophenol, in a decrease in p-nitrophenol concentrations in samples col• lected 20 hours after application (Kummer and Van Sittert, 1986).

Table 4 Biological monitoring parameters of endrin exposure and parameters of hepatic enzyme induction

Urinary Parameters of hepatic enzyme induction anti-12-hydroxyendrin (mean values and range) (mg/g creatinine) Endrin in No of subjects Reference blood 6B-hydroxycortisol (ng/mL) pp'DDE D-glucaric acid Urinary mean min max. in blood 17-hydroxycorticos- in urine PNP (ng/mL) teroids in urine (mg/g creat.) (mg/g creat.) (^g/mg)

130 (Controls) <5 — — — 13.2 (5-58) — — — 29D <5 — — — 4.5 (<5-18) — — — Jager, 1970 20 (Controls) <5 — — — — 26.6 (6.7-82) — — 8D <5 — — — — 87.3 (59-126) — — 21 (Controls) <5 — — — — — 9 (3-14)3) — Hunter et al., 9D <5 — — — — 27 (17-42)3) — 1972 121> (I) <5 0.064 0.016 0.118 — — 12 (1.3-49) — Ottevanger 121> (II) <5 0.164 0.025 0.325 — — 32 (9-68) — and Van 121) (III) <5 0.063 0.018 0.181 — — 24 (5-58) — Sittert, 1979 62> (A) — 0.06 0.001 0.12 — — — 0.08 (0.05-0.20) Kummer and 102> (B) — 0.55 0.21 1.4 — — — 0.13(0.06-0.44) Van Sittert, 222> (C) — 0.25 0.03 0.59 — — — 0.38 (0.04-1.38) 1986

1) Manufacturing and formulation workers I Urine samples collected after 6 weeks shut-down and maintenance of plant II Urine samples collected after 7 days production/formulation III Urine samples collected 3 days without exposure to endrin 2) Spraymen who sprayed a methylparathion/endrin/DDT formulation

A Supervisors of spraymen ur¡ne sampies were collected 20 hours after the application B Spraymen, after their 3rd application C Spraymen, after their 1st and 2nd application 3) Expressed as glucarolactone (uM/g creatinine)

Conclusions

For the measurement of short-term overexposure to endrin, the concentration of endrin In blood should be measured, while for the measurement of normal low occupational ex• posures, the determination of the urinary antl-12-hydroxyendrin concentration is the most useful test. Because of the absence of a controlled human volunteer study in which a known dose Is related to the urinary anti-12-hydroxyendrin concentration, the urinary concentrations in 74 exposed workers cannot be interpreted in terms of ingested dose. Neither has a relation• ship been established between urinary anti-12-hydroxyendrin concentrations and acute ef• fects. The highest urinary anti-12-hydroxyendrin concentration ever measured (1.4 mg/g creatinine) was In a sample collected 20 hours after the end a hand-held ULV application. At this concentration no clinical signs or symptoms were observed. Hepatic enzyme induction has been demonstrated in workers exposed to endrin, using several parameters. No enzyme induction (i.e. D-glucaric acid concentrations in urine above 20 mg/g creatinine) occurred at a urinary anti-12-hydroxyendrin concentration below 0.130 mg/g creatinine in samples collected at the end of seven days production and formulation of endrin. With regard to the determination of endrin in blood, this test is only useful after short-term overexposures to endrin. Because of the short half-life of the compound ¡n blood, samples have to be collected Immediately after the end of exposure to ascertain whether or not there is any risk for imminent intoxication. No cases of intoxication due to short-term en• drin overexposure have been reported at endrin concentrations in blood below 50-100 ng/mL, provided that samples were collected at the end of exposure. Electroencephalography ¡s an obsolete method for monitoring endrin exposure. The test still can be used for confirming a diagnosis of endrin intoxication and to determine the suspen• sion period. The latter cannot be done by the determination of endrin in blood or serum because of the rapid rate of decrease of endrin in blood.

Research Needs There is no data available on the relationship between the urinary concentration of anti-12-hydroxyendrin and a defined oral or dermal dose. This relationship can only be determined in a study under closely controlled conditions in human volunteers, relating defined oral and dermal doses to the parameters used in biological monitoring of endrin, i.e. the urinary concentration of anti-12-hydroxyendrln, the urinary D- glucaric acid concentration or other tests for hepatic enzyme induction. Such a study should be carried out in accordance with legal and ethical requirements and should observe rele• vant national and International codes of conduct. However, at present there ¡s little need for such a study as the use of endrin is largely being replaced by other products. 75

