This report contains the collective views of an international group of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

First draft prepared by Drs A. Boehncke, J. Kielhorn, G. Könnecker, C. Pohlenz-Michel, and I. Mangelsdorf, Fraunhofer Institute of Toxicology and Aerosol Research, Drug Research and Clinical Inhalation, Hanover, Germany

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization

Geneva, 2003

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

4-Chloroaniline.

(Concise international chemical assessment document ; 48)

1.Aniline compounds - adverse effects 2.Risk assessment 3.Environmental exposure

4.Occupational exposure I.International Programme on Chemical Safety II.Series

ISBN 92 4 153048 0              (NLM Classification: QV 632)

ISSN 1020-6167

©World Health Organization 2003

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TABLE OF CONTENTS

FOREWORD

Concise International Chemical Assessment Documents (CICADs) are the latest in a family of publications from the International Programme on Chemical Safety (IPCS) — a cooperative programme of the World Health Organization (WHO), the International Labour Organization (ILO), and the United Nations Environment Programme (UNEP). CICADs join the Environmental Health Criteria documents (EHCs) as authoritative documents on the risk assessment of chemicals.

International Chemical Safety Cards on the relevant chemical(s) are attached at the end of the CICAD, to provide the reader with concise information on the protection of human health and on emergency action. They are produced in a separate peer-reviewed procedure at IPCS. They may be complemented by information from IPCS Poison Information Monographs (PIM), similarly produced separately from the CICAD process.

CICADs are concise documents that provide summaries of the relevant scientific information concerning the potential effects of chemicals upon human health and/or the environment. They are based on selected national or regional evaluation documents or on existing EHCs. Before acceptance for publication as CICADs by IPCS, these documents undergo extensive peer review by internationally selected experts to ensure their completeness, accuracy in the way in which the original data are represented, and the validity of the conclusions drawn.

The primary objective of CICADs is characterization of hazard and dose–response from exposure to a chemical. CICADs are not a summary of all available data on a particular chemical; rather, they include only that information considered critical for characterization of the risk posed by the chemical. The critical studies are, however, presented in sufficient detail to support the conclusions drawn. For additional information, the reader should consult the identified source documents upon which the CICAD has been based.

Risks to human health and the environment will vary considerably depending upon the type and extent of exposure. Responsible authorities are strongly encouraged to characterize risk on the basis of locally measured or predicted exposure scenarios. To assist the reader, examples of exposure estimation and risk characterization are provided in CICADs, whenever possible. These examples cannot be considered as representing all possible exposure situations, but are provided as guidance only. The reader is referred to EHC 170.¹

While every effort is made to ensure that CICADs represent the current status of knowledge, new information is being developed constantly. Unless otherwise stated, CICADs are based on a search of the scientific literature to the date shown in the executive summary. In the event that a reader becomes aware of new information that would change the conclusions drawn in a CICAD, the reader is requested to contact IPCS to inform it of the new information.

Procedures

The flow chart on page 2 shows the procedures followed to produce a CICAD. These procedures are designed to take advantage of the expertise that exists around the world — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and other data that are necessary for assessing risks to human health and/or the environment. The IPCS Risk Assessment Steering Group advises the Coordinator, IPCS, on the selection of chemicals for an IPCS risk assessment based on the following criteria:

• there is the probability of exposure; and/or

• there is significant toxicity/ecotoxicity.

Thus, it is typical of a priority chemical that

• it is of transboundary concern;

• it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;

• there is significant international trade;

• it has high production volume;

• it has dispersive use.

The Steering Group will also advise IPCS on the appropriate form of the document (i.e., EHC or CICAD) and which institution bears the responsibility of the document production, as well as on the type and extent of the international peer review.

The first draft is based on an existing national, regional, or international review. Authors of the first draft are usually, but not necessarily, from the institution that developed the original review. A standard outline has been developed to encourage consistency in form. The first draft undergoes primary review by IPCS to ensure that it meets the specified criteria for CICADs.

The second stage involves international peer review by scientists known for their particular expertise and by scientists selected from an international roster compiled by IPCS through recommendations from IPCS national Contact Points and from IPCS Participating Institutions. Adequate time is allowed for the selected experts to undertake a thorough review. Authors are required to take reviewers’ comments into account and revise their draft, if necessary. The resulting second draft is submitted to a Final Review Board together with the reviewers’ comments. At any stage in the international review process, a consultative group may be necessary to address specific areas of the science.

The CICAD Final Review Board has several important functions:

• to ensure that each CICAD has been subjected to an appropriate and thorough peer review;
• to verify that the peer reviewers’ comments have been addressed appropriately;
• to provide guidance to those responsible for the preparation of CICADs on how to resolve any remaining issues if, in the opinion of the Board, the author has not adequately addressed all comments of the reviewers; and
• to approve CICADs as international assessments.

Board members serve in their personal capacity, not as representatives of any organization, government, or industry. They are selected because of their expertise in human and environmental toxicology or because of their experience in the regulation of chemicals. Boards are chosen according to the range of expertise required for a meeting and the need for balanced geographic representation.

Board members, authors, reviewers, consultants, and advisers who participate in the preparation of a CICAD are required to declare any real or potential conflict of interest in relation to the subjects under discussion at any stage of the process. Representatives of nongovernmental organizations may be invited to observe the proceedings of the Final Review Board. Observers may participate in Board discussions only at the invitation of the Chairperson, and they may not participate in the final decision-making process.

This CICAD on 4-chloroaniline (p-chloroaniline) was prepared by the Fraunhofer Institute of Toxicology and Aerosol Research, Hanover, Germany. It is based on reports prepared by the German Advisory Committee on Existing Chemicals of Environmental Relevance (BUA, 1995), the German MAK Commission (MAK, 1992), and the US National Toxicology Program (NTP, 1989). A comprehensive literature search of relevant databases was conducted in March 2001 to identify any relevant references published subsequent to those incorporated in these reports. Information on the preparation and peer review of the source documents is presented in Appendix 1. Information on the peer review of this CICAD is presented in Appendix 2. This CICAD was approved as an international assessment at a meeting of the Final Review Board, held in Ottawa, Canada, on 29 October – 1 November 2001. Participants at the Final Review Board meeting are listed in Appendix 3. The International Chemical Safety Card on 4-chloroaniline (ICSC 0026), produced by the International Programme on Chemical Safety (IPCS, 1999), has also been reproduced in this document.

Anilines chlorinated at the 2, 3, and 4 (ortho, meta, and para) positions have the same use patterns. All chloroaniline isomers are haematotoxic and show the same pattern of toxicity in rats and mice, but in all cases 4-chloroaniline shows the most severe effects. 4-Chloroaniline is genotoxic in various systems (see below), while the results for 2- and 3-chloroaniline are inconsistent and indicate weak or no genotoxic effects. This CICAD therefore focuses only on 4-chloroaniline as the most toxic of the chlorinated anilines.

4-Chloroaniline (in the following called PCA) (CAS No. 106-47-8) is a colourless to slightly amber-coloured crystalline solid with a mild aromatic odour. The chemical is soluble in water and in common organic solvents. PCA has a moderate vapour pressure and n-octanol/

water partition coefficient. It decomposes in the presence of light and air and at elevated temperatures.

PCA is used as an intermediate in the production of a number of products, including agricultural chemicals, azo dyes and pigments, cosmetics, and pharmaceutical products. Thus, releases of PCA into the environment may occur from a number of industrial sources (e.g., production, processing, dyeing/printing industry).

The main environmental target compartment of the chemical can be predicted from its use pattern to be the hydrosphere. Measured concentrations in, for example, the river Rhine and its tributaries are roughly between 0.1 and 1 µg/litre. In the hydrosphere, PCA is rapidly degraded under the influence of light (measured half-lives 2–7 h). The calculated half-life of the chemical in air for the reaction with hydroxyl radicals is 3.9 h. Numerous studies on the biodegradation of PCA indicate it to be inherently biodegradable in water under aerobic conditions, whereas no significant mineralization was detected under anaerobic conditions.

