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 Mr P.D. Howe, Centre for Ecology & Hydrology, Monks Wood, United Kingdom, and Dr C. Melber, Dr J. Kielhorn, and Dr I. Mangelsdorf, Fraunhofer Institute for Toxicology and Aerosol Research, 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, 2001

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.

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WHO Library Cataloguing-in-Publication Data

Chlorinated naphthalenes.

(Concise international chemical assessment document ; 34)

1. Naphthalenes – toxicity

2. Risk assessment

3. Environmental exposure

4. Occupational exposure

I. International Programme on Chemical Safety

II. Series

ISBN 92 4 153034 0

(NLM Classification: QV 241)

ISSN 1020-6167

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This CICAD on chlorinated naphthalenes was prepared by the Centre for Ecology & Hydrology, Monks Wood, United Kingdom, and the Fraunhofer Institute for Toxicology and Aerosol Research, Hanover, Germany. It is based on the United Kingdom’s Environ mental hazard assessment: Halogenated naphthalenes (Crookes & Howe, 1993) and the work of the German Commission for the Investigation of Health Hazards of Chemical Compounds in the Work Area (Greim, 1997), supplemented by a literature search (June 2000). Information on the nature of the peer review and availability 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 Geneva, Switzerland, on 8–12 January 2001. Participants at the Final Review Board meeting are listed in Appendix 3. The International Chemical Safety Cards for trichloronaphthalene (ICSC 0962), tetrachloronaphthalene (ICSC 1387), pentachloronaphthalene (ICSC 0935), hexachloronaphthalene (ICSC 0997), and octachloronaphthalene (ICSC 1059), produced by the International Programme on Chemical Safety (IPCS, 1993a,b,c, 1999a,b), have also been reproduced in this document.

There are 75 possible congeners of chlorinated naphthalenes. Commercial products are generally mixtures of several congeners and range from thin liquids to hard waxes to high melting point solids. Their main uses have been in cable insulation, wood preservation, engine oil additives, electroplating masking compounds, capacitors, and refractive index testing oils and as a feedstock for dye production.

The major sources of release of chlorinated naphthalenes into the environment are likely to be from waste incineration and disposal of items containing chlorinated naphthalenes to landfill.

Chlorinated naphthalenes are expected to adsorb onto soil and sediments to a large extent. Predicted soil organic carbon/water partition coefficients show an increase as the degree of chlorination in the chlorinated naphthalene increases. Thus, the lower chlorinated congeners are likely to show a moderate sorption tendency, and the higher chlorinated congeners are likely to show a strong sorption tendency.

Chlorinated naphthalenes have been shown to be highly bioaccumulative in fish, but less so in shrimp and algae. The amount of bioaccumulation observed increases with the degree of chlorination of the chlorinated naphthalenes, but the most highly chlorinated naphthalenes (e.g., octachloronaphthalene) do not appear to bioaccumulate due to their very limited absorption.

Monochloronaphthalenes appear to be readily degradable by soil and water microorganisms under aerobic conditions. No information was found on the biodegradation of higher chlorinated congeners by microorganisms.

One report gave an atmospheric half-life of 2.7 days for 1,4-dichloronaphthalene. No other information was found regarding the atmospheric fate of other chlorinated naphthalenes. As all chlorinated naphthalenes absorb light at environmentally relevant wavelengths, direct photolysis reactions may occur in water, in air, or on soil.

In the past, chlorinated naphthalene concentrations of up to 14.5 mg/m³ have been measured in the work place, while levels of 25–2900 ng/m³ have been recorded in outdoor air in the vicinity of manufacturing sites. More recently, monitoring studies have revealed chlorinated naphthalene concentrations of up to 150 pg/m³ at "semirural" sites and 1–40 pg/m³ at remote sites. Predominant congeners in outdoor air were tri- and tetrachloronaphthalenes. In the 1970s, surface water concentrations of up to 5.5 µg/litre were measured near chlorinated naphthalene manufacturing plants, with higher levels recorded in groundwater. Recent studies have found surface water levels in the low ng/litre range. A single study on chlorinated tap water revealed chlorinated naphthalene concentrations of up to 0.15 ng dichloronaphthalene/litre and up to 0.44 ng monochloronaphthalene/litre. Sediment levels of up to 100 mg/kg have been recorded in the past; however, recent results show levels of 0.2 µg/kg at unpolluted sites and 250 µg/kg at polluted sites. Similarly, soil levels of up to 1300 mg/kg were measured at contaminated sites in the early 1980s compared with a more recent value for a former chlor-alkali plant of 18 mg/kg dry weight. Chlorinated naphthalene concentrations in fish range up to a maximum of around 300 µg/kg lipid weight. Tetra- and pentachloronaphthalene congeners tend to predominate in biota. Monitoring studies with seabird eggs have revealed a decrease in chlorinated naphthalene levels between 1974 and 1987.

Chlorinated naphthalenes, especially dioxin-like congeners, have been detected in adipose tissue, liver, blood, and breast milk samples of the general population at concentrations in the ng/kg lipid range. The chlorinated naphthalene congener/isomer pattern found in human specimens was significantly different from that in commercial chlorinated naphthalene mixtures. The dominating congeners in almost all human specimens examined were two penta- and two hexa-isomers, namely 1,2,3,5,7/1,2,4,6,7-pentachloronaphthalene and 1,2,3,4,6,7/1,2,3,5,6,7-hexachloronaphthalene, and to a lesser extent some tetra-isomers.