References

ANON. Acute convulsions associated with endrin poisoning - Pakistan. (1985) J. Am. Med. Assoc. 253(3): 334-335. BALDWIN, M.K.; HUTSON, D.H. (1980) Analysis of human urine for a metabolite of endrin by chemical oxidation and gas-liquid chromatography as an indicator of exposure to endrin. Analyst 105:60-65. BALDWIN, M.K.; CRAYFORD, J.V.; THORBURN BURNS, D. (1976) The metabolism and residues of l4C-endrin in lactating cows and laying hens. Pestic. Sci. 7(6):575-594. BEDFORD, CT.; HARROD, R.K.; HOADLEY E.C.; HUDSON, D.H.; (1975) The metabolic fate of endrin in the rabbit. Xenobiotica 5(8):485-500. BEDFORD, CT.; HUTSON, D.H. (1976) The comparative metabolism in rodents of the isomeric insecticides dieldrin and endrin. Chem. Ind. pp. 440-447. COBLE, Y.; HILDEBRANDT, P.; DAVIS, J. RAASCH, F.; CURLEY, A. (1967) Acute endrin poisoning. J. Am. Med. Assoc. 202:489-493. CURLEY, Α.; JENNINGS, R.W.; MANN, H.T.; SEDLAK, V. (1970) Measurement of endrin following epidemics of poisoning. Bull. Environ. Contam. Toxicol. 5(1):24-29. DAVIES, G.M.; LEWIS, I. (1956) Outbreak of food poisoning from bread made of chemically contaminated flour. Br. Med. J. 2:393-398. DITRAGLIA, D.; BROWN, D.P.; NAMEKATA, T.; IVERSON, Ν. (1981) Mortality study of workers employed at organochlorine pesticide manufacturing plants. Scand. J. Work. Environ. Health 7(suppl. 4):140-146. DONOSO, J.; DORIGAN, J.; FULLER, B. (1979) Reviews of the environmental effects of pollutants. XIII. Endrin. Oakridge, Nat. Lab. Oakridge, Tennessee, U.S. Dept. of Commerce, Nat. Techn. Inform. Service N.T.I.S., Aug. 1979 HAYES Jr., W.J.; CURLEY, A. (1968) Storage and excretion of dieldrin and related compounds. Effect of occupational exposure. Arch. Environ. Health 16(12):155-162. HINE, C.H.; ANDERSON, H.H.; KODOMA, J.F. AND GUTENBERG, E.F. (1968) Class Β evaluation of endrin compositions. San Francisco, The Hine laboratories, Report no. 7. HOOGENDAM, I.; VERSTEEG, J.P.J.; VLIEGER, M. de (1962) Electroencephalograms in insecticide toxicity. Arch, environ. Health 4:86-94. HOOGENDAM, I.; VERSTEEG, J.P.J.; VLIEGER, M. de (1965) Nine years toxicity control in insecticide plants. Arch, environ. Health 10:441-448.

HUNTER, J.; MAXWELL, J.D.; STEWART, D.A.; WILLIAMS, R.; ROBINSON, J.; RICHARD­ SON, A. (1972) Increased hepatic microsomal enzyme activity from occupational exposure to certain organochlorine pesticides. Nature 237:399-401. HUTSON, D.H.; BALDWIN, M.K.; HOADLEY, EC (1975) Detoxication and bioactivation of endrin in the rat. Xenobiotica 5:697-714. JACOBZINER, H.; RAYBIN, H.W. (1959) Brielfs on accidental chemical poisonings in New York City. Poisoning by insecticide (Endrin). N.Y. State J. Med. 59:2017-2022. 76

JAGER, K.W. (1970) Aldrin, dieldrin, endrin and telodrin. An epidemiological study of long-term occupational ex• posure. Amsterdam, Elsevier Pubi. Comp.. KARPLUS, M. (1971) (Endrin polsoning in children.) Harefuah 81(3):113-116, 1971 (in Hebrew) Abstracts on Hygiene 46(12): 1142. KLEIN, W.; MUELLER, W.; KORTE, F, (1968) Insektizide im Stoffwechsel, XVI. Ausscheidung, Verteilung und Stoffwechsel von Endrin-14C in Ratten. Liebigs Ann. Chem. 713:180-185. KORTE, F.; KLEIN, W.; WEISGERBER, I. (1970) Recent results in studies on the fate of chlorinated insecticides. In: Deichman, W.B.; Radomski, J.L.; Penalver, R.A. (eds) Proceedings Sixth Conf. Toxicol. and Occup. Med. 1968, Pesticide Symposia pp. 51-56. Miami, Halos and Assoc. Inc. KUMMER, R.; SITTERT, N.J. VAN (1986) Field studies on health effects from the application of two organophosphorus insecticide formulations by hand-held ULV to cotton. Toxicology Letters 33: 7-24. OTTEVANGER, CF.; SITTERT, N.J. VAN (1979) Relation between anti-12-hydroxy endrin excretion and enzyme induction in workers involved in the manufacture of endrin. In: Strik, J.J.T.W.A. and Koeman, J.H. (eds), Chemical Porphyria in Man, pp. 123-129. Elsevier/North Holland Biomedical Press. REDDY, D.B.; EDWARD, V.D.; ABRAHAM, G.J.S.; RAO, K.K. (1966) Fatal endrin poisoning. A detail autopsy, histopathological and experimental study. J. Indian M.A. 46(3):121-124. RIBBENS, P.H. (1985) Morality study of industrial workers exposed to aldrin, dieldrin and endrin. Int. Arch. Occup. Environ. Health 56:75-79. RICHARDSON, LA; LANE, J.R.; GARDNER, W.S.; PEELER, J.T.; CAMPBELL, J.E. (1967) Relationship of dietary intake to concentration of dieldrin and endrin in dogs. Bull, environ. Contam. Toxicol. 2(4):207-219.

SITTERT, N.J. VAN (1985) Biologische monitoring in de praktijk van de bedrijfsgezondheidszorg. De mens als meetinstrument op het werk? Coronel-PAOG naschollngssymposium. Amsterdam, Vrije Universiteit, February 1985 TREON, J.F.; CLEVELAND, F.P.; CAPPEL, J. (1955) Toxicity of endrin for laboratory animals. J. agrie. Food Chem. 3:842-848.

WANG. H.H.; MacMAHON, B. (1979) Mortality of workers employed in the manufacture of chlordane and heptachlor. J. Occ. Med. 21(11):745-748. WEEKS, D.E. (1967) Endrin food-poisoning. A report on four outbreaks caused by two separate shipments of endrln-contaminated flour. Bull. Wld. Hlth. Org. 37:499-512. Biological indicators for the assessment of human exposure to industrial chemicals