Soil sorption coefficients in a variety of soil types determined according to the Freundlich sorption isotherm indicate only a low potential for soil sorption. In most experiments, soil sorption increased with increasing organic matter and decreasing pH values. As a consequence, under conditions unfavourable for abiotic and biotic degradation, leaching of PCA from soil into groundwater, particularly in soils with a low organic matter content and elevated pH levels, may occur. The available experimental bioconcentration data as well as the measured n-octanol/water partition coefficients indicate no bioaccumulation potential for PCA in aquatic organisms.

PCA is rapidly absorbed and metabolized. The main metabolic pathways of PCA are as follows: a) C-hydroxylation in the ortho position to yield 2-amino-5-chlorophenol followed by sulfate conjugation to 2-amino-5-chlorophenyl sulfate, which is excreted per se or after N-acetylation to N-acetyl-2-amino-5-chlorophenyl sulfate; b) N-acetylation to 4-chloroacetanilide (found mainly in blood), which is further transformed to 4-chloroglycolanilide and then to 4-chlorooxanilic acid (found in the urine); or c) N-oxidation to 4-chlorophenylhydroxylamine and further to 4-chloronitrosobenzene (in erythrocytes).

Reactive metabolites of PCA bind covalently to haemoglobin and to proteins of liver and kidney. In humans, haemoglobin adducts are detectable as early as 30 min after accidental exposure, with a maximum level at 3 h. Slow acetylating individuals have a higher potency to form haemoglobin adducts compared with fast acetylators.

Excretion in animals or humans occurs primarily via the urine, with PCA and its conjugates appearing as early as 30 min after exposure. Excretion takes place mainly during the first 24 h and is almost complete within 72 h.

Oral LD₅₀ values of 300–420 mg/kg body weight for rats, 228–500 mg/kg body weight for mice, and 350 mg/kg body weight for guinea-pigs are reported. Similar values have been found for intraperitoneal and dermal application of PCA to rats, rabbits, and cats. An LC₅₀ value for rats was given as 2340 mg/m³. The prominent toxic effect is methaemoglobin formation. PCA is a more potent and faster methaemoglobin inducer than aniline. PCA also exhibits a nephrotoxic and hepatotoxic potential.

PCA was found to be non-irritating to rabbit skin and slightly irritating to rabbit eyes. A weak sensitizing potential was demonstrated with several test systems.

Repeated exposure to PCA leads to cyanosis and methaemoglobinaemia, followed by effects in blood, liver, spleen, and kidneys, manifested as changes in haematological parameters, splenomegaly, and moderate to heavy haemosiderosis in spleen, liver, and kidney, partially accompanied by extramedullary haematopoiesis. These effects occur secondary to excessive compound-induced haemolysis and are consistent with a regenerative anaemia. The lowest-observed-adverse-effect levels (LOAELs; no-observed-effect levels, or NOELs, are not derivable) for a significant increase in methaemoglobin levels in rats and mice are, respectively, 5 and 7.5 mg/kg body weight per day for a 13-week oral administration of PCA by gavage (5 days/week) and 2 mg/kg body weight per day for rats administered PCA by gavage (5 days/week) at 26, 52, 78, and 103 weeks’ exposure. Fibrotic changes of the spleen were observed in male rats, with a LOAEL of 2 mg/kg body weight per day, and hyperplasia of bone marrow was observed in female rats, with a LOAEL of 6 mg/kg body weight per day (103-week gavage).

PCA is carcinogenic in male rats, with the induction of unusual and rare tumours of the spleen (fibrosarcomas and osteosarcomas), which is typical for aniline and related substances. In female rats, the precancerous stages of the spleen tumours are increased in frequency. Increased incidences of pheochromocytoma of the adrenal gland in male and female rats may have been related to PCA administration. There was some evidence of carcinogenicity in male mice, indicated by hepatocellular tumours and haemangiosarcoma.

PCA shows transforming activity in cell transformation assays. A variety of in vitro genotoxicity tests (e.g., Salmonella mutagenicity test, mouse lymphoma assay, chromosomal aberration test, induction of sister chromatid exchange) indicate that PCA is possibly genotoxic, although results are sometimes conflicting. Due to lack of data, it is impossible to make any conclusion about PCA’s in vivo genotoxicity.

No studies are available on reproductive toxicity.

Data on occupational exposure of humans to PCA are mostly from a few older reports of severe intoxications after accidental exposure to PCA during production. Symptoms include increased methaemoglobin and sulfhaemoglobin levels, cyanosis, the development of anaemia, and changes due to anoxia. PCA has a strong tendency to form haemoglobin adducts, and their determination can be used in biomonitoring of employees exposed to 4-chloroaniline in the workplace.

There are reports of severe methaemoglobinaemia in neonates from neonatal intensive care units in two countries where premature babies were exposed to PCA as a breakdown product of chlorohexidine; the chlorohexidine, which had been inadvertently used in the humidifying fluid, broke down to PCA upon heating in a new type of incubator. Three neonates in one report (14.5–43.5% methaemoglobin) and 33 of 415 neonates in another report (6.5–45.5% methaemoglobin during the 8-month screening period) were found to be methaemoglobin positive. A prospective clinical study showed that immaturity, severe illness, time exposed to PCA, and low concentrations of NADH reductase probably contributed to the condition.

From valid test results available on the toxicity of PCA to various aquatic organisms, PCA can be classified as moderately to highly toxic in the aquatic compartment. The lowest no-observed-effect concentration (NOEC) found in long-term studies with freshwater organisms (Daphnia magna, 21-day NOEC 0.01 mg/litre) was 10 times higher than maximum levels determined in the river Rhine and its tributaries during the 1980s and 1990s. Therefore, a possible risk to aquatic organisms, particularly benthic species, cannot be completely ruled out, particularly in waters where significant amounts of particulate matter inhibit rapid photomineralization. The only benthic species tested, however, exhibited no significant sensitivity (48-h EC₅₀ 43 mg/litre), and experiments with Daphnia magna revealed significantly reduced toxicity with increasing concentrations of dissolved humic materials in the medium, possibly caused by reduced bioavailability of PCA from adsorption to dissolved humic materials. Furthermore, bioaccumulation in aquatic species was reported to be very low. Therefore, from the available data, a significant risk associated with exposure of aquatic organisms to PCA is not to be expected.

The data available on microorganisms and plants indicate only a moderate toxicity potential of PCA in the terrestrial environment. There is a 1000-fold safety margin between reported effects and concentrations in soil.

Assessment of consumer exposure to PCA via a number of possible routes leads to total exposure of a maximum 300 ng/kg body weight per day, assuming only 1% penetration by clothes. Considering only non-neoplastic effects (i.e., methaemoglobinaemia), these possible human exposures are within an order of magnitude of the calculated tolerable daily intake of 2 µg/kg body weight per day. Acute incidental exposure to high concentrations of PCA can be fatal.

Further effects that give reason for concern are carcinogenicity and possibly skin sensitization.

Residual levels of PCA in consumer products should be further reduced or entirely eliminated.

4-Chloroaniline (CAS No. 106-47-8) is a colourless to slightly amber-coloured crystalline solid aniline derivative with a mild aromatic odour. It has the chemical formula C₆H₆ClN, and its relative molecular mass is 127.57. Its molecular structure is shown in Figure 1. Its IUPAC name is 1-amino-4-chlorobenzene; other names include PCA, p-chloroaniline, 1-chloro-4-aminobenzene, 4-chloro-1-aminobenzene, 4-chlorobenzenamine, 4-chloroaminobenzene, and 4-chlorophenylamine. Depending on the purity of the product, the substance melts between 69 and 73 °C. Its boiling point is given as 232 °C (BUA, 1995).

Figure 1: Molecular structure of 4-chloroaniline

The water solubility of PCA is given as 2.6 g/litre at 20 °C (Scheunert, 1981) and 3.9 g/litre (Kilzer et al., 1979). PCA dissociates in water, as it is a weak acid (measured pK_a 4.1–4.2 at 20 °C; BUA, 1995). The chemical is furthermore readily soluble in most organic solvents (BUA, 1995). Numerous measured data on vapour pressure are available. The vapour pressure is 0.5 Pa at 10 °C and between 1.4 and 2.1 Pa at 20 °C (BUA, 1995). The n-octanol/water partition coefficients measured by high-performance liquid chromatographic (HPLC) or gas chromatographic (GC) standard methods are 1.83 and 2.05, respectively (Kishida & Otori, 1980; Kotzias, 1981; Garst & Wilson, 1984). From its water solubility and vapour pressure at 20 °C, a Henry’s law constant of about 0.1 Pa·m³/mol (air/water partition coefficient = 4.1 × 10^–5) can be calculated for PCA.