Chlorinated naphthalenes can be absorbed via oral, inhalative, and dermal routes, with absorption and dis tribution over the whole body after oral administration. The main target organs are liver and fat tissue (besides kidney and lung), both showing a high retention, especially for higher chlorinated congeners such as 1,2,3,4,6,7/1,2,3,5,6,7-hexachloronaphthalene. Half-lives of 1,2,3,4,6,7/1,2,3,5,6,7-hexachloronaphthalene were calculated to be 41 days in adipose tissue and 26 days in the liver of rats. Calculations with monitoring data from human blood samples suggested half-lives of 1.5– 2.4 years for these hexa-isomers in humans. Hydroxy metabolites have been identified mostly for the lower chlorinated naphthalenes (mono- to tetra-) in experimental animals. There are also preliminary indications for the occurrence of methylthio- or methylsulfoxide chloronaphthalene metabolites in faeces of rats. Elimination of parent compounds and/or metabolites occurs via faeces and urine. There was also a transfer of 1,2,3,4,6,7-hexachloronaphthalene to offspring of rats via placental and lactational routes.

LD₅₀ values of some chlorinated naphthalenes ranged from >3 (2,3,6,7-tetrachloronaphthalene) to 1540(1-monochloronaphthalene) mg/kg body weight. Short-term exposure to higher chlorinated naphthalenes resulted in mortality, liver damage, degeneration of kidneys, etc., in rats, rabbits, and cattle. Cattle developed severe systemic disease (bovine hyperkeratosis) during a 5- to 10-day oral exposure to 1.7–2.4 mg/kg body weight per day of penta-, hexa-, hepta-, or octachlorinated naphthalenes. Similar symptoms (death, severe weight loss, and liver damage) have also been observed during medium-term oral or inhalative exposures of laboratory and domestic animals. The higher chlorinated congeners appeared to be more toxic than the lower chlorinated ones. Inhalation of 1.4 mg/m³ (8 h/day) of a penta/hexachlorinated naphthalene mixture for 143 days resulted in slight to moderate histological liver damage in rats.

Long-term and carcinogenicity studies with chlorinated naphthalenes have not been performed.

The few chlorinated naphthalenes tested for mutagenicity — 1-monochloronaphthalene and 1,2,3,4-tetrachloronaphthalene — have proved to be not mutagenic in the Salmonella Ames test.

1,2,3,4,6,7-Hexachloronaphthalene has been found to accelerate the onset of spermatogenesis in male offspring of rats when given to dams at 1 µg/kg body weight per day on days 14–16 of gestation.

Like related compounds, chlorinated naphthalenes have been demonstrated to be inducers of the cytochrome P-450 (CYP)–dependent microsomal enzymes. Two very persistent (and frequently identified in human and environmental samples) hexachlorinated naphthalene isomers (i.e., 1,2,3,4,6,7/1,2,3,5,6,7-hexachloronaphthalene) caused induction of CYP1A1 — typical for dioxin-like compounds — in several in vitro and in vivo test systems. Chlorinated naphthalenes were also found to change lipid peroxidation and antioxidant enzyme activities in rats in a manner indicative of increased oxidative stress. At least some of the biological and toxic responses of chlorinated naphthalenes are believed to be mediated via the cytosolic Ah receptor, resembling those of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related compounds.

All chlorinated naphthalenes tested caused skin irritations, and the penta- and hexachlorinated naphthalenes showed hyperkeratotic activity in the rabbit ear test and in hairless mice, consistent with findings in cattle (bovine hyperkeratosis or X-disease) and humans (chloracne).

Severe skin reactions (chloracne) and liver disease have both been reported after occupational exposure to chlorinated naphthalenes. Chloracne was common among workers handling chlorinated naphthalenes in the 1930s and 1940s.

Other symptoms described in workers exposed to chlorinated naphthalenes included irritation of the eyes, fatigue, headache, anaemia, haematuria, impotentia, anorexia, nausea, vomiting, and occasionally severe abdominal pain. At least 10 deaths were reported from acute atrophy of the liver. Systemic effects resulting in liver disease have been reported only from the inhalation of chloronaphthalenes.

After dermal application of various Halowax samples to adult subjects, only Halowax 1014, containing penta- and hexachloronaphthalenes, produced chloracne; Halowaxes containing mono-, di-, tri-, tetra-, hepta-, and/or octachloronaphthalenes did not.

A cohort mortality study on workers exposed to chlorinated naphthalenes at a cable manufacturing plant found an excess of deaths from cirrhosis of the liver. However, individuals who had shown symptoms of chloracne did not show a higher mortality due to liver cirrhosis compared with other workers. The mortality from all cancers was slightly but significantly elevated among all exposed men (standardized mortality ratio = 1.18) but was not more elevated in the subcohort with chloracne. This subcohort showed statistically significant excess mortality from cancer of the oesophagus and from "benign and unspecified neoplasms."

There are only a few miscellaneous reports on the effects of incidental exposure to chlorinated naphthalenes on the general population. With one exception, they involve ingestion of oil contaminated with other chemicals as well as chlorinated naphthalenes, resulting in general systemic symptoms followed by chloracne.

Chlorinated naphthalenes appear to be of moderate to high acute toxicity to aquatic organisms.

Chlorinated naphthalenes are a group of compounds based on the naphthalene ring system, but where one or more hydrogen atoms have been replaced by chlorine. The generic molecular formula is C₁₀H_8ånCl_n, where n = 1–8. There are 75 possible chlorinated naphthalenes, and they are usually identified using the numbering system shown below:

Chlorinated naphthalenes are often called polychlorinated naphthalenes, or PCNs.

Most of the industrially produced PCNs are not pure materials, but are usually a mixture of several congeners. The commercial products range from thin liquids to hard waxes to high melting point solids, with melting points ranging from -40 to 180 °C. Liquid PCNs are soluble in most organic solvents, whereas the waxy or solid PCNs are soluble in chlorinated solvents, aromatic solvents, and petroleum naphthas and can be mixed with petroleum waxes, chlorinated paraffins, polyisobutylates, and plasticizers. PCNs have low flammability and are of medium to low volatility, volatility decreasing with increasing chlorination.