Vanadium K.H. Schaller and G. Triebig

79

Summary

Vanadium (V) is discussed to be an essential trace element for man. Adverse health ef• fects of vanadium are mainly observed in connection with occupational exposure to vanadium pentoxide ¡n the form of dusts and fumes. Predominately the exposure causes toxic-irritative effects to skin, mucous membranes of the eyes and of upper and lower respiratory tract. A green discoloured tongue may be an indicator of acute or chronic exposure to vanadium. At present there are no findings indicating carcinogenic effects of vanadium in humans. For biological monitoring, V in blood and urine can be used. Vanadium levels in blood and urine samples of occupationally nonexposed persons can be regarded below 1 ,ug/l. The most suitable indicator of exposure is urinary vanadium. Vanadium in urine has - in com• parison to vanadium in blood - enough biological sensitivity to allow biological monitoring also at low levels of exposure. Currently, for the analysis of V in biological materials spectrophotometric methods, plasma atomic emission spectrometry, neutron activation analysis and electrothermal atomic ab• sorption spectrometry are relevant. Neutron activation determination seems to be the most reliable method for the analysis of biological specimens taken from unexposed subjects. To avoid contamination in the preanalytical phase (sampling, transport, storage) it is necessary to follow recommended procedures. At present, no dose-effect/response-relationshlps between an occupational exposure to vanadium dust and fumes on the one hand and early adverse effects of the respiratory system on the other could be established. Furthermore there is no specific biological test indicating early effects of an increased vanadium exposure.

81

Vanadium

Introduction Chemical and physical properties

Vanadium (V) is a steel-gray, ductile and corrosion-resistent metal with atomic number of 23, atomic weight of 50.9, density 6.11. Vanadium is an element of the "Va group" of the periodic system and belongs to the first transition series. It forms compounds mainly in oxidation states +3, +4, +5. The most stable oxidation state is + 4, in which V forms oxovanadium (IV) ions. Vanadium pentoxlde dissolves in acids forming dioxovanadium (V) ion, and with bases it may form vanadates (e.g. Na3V04). The so-called metavanadates are compounds of the type NaV03. The physical and chemical properties of V and some of Its compounds which are important from the point of occupational exposure, are shown in table I.

Table I: Properties of vanadium and its compounds (WEAST 1972) insol = insoluble, al sol = alcohol soluble, aq. reg. = aqua regia, dec = decomposition

Sok bility g/100_ ;m3 Melting Boiling Molecular Compound point point Other solvents formula Cold •C •C water Soluble Insoluble

Vanadium V 1890 3380 insol aq. reg, HCl HNO3, alkali H2S04 HF Vanadium (IV) VC 2810 3900 insol HN03 HCl carbide KN03 H2S04 Vanadium (II) VO or ignites — insol acid — oxide V202 Vanadium (III) V203 1970 — al sol HN03, HF — oxide alkali Vanadium (IV) V02 or 1967 — insol acid — oxide V204 alkali Vanadium (V) V205 690 1750 0.8 acid abs. oxide dec alkali alcohol Ammonium NH4VO3 200 — 0.52 — — metavanadate (V) Sodium meta• NaV03 630 — 21.1 — — vanadate (V) 82

Principal occupational exposures

Vanadium is found in the earth's crust of about in a proportion 0.2% (CLARK 1975). It is quite evenly distributed in minerals. A few commercial deposits contain more than 3% vanadium pentoxides but normal concentrations are 0.1-1 % (NAS 1974). The main sources of vanadium are vanadium sulphide (patronite), carnotite and titanomagnetite ores. Many crude oils contain considerable amounts (even about 0.1 %) of vanadium, notably those from Venezuela. The ash obtained from burning vanadium containing oils may have a high percentage of vanadium. Vanadium can be extracted from fuel ashes.

A considerable amount of vanadium production is based on the extraction of converter slag in the steel Industry, which can contain 2-10% (MICHELS 1973) or even 25% (NAS 1974) vanadium pentoxide. Vanadium is usually manufactured by converting the vanadium minerals to a water-soluble form (LEVANTO 1969). The world production of vanadium was about 53000 tons in 1978, the largest producers being South Africa, the Soviet Union, the USA and Finland (KIVILUOTO 1981). The main use of vanadium Is in ferrous metallurgy (STOCKINGER 1981), where vanadium is added to steel in the form of either ferrovanadium or vanadium carbide. Vanadium is also used in non-ferrous alloys in nuclear technology (CLARK 1975), aircraft construction and space technology. In addition to the metallurgical use, vanadium compounds are employed as catalysts in the chemical Industry, I.e., in the production of acetic and sulfuric acids. Small amounts of vanadium are used e.g. in ceramics, and the electronics industry (NAS 1974). V generally occurs in air at workplaces as vanadium pentoxide dust or fume.

Effects on humans

The question of whether vanadium is an essential trace element for man Is still unanswered. The humans dietary requirement for vanadium could not be estimated (WHO 1973). Vanadium can cause acute as well as chronic effects in humans. It is noted that toxicity depends on the oxidation state and augments with increasing states, the most toxic being vanadium (V) such as V205 or V02CI. Acute exposure to vanadium compounds causes irritation of the eyes and the upper and lower respiratory tract. Mild effects are characterized by sneezing rhinitis, cough, chest pain and conjunctivitis (SYMANSKI 1939, KOELSCH 1966, BAADER 1961, SJÖBERG 1950, ROSKIN 1968). These effects seem to be concentration-dependent and usually reversible. Exposure of healthy volunteers to vanadium pentoxide dust (1 mg/m3) produced respiratory irritation and a persistent cough, lasting 8 days; this occured after about 12 hours. Exposure to a lower concentration (0.2 mg/m3) had the same effect, but the induction time for the persistent cough was longer (about 20 hours). Concentrations of 0.1 mg/m3 caused no ir• ritation. On re-exposure the respiratory irritation occured at this concentration (ZENZ and BERG, 1967). In severe cases, which are mainly also reversible, bronchitis, obstructive bronchitis and bronchopneumonia may resuit (ZENZ et al., 1962, ROSKIN 1968, MUSK and TEES 1982). Reversible reduction of forced vital capacity and forced exspiratory volume was found in boiler cleaners exposed to V containing respirable dust (LEES 1980). A green-black discoloured tongue can be regarded as an Indicator of vanadium exposure, but may be absent even with long-term exposure. Vanadium has allergenic properties - probably as a hapten - and can therefore cause eczema as well as asthma in hypersensitized persons (SJÖBERG 1950, MUSK and TEES 1982). Chronic exposure to vanadium may also cause rhinitis, pharyngitis and bronchitis, but the amount of exposure ¡s not conclusive. Evidence of systematic effects in workers exposed to vanadium compounds is sketchy. In most cases, no systematic effects have been reported (SYMANSKY 1939, WILLIAMS 1952, LEWIS 1959, ZENZ et al., 1962, EISLER et al., 1968). Other reports describe only vague or unspecific signs and symtoms such as weakness, ringing in the ears, nausea, vomiting, and headache (ROSHCHIN 1962). KIVILUOTO et al. (1981) found no abnormalities in hematologic tests in workers exposed 3 to 0.01 - 0.04 mg V205/m . This study did not reveal any decrease in serum cholesterol or increase in serum triglycerides, in contrast to earlier studies (CURRAN 1959). Table II summarizes some relevant studies on effects following ocupational exposure. At present, there are no consistent data about upper V205-concentrations which produce 83