PCA decomposes in the presence of light and air and at elevated temperatures (decomposition temperature 250–300 °C; BUA, 1995). Very vigorous reactions may occur with strong oxidants (Hommel, 1985). The decomposition in the presence of light is due to direct photolysis. In ethanolic solution, PCA shows a strong absorption maximum at about 300 nm (concentration not given; absorption coefficient log epsilon [from graphical presentation] = 3.3; Kharkharov, 1954).

The conversion factors² for PCA in air (20 °C, 101.3 kPa) are as follows:

1 mg/m³ = 0.189 ppm

1 ppm = 5.30 mg/m³

Additional physicochemical properties for PCA are presented in the International Chemical Safety Card (ICSC 0026) reproduced in this document.

PCA can be analysed by either GC or HPLC methods. Analysis by GC (usually capillary columns) is frequently combined with a preceding derivatization step (e.g., diazotation/azo coupling, bromination). All common detectors (flame ionization, phosphorus–nitrogen, thermoionic, and electron capture) are used, with the electron capture detector showing the highest sensitivity. HPLC is in general carried out with reversed phases, ultraviolet (UV) detection being most important. The photodiode-array detector and the electrochemical detector are also applied. For GC and HPLC, mass spectrometric detection can be used for the identification of PCA. Thin-layer chromatographic methods (e.g., for screening purposes) are also described (see, for example, Ramachandran & Gupta, 1993). A detailed compilation of common detection methods is given in BUA (1995). Furthermore, the analytical methods that are described for the isomers 2-chloroaniline and 3-chloroaniline (BUA, 1991) may be used for the detection of PCA.

PCA has been shown to adsorb to several laboratory plastics (silicone, 44%; soft polyvinyl chloride, 30%; rubber, 25%; and polyvinyl acetate, 33%) with a test period of 4 h and a starting concentration of 50 mg/litre (Janicke, 1984). This may affect the recovery rates and sensitivity of the analytical determination of PCA.

The number of methods for the determination of PCA in air is small. Some methods for workplace surveillance (GC and HPLC) are described. However, a thorough validation of these methods is lacking (BIA, 1992; OSHA, 1992). The detection limit is given as 1.1 mg/m³ (OSHA, 1992).

Numerous methods are available for analysis of PCA in the water compartment (BUA, 1995; Holm et al., 1995; Börnick et al., 1996; Götz et al., 1998). Micro-extraction methods are also described (Müller et al., 1997; Fattore et al., 1998). Very low detection limits of up to 0.002 µg/litre can be achieved, especially with GC methods. With HPLC methods, the detection limits are in the range of 0.04–100 µg/litre. The recovery rates are in general near 100%. Also, some methods are available for the determination of PCA in wastewater (Riggin et al., 1983; Gurka, 1985; Onuska et al., 2000). Although no validated chromatographic method is available for the determination of PCA in drinking-water samples, the methods used for groundwater and surface water should be applicable.

For the detection of PCA in soil, a very sensitive GC method was reported (detection limit 1 µg/kg, recovery rate >90%; Wegman et al., 1984).

Combined with an appropriate enrichment technique (e.g., acid hydrolysis), the HPLC and GC methods can also be used in the determination of PCA in biological material such as urine (Lores et al., 1980; Hargesheimer et al., 1981). Recovery rates between 93 and 104% and a detection limit of <5 µg/litre were determined (Lores et al., 1980).

Some HPLC and GC methods are available for the detection of PCA in consumer products such as handwash and mouthwash products and dyed papers and textiles (Kohlbecker, 1989; Gavlick, 1992; Gavlick & Davis, 1994; BGVV, 1996). The method of BGVV (1996) is the official method in Germany for controlling the PCA content in dyed textiles and papers. Recovery rates of about 97% and detection limits up to 43 µg/kg are reported.

There are no known natural sources of PCA.

PCA is produced by low-pressure hydrogenation of 4-chloronitrobenzene in the liquid phase in the presence of noble metal and/or noble metal sulfide catalysts. The addition of metal oxides helps to avoid dehalogenation. The yield is about 98% (Kahl et al., 2000). The two German manufacturers produce PCA either continuously or in a batch process at temperatures of 50–60 °C, at pressures between 1000 and 10 000 kPa, and with toluene or isopropanol as solvent. The raw product is purified by distillation. Catalyst and solvent are recycled into the reactor (BUA, 1995).

In 1988, the global annual production figure was 3500 tonnes (Srour, 1989). A more recent global production figure is not available. In 1990, about 1350 tonnes of PCA were manufactured in the former Federal Republic of Germany, of which about 350 tonnes were exported; about 850 tonnes were processed by the manufacturers themselves. France has registered PCA in the high production volume chemicals programme of the Organisation for Economic Co-operation and Development (OECD), which indicates that the produced amount in this country is >1000 tonnes/year (OECD, 1997). For 1995, combined Western European and Japanese PCA production is given as 3000–3300 tonnes. In India and China, another 800–1300 tonnes/year are produced (Srour, 1996). For 1991, the annual US production quantity is estimated to be 45–450 tonnes (IARC, 1993). More recent data are not available.

PCA is used as an intermediate in the production of several urea herbicides and insecticides (e.g., monuron, diflubenzuron, monolinuron), azo dyes and pigments (e.g., Acid Red 119:1, Pigment Red 184, Pigment Orange 44), and pharmaceutical and cosmetic products (e.g., chlorohexidine, triclocarban [3,4,4'-trichlorocarbanilid], 4-chlorophenol) (Srour, 1989; BUA, 1995; Herbst & Hunger, 1995; Hunger et al., 2000; IFOP, 2001). In 1988, about 65% of the global annual production was processed to pesticides (Srour, 1989). In Germany, in 1990, about 7.5% was used as dye precursors, 20% as intermediates in the cosmetics industry, and 60% as pesticide intermediates. The use for the remaining 12.5% of the production quantity was not specified (BUA, 1995). More recent data on the use pattern of PCA are not available.

The PCA-based azo dyes and pigments are especially used for the dyeing and printing of textiles (Herbst & Hunger, 1995; Hunger et al., 2000). Triclocarban is a bactericide in deodorant soaps, sticks, sprays, and roll-ons (Srour, 1995), and chlorohexidine is used in mouthwashes (BUA, 1995) and spray antiseptics. 4-Chlorophenol is also listed as an antimicrobial agent for cosmetic products in the European Inventory of Cosmetic Ingredients (EC, 2001). However, no information is available on the products in which it is used. All of these products may contain residual PCA, or PCA may emerge during their degradation (see sections 6 and 11).

The marketing and use of products containing PCA-based azo dyes were recently banned by the European Union (EU) (EC, 2000).

The global release of PCA cannot be estimated with the available data.

The releases from the production of PCA at the German manufacturers in 1990 were as follows: <20 g/tonne produced released into air at each site (derived from the registry limit of the emission declaration of 25 kg/year), and 13 g/tonne produced released into surface water. The annual wastes are estimated to be a maximum of 400 g PCA/tonne produced. These wastes are disposed of in special company incinerators (BUA, 1995).

The releases from the processing of PCA at the German manufacturers in 1990 were as follows (assuming a processed quantity of approximately 1000 tonnes/

year): <25 g/tonne processed released into air at each site (derived from the registry limit of the emission declaration of 25 kg/year), and 240 g/tonne processed released into surface water (estimated degradation in industrial wastewater treatment plant about 85%). The annual wastes are estimated to be a maximum of 695 g PCA/tonne processed. These wastes are disposed of in special company incinerators (BUA, 1995).

Total releases of PCA for 1995, 1998, and 1999 in the USA were given as 500, 2814, and 212 kg, respectively (US Toxics Release Inventory, 1999).