Some relevant physical and chemical properties of PCNs and commercial PCNs are listed in Tables 1 and 2, respectively. Additional physical/chemical properties are presented in the International Chemical Safety Cards reproduced in this document.

Highly sensitive and specific analytical techniques are necessary for the measurement of PCNs because of their complexity. Another analytical complication is the co-occurrence of bulk quantities of polychlorinated biphenyls (PCBs) or organochlorine pesticides in environmental matrices when typical gas chromatography– electron capture detector methods are used (Falandysz, 1998).

Various methods have been used to overcome this difficulty, typically to perchlorinate the PCNs and PCBs to give octachloronaphthalene and decachlorobiphenyl, respectively, or to hydrodechlorinate the PCNs and PCBs to give naphthalene and biphenyl, respectively. The products from these reactions can be more easily quantified by gas chromatographic techniques, but the methods suffer from interference from naphthalene already present in the samples; also, a lot of information about the individual PCNs originally present in the samples is lost. Recent advances in the analysis of PCNs include the use of mass spectrometric detection, which allows individual compounds in complex analyses to be identified and quantified, but this is possible only if authentic standard material is available for all the possible PCNs (Crookes & Howe, 1993).

All 75 PCN congeners have been synthesized, although sometimes only in mixtures (Nikiforov et al., 1992, 1993; Auger et al., 1993; Imagawa et al., 1993; Imagawa & Yamashita, 1994, 1997; Takasuga et al., 1994). Quantification of all PCN congeners with gas chromatography–mass spectrometry using the molar response of electron impact ionization and based on one or two reference compounds is possible (Falandysz, 1998).

For congener-specific determination of PCNs in environmental matrices, a contaminant enrichment procedure using an activated carbon column is required, coupled with final separation and quantification using high-resolution capillary gas chromatography and electron capture negative ionization mass spectrometry (Järnberg et al., 1993, 1997; Haglund et al., 1995; Schlabach et al., 1995; Falandysz & Rappe, 1996, 1997; Falandysz et al., 1996b).

Falandysz (1998) states that a further problem in the analysis of samples for PCNs is the co-elution of some PCN congeners when using a single column in capillary gas chromatographic separation. However, advances in the analytical separation of PCN congeners are being identified. For example, Helm et al. (1999) reported the complete resolution of all 14 pentachloronaphthalenes and all 10 hexachloronaphthalenes. No individual published method reports the ability to discriminate all individual congeners; for this reason, comparison between results from different research groups is difficult (see later sections).

a * indicates estimated value.

b Schoene et al. (1984).

c Mackay et al. (1982).

d Budavari et al. (1996).

e Harner & Bidleman (1998).

f Burreau et al. (1997).

g Opperhuizen et al. (1985).

Table 2: Physical/chemical properties of commercial chlorinated naphthalenes.

Chlorinated naphthalene	CAS No.	Chlorine content (%)	Chlorinated naphthalene composition (% weight)	Boiling point (°C)	Melting point (°C)	Vapour pressure (kPa)	Aqueous solubility	Henry’s law constant (Pabetam3/ mol)
Halowaxes
Halowax 1031	25586-43-0	22	95% mono-, 5% di-	250 a	-25	1.9 × 10–3 b	Insoluble a	31.9
Halowax 1000	58718-66-4	26	60% mono-, 40% di-	250 a	-33		Insoluble a
Halowax 1001	58718-67-5	50	10% di-, 40% tri-, 40% tetra-, 10% penta-	308 a	98		Insoluble a
Halowax 1099	39450-05-0	52	10% di-, 40% tri-, 40% tetra-, 10% penta-	315 a	102		Insoluble a
Halowax 1013	12616-35-2	56	10% tri-, 50% tetra-, 40% penta-	328 a	120		Insoluble a
Halowax 1014	12616-36-3	62	20% tetra-, 40% penta-, 40% hexa-	344 a	137		Insoluble a
Halowax 1051		70	10% hepta-, 90% octa-		185
Nibren waxes
D88					90
D116N					113
D130					135
Seekay waxesc
68 (R Grade)		46.5
93 (R Grade)		50
123 (R Grade)		56.5
700 (R Grade)		43
93 (RC Grade)		50
123 (RC Grade)		56.5
Clonacire waxes
90					90
115					115
130					130

a Brinkman & De Kok (1980).

b Estimated value.

c R Grade = refined or white wax; RC Grade = electrical grade.

Detection limits of 0.1 ng/g for fly ash, 1 ng/g dry weight for sediment, 0.2 pg/g wet weight for biota, and 0.01 ng/g fat for adipose tissue have been reported (Wiedmann & Ballschmiter, 1993; Williams et al., 1993; Kannan et al., 2000a,b).

4. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

No information was found relating to possible natural sources of PCNs.

Past sources of release such as PCN manufacturing sites (US EPA, 1977; Erickson et al., 1978a,b) and sites using PCN pesticides (Kauppinen, 1986) have been identified.

Major current sources of release of PCNs are likely to be emissions from both municipal and special waste incinerators (Ross & Whitmore, 1984; Tong et al., 1984; Rubey et al., 1985; Oehme et al., 1987; Janssens & Schepens, 1988; Alarie et al., 1989; Benfenati et al., 1991; Schneider et al., 1998; Abad et al., 1999) and disposal of items containing PCNs to landfill (De Kok et al., 1983; Weistrand et al., 1992; Espadaler et al., 1997; Martí & Ventura, 1997).

PCNs have been detected in water and sediments receiving both industrial and municipal sewage dis charges (Kuehl et al., 1984b; Vogelgesang, 1986; Furlong et al., 1988) or via leaching from hazardous waste sites (Elder et al., 1981; Kaminsky et al., 1983; Jaffe & Hites, 1984). A characteristic profile of PCNs in soil, sediment, and biota samples collected near a chlor-alkali plant suggests the formation of PCN congeners during the chlor-alkali process (Järnberg et al., 1997; Kannan et al., 1998).