no adverse effects on the respiratory system. Furthermore it is difficult to differentiate bet• ween dust and fumes with respect to aerodynamic diameter and distribution of the size of particles. For this reason, the German Commission for MAK-values recommends a MAK- 3 value for V205 dust of 0.05 mg/m for particles with a diameter smaller than 5/¿m (DFG 1985).

Table Findings of adverse health effects in occupationally exposed subjects.

Duration and amount Number of Main symptoms and Type of exposure References of exposure persons findings

Slag dust containing 2 to 3 years 36 Rhinitis, cough, asthma, Sjöberg 1950 3 17% V203 max. 6.5 mgV/m allergic eczema, broncho• pneumonia.

Slag dust, V205-dust Mostly 4 to 7 years 62 Conjunctivitis, bronchitis, Massmann and Opitz 1954

NaV03-dust headaches, tiredness. No lung fibrosis or chronic bronchitis.

V205-dust More than 10 months 39 Irritations of eyes, cough, Vintinner et al. 1955 0.02-58.8 mgV/m3 expectoration, pneumonia.

V205-dust More than 6 months 24 Irritations of eyes and nose, Lewis 1959 5-10% of the particle (average: 2-5 years) asthma, green discoloured 3 5/jm V205 below 0.5 mg/m tongue.

Pure V205 3 weeks 18 Conjunctivitis, pharyngitis, Zenz et al. 1962 3 0.1-10/im V205 above 0.5 mg/m cough, asthma, increasing complaints with duration of exposure.

V205-dust Average 11 years 63 Wheezing. No evidence of Kilivuoto 1981 20% of particle 5//m 0.2-0.5 mg/m3 chronic pulmonary disease. Ammonium vanadate Hours and days, resp. 4 Upper respiratory symptoms, Musk and Tees 1982 asthma, green discoloured tongue.

Metabolism Absorption

In occupational exposure the main route of absorption is by inhalation of dust and fumes containing vanadium or vanadium compounds. The extent of absorption of different vanadium compounds in the lungs has not been ade• quately, determined although It is estimated that about 25% of soluble vanadium compounds are absorbed (ICRP, 1975). Vanadium pentoxide (V205) as well as vanadiumoxychlorlde (V02CI) are absorbed nearly totally via the lungs (CONKLIN et al. 1982).

In acute exposure of animals there is a rapid complete clearance from the lungs of V205 within 1-3 days, but this is probabely not relevant in chronic exposure (STOKINGER 1967). Ingested vanadium compounds are, In general, poorly absorbed. After intragastral administra• tion only a low percentage of the dose could found in urine (CURRAN et al. 1959, CON• KLIN et al. 1982). The daily V-intake by food was calculated to be in the range of 8.2-13.2/¿g, with a mean of 10.7/ig (SCHALLER et al. 1980). In this study a uniform diet was administered for two weeks to 10 adults who had no occupational exposure to vanadium. The concentrations of V in the food and the feces were analyzed by atomic absorption spectrometry after wet oxidative digestion of the samples. SCHROEDER et al. 1963 reported an average daily V- ¡ntake of 2 mg for the general population which is certainly too high and reflects an analytical• ly unreliable method. BYRNE and COSTA (1978) studied the V content of various foodstuffs and concluded that the daily V Intake by food is several tens of micrograms. In animal experiments, soluble vanadium compounds may penetrate the skin (STOKINGER 1967). Cutaneous absorption seems to be of no significance in man during exposure at the workplace. 84

Distribution

Absorbed vanadium is transported mainly in the plasma and is widely distributed in body tissues. Short-term localization in liver, spleen, kidney and lung was reported by ROSKIN (1968) In animal feeding experiments. In animal experiments vanadium was accumulated especially in bone, but also in liver, lung, kindney and placenta (CONKLIN et al. 1982, TALVITIE et al. 1954). No data are yet available concerning the distribution of vanadium in man after chronic occupational exposure. V easily accumulates in the tissues of exposed subjects and is slowly released by the organism, as previous animal and human Investigations showed (MARONI et al. 1981, KIVILUOTO et al. 1981).

Excretion

Elimination of vanadium from the body occurs mainly via the urine. The amount was estimated to be about 40 to 60% within 1 to 3 days after resorption (CONKLIN et al. 1982, TALVITIE et al. 1954, ROSKIN 1968). A minor portion of about 10% is exceted via the feces.