In wastepaper and wastewater samples from the industrial wastepaper de-inking process in Germany, PCA (seven samples each) was not detected before or after the bleaching process at a detection limit of 1 mg/kg for solid and 1 mg/litre for liquid samples (Hamm & Putz, 1997). Data on releases from other countries or other industries (dyeing, printing) are not available.

PCA releases into the hydrosphere may furthermore result from the use of pesticides that contain residual PCA and/or form PCA as a degradation product. In some laboratory experiments in the anaerobic water layer over aquarium soil and in water samples of pasture, both being treated with diflubenzuron, PCA concentrations of 0.1 to about 4 µg/litre were detected a few days after the application (Booth & Ferrell, 1977; Schaefer et al., 1980). However, PCA was not detected in the drainage and in the groundwater after the use of diflubenzuron under field conditions in a Finnish forest area (Mutanen et al., 1988).

PCA could, in principle, be released into surface waters from the use of dyed textiles and printed papers. A residual PCA content of <100 mg/kg is reported for German dye products (BUA, 1995). A quantification of the releases from this source is not possible with the available data. However, as noted above, the marketing and use of PCA-based azo dyes have recently been restricted in the EU (EC, 2000).

The releases of PCA into the hydrosphere from the use of pharmaceuticals and cosmetics (e.g., chlorohexidine-based mouthwashes, triclocarban-containing soaps) with residual PCA are also not quantifiable with the available data. The residual PCA content in chlorohexidine is given as <500 mg/kg (<0.05%)³; in triclocarban, it is given as <100 mg/kg (<0.01%).⁴ If solutions of chlorohexidine are stored for prolonged periods (2 years or more) at high (tropical) ambient temperatures or if they are inadvertently heat sterilized, the PCA content may reach 2000 mg/litre (0.2%) (Scott & Eccleston, 1967; Hjelt et al., 1995).

In 1985, 6.1 tonnes of monochloroanilines (sum of 2-, 3-, and 4-chloroaniline), coming completely from industrial processes, were estimated to be released to the river Rhine (IAWR, 1998).

The application of pesticides (mainly phenylureas) may lead to releases of PCA into soils. Monolinuron is reported to contain an average of 0.1% PCA. The insecticide diflubenzuron and the herbicides monolinuron, buturon, propanil, chlorofenprop-methyl, benzoylpropmethyl, chloroaniformmethane, chlorobromuron, neburon, and oxadiazon can release PCA as a degradation product; this has been confirmed in laboratory experiments with some of these pesticides (diflubenzuron, monolinuron) using radioactive labelling. However, the reported concentrations vary widely. PCA can be released from 3,4-dichloroaniline only under anaerobic conditions. Under aerobic conditions, a complete mineralization without the intermediate synthesis of PCA was observed (BUA, 1995). In general, the results from laboratory tests are supported by a field study on PCA concentrations in agricultural soils in Germany. In 54 of 354 soil samples, PCA was detected with a maximum concentration of 968 µg/kg (Lepschy & Müller, 1991) (see section 6.1). An estimation of the total amount of PCA released from the use of pesticides cannot be made with the available data. In Germany, phenylurea pesticides, which may form PCA during their degradation, are no longer marketed.

Releases of PCA into the biosphere from the application of pesticides also appear to be possible. However, PCA concentrations in wild mushrooms, blueberries, and cranberries were below the detection limit of 10–20 µg/kg after diflubenzuron treatment in a forest area in Finland in 1984 (Mutanen et al., 1988). PCA was also not detected in spinach plants that were cultivated in soil treated with monolinuron or in the subsequently cultivated cultures of cress and potatoes (Schuphan & Ebing, 1978). In contrast, PCA concentrations between 0.9 and 1.3 µg/kg were detected in tissue samples of the bluegill (Lepomis macrochirus) 19 days after the application of diflubenzuron to an artificial pond (detection limit 0.8 µg/kg; calculated diflubenzuron concentration in pond 200 µg/litre; Schaefer et al., 1980).

PCA has a moderate vapour pressure (see section 2). From this, significant adsorption of the substance onto airborne particles is not to be expected. Releases of the chemical into air may, however, be washed out of the atmosphere by wet deposition (e.g., fog, rain, snow). Measured data on this are not available.

From PCA’s water solubility and vapour pressure, a Henry’s law constant of about 0.1 Pa·m³/mol can be calculated (see also section 2), suggesting a low volatility from aqueous solutions (Thomas, 1990). The measured half-life of PCA in water, measured according to a draft OECD Guideline from 1980 (unspecified; probably the Test Guideline for the Determination of the Volatility from Aqueous Solution of February 1980), is 151 days at a water depth of 1 m and a temperature of 20 °C (Scheunert, 1981). From this and its use pattern (see section 4), the hydrosphere is expected to be the main target compartment for PCA.

Evaporation from soil was found to be in the range of 0.11–3.65% of applied PCA, depending on soil type and sorption capacity (Fuchsbichler, 1977; Kilzer et al., 1979).

PCA is hydrolytically stable, as measured according to OECD Guideline A-79.74 D (25 °C; pH 3, pH 7, pH 9) (Lahaniatis, 1981). This is confirmed by measurements at 55 °C and pH 3, pH 7, and pH 11, which gave a half-life of about 3 years (initial concentration 129 mg/litre; Ekici et al., 2001).

From the UV spectrum of PCA (see section 2), direct photolysis of the chemical in air and water appears to be likely. However, as far as the atmosphere is concerned, the main degradation pathway is the reaction of PCA with hydroxyl radicals. The rate constant for this reaction was measured in flash photolysis/resonance fluorescence model experiments to be 8.2 ± 0.4 × 10^–11 cm³/molecule per second (Wahner & Zetzsch, 1983). Calculated figures are in the same order of magnitude (BUA, 1995). Assuming a mean global hydroxyl radical concentration of 6 × 10⁵ molecules/cm³ (BUA, 1993), the half-life of PCA in the troposphere can be calculated to be 3.9 h. From this, long-distance transport of PCA in ambient air is assumed to be negligible.

Irradiation of an aqueous PCA solution with light of wavelengths >290 nm (emission maximum 360 nm) led to a half-life of 7.25 h (Kondo et al., 1988) or to total disappearance of the substance after 6 h (Miller & Crosby, 1983). 4-Chloronitrobenzene and 4-chloronitrosobenzene were detected as primary degradation products. 4-Chloronitrobenzene was stable over the irradiation period of 20 h (Miller & Crosby, 1983). Furthermore, very short half-lives of 2 h (summer, 25 °C) and 4 h (winter, 15 °C) were measured in irradiation model experiments with sterilized natural river water (Hwang et al., 1987). From this, it can be concluded that PCA in aqueous solutions is rapidly degraded by direct photolysis.

Numerous tests have been performed on the biodegradability of PCA in various media. Tests performed according to internationally accepted standard procedures under aerobic conditions are summarized in Table 1. Whereas no degradation of PCA was found in tests on ready biodegradability (closed bottle test), >60% removal was observed in most tests on inherent biodegradability. In two of the latter tests (e.g., Zahn- Wellens test), however, nearly half of the elimination could be attributed to adsorption (Rott, 1981b; Haltrich, 1983). In non-standardized tests using spiked test material, 14.5 and 23% of the applied PCA concentration of 25 g/litre were mineralized by activated sludge within 5 days (Rott et al., 1982; Freitag et al., 1985). Thus, under conditions not particularly favouring abiotic removal during sewage treatment, PCA may be applied to agricultural soils in sludges.

Table 1: Elimination of PCA in standard biodegradation tests under aerobic conditions.^a

Non-standardized experiments with mixed microbial inocula from natural surface water (eutrophic pond, river estuary) showed no significant microbial decomposition of PCA. Observed elimination could be attributed to photodegradation, evaporation, and/or autoxidation (Lyons et al., 1985; Hwang et al., 1987). Thus, under conditions not favouring biotic or abiotic removal in surface waters, adsorption of PCA to sediment particles can be expected.