PCNs have been shown to be formed following the use of chlorine to treat drinking-water supplies (Lin et al., 1984; Shiraishi et al., 1985).

It has been estimated that a world total of about 9000 tonnes of PCNs was produced annually in the 1920s. Between the 1930s and 1950s, PCNs were used extensively in the manufacture of electrical insulation; in 1956, it was estimated that approximately 3200 tonnes of PCNs were produced in the USA. By 1978, production in the USA had fallen to about 320 tonnes/year due to the replacement of PCNs by a variety of substitutes. Production of PCNs by Koppers Company, Inc. (manufacturers of Halowaxes) ceased in the USA in 1977 (Kirk-Othmer, 1980), and the last US producer of PCNs (Chemisphere) had stopped manufacture by 1980. Small amounts of PCNs were still being imported into the USA in 1981, around 15 tonnes/year, mainly for use in refractive index testing oils and capacitor dielectrics (US EPA, 1983).

There are no known commercial uses for purified individual isomers of di-, tri-, tetra-, penta-, hexa-, or heptachloronaphthalene. Monochloronaphthalenes and mixtures of mono- and dichloronaphthalenes have been used for chemical-resistant gauge fluids and instrument seals, as heat exchange fluids, as high boiling speciality solvents, for colour dispersions, as engine crankcase additives, and as ingredients in motor tune-up compounds. Monochloronaphthalenes have also been used as a raw material for dyes and as a wood preservative with fungicidal and insecticidal properties (Crookes & Howe, 1993).

The tri- and higher chlorinated naphthalene prod ucts have been used as impregnants for condensers and capacitors and dipping encapsulating compounds in electronic and automotive applications, as temporary binders in the manufacture of ceramic components in paper coating and impregnation, in precision casting of alloys, in electroplating stop-off compounds, as additives in gear oils and cutting compounds, in flame-proofing and insulation of electrical cable and conductors, as moisture-proof sealants, as separators in batteries, in refractive index testing oils, as masking compounds in electroplating, and in grinding wheel lubricants (Kirk-Othmer, 1980; US EPA, 1983).

The most important uses, in terms of volume, have been in cable insulation, wood preservation, engine oil additives, electroplating masking compounds, feedstocks for dye production, dye carriers, capacitors, and refrac tive index testing oils (US EPA, 1983). The use of PCNs as wood preservatives was popular in the 1940s and 1950s, but they are no longer used for this purpose in the USA (US EPA, 1975).

The US Environmental Protection Agency stated that in the USA, only very small amounts of PCNs (about 15 tonnes/year in 1981) were still being used, mainly as refractive index testing oils and as capacitor dielectrics. It did note that the most likely possible new uses for PCNs would be as intermediates for polymers and as flame retardants in plastics (US EPA, 1983).

Popp et al. (1997) reported that PCNs were used in a German plant producing models and tools for car manufacturing and mining until 1989. The production of waxes containing PCNs ceased in the mid-1980s.

5. ENVIRONMENTAL TRANSPORT, DISTRIBUTION, AND TRANSFORMATION

Most of the possible sources of PCN release to the environment are likely to result in emissions to air (possibly adsorbed onto particulate matter), water, and soil. There is some evidence that the emissions of PCNs from incineration processes are associated with particulate matter. This means that there is the possibility that PCNs adsorbed to particulate matter may be removed from the atmosphere by rain. PCNs have moderate to low vapour pressures. The vapour pressure is estimated to decrease as the degree of chlorination increases. This means that volatilization of the more highly chlorinated naphthalenes from water and soil is likely to be small, but volatilization may be important for the less highly chlorinated congeners (Crookes & Howe, 1993). Atmospheric PCN concentrations are controlled by air–surface exchange and advection even during periods of stable conditions and high temperatures, suggesting that there are ongoing emissions affecting ambient concentrations (Lee et al., 2000). There is evidence for the long-range transport and atmospheric stability of PCNs based on the fact that they have been reported in remote areas such as the Arctic (Harner et al., 1998).

PCN fluxes were measured in a dated core from freshwater sediment in the United Kingdom. The vertical profile showed that flux remained fairly constant at 0.4–0.6 µg/m² per year until the early 1940s, rising sharply to a subsurface maximum of around 12 µg/m² per year in the late 1950s to mid-1960s followed by a 4-fold decrease to the sediment–water interface. There was no significant difference in homologue profiles with time (Gevao et al., 2000).

The high octanol/water partition coefficients measured for PCNs indicate that adsorption onto soil or sediment may be significant. Soil organic carbon/water partition coefficients have been estimated for several PCNs using molecular connectivity regression equations ranging from 2.97 for monochloronaphthalene to 5.38 for octachloronaphthalene (Koch & Nagel, 1988). The estimated partition coefficients increase as the degree of chlorination increases, indicating that the lower chlorinated naphthalenes are likely to show moderate sorption tendency from water onto soil and sediments and that the higher chlorinated naphthalenes are likely to show a high sorption tendency from water onto soil and sediment.

Photolysis of PCNs has been carried out in methanol solution at 30 °C in the presence of atmospheric amounts of oxygen. A light source with peak energy output at 300 nm and an ultraviolet cut-off at 285 nm was used. Dechlorination and dimerization were the major reaction pathways observed, with traces of methoxylated naphthalenes being formed by reaction with the solvent. PCNs with vicinal or peri-substituted chlorine atoms gave mostly dechlorinated products, while the more unhindered PCNs gave mostly dimer products. The reaction was found to be slower for the more highly chlorinated naphthalenes due to increased stabilization of the radical intermediate with increased chlorination (Ruzo et al., 1975a). Similar reactions may occur in the environment, although the intensity of natural light is likely to be lower than that used in this experiment. More recently, in experiments to test the feasibility of using solar photons for the destruction of waste, photothermal oxidative destruction of chloronaphthalene was demonstrated using ultraviolet photons (Nimlos et al., 1994). Järnberg et al. (1999) found a general shift towards lower chlorinated congeners during exposure of Halowax 1014 in methanol to sunlight.