Biological half-time

Biological half-time in man have not been adequately determined. The ICPP (1960) estimates the whole body retention of vanadium in man as 42 aays (LAGERKVIST et al. 1986). Studies on workers exposed to V in workroom air report that blood and urinary values of V drop to half of the initial value within a few days after cessation of exposure (THÜRAUF et al. 1979, KIVILUOTO et al. 1981). In a recent investigation (SABBIONI and MARONI 1983, MARONI et al. 1983), the kinetics of urinary elimination of vanadium was studied by evaluating the behaviour of the mean levels of excretion after cessation of exposure in two groups of workers with different degrees of vanadium exposure. In the subjects with "high" exposure (who showed mean urinary vanadium values between 60 and 70/tg/g creatinine at the end of exposure), the half-life was about 20 hours. A period of about 600 hours after cessation of exposure was required to reach urinary vanadium levels of 2/¿g/g creatinine (upper value in the control group). In the subjects with "low" exposure (who showed mean urinary vanadium levels of 4.3 /ig/g creatinine) the half-life was about 15 hours and about 18 hours after cessation of ex• posure were required to reach excretion levels of 2/ig/g creatinine. There are only limited data indicating that excretion may occur ¡n at least two phases. In a study on rats by HOPKINS and TILTON (1966), 55% of intravenously injected 48V was excreted within 97 hours. For further information see the review by LAGERKVIST et al. 1986.

Biological indicators

Indicators of internal dose

The biological indicators for evaluation of the internal dose in occupational exposure are vanadium in whole blood/plasma/serum and vanadium in urine.

Vanadium concentrations in whole blood, serum, plasma and urine of non-occupationally exposed subjects

Table III shows the vanadium levels in blood found by various investigators in non-exposed subjects. The values decline with time. This is commonly observed, as analytical methodologies have become more sensitive and specific. It may be that the wide range of values reflects analytical error rather than real variation. The data, obtained by neutron activation analysis are especially very low. In this case, it must be considered, that the sampling was performed in a manner hardly feasible in routine (plastic-over-the-needle catheter, especially cleaned containers, clean benches, etc.). The available data on urinary vanadium values in non-exposed persons show a wide varia• tion, as can be seen In table IV. 85

Table III: Vanadium concentrations in blood/serum samples of non­occupationally exposed subjects 1 = Milan residents, 2 = Unexposed workers, 3 = /ig/kg

References Mean/ig/l Range/ig/l Material Method

Heydorn and Lukens 1966 4.6 serum NAA Sabbioni et al. 1978 6.6 3.8­11.3 serum NAA Byrne and Kosta 1978 0.5 blood NAA Cornells et al. 1978 0.024 0.017­0.031 blood NAA Keim and Schaller 1978 2.25 1­11.4 blood Photometry Thürauf et al. 1978 2.5 blood Photometry Gylseth et al. 1979 0.78 blood NAA Cornells et al. 1980 0.033 0.17­0.939 blood NAA Maroni et al. 1983 0.94 I 0.651 blood NAA 1.86Ì2.392 Naylor et al. 1984 7.14 ï 1.083 blood NAA 4.13Í 1.133 serum NAA Simonoff et al. 1984 0.670 0.26­1.3 blood NAA

Table IV: Vanadium concentrations in urine samples of non­occupationally exposed subjects 1 = Milan residents, 2 = Unexposed workers, 3 = //g/kg

Reference Mean/ig/l Range/ig/l Method

Watanabe et al. 1966 6.9 3.5­11.0 UV­V1SSP Holzhauser et al. 1977 1.4 0.3­2.45 Photometry Thürauf et al. 1978 0.8 0.5­1.9 Photometry Byrne and Kosta 1978 0.3 NAA Kiviluoto et al. 1979 2.0 AAS Ishizaki and Ueno 1979 0.26 AAS Buchet et al. 1982 0.44 0.3­0.7 AAS Stroop et al. 1982 7.6 3.8­12 AAS Maroni et al. 1983 0.81 i 0.721 NAA 1.21 ± 0.582 Naylor et al. 1984 6.62 i 3.153 NAA Schaller et al. 1986 0.9i 0.4 0.4­1.9 AAS­Zeeman

Vanadium concentrations in blood and urine of occupationally ex­ posed subjects

In the older literature several authors determined the V concentration in urine and blood of exposed workers (VINTINNER et al. 1955, MOUNTAIN et al. 1955, LEWIS 1959, JARACZEWSKA and JAKULOWSKI 1963, WATANABE et al. 1966, KÖHLER 1970). In general, the authors did not find a statistically significant correlation between V exposure and urinary V concentration. Due to the older analytical methods, the results for the V con­ centrations in biological materials appear too high in comparison to more recent data and should not therefore considered in this review. 3 ZENZ and BERG 1967 exposed two healthy volunteers to V205 dust (1 mg/m , 8h) and 5 persons to a lower concentration (0.2 mg/m3, 98% particles 5/im, 8h). The greatest amount of V was found in the urine three days after exposure, i.e., 130/ig/l, and none was detectable one week following exposure. Unfortunately the authors do not quote the values found before the exposure. Due to detection limit of the analytical method used no vanadium in blood was detectable. KIVILUOTO et al. 1979 measured serum and urinary vanadium of 60 V­exposed workers (process workers, dayworkers, repairmen). Measurements of the vanadium concentrations ¡n the workroom air and the respiratory zones in the vanadium factory gave values from 0.01 ­ 0.04 mg/m3 The mean urinary vanadium excretion of the exposed workers was 13.3 ­ 9/ig V per 18 hours, and the mean vanadium concentration in the serum was 11.2 ± 7/xg/l (18 hours after exposure). The concentration of vanadium ¡n the urine of the con­ trols was below 2μ$\, which was the detection limit of the analytical method. No significant correlation could be established here between the vanadium concentration in the factory air and the serum level and urinary excretion of vanadium. The lack of cor­ relation between the low vanadium concentration in the air (0.01 ­ 0.04 mg/m3) and serum 86

and urinary excretion of vanadium could have bean due to the non­uniform distribution of vanadium dust particle sizes and inter­individual differences in inhalation and absorption, even though personal dust sample collection was used, or even to analytical problems. The detection limit of the method used was just below the concentration measured, and this may explain the high coefficient of variation, which would obscure any possible cor­ relation. Under conditions of higher vanadium exposure (0.02 ­ 0.05 mg/m3) ­ the vanadium concentration in the air inhaled remaining unknown due to the use of dust masks ­ urinary vanadium excretion and serum vanadium levels decreased significantly with exposure­free time.