In several experiments with microbial cultures from soil, removal of PCA was in the range of 0–17% when non-acclimated inocula were used (Alexander & Lustigman, 1966; Fuchsbichler, 1977; Bollag et al., 1978; Süß et al., 1978; Kloskowski et al., 1981a; Cheng et al., 1983). Significant removal of >50% was observed after incubation periods exceeding 8 days only when cultures had been previously cultivated with the herbicide propham (McClure, 1974).

In tests on the incorporation and metabolism of PCA applied to cell suspension cultures of monocotyledon and dicotyledon plants, Harms & Langebartels (1986a,b) observed the formation of considerable amounts of polar metabolites in the cell extracts of soybean (13.5%) and wheat (6.1%).

Under anaerobic conditions, no significant biodegradation was found in sludge (US EPA, 1981; Wagner & Bräutigam, 1981) or aquifer samples (Kuhn & Suflita, 1989).

Under aerobic conditions, PCA released to soil may covalently bind to soil particles, particularly in the presence of high amounts of organic material and/or clay and under low pH levels. However, soil sorption coefficients in a variety of soil types, determined according to the Freundlich sorption isotherm, were in the range of 1.5–50.4, with the highest levels found in soils containing the highest amounts of organic carbon (Fuchsbichler, 1977; van Bladel & Moreale, 1977; Müller-Wegener, 1982; Rippen et al., 1982; Quast, 1984; Scheubel, 1984; Gawlik et al., 1998). In most experiments, soil sorption increased with increasing organic matter and decreasing pH values (Fuchsbichler, 1977; van Bladel & Moreale, 1977). Sorption to the clay fraction was less pronounced (Worobey & Webster, 1982). As a consequence, under conditions unfavourable for abiotic and biotic degradation, leaching of PCA from soil into groundwater, particularly in soils with a low organic matter content and elevated pH levels, may be expected.

Accumulation factors determined for PCA in various aquatic species are summarized in Table 2. For activated sludge and the green alga Chlorella fusca, levels were in the range of 240–280 when based on fresh weight, whereas factors based on dry weight were reported to be up to 1300. However, considering the sorption behaviour of PCA observed in biodegradation tests, the accumulation factors found in sludge and algae are expected to be caused by adsorption to surfaces rather than by bioaccumulation. Bioconcentration factors determined for fish in static and semistatic test systems were considerably lower, ranging from 4 to 20, even at exposure concentrations up to 5 mg/litre. Dissolved humic materials added to the exposure medium had no significant influence on the bioconcentration of PCA in Daphnia magna (Steinberg et al., 1993).

The available experimental bioconcentration data as well as the measured n-octanol/water partition coefficients (1.83 and 2.05) indicate no bioaccumulation potential for PCA in aquatic organisms.

Table 2: Bioaccumulation of PCA in aquatic species.^a

The uptake and incorporation of PCA from soil into cultivated plants have been shown in several experiments (Fuchsbichler, 1977; Kloskowski et al., 1981a,b; Freitag et al., 1984; Harms & Langebartels, 1986a,b; Harms, 1996). PCA was predominantly incorporated in the roots. A translocation into shoots was also detected, with the amount mainly depending on the applied PCA concentration and the growth stage of the exposed plants (Fuchsbichler, 1977). In tests on the incorporation of PCA in cell suspension cultures of monocotyledon plants (corn, wheat), Pawlizki & Pogany (1988) found residual PCA and its metabolites mainly bound to the cell wall in the lignin and pectin fractions; in dicotyledon cell cultures (tomato), it was detected in the starch, protein, and pectin fractions.

Data on the concentrations of PCA in ambient and indoor air are not available.

The concentrations of PCA in surface water from the German and Dutch parts of the river Rhine and its tributaries in the 1980s and 1990s were reviewed in BUA (1995); the levels were roughly between 0.1 and 1 µg/litre. A temporal trend cannot be derived from the data, as the measured concentrations are in the range of the detection limit. In the same period, measured PCA concentrations in surface water in Japan were between 0.024 and 0.39 µg/litre, the chemical being detected in 9 of 128 samples (Office of Health Studies, 1985). In 1992, PCA was not detected at a detection limit of 0.002 µg/litre at two sampling sites in the river Elbe upstream and downstream of Hamburg harbour (Götz et al., 1998). In 1995, the PCA concentrations in the river Rhine and its main tributaries were below the detection limit of 0.5 µg/litre. In the river Emscher, a concentration of 0.84 µg/litre was measured in 1995 (LUA, 1997). More recent data are not available.

In the 1980s and 1990s, PCA was also found in German drinking-water samples at concentrations between 0.007 and 0.013 µg/litre (BUA, 1995). More recent data are not available.

PCA was not detected (detection limit 0.2 µg/litre) in groundwater after the application of the insecticide diflubenzuron in 1984 in Finland (Mutanen et al., 1988). In the groundwater below a Danish landfill site containing domestic wastes and wastes from pharmaceutical production, PCA was detected at concentrations between <10 µg/litre (depth 5.5 m) and 50 µg/litre (depth 8.5 m) (Holm et al., 1995). It is supposed by Holm et al. (1995) that PCA was formed from the wastes from pharmaceutical production (e.g., sulfonamides). In 1995–96, PCA was found in groundwater from three sites in an industrialized area near Milan, Italy, at concentrations between 0.01 and 0.06 µg/litre (positive results in four of seven wells) (Fattore et al., 1998).

In a German agricultural soil measurement programme detecting the degradation products from phenylurea herbicide application, including PCA (see also section 4.4), the following concentrations were found (Lepschy & Müller, 1991):

A maximum of 30 µg/kg was detected in the upper soil horizons of meadows not being treated with phenylurea herbicides (BUA, 1995).

In 1976, PCA was detected in 39 of 121 Japanese sediment samples at concentrations between 1 and 270 µg/kg (detection limit 0.5–1200 µg/kg; Office of Health Studies, 1985). More recent data are not available.

PCA was not detected (detection limit 1000 µg/kg) in two not further specified Japanese fish samples from 1976 (Office of Health Studies, 1985). More recent data concerning the occurrence of PCA in biological material are not available.

6.2.1 Workplaces

Workplace exposure to PCA may occur during production and processing and in industrial dyeing and printing processes. The exposure can be by inhalation of dust containing PCA or by dermal contact with either PCA itself or products containing residual PCA.

For the production of PCA, only some older exposure data from a Hungarian production facility are given by Pacséri et al. (1958), with average values of 58 (range 37–89) and 63 (range 46–70) mg/m³ (two sites measured in one facility). More recent data were not available.

For the processing of PCA, only some older data from a Russian monuron manufacturing facility are available, with concentrations ranging from 0.2 to 2.0 mg/m³ (Levina et al., 1966). More recent data on inhalation exposure were not available.

Studies in a number of US textile dyeing factories, especially during weighing/mixing operations, gave concentration ranges of the colorant between 0.007 and 0.56 mg/m³ (personal sampling of 24 textile weighers; 95th percentile 0.27 mg/m³) for long-term measurements (8-h time-weighted average; US EPA, 1990). Assuming a PCA content in the corresponding dyes and pigments of <100 mg/kg (see section 4.4), PCA levels in the air of these workplaces are estimated to be below 27 ng/m³. Assuming an uptake by inhalation of 100%, an 8-h inhalation volume of about 10 m³, and a body weight of 64 kg (IPCS, 1994), a workshift inhalation intake of PCA of <4 ng/kg body weight per day can be calculated.

Measured or estimated data on dermal exposure to PCA at the different workplaces are not available.

6.2.2 Consumer exposure

The general public may be exposed to PCA from the use of PCA-based dyed/printed textiles and papers and cosmetic and pharmaceutical products. The exposure can result from residual PCA in the commercial product or from degradation of this product to PCA during use. It can be dermal (wearing of clothes, use of soaps or mouthwashes), oral (small children sucking clothes and other materials, use of mouthwashes), or by direct entry into the bloodstream (e.g., through breakdown products of chlorohexidine in spray antiseptics).