The reaction of 1,4-dichloronaphthalene with hydroxyl radicals has been studied in smog chamber experiments. At 300 °K, the rate constant for the reaction was found to be 6 × 10^–12 cm³/molecule per second, using nitrous acid photolysis as a source of hydroxyl radicals. Assuming a typical atmospheric hydroxyl radical concentration of 5 × 10⁵ molecules/cm³, this corresponds to an atmospheric half-life of 2.7 days (Klöpffer et al., 1988).

Little information appears to be available concerning the biodegradation of PCNs. Walker & Wiltshire (1955) found that two species of bacteria isolated from soil were capable of using 1-chloronaphthalene as the sole carbon source. Morris & Barnsley (1982) showed that both 1- and 2-chloronaphthalene were metabolized by pseudomonads grown on naphthalene as the sole source of carbon and energy. Using a sewage sludge inoculum grown on naphthalene, it was shown that 1- and 2-chloronaphthalene were both degraded on incubation with the inoculum (Okey & Bogan, 1965). Bacteria of the genera Pseudomonas, Alcaligenes, and Moraxella from the River Rhine have been shown to metabolize 2-chloronaphthalene in the presence of 1,2-dichlorobenzene or 4-chlorophenol. Two metabolites were identified, a hydroxy compound and 1-oxy-3-carboxymethyl-5(6)-chloro-isocumarin (Springer & Rast, 1988). Half-lives for degradation of 2-chloronaphthalene by soil microorganisms of 38 days in waste sludge, 59 days in slop oil sludge, and 70–104 days in wood preserving sludge were obtained (Kincannon & Lin, 1985). Järnberg et al. (1999) did not find any measurable change in the congener composition of tetra- to hexachlorinated naphthalenes (Halowax 1014) in a 28-day aerobic degradation experiment. They stated that lower chlorinated congeners (mono- to trichloronaphthalenes) may have been affected but were not determined.

No information appears to be available on the degradation of PCNs under anaerobic conditions.

The large octanol/water partition coefficients measured for PCNs (see Table 1) indicate that bioaccumulation may be significant. The general trend among the PCNs is for the bioconcentration factor (BCF) to increase as the degree of chlorination increases. This follows closely the trend observed in the octanol/water partition coefficients of PCNs (Table 1). BCFs have been measured in fish for a range of PCNs and are shown in Table 3. The measured BCFs in fish indicate that bioaccumulation is likely to occur to a large extent with PCNs up to and including the hexachlorinated naphthalenes, but is not likely to occur for the hepta- or octachlorinated naphthalenes. However, it should be noted that heptachloronaphthalene residues have been measured in some fish species (see section 6). The increase in log BCF occurs up to a maximum of about 4.5 in the experiments with the guppy (Poecilia reticulata), whereas no uptake and hence no accumulation occurred with heptachloronaphthalenes or octachloronaphthalene (Opperhuizen et al., 1985; Opperhuizen, 1986). This has been explained by the fact that loss of membrane permeation occurs for large molecules, so the chemical cannot pass from water into the cell. The cross-sectional size of the molecule for which this phenomenon occurs has been estimated at around 1 nm (Opperhuizen et al., 1985; Opperhuizen, 1986; Anliker et al., 1988).

Table 3: Bioconcentration factors for chlorinated naphthalenes in fish.

Chlorinated naphthalene	Species	Exposure concentration (µg/litre)	BCF	Reference
Monochloronaphthalene	Cyprinus carpio		191	Matsuo (1981)
2-Chloronaphthalene	Poecilia reticulata	100–1000 a	4 266	Opperhuizen et al. (1985)
1,4-Dichloronaphthalene	Poecilia reticulata	10–1000 a	2 291	Opperhuizen et al. (1985)
1,4-Dichloronaphthalene	Oncorhynchus mykiss	1.7 × 10–3	5 600	Oliver & Niimi (1984)
1,8-Dichloronaphthalene	Poecilia reticulata	10–100 a	6 166	Opperhuizen et al. (1985)
2,3-Dichloronaphthalene	Poecilia reticulata	10–100 a	10 965	Opperhuizen et al. (1985)
2,7-Dichloronaphthalene	Poecilia reticulata	10–100 a	10 965	Opperhuizen et al. (1985)
Trichloronaphthalene	Cyprinus carpio		4 677	Matsuo (1981)
1,3,7-Trichloronaphthalene	Poecilia reticulata	1–100 a	26 915	Opperhuizen et al. (1985)
Tetrachloronaphthalene	Cyprinus carpio		8 710	Matsuo (1981)
1,2,3,4-Tetrachloronaphthalene	Poecilia reticulata	0.1–10 a	33 113 b	Opperhuizen et al. (1985)
1,2,3,4-Tetrachloronaphthalene	Oncorhynchus mykiss	5.6 × 10–3	5 100	Oliver & Niimi (1985)
1,3,5,7-Tetrachloronaphthalene	Poecilia reticulata	0.1–1 a	33 884 b	Opperhuizen et al. (1985)
1,3,5,8-Tetrachloronaphthalene	Poecilia reticulata	1–10 a	25 119 b	Opperhuizen et al. (1985)
Pentachloronaphthalene	Cyprinus carpio		10 000	Matsuo (1981)
Heptachloronaphthalene	Poecilia reticulata		0	Opperhuizen et al. (1985)
Octachloronaphthalene	Poecilia reticulata		0	Opperhuizen et al. (1985)
Octachloronaphthalene	Oncorhynchus mykiss	1.3 × 10–2	330	Oliver & Niimi (1985)

a Exposure concentrations are estimated ranges from a graphical presentation of results.

b Equilibrium was not reached within the duration of the experiment.