As vanadium concentrations in the serum and urine only decrease significantly 42 ­ 66 hours after the termination of exposure, biological monitoring ought no: to be carried out immediately after the cessation of exposure (KIVILUOTO 1981). In a more recent study KIVILUOTO and co­workers (1981) investigated serum and urinary

V concentrations of 8 men exposed to V205 dust (smelting, packing and filtering workers). The time­weighted levels of exposure to vanadium of the filtering workers were 0.03 ­ 0.77 (mean 0.28) mg/m3. The levels for the workers who smelted and packed were 0.07 ­ 0.15 (mean 0.10) mg/m3. The mean respirable fraction of the vanadium dust was 58% (range 41 ­ 71 %). The creatinlne­adjusted urinary vanadium concentrations were found to correlate with serum vanadium concentrations (r = 0.81), but not with the atmospheric concentration of vanadium in the factory. The urinary vanadium excretion decreased significantly with the decreasing exposure time. At the beginning of the summer holidays, the serum vanadium concentra­ tion of the workers was 20 ± 11.4/ig/l and the urinary excretion of vanadium 32.7 ± 22.5/ig/g creatinine. Three days after exposure the urinary excretion of vanadium was 20.7 +10.8 /ig/g creatinine. On the 16th day of the holidays vanadium could be detected in the serum 11.5 ± 4.3/ig/l) and in the urine (21.6 ±.11.7/¿g/g creatinine) of the workers (Fig. 1 and 2). These results suggest that most of the absorbed vanadium is excreted in the urine within one day after a long­term moderate exposure to vanadium dust. In this study the normal creatinine adjusted excretion was very high for the non­occupationally exposed persons in relation to the values reported by other groups (see table IV). This difference could be due to dietary habits or to the geographical location.

180 ­I

160

~ 140 fo 120 E _E õ 100 ..JO (56)

80 D < Σ ρ(61) < 'cTfSOL > 60 ­ rr < 40 _ ­O (20) rr "TT­ — 0(28) Ζ) (24)_ ­0(19) 20 _ ""Of 19)

—r-//- 16 24 48 72 10 20 30 TIME (hours) (days)

Figure 1 Excretion pattern of urinary vanadium. Significance of the differences between excretions: 72­8h = Ρ <0.01, 72­16h = Ρ < 0.025, 72­24h = Ρ < 0.05, 72­48h = Ρ < 0.05, and 16d­72h = Ρ <0.1. — — indicates the holiday values, which are significantly (P <0.01 ) higher than the levels of the respective family members, expressed in parentheses (KIVILUOTO et al. 1981) 87

700 _

600 ­

­ 500

I 400 _ Q < < > 300 _

rr 200 ­

100 ­

"Γ 10 15 20 25 30

DAYS AFTER EXPOSURE

Figure 2 Concentration of serum vanadium. At the end of a work shift the values were not signifi­ cantly higher (P>0.1) than those during the holiday (KIVILUOTO et al. 1981)

THÜRAUF et al. (1979) carried out a study in a metallurgie plant and analyzed the V levels in blood and urine samples of 54 workers exposed to vanadium compounds. On the basis of sporadic dust measurements it can be assumed that the results obtained were, as a rule, clearly lower than the MAK value for vanadium pentoxide dust. The median vanadium concentration in whole­blood was 2.9/zg/l. This indicates that the exposed persons differed significantly from the control group ( < 2.5/tg/l, which was the detection limit of the analytical method).

The median vanadium concentration measured in urine was 37.8/¿g/l. This means, that there was a significant difference in comparison with the control group (0.8/¿g/l). Conver­ sion of the results to creatinine yields a median vanadium concentration of 33.9 and 0.6 /¿g/g creatinine for the exposed workers and normal persons respectively.

The median vanadium concentration measured in urine was 37.8/¿g/l. This means that there was a significant difference ¡n comparison with the control group (0.8/¿g/l). Conver­ sion of the results to creatinine yields a median vanadium concentration of 33.9 and 0.6 /¿g/g creatinine for the exposed workers and normal persons, respectively. After a weekend without exposure the vanadium concentrations in blood and urine drop­ ped. In general, this decrease was the more pronounced the higher the initial value (Table V). GLYSETH et al. (1979) performed a study in the ferrovanadium production Industry on workers exposed to siag dust, containing mainly V205; other oxides may, however, have been present (V02, V203). During the reduction process the potmen were exposed to a mixture of metal and oxide fumes of iron and vanadium. The crushing and packing men were exposed to ferrovanadium dust. Seventeen workers, 6 persons with zero to low exposure, and 11 persons with moderate to high exposure, participated in the Investigation. The daily V exposure of the 11 subjects recorded with personal air sampling, varied from 0.02 ­ 1.5 mg/m3. The mean V con­ centration In blood for the low exposure group was 1.0 /¿g/l, for the high exposure group 1.8/ig/l; the V concentration in urine adjusted for creatinine ranged from 1.6 to 6.9/¿g/g creatinine. No statistically significant difference was found between the pre­and postshlft values. It was shown, that for the moderate­to­high exposure group there was a slight in­ 88

Table V: Statement of the Vanadium concentrations measured in the blood and urine of 13 expos• ed subjects before and after a weekend (Friday-Monday) without exposure (THÜRAUF et al. 1979).

V-blood V-urine Exposed Monday persons Friday Monday Friday No. /¿g/i ¿¿g/l /¿g/i //g/g creat. /¿g/i /ig/g creat.