Estimates of exposure to PCA from the wearing of dyed textiles can be carried out on the basis of estimates made by the United Kingdom Laboratory of the Government Chemist (LGC, 1998). For direct dyes, for example, a dye weight of 0.5 g/m², a weight fraction of 0.8 at 4% depth of shade, and a migration rate of 0.01%/h are assumed. If it is also assumed that the exposure time is 10 h/day, the exposed surface area is 1.7 m², the percutaneous penetration of the dye is 1%, and the extent of azo cleavage via metabolism is 30% (according to Collier et al., 1993), the effective dermal body dose for PCA can be calculated to be 27 ng/kg body weight per day (average body weight 64 kg; see IPCS, 1994). Assuming 100% percutaneous penetration of the dye, the body dose is estimated at 2.7 µg/kg body weight per day.

The sucking of dyed clothes by small children may lead to oral exposure to PCA. This can also be estimated according to LGC (1998) for, for example, a direct dye (for dyeing and migration assumptions, see above). Assuming an exposure time of 6 h/day, a sucked area of 0.001 m², and five sucking bursts per minute, with three sucks per burst, oral doses between about 1 µg/kg body weight per day (1% azo cleavage) and 130 µg/kg body weight per day (100% azo cleavage) can be calculated (body weight of young child is 10 kg, according to LGC, 1998).

In contrast, dermal and oral exposure to PCA due to the metabolism of azo compounds on printed textiles can be assumed to be negligible, as pigments are used. The bioavailability of these water-insoluble products is assumed to be low. Therefore, only the exposure from residual PCA has to be considered. Pigment dispersions for textile printing contain between 25 and 50% pigment (Koch & Nordmeyer, 2000). As no further data are available on the pigment application during textile printing, an estimation of dermal exposure from this source is not possible at this time.

From the use of triclocarban-containing deodorant products (see section 4), dermal exposure to residual PCA concentrations can be estimated as follows. In the EU, the maximum permitted amount of triclocarban in cosmetic products is 0.2% (EC, 1999). According to a German manufacturer, commercial triclocarban contains <100 mg PCA/kg.⁵ This would result in about 0.2 mg PCA/kg cosmetic product at a maximum. Assuming, for example, one application of a triclocarban-containing roll-on antiperspirant per day, a total amount of 0.5 g antiperspirant per application (SCCNFP, 1999), and a dermal absorption of PCA of 100%, this would result in a body dose of 1.6 ng/kg body weight per day at a maximum (average body weight 64 kg; see IPCS, 1994).

From the use of chlorohexidine-containing mouthwashes (see section 4), the following oral and dermal (via mucous membrane) exposure concentrations can be estimated. In the EU, the maximum permitted amount of chlorohexidine in cosmetic products is 0.3% (EC, 1999). According to a German manufacturer, chlorohexidine contains <500 mg PCA/kg,⁶ resulting in about 1.5 mg PCA/litre commercial chlorohexidine solution at a maximum. PCA concentrations between 0.5 and 2.4 mg/litre were detected in chlorohexidine preparations (chlorohexidine content 0.2%). Assuming two mouthwashes per day with 10 ml of this chlorohexidine solution each, the mucous membrane is exposed to between 10 and 48 µg PCA (Kohlbecker, 1989). About 30% of the chlorohexidine is retained in the oral cavity, and about 4% is swallowed (Bonesvoll et al., 1974). Therefore, uptake of PCA from mouthwash is from 50 to 255 ng/kg body weight (average body weight 64 kg, according to IPCS, 1994).

As no data are available on the use of 4-chlorophenol in cosmetics (see section 4), a quantitative exposure assessment concerning residual or degradation product PCA is not possible.

There is some evidence that PCA may be formed during the chlorination of drinking-water (Stiff & Wheatland, 1984). On the basis of the drinking-water concentrations measured in Germany in the 1980s and 1990s (see section 6.1), body doses of about 0.2–0.4 ng/kg body weight per day can be calculated (drinking-water consumed per day, 2 litres; average body weight, 64 kg; IPCS, 1994).

PCA is rapidly absorbed from the gastrointestinal tract. After nasogastric intubation of rhesus monkeys (20 mg [¹⁴C]PCA/kg body weight), the maximal ¹⁴C plasma concentration was reached within 0.5–1 h (Ehlhardt & Howbert, 1991).

Acute toxicity studies in rats show that PCA is well absorbed through the skin. LD₅₀ values (see section 8.1; BUA, 1995) as well as methaemoglobin levels (see Table 3; Scott & Eccleston, 1967) are similar for oral, dermal, and intraperitoneal administration. In rats, dermal uptake of PCA seems to be greater than uptake via inhalation exposure (see section 8.1; Kondrashov, 1969b). Dermal absorption also plays a predominant role for humans (Linch, 1974).

Table 3: Methaemoglobin formation after a single dose of PCA.^a

^a max = maximal reaction; n.g. = not given; f = female; m = male.

Percutaneous absorption of PCA was studied using in vivo microdialysis in female hairless mutant rats. Dialysates collected from the dermis and the jugular veins were analysed by HPLC, and PCA was found in both dialysates. The highest concentration was found 3 h after the topical application and gradually decreased with a half-life of about 20 h, showing that PCA was able to cross the skin (El Marbouh et al., 2000).

This confirmed results from in vitro studies with PCA using hairless rat skin, which showed that PCA was able to penetrate rat skin to a much greater extent than other aromatic amines tested (Levillain et al., 1998). Another in vitro study had previously shown the penetration of human skin by PCA (Marty & Wepierre, 1979).

After a single intravenous administration of 3 mg [¹⁴C]PCA/kg body weight to rats, most of the radioactivity was found within 15 min post-administration in the following tissues (% of dose): liver (8%), muscle (34%), fat (14%), skin (12%), blood (7%), and small intestine and kidney (each about 3%). These tissue levels decreased to less than 0.5% within 72 h.

Elimination from all tissues followed bi-exponential kinetics, with initial elimination half-lives between 1.5 and 4 h (Perry et al., 1981; NTP, 1989). The red blood cell to plasma ratio was 2:1 at 2 h, 20:1 at 12 h, and 74:1 after 2 days, indicating that PCA metabolites were rapidly bound to erythrocytes. After 7 days, radioactivity was found only in the erythrocytes (0.85–2.3% of dose).

After intraperitoneal application of 0.5 or 1.0 mmol [¹⁴C]PCA/kg body weight to male Fischer 344 rats, radioactivity was measured in blood, spleen, kidney, and liver. For the low-dose animals at 3 h post-application, the highest tissue concentrations were measured in liver and kidney medulla, followed by spleen (11.04, 9.05, and 4.19 µmol/g tissue, respectively), with 94% of the total dose located in liver. Doubling of the dose to 1.0 mol/kg body weight increased these tissue levels by 65, 83, and 50% to 18.19, 16.61, and 6.26 µmol/g tissue, whereas the percentage of the total dose located in these tissues decreased (e.g., 85% in liver) at 3 h post-application. However, 21 h later, the tissue distribution had changed: tissue concentrations were highest in kidney medulla, followed by spleen and then liver (25.55, 16.7, and 14.91 µmol/g tissue, respectively, corresponding to 3.16, 2.63, and 70% of the total dosage). In the kidney, the concentrations were lower in the cortex than in the medulla. The plasma concentration tended to increase with dose and post-application time, whereas the erythrocyte concentration showed minor changes. Subcellular distribution studies in the renal cortex demonstrated a preferential localization in the cytosolic compartment; in liver, distinct amounts were also found in the microsomal and nuclear fraction. Covalent binding of radioactivity to microsomal and cytosolic proteins of kidney and liver was shown, with little influence of time or dose, however (Dial et al., 1998).