Slight to moderate accumulation occurred in algae (Chlorococcum sp.) after exposure to Halowax (1000, 1013, and 1014) for 24 h, with BCFs ranging from 25 to 140. Accumulation increased as the chlorine content of the Halowax increased (Walsh et al., 1977). BCFs of 257 for Halowax 1099, 187 for Halowax 1013, and 63 for Halowax 1000 were measured in grass shrimp (Palaemonetes pugio) exposed to 40 µg/litre for 15 days (Green & Neff, 1977). A BCF of 21 000 was found for worms (Tubifex tubifex and Limnodrilus hoffmeisteri) maintained in spiked sediment (1300 ng 1,2,3,4-tetrachloronaphthalene/g) for up to 79 days, and a depuration half-life of 30 days was measured (Oliver, 1987).

BCFs ranging from 0.73 to 2.5 were reported for tetra-, penta-, hexa-, and heptachloronaphthalenes fed to salmon (Salmo salar) in their diet (0.1–10 µg Halowax 1001, 1014, and 1051/g food) for up to 41 weeks (Tysklind et al., 1998; Åkerblom et al., 2000).

6. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

6.1 Environmental levels

Crookes & Howe (1993) stated that relatively few environmental levels of PCNs had been reported by the early 1990s. The authors found this surprising, since their uses and environmental releases were similar to those of PCBs, for which numerous reports of contamination of the environment existed. One reason for this may have been due to the analytical methodology used in determining PCN residues. It was reported that in many analytical techniques, particularly gas chromatography with electron capture detection, PCNs and PCBs interfered with each other, and so determination of one class of compound in the presence of the other was extremely difficult (Cooke et al., 1980; Kennedy et al., 1982).

During recent years, substantial progress has been made in both the synthesis and analysis of PCNs, enabling the identification of individual congeners in wildlife and abiotic environmental matrices (Falandysz, 1998). For example, such studies have shown a similarity in the pattern of tetra- to heptachlorinated naphthalenes in such abiotic environmental matrices as gas-phase air and fresh water collected in Sweden (Järnberg et al., 1997), riverine sediments in Poland, and subsurface marine plankton (Falandysz & Rappe, 1996; Falandysz et al., 1996a). However, apart from very well defined local situations, it can be very difficult or impossible to relate the pattern of PCNs found in abiotic and biological matrices to any particular source of environmental pollution (Falandysz, 1998). In the following sections, it has been indicated, wherever possible, which of the congeners have been identified. Care should be taken in interpreting the data, particularly those quoted as levels of total PCNs, as it is not always clear if all possible congeners were looked for in the original samples. Data have, in some cases, been summarized into classes of PCNs; this reflects the volume of data if all individual congeners were to be reported and also the differences in resolution between congeners in specific methodologies from different laboratories.

Mean atmospheric PCN concentrations at an urban site (Chicago, USA) and a semiurban site (Toronto, Canada) were 68 and 17 pg/m³, respectively. For urban air, approximately 40% was identified as 1,4,6-trichloronaphthalene (Harner & Bidleman, 1997). Similarly, Dorr et al. (1996) reported PCN concentrations of 60 pg/m³ for urban air (Augsburg, Germany) and 24 pg/m³ for a rural area. Concentrations in urban air ranging up to 98 pg/m³ were reported by Helm et al. (2000), whereas concentrations at the Great Lakes (Canada/USA) ranged from 3 to 27 pg/m³. More than 85% of the PCNs in the air samples were tri- and tetra-isomers. Lee et al. (2000) reported a mean PCN concentration of 152 pg/m³ for a semirural site at Lancaster, United Kingdom. They found that tri- and tetrachloronaphthalenes contributed >95% of the total. Tri- and tetrachloronaphthalenes also formed >90% of the mean PCN concentrations in Arctic air. Mean concentrations were 40 pg/m³ for the Barents Sea, 11.6 pg/m³ for the eastern Arctic Ocean, 7.1 pg/m³ for the Norwegian Sea, 3.5 pg/m³ for Ellesmere Island, Canada, and 0.84 pg/m³ for Dunai Island, Siberia (Harner et al., 1998).

PCNs have been detected at levels of 0.08 µg/m³ (n = 2) and 3.4 µg/m³ (n = 1) in the ambient air of household basements in Niagara Falls, New York, USA. Highly elevated levels of a wide range of organic chemicals were found in the basement, with the higher PCN levels suggesting major contamination. The authors stated that there were several known toxic waste dumps in the area (Pellizzari, 1982).

Levels of PCNs in air have been measured at various manufacturing sites in the USA where PCN use was suspected (US EPA, 1977; Erickson et al., 1978a,b). PCN levels of between 25 and 2900 ng/m³ (n = 7) were measured near a PCN manufacturing site, the congeners detected being mainly mono- (27%), di- (31%), and trichloronaphthalenes (37%), but other congeners were also detected. PCN levels of not detected to 33 ng/m³ (3 of 16 samples were below the detection limit of 0.3 ng/m³) were measured near two capacitor manufacturing plants, and levels of not detected to 3.1 ng/m³ were measured near a paper manufacturing plant.

2-Chloronaphthalene has been detected in fly ash from municipal incinerators in the USA at levels up to 3 µg/kg (Alarie et al., 1989). A concentration of up to 19.6 µg 2-chloronaphthalene/m³ was detected at the scrubber inlet during incineration of sewage sludge (corresponding to an emission of 0.0011 kg/h), but no 2-chloronaphthalene was detected in the scrubber outlet gases (Gerstle, 1988). 2-Chloronaphthalene and 1,2,3,4-tetrachloronaphthalene have been detected in the emissions from the incineration of hexachlorobenzene (Ross & Whitmore, 1984). Similarly, tri- and tetra chloronaphthalenes have been shown to be formed in the high-temperature degradation of 2,3’,4,4’,5-pentachlorobiphenyl (Rubey et al., 1985). Levels of 0.10 µg monochloronaphthalene/m³, 2 µg dichloronaphthalene/m³, and 0.10 µg trichloronaphthalene/m³ have been measured in flue gas samples from a municipal incinerator (Eklund & Strömberg, 1983).