1 4.5 ,5.0 4.4 6.0 2.4 2 6.0 4.5 5.3 '8.7 5.4 7.3 3 6.5 — 11.2 8.0 23.8 5.2 4 4.0 2.5 20.2 — 11.4 5.7 5 25.0 6.0 20.8 13.0 12.6 8.4 6 2.5 5.0 12.0 4.8 10.8 4.2 7 2.5 4.5 5.6 10.2 7.4 2.4 8 9.0 3.0 12.6 25.2 7.0 6.9 9 21.0 10.0 9.6 29.1 11.6 — 10 3.0 3.0 14.0 11.1 6.4 9.6 11 183.0 121.0 43.8 43.0 41.8 65.3 12 36.0 19.0 11.6 64.4 30.6 16.1 13 32.0 26.0 22.0 26.2 17.3 —

crease ¡n V excretion during the periods of investigation, while the blood V values revealed no significant alterations. The V excretion adjusted to the creatinine concentration shows a slightly higher correlation with V exposure than the liter-related value. MARONI et al. (1983) examined workers employed in maintenance operations of a fuel-oil boiler at a power station and 2 groups of non-occupationally exposed persons (N = 5 and 14). The exposed group of 16 workers, who carried out the maintenance operations, i.e. cleaners, scaffolders, welders and burner operators. V-concentratlons in air were estimated, to be between 0.1-5 mg/m3 for total dust and 0.05-1.5 mg/m3 for the respirable fraction. The mean blood V levels of the exposure groups varied in the preshift samples from 1.6 to 2.8/¿g/l, in the postshift samples from 0.8 to 5.5/ig/l. The mean urinary V excretions, adjusted for creatinine, ranged from 0.8-2.7/¿g/g creatinine, (preshift), to 2-53/ig/g creatinine (postshift). Interestingly, while urinary V excretion showed a several-fold increase when compared to the baseline values, only a moderate Increase In blood V concentration was observed in all cases. This indicated that the blood V concentration is not an indicator of V exposure as sensitive and reliable as urinary excretion, at least with the analytical methods currently available. Regarding the validity of the two indicators of internal dose the authors of the above studies conclude that urinary vanadium reflects the absorption of vanadium and the relationship between exposure and level of biological indicator allows urinary vanadium to be used as an exposure test. However, according to KIVILUOTO et al. (1979) no significant correlation could be established between vanadium in air and in the urine, any dose-effect relationship remains obscure. In a more recent study by these authors, (KIVILUOTO et al., 1981) there were also no signifi• cant correlations between the V concentrations in the factory air and serum V concentra• tions (r = 0.46) or creatinine adjusted urinary V excretion values (r = 0.41). Significant correlations for the relationship between V in air and creatinine adjusted urinary V concentrations, reported by THÜRAUF et al. (1979) and GLYSETH et al. (1979), were evaluated ¡n a higher exposure range. According to THÜRAUF et al. (1979), the V concen• trations in blood and urine reflect the extent of external V exposure but :he urinary V con• centration is the more sensitive parameter. GLYSETH et al. (1979) concluded that at the exposure levels encountered this study, the blood-V and urinary-V values for a single worker give no clear information about the exposure level. However, on a grcup basis, blood-V and urinary-V analysis can be used to reveal exposure, and under certain conditions the urinary V analysis may give an indication of the exposure level.

According to the results of MARONI et al. (1983), urinary V excretion is the most suitable indicator of exposure and is sensitive enough to permit biological monitoring also at low levels of exposure, whereas blood V is not as sensitive and reliable (ALESSIO et al. 1986). The Institute of Occupational Health in Finland has used the values of 5.1 /¿g/l as a reference value which is rather high for un-exposed subjects, and 30/ig/l as a BTLV for exposed. LAUWERYS et al. (1980) proposed a tentative biological threshold limit value of 50 ¿¿g/g 89 creatinine. For both BTLV proposals neither the basis of evaluation nor the type of vanadium compound concerned is Indicated.

Vanadium concentrations in hair

Determination of the quantity of vanadium in the hair in the general population as well as In patients with maniac-depressive syndromes was performed by NAYLOR et al. (1984). There is considerable evidence, that in maniac-depressive psychosis there is a change in membrane sodium transport. It has been suggested that this change ¡s of aetiologlcal im• portance and is secondary to changes in vanadium metabolism. Vanadium is a potent in• hibitor of the sodium pump. The results suggest that maniac patients have significantly raised vanadium levels In the hair which fall to control levels with recovery, but there are no significant differences ¡n the mean vanadium content of whole blood and serum. In contrast, depressed patients have raised levels of vanadium in whole blood and serum which appear to fall with recovery. Levels of vanadium in serum correlate strongly with those in whole blood, but neither show significant correlation with vanadium in hair for either patients or controls. Hair and blood probably serve as Indicator tissues for differing aspects of vanadium metabolism. No data are available of the values of vanadium in hair ¡n occupationally exposed subjects.

Indicators of effect

At present, there are not data on indicators of an early specific biological or biochemical effect related to vanadium exposure. This fact may be connected to the still unanswered question, I.e. whether vanadium Is an essential trace element for man. It should be noted, that In vivo vanadium inhibits Na + -K + - adenoslntriphosphatase, the synthesis of cholesterol, phospholipids, and L-ascorbic acid, as well as (at higher concen• tration) serotonin oxidation (CURRAN et al. 1956 und 1959, SCHROEDER et al. 1963, LEWIN et al. 1959). On the other hand, in vitro vanadium induces the syntheses of DNA in human fibroblasts (CARPENTER et al. 1981). The decrease of cystine content in fingernails and/or hair has been reported to be an in• dicator of long-term exposure to vanadium (THÜRAUF et al. 1979, MOUNTAIN et al. 1953). But these findings were not confirmed in another study (KIVILUOTO et al. 1980). The cystine content found in the fingernails of persons occupationally exposed to vanadium (8.9 mg/100 mg) was significantly reduced with respect to the control groups (9.9 mg/100 mg). No correlations exist between the cystine content of the fingernails and the age of the persons nor between the V concentrations in blood and urine.