PCA is rapidly metabolized. The main metabolic pathways of PCA are as follows (see Figure 2): a) C-hydroxylation in the ortho position to yield 2-amino-5-chlorophenol followed by sulfate conjugation to 2-amino-5-chlorophenyl sulfate, which is excreted per se or after N-acetylation to N-acetyl-2-amino-5-chlorophenyl sulfate; b) N-acetylation to 4-chloroacetanilide (found mainly in blood), which is further transformed to 4-chloroglycolanilide and then to 4-chlorooxanilic acid (found in the urine); or c) N-oxidation to 4-chlorophenylhydroxylamine and further to 4-chloronitrosobenzene (in erythrocytes). After a single intravenous administration of 3 mg [¹⁴C]PCA/kg body weight to rats, PCA in most tissues (except adipose tissue and small intestine) disappeared with bi-exponential decay kinetics, with initial half-lives of about 8 min and terminal half-lives of 3–4 h. However, the PCA concentration was higher at 1 h post-administration in blood, muscle, fat, and skin than at other time points. PCA is rapidly N-acetylated to 4-chloroacetanilide. The highest levels of this metabolite are found in muscle, skin, fat, liver, and blood, but it is not excreted in the urine. 4-Chloroacetanilide had an appearance half-life of approximately 10 min and an elimination half-life of 3 h (NTP, 1989).

Figure 2: Scheme of metabolic pathways for 4-chloroaniline (MAK, 1992)

In parallel studies in which ¹⁴C-labelled PCA at a concentration of 20 mg/kg body weight was administered to three male Fischer rats and six female C3H mice (dosed intragastrically) and two male rhesus monkeys (dosed by nasogastric intubation), 2-amino-5-chlorophenyl sulfate was by far the main excretion product (54, 49, and 36% for rat, mouse, and monkey, respectively) in the 24-h urine, followed by 4-chlorooxanilic acid (11, 6.6, and 1.0%). The parent compound accounted for 0.2, 1.7, and 2.5%, respectively, of the radioactivity in urine. Minor urinary metabolites were N-acetyl-2-amino-5-chlorophenyl sulfate (7.0, <0.1, and 2.0%) and 4-chloroglycolanilide (<1%), whereas 4-chloroacetanilide was not detected in urine. Unknown metabolites accounted for 22, 14, and 14%. Total radioactivity was 80, 88, and 56%, respectively, in 0- to 24-h urine and 4, 7, and 1% in the faeces. In 24- to 48-h urine, 2, 3, and 16% of the radioactivity were found for mouse, rat, and monkey. In the monkey, 6% and 5% of the radioactivity were detected in the urine between 48 and 72 h and between 72 and 96 h, respectively, whereas no more notable amounts of radioactivity were excreted by rat and mouse after 48 h. Therefore, the monkey retains PCA metabolites in the body much longer than the mouse and rat (Ehlhardt & Howbert, 1991).

In a separate study in the monkeys dosed as above (Ehlhardt & Howbert, 1991), the major circulating metabolite in plasma at 1 h post-administration was 2-amino-5-chlorophenyl sulfate (27% of plasma radiocarbon). 4-Chloroacetanilide, which was the major circulating metabolite in the rat (Perry et al., 1981; NTP, 1989), appeared more slowly in monkey plasma, accounting for 26% of the circulating radiolabel at 1 h post-administration, but was the major plasma metabolite (>90%) at 24 h (Ehlhardt & Howbert, 1991).

A study on the disposition of ¹⁴C-labelled PCA or its hydrochloride in F344 rats, mongrel dogs, and A/J and Swiss Webster mice (route not given) showed that the initial decay constants for PCA clearance from whole blood in both strains of mice were 10 times greater than those in dogs and rats. The PCA clearance in mice was too rapid to permit calculation of kinetic parameters (NTP, 1989).

Early studies in rabbits given a single oral dose of 100 mg/kg body weight reported the presence of 2-amino-5-chlorophenol in urine (Bray et al., 1956). After a single intraperitoneal administration of 50 mg PCA/kg body weight to rabbits, the metabolites 4-chloroglycolanilide and 4-chlorooxanilic acid (only analysed metabolites) were quantified in equal amounts of 3% of the dosage in the 24-h urine. These metabolites are the excretion products of the primary metabolite 4-chloroacetanilide, as had been demonstrated by direct administration of this compound (Kiese & Lenk, 1971). In similar experiments with pigs (intraperitoneal injection of 20–50 mg PCA/kg body weight), 4-chlorooxanilic acid was not detectable in the urine of pigs (Kiese & Lenk, 1971).

From a case of acute PCA poisoning in humans (no details of exposure/dose), PCA (0.5% free, 62% total), 2-amino-5-chlorophenol (36%), and 2,4-dichloroaniline (1.7%; not reported in other studies), all in free and conjugated form, were detected (using HPLC) as excretory products in the urine (Yoshida et al., 1991). The biphasic elimination of the metabolites 2-amino-5-chlorophenol and 2,4-dichloroaniline was faster (half-lives of 1.7 h for both metabolites in the first phase [T1] and 3.3 and 3.8 h for the two metabolites, respectively, in the second phase [T2]) than that of PCA (all forms: half-lives T1 2.4 h, T2 4.5 h). PCA and 2-amino-5-chlorophenol were still detectable in the urine on days 3 and 4 (Yoshida et al., 1992a,b).

4-Chloronitrosobenzene (plus 4-chlorophenylhydroxylamine) was detected in the blood of dogs after a single intravenous injection of 25 or 100 mg PCA (Kiese, 1963). The relationships between rate of haemoglobin formation, the concentration of 4-chloronitrosobenzene, and time after injection were similar for both doses.

PCA undergoes N-oxidation to 4-chlorophenylhydroxylamine and 4-chloronitrosobenzene in rat liver microsomal preparations (Ping Pan et al., 1979). Further in vitro studies demonstrated that besides the microsomal monooxygenase, haemoglobin, prostaglandin synthetase, and products of lipid peroxidation can also be involved in this reaction (BUA, 1995). Of the possible metabolic reactions of PCA that could be catalysed by the cytochrome P450-dependent monooxygenase system, C2- and N-hydroxylation with 2-amino-5-chlorophenol and 4-chlorophenylhydroxylamine as products are favoured in vitro (values for apparent V_max 0.54, 2.93, and 4.35 nmol product/min per nanomole cytochrome P450, respectively). Dechlorination followed by C-hydroxylation to 4-aminophenol plays a minor role (Cnubben et al., 1995).

It has been shown for aniline that the N-oxidation pathway is inconsequential in rat liver, as liver rapidly reduces N-oxidized metabolites back to the parent compound. In the erythrocytes, however, N-phenylhydroxylamine is rapidly oxidized by oxyhaemoglobin to nitrosobenzene, with concurrent formation of methaemoglobin (Bus & Popp, 1987). The mechanism and pattern of erythrocyte toxicity of PCA, an analogue of aniline, could be similar (NTP, 1989). It should be remembered that radioactivity was rapidly bound to erythrocytes of rats (Perry et al., 1981).

The methaemoglobin can be reduced to haemoglobin in mammalian species by an NADH-dependent methaemoglobin diaphorase located in the erythrocytes. Enzymatic activity in rat and mouse erythrocytes is 5 and 10 times higher, respectively, than that in human erythrocytes (Smith, 1986), suggesting that humans are more susceptible to this toxic effect.

Besides covalent binding of PCA to proteins of kidney and liver (see section 7.2; Dial et al., 1998), the formation of haemoglobin adducts has been investigated in both rats and exposed humans.

Among 12 chloro- or methyl-substituted anilines, PCA had the strongest potential to bind covalently to haemoglobin. The haemoglobin binding index was 569 in female Wistar rats and 132 in female B6C3F₁ mice (dosages: 0.6 and 1 mmol/kg body weight, respectively, by gavage), compared with values of 22 and 2.2 for aniline (dosages: 0.47 and 2 mmol/kg body weight, respectively) (Birner & Neumann, 1987, 1988). The active metabolite responsible for covalent binding is 4-chloronitrosobenzene, which forms a hydrolysable sulfinic acid amide adduct (93% of total haemoglobin adducts). It is predominantly formed in the erythrocytes, as the ratio of the covalent binding index to haemoglobin and plasma proteins is 29.3 (Neumann et al., 1993).