PCNs have been detected at 19 Finnish plywood plants. The source of the PCNs was the pesticide Basil eum SP-70, which contains approximately 80% PCNs (mainly mono- and dichlorinated isomers) and 4% tributyltin oxide. The pesticide was mixed into the glues used to make the plywood, and monochloronaphthalene and dichloronaphthalene concentrations of 0.2–8 mg/m³ were detected in the glueing department (Kauppinen, 1986).

A laboratory investigation was initiated in order to investigate the possible individual exposure levels at the workplace for those using Beranit^® for casting moulds (Popp et al., 1997). After a 1-h emission, assuming a working room volume of 100 m³ and no ventilation, mean air concentrations of halogenated compounds were 14.5 mg/m³ for total PCNs, 4.9 mg/m³ for trichloronaphthalene, 1.0 mg/m³ for pentachloronaphthalene, 0.025 mg/m³ for PCBs, and 2.2 pg international toxicity equivalency factor (I-TEF)/m³ for polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) (NATO/CCMS, 1988).

Levels of PCNs in water have been measured at various manufacturing sites in the USA where PCN use was suspected (US EPA, 1977; Erickson et al., 1978b). PCN levels of 0.6 and 1.4 µg/litre were measured in two water samples near a PCN manufacturing plant. Levels of not detected to 5.5 µg/litre (four of seven samples were below the detection limit of 0.2 µg/litre) were measured in water near two capacitor manufacturing sites, and PCNs were generally not detected near a paper manufacturing plant.

Monochloronaphthalene at levels of 650– 750 ng/litre and dichloronaphthalene at levels of 150– 260 ng/litre have been detected in the River Besós and River Llobregat, Barcelona, Spain. Both rivers receive a wide spectrum of waste discharges, including domestic, industrial, and agricultural wastes (Gomez-Belinchon et al., 1991). In groundwater, total PCNs (expressed as Halowax 1099 equivalents) ranged from <0.5 ng/litre to 79.1 µg/litre for the Llobregat aquifer, with tri- and tetrachloronaphthalenes the major groups of congeners identified. The authors reported that the higher levels probably originated from the poor disposal of illegal landfills closed during the 1970s (Espadaler et al., 1997; Martí & Ventura, 1997). Total PCN concentrations of 0.89 and 2.6 ng/litre were reported for a PCB-polluted river and percolating water at a city dump site (Stockholm, Sweden), respectively (Järnberg et al., 1997).

Levels of chloronaphthalene and dichloronaphthalene have been measured in two samples of Tsukuba (Japan) tap water after chlorination. The detection limit in the experiment was 0.003 ng/litre, and the levels of both chloronaphthalene and dichloronaphthalene were below the detection limit in the raw water before chlori nation. After chlorination, levels of 0.03–0.44 ng/litre for chloronaphthalene and levels of not detected to 0.15 ng/litre for dichloronaphthalene were measured (Shiraishi et al., 1985).

Octachloronaphthalene was detected in sediments from Bayou d’Inde (Louisiana, USA), near an industrial outfall. Octachloronaphthalene levels expressed in terms of organic carbon were 12 mg/kg in bottom sediments and 0.8 mg/kg in suspended sediments. No octachloronaphthalene was detected in water from the same area (Pereira et al., 1988). PCN concentrations of up to 23 mg/kg dry weight were found in sediments collected in an area contaminated by disposal of wastes from the chlor-alkali process. Hexa- and heptachloronaphthalenes were the most abundant congeners, accounting for >70% of the total; a characteristic profile of PCNs suggested the formation of congeners during the chlor-alkali process (Kannan et al., 1998). Octachloronaphthalene has been detected in estuarine sediment samples from the USA at levels of 104 mg/kg dry weight (Rostad & Pereira, 1989).

Levels of 1-chloronaphthalene ranging up to 100 µg/kg dry weight have been measured in marine sediments of Cortiou Creek in the sewage outfall area of Marseilles, France (Milano et al., 1985; Milano & Vernet, 1988). Surface sediment samples from Venice and Orbetello lagoons, Italy, collected during 1995 contained total PCN levels ranging from 0.03 to 1.51 µg/kg dry weight. Tetra- and pentachlorinated naphthalenes were the predominant congeners (Eljarrat et al., 1999). Falandysz et al. (1996a) reported that >80% of the total PCNs found in surface sediment (Gdansk Basin, Baltic Sea) at a concentration of 6.7 µg/kg dry weight was tetrachloronaphthalene, with the congeners 1,2,4,6-, 1,2,4,7-, 1,2,5,7-, 1,2,5,8-, and 1,2,6,8- being the most predominant. Similarly, Ishaq et al. (2000) found that tetrachloronaphthalenes (65%) were the predominant congeners in Baltic sediment, followed by pentachloronaphthalene (27%).

Pentachloronaphthalenes have been detected at levels of 1.3 µg/kg wet weight in sediments from Lake Järnsjön on the River Emån, Sweden. Hexachloro- and heptachloronaphthalenes were not detected in the same samples (Asplund et al., 1990a). Similarly, Järnberg et al. (1997) found a total PCN level of 0.23 µg/kg dry weight (tetra- to hepta-congeners) in an unpolluted Swedish river, compared with concentrations of up to 260–270 µg/kg near chlor-alkali plants and in PCB-polluted rivers. Kannan et al. (2000a) found total PCN levels ranging from 0.08 to 187 µg/kg dry weight for surface sediments from the Detroit and Rouge rivers, Michigan, USA, with penta- and hexachloronaphthalenes the predominant congeners.