Conclusions

In summary, the indicator proposed for the biological monitoring of workers exposed to vanadium and/or vanadium compounds (e.g. vanadium pentoxide) is vanadium in urine. This test mainly reflects the real internal exposure. The sampling of the urine specimen should be carried out at the end of the shift, preferably at the end of the week. The vanadium concentration in blood/plasma/serum is not sufficiently sensitive for the assess• ment of the lower exposure range (mainly below current TLV-values). At present no dose-effect/response-relationships can be established regarding internal ex• posure and early effects on the respiratory system in occupationally exposed persons. Furthermore, there are no specific and sensitive biological indicators of effect for long- term exposure to vanadium.

Need for further research Further research is required In the following areas: - critical assessment and standardization of the analytical methods for the determination of V in biological fluids as well as the sampling of the specimens; the Implementation of quality control programmes is recommended - the internal-external exposure relationship should be assessed by simultaneous biological and environmental monitoring in workers exposed to different types of vanadium- 90

containing dusts and fumes from different types of vanadium compounds. Furthermore, in these studies the assessment of the external exposure should consider the physical state of vanadium, e.g., fumes or dust evaluation of dose-response-relationships for local irritation by vanadium and Its com• pounds on the respiratory tract investigations to evaluate possible systemic and carcinogenic effects after long-term exposure by inhalation.

Annex

Analytical methods

Reviews on analytical aspects of vanadium in biological material have been published by NAS 1974, BERMAN 1980, PYY 1984. Numerous methods have been described. Nowadays for the analytical assessment of V in biological material spectrophotometric methods, plasma atomic emission spectrometry, neutron activation analysis and atomic absorption spec• trometry are relevant. The extremely low V concentration in the human body permit only limited recording of V by conventional chemical methods. Very sensitive catalytic reactions have been used to determine vanadium in biological materials (FISHMAN and SKOUGSTAD 1964, WELCH and ALLA WAY 1972, KELM and SCHALLER 1978, HOLZHAUSER and SCHALLER 1977, THÜRAUF et al. 1979). Plasma atomic emission spectrometry, e.g. DCP-AES, represents a new technique of In• strumental methods for analytical (PYY et al. 1983). DCP-AES procedure for animal tissues have been published, in which the lower limit of the working range for V is 3 g/l (FRANK and PETERSSON 1983). Urine could be analyzed directly for V levels of 10 g/l, if the method of standard addition is used. For the accurate determination of low V concentrations in urine and serum, the use of an extraction method with ammonium 1-pyrrolidinecarbodithloate (APDC) into 4-methylpentane-2 one was necessary. The detection limit was about 2.5 g/l (PYY et al. 1983). Neutron activation determination seems to be the most reliable method for the analytical determination of V in biological specimens taken from occupationally nonexposed and ex• posed people (ALLEN and STEINNES 1978, GYLSETH et al. 1979). In biological samples, destructive techniques using post-Irradiation radiochemical separation, are capable of achiev• ing an absolute detection limit of about 0.2 ng vanadium (BYRNE and KOSTA 1978). Cor• nells et al. 1980 and Simonoff et al. 1984 also used the neutron activation analysis for the determination of V In blood. They concluded that this technique appears to be the most sensitive and most nearly accurate available for the blood V analysis. In the literature, there are several reports regarding the electrotherma AAS methods for the determination of vanadium in biological samples (KRISHNAN et al. 1976, STROOP et al. 1982, BUCHET et al. 1982, PYY et al. 1984). The analysis of urine and serum samples with no pretreatment is described in the literature (PYY et al. 1984, STROOP et al. 1982). To improve the detection limits and to avoid matrix interferences the preparation of biological samples usually Involves mineralization through dry wet ashing and liquid-extraction. Mineralization with nitric acid Is preferred (Buchet et al. 1982). Cupferron gave the best results in comparison to APDC and dichloroxine (5, 7 - dichloro- 8- quinolinol). The major difficulties reported have involved problems of contamination dur• ing specimen sampling, transport, storage and sample pretreatment. Therefore, especially for vanadium, the preanalytical phase is increasingly important for the accuracy of the results (ANGERER et al. 1983). To avoid contamination, it is necessary to follow recommended procedures (WHO 1981, AITIO and JÄRVISALO 1984, ANGERER and SCHALLER 1985). This is important due to the fact that V is among the least concen• trated elements in biological materials. To determine the "real" V concentration in blood and urine of non-occuoationally exposed subjects, sophisticated efforts were made to avoid contamination (BYRNE and KOSTA 1978). For practicable monitoring in the environment and in the workplace these efforts can not be made. Therefore data elaborated under practicable conditions, which may be a little higher than the "real" V-concentrations, must be sufficient for biological monitoring in oc• cupational medicine. This important aspect, which will become more relevant In the future, needs further discussion. 91

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European Communities - Commission

EUR 11135 — Biological indicators for the assessment of human exposure to industrial chemicals

L. Alessio, A. Berlin, M. Boni, R. Roi

Luxembourg: Office for Official Publications of the European Communities

1987 - VI, 96 pp. — 21.0 χ 29.7 cm

Series: Industrial health and safety

EN

The Biological Monitoring is one of the most promising tools for performing an ef­ fective program of prevention from the potentially toxic effects of chemicals in oc­ cupational environment. Its most interesting feature consists in the possibility of predicting toxicological diseases at a very early stage by means of measurement of indicators of dose and effects. For Industrial chemicals such as Aldrin and Dieldrin, Arsenic, Cobalt, Endrin, Vanadium a description of the types and characteristics of indicators is given. The limitation and difficulties inherent biological monitoring are analyzed, indicating the goals for further research in this field.

CDNA11135ENC