Covalent binding of PCA to haemoglobin was detected in humans with accidental exposure to PCA and aniline (no information on exposure conditions or dose) as early as 30 min after exposure. Maximum haemoglobin adduct levels of PCA were detected 3 h after the accident, which also correlated with the time course of methaemoglobinaemia, whereas the maximum was delayed to 16 h for aniline adducts. For both substances, haemoglobin adducts were detectable up to 7 days post-administration and fell below the detection limit of 10 µg/litre within 12 days (Lewalter & Korallus, 1985). Biomonitoring of workers employed in the synthesis and processing of aniline and PCA (no information on dose or exposure conditions, dermal absorption assumed to prevail) and grouped for smoking habits and acetylator status showed that haemoglobin adduct levels of PCA (22–26 workers) were mostly higher than those of aniline (45 workers). For PCA, haemoglobin adduct levels were not significantly different for smokers (mean 975 ng/litre; range 500–1700 ng/litre) and non-smokers (mean 1340 ng/litre; range 500–2500 ng/litre). However, slow acetylators showed a significantly increased adduct level compared with fast acetylators for the whole group (1443 vs. 663 ng/litre; P = 0.0001) as well as in the subgroup of smokers (1575 vs. 725 ng/litre; P = 0.0052). There was no correlation between haemoglobin adduct level and urinary excretion (see section 7.5) of PCA (Riffelmann et al., 1995).

Excretion of PCA takes place primarily via the urine. Within 24 h, the urinary excretion of rats and mice (intragastric application) and monkeys (nasogastric intubation) of a 20 mg [¹⁴C]PCA/kg body weight dosage accounted for 93, 84, and 50–60% of the radiocarbon, respectively, and faecal excretion accounted for 6.9, 4.5, and 0.5–1.0%, respectively. Elimination was complete in the rat (98%) within 48 h and in the mouse (89%; 91% recovered in total) within 72 h, whereas the monkey still excreted considerable amounts, 3.1–5.8%, from 72 to 96 h post-administration (Ehlhardt & Howbert, 1991).

After oral application (gavage) of doses of 0.3, 3, or 30 mg [¹⁴C]PCA/kg body weight to rats, 77% of the radiocarbon was excreted in the urine and 10% in the faeces within 24 h, independent of dose. Within 72 h, excretion was almost complete. Consequently, doses up to 30 mg/kg body weight apparently did not saturate the metabolic and excretory pathways (Perry et al., 1981; NTP, 1989).

After intraperitoneal application of 0.5 or 1.0 mmol [¹⁴C]PCA/kg body weight to male Fischer 344 rats, excretion was primarily via the urine (5.2 and 1.2% of dose, respectively, within 3 h) compared with the faeces (0.01% for both doses). Within 24 h, urinary excretion rose to 30% of the injected dose. It was argued that excretion could have been retarded relative to oral administration because the parent compound enters the general circulation first before it is metabolized in the liver (Dial et al., 1998).

In humans with accidental exposure to PCA and aniline (no information on exposure conditions or dose), urinary elimination of PCA and aniline in free and conjugated form reached its maximum as early as 30 min after exposure. Detection was possible up to 16 h post-administration (50 µg PCA/g creatinine, 100 µg aniline/g creatinine), but not after 3 days (<10 µg/g) (Lewalter & Korallus, 1985).

Among 22–26 workers employed in the synthesis and processing of aniline and PCA (no information on dose or exposure conditions, dermal absorption assumed to prevail), urinary excretion of PCA (free and conjugated) was similar for smokers and non-smokers, whereas it tended to be higher for slow acetylators than for fast acetylators (no significant increase). There was no correlation between haemoglobin adduct level (see section 7.4) and urinary excretion (Riffelmann et al., 1995).

8.1.1 Inhalation

The LC₅₀ for rats (4-h inhalation, head-only exposure) was calculated as 2340 mg PCA/m³ (vapour/aerosol mixture: respirable fraction 57–95%) (for further details of the study, see Table 4) (BUA, 1995).

Table 4: Acute toxicity of PCA (for methaemoglobin induction, see Table 3).

Mice, cats, and albino rats were exposed by inhalation for 4 h to PCA, and an increase of Heinz bodies in the erythrocytes was observed. The lowest doses at which an increase was seen were 22.5, 21.4, and 36 mg/m³, respectively (Kondrashov, 1969a). In a further inhalation study in albino rats, which applied an experimental device allowing exposure of either the head or the back half of the body (shaved skin) to a PCA-containing atmosphere, the lowest doses at which an increase of Heinz bodies in the erythrocytes was observed were 22 mg/m³ for "body-only" and 36 mg/m³ for head-only exposure, showing that dermal absorption of PCA contributes to the body burden to a greater extent than absorption by the lung (Kondrashov, 1969b).

8.1.2 Oral, intraperitoneal, and dermal application

Compilations of LD₅₀ values are given in BUA (1995) and NTP (1998). Oral LD₅₀ values of 300–420 mg/kg body weight for rats, 228–500 mg/kg body weight for mice, and 350 mg/kg body weight for guinea-pigs are reported. Similar values have been found for intraperitoneal and dermal application of PCA to rats, rabbits, and cats. Signs of intoxication included excitation, tremors, spasms, and shortness of breath (BUA, 1995). Cyanosis, methaemoglobinaemia, and mild hepatotoxic and nephrotoxic changes were reported after acute exposure to PCA (see Tables 3 and 4).

PCA is a more potent methaemoglobin inducer than aniline, as demonstrated in mice, rats, and cats (compilation of studies in Table 3). Very high methaemoglobin formation (60–70% methaemoglobin) occurs as early as 10 min after intraperitoneal application of 64 mg/kg body weight to mice (Nomura, 1975). After oral and dermal administration to rats and after dosing of pregnant rats, similar methaemoglobin levels of 3–15% after a 13–40 mg/kg body weight dose were induced. Rats and monkeys were of comparable sensitivity, whereas dogs were much more sensitive (similar methaemoglobin levels with 16% of the rat dose). Also, in an in vitro assay with whole blood, the dog was the most sensitive species (dog >> human and monkey > rat) (Scott & Eccleston, 1967). For comparison, significant increases of methaemoglobin have been reported for medium-term exposure by gavage, with LOAELs of 5 mg/kg body weight in rats and 7.5 mg/kg body weight in mice (lowest investigated doses) (see Table 5; NTP, 1989).

After intraperitoneal application of PCA to rats in dosages of 128–191 mg/kg body weight, a nephrotoxic and hepatotoxic potential was derived from changes in urine volume and urine and blood chemistry and from slight morphological alterations. However, these dosages lie in the range of 50% of the LD₅₀ and were reported to cause a severe depression of water and food intake (for details, see Table 4).

Table 5: Toxicity studies with repeated oral administration of PCA.^a

In OECD guideline studies, PCA was found to be non-irritating to rabbit skin and slightly irritating to rabbit eyes (grade 1–2 effects). Older studies reported an absence of irritant effects on rat skin in contrast to inflammatory skin reactions of rabbits and cats, whereas effects on mucous membranes in these studies were characterized as slight to severe (limited validity of data because of insufficient documentation) (for details, see BUA, 1995).

In guinea-pigs, PCA was tested by three different testing procedures with the following classifications: moderate sensitizer in the maximization test (50% positive response), very weak sensitizer in the single injection adjuvant test (30% positive response), and no sensitizing potential in a modified Draize test (0% positive response) (Goodwin et al., 1981). Another group compared the sensitizing potential of PCA in the guinea-pig maximization test according to Magnusson & Kligman (1969) and the local lymph node assay (Kimber et al., 1986). In the guinea-pig maximization test (vehicle ethanol, intradermal induction with 0.3% PCA, topical induction with 10.0% PCA, and challenge with 2.5%), 50–60% of the animals responded with a positive reaction, so that PCA was classified as moderately sensitizing. Test results of the local lymph node assay from four independent laboratories (test concentrations 2.5, 5.0, and 10.0% in acetone–olive oil 4:1) pointed to a sensitizing potential, as test results in one laboratory were slightly positive and were confirmed by concentration-dependent increases (suggestive of sensitization) in two other laboratories. It was speculated that positive responses would have been obtained if testing with higher concentrations had been possible, but the high toxicity of PCA was limiting (Basketter & Scholes, 1992; Scholes et al., 1992). A weak sensitizing potential was reported in a further maximization test (BUA, 1995). From these data, PCA can be considered to be a skin sensitizer.

A 2-week inhalation exposure of rats from the lowest concentration of