Levels of 20 mg chloronaphthalene/kg, 8 mg dichloronaphthalene/kg, and 6 mg trichloronaphthalene/ kg have been measured in sediments from the Niagara River, New York, USA, near a dump site (Elder et al., 1981). Another report on the same area gave sediment levels (dry weight) of 5 mg chloronaphthalene/kg, 10 mg dichloronaphthalene/kg, and 4.4 mg trichloronaphthalene/kg. These levels were found at the surface of the sediment. The levels were found to decrease markedly with depth. Lower levels were measured at other sites in the area, and the source of contamination was thought to be a storm sewer outfall from a toxic dump site (Jaffe & Hites, 1984). Levels of PCNs in 33 sediment samples from the Trenton Channel of the Detroit River, Michigan, USA, have been measured. The channel receives waste discharges from several chemical manufacturers. PCNs with 2–8 chlorine atoms were identified, and total PCN levels between not detected and 61 mg/kg dry weight were reported (Furlong et al., 1988).

Levels of PCNs have been measured in soils near various manufacturing plants in the USA where PCNs are thought to have been used (US EPA, 1977; Erickson et al., 1978b). Near a PCN manufacturing plant, levels of 130–2300 ng/kg were measured, made up of mainly tri-, tetra-, and pentachloronaphthalenes. PCN levels of between not detected and 21 µg/kg and between not detected and 470 µg/kg were measured near two capacitor manufacturing facilities, and levels ranging from not detected to 34 µg/kg were measured near a paper manufacturing plant (detection limit 0.05 µg/kg).

PCNs have been detected in contaminated soil samples from areas in the Netherlands that have been used for municipal waste disposal. The distribution of congeners was the same as that for Halowax 1013, suggesting that this was the source of contamination and that the composition of the PCNs had not changed, despite being buried in the landfill for 10–15 years. PCN levels of 31–38 mg/kg dry soil and 1180–1290 mg/kg dry soil were measured in two soils; a third soil contained no PCNs (De Kok et al., 1983). Kannan et al. (1998) found PCN concentrations of 17.9 mg/kg dry weight in soil near a former chlor-alkali plant; hexa- and heptachloronaphthalene congeners accounted for >70% of the total concentration. Harner et al. (2000) analysed rural soils in the United Kingdom dating back to the 1940s. They found a peak level of 12 µg/kg dry weight in the 1960s, falling to 0.5–1 µg/kg in 1990. More detailed analysis revealed that tetra- and pentachloronaphthalenes reached a peak in the 1950s, whereas peak values for trichlorinated isomers were recorded during the 1970s.

During 1968, commercial rice oil was found to be contaminated with PCBs containing total PCN concentrations of 2.6 µg/g. Penta-, hexa-, and heptachloronaphthalenes were the principal isomers (Haglund et al., 1995).

Residues of PCN congeners in biota from the Baltic are summarized in Figure 1. Tetra- and pentachloronaphthalenes predominate at all trophic levels. Because residues in lower organisms (plankton, invertebrates, and fish) are expressed as whole body and residues in higher organisms (birds and mammals) are expressed in specific tissues, it is not possible to directly compare trophic levels. When moving upwards in the food-chain, the homologue distribution does become richer in the higher chlorinated homologues. Ishaq et al. (2000) found that the predominant congeners in isopods were pentachloronaphthalenes (53%), whereas in four horned sculpins (Myoxocephalus quadricornis), which feed on isopods and amphipods, hexachloronaphthalenes comprised 42% of the total. Furthermore, it was found that polychlorinated congeners that lack two unsubstituted carbon atoms adjacent to each other bioaccumulate to a greater extent than other congeners.

Levels of 2-chloronaphthalene have been measured in oysters (Crassostrea virginica) and clams (Rangia cuneata) from Lake Pontchartrain, USA. The levels were 34 µg/kg wet weight in oysters and 140 and 970 µg/kg wet weight in clams (McFall et al., 1985). PCN levels of 39 µg/kg in fish have been measured near a PCN manufacturer in the USA (US EPA, 1977; Erickson et al., 1978b). Levels of tetra- to hexachlorinated naphthalenes in one species of crab and two species of estuarine fish ranged from non-detectable to 0.3 µg/kg wet weight in the vicinity of a coastal former chlor-alkali plant (Kannan et al., 1998).

Congeners of PCNs have been detected in fish from 16 out of 18 watersheds sampled in the Great Lakes. It was concluded that PCNs were widely distributed in the fish samples, although not all congeners were found in all 16 samples (Kuehl et al., 1984a). Freshwater fish sampled in the Great Lakes and their US catchments (1996–1997) showed total PCN levels ranging from 0.04 to 31.4 µg/kg wet weight (Kannan et al., 2000b).

PCNs have been reported to be present at levels of <1 µg/kg in marine fish from Japan; however, some samples taken at a river mouth contained several hun dred µg/kg (Takeshita & Yoshida, 1979a). Total PCN concentrations ranging from 6.3 to 260 µg/kg lipid weight have been recorded for fish in the Baltic Sea (Falandysz et al., 1996a). Sinkkonen & Paasivirta (2000) found PCN concentrations in Arctic cod (Cadus callarias) liver ranging from 0.078 to 0.78 µg/kg lipid weight for the penta-congeners and from 0.05 to 0.48 µg/kg for the hexa-congeners during the period 1987–1998.

PCNs have been detected in the livers of gulls from the Mediterranean at levels of up to 62.5 mg/kg wet weight (Vannucchi et al., 1978). Guillemot (Uria aalge) eggs sampled from the same site on the Baltic coast of Sweden between 1974 and 1987 showed decreasing trends in tetra-, penta-, and hexachloronaphthalene residues (Järnberg et al., 1993).

Total PCNs have been detected in British birds of prey (Eurasian kestrel Falco tinnunculus, sparrowhawk Accipiter nis