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.

Concise International Chemical Assessment Document 62

First draft prepared by Drs Christine Melber, Janet Kielhorn, and Inge Mangelsdorf, Fraunhofer Institute of Toxicology and Experimental Medicine, 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, 2004

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

Coal tar creosote.

(Concise international chemical assessment document ; 62)

1.Coal tar - toxicity 2.Creosote - toxicity 3.Risk assessment 4.Environmental

exposure 5.Occupational exposure I.International Programme on Chemical Safety

II.Series

ISBN 92 4 153062 6         (LC/NLM Classification: QV 633)

ISSN 1020-6167

©World Health Organization 2004

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Technically and linguistically edited by Marla Sheffer, Ottawa, Canada, and printed by Wissenchaftliche Verlagsgesellschaft mbH, Stuttgart, Germany

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 usually 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., a standard CICAD or a de novo 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 usually based on an existing national, regional, or international review. When no appropriate source document is available, a CICAD may be produced de novo. 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. When a CICAD is prepared de novo, a consultative group is normally convened.

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.

The first draft of this CICAD was prepared by the Fraunhofer Institute of Toxicology and Experimental Medicine, Hanover.² A comprehensive literature search of relevant databases was performed in June 2002. The first draft of this document was circulated for a limited peer review, and a Consultative Group was convened to finalize the document and verify that the peer review comments had been adequately dealt with. The members of the Consultative Group, who were participants in this peer review, are provided in Appendix 2. The final draft was then sent for peer review to IPCS Contact Points and Participating Institutions, as well as to further experts identified in collaboration with the IPCS Risk Assessment Steering Group. Information on the peer review of this CICAD is presented in Appendix 3. This CICAD was approved as an international assessment at a meeting of the Final Review Board, held in Varna, Bulgaria, on 8–11 September 2003. The members of the Final Review Board are listed in Appendix 4. The International Chemical Safety Card for creosote (ICSC 0572), produced by the International Programme on Chemical Safety (IPCS, 2002), has also been reproduced in this document.

This CICAD is on coal tar creosote. Wood creosote is a different product that is used mainly in pharmaceutical preparations and is not covered in this document.

Coal tar creosote is a brownish-black/yellowish-dark green oily liquid with a characteristic odour, obtained by the fractional distillation of crude coal tars. The approximate distillation range is 200–400 °C. The chemical composition of creosote is influenced by the origin of the coal and also by the nature of the distilling process; as a result, the creosote components are rarely consistent in their type and concentration.

Creosote is a mixture of several hundred, probably a thousand, chemicals, but only a limited number of them are present in amounts greater than 1%. There are six major classes of compounds in creosote: aromatic hydrocarbons, including polycyclic aromatic hydrocarbons (PAHs) and alkylated PAHs (which can constitute up to 90% of creosote); tar acids / phenolics; tar bases / nitrogen-containing heterocycles; aromatic amines; sulfur-containing heterocycles; and oxygen-containing heterocycles, including dibenzofurans. Creosote may be sold as diluted preparations, which may contain carrier oil or solvents. The composition and use of creosote are regulated in some countries; the regulations usually focus on the content of benzo[a]pyrene (BaP) and phenolics.

Creosote is only slightly soluble in water and soluble in a variety of organic solvents. However, the physical and chemical properties of the individual components of creosote vary widely; some, for example, are highly soluble in water.

The analysis of creosote is complex. Different profiles of creosote chemicals are found in the different matrices: the most volatile are found in air, the most soluble in water, and those with greater sorptive capacity in sediment/soil. Depending upon the matrix (e.g., air, water, soil/sediment, biological materials) from which the sample is taken, suitable cleanup and extraction are necessary. High-resolution gas chromatography (HRGC) with a flame ionization detector (FID) or mass spectrometric (MS) detection or reversed-phase high-performance liquid chromatography (HPLC) with a fluorescence detector (FL) have been the separation and determination methods most commonly used.

Occupational exposure to airborne creosote particles has been previously monitored as coal tar pitch volatiles (CTPV). However, the CTPV method is not sensitive enough to measure low concentrations of creosote fumes. Important components such as airborne PAHs can be sampled on a polytetrafluoroethylene (PTFE) filter connected to a sorbent tube and analysed after extraction by HRGC or HPLC. Other volatile compounds from creosote can be sampled on sorbent tubes.

The urinary PAH metabolites 1-pyrenol (1-hydroxypyrene) and 1-naphthol (1-hydroxynaphthalene) have been used in the assessment of creosote exposure.

Coal tar creosote is a wood preservative and water-proofing agent for structures on land and in marine and fresh waters and for railway crossing timbers and sleepers (railroad ties), bridge and pier decking, poles, log homes, fencing, and equipment for children’s playgrounds.

The majority of creosote used in the European Union (EU) is for the pressure impregnation of wood. In the USA and many other countries, the use of coal tar creosote is limited to certified applicators.

Non-wood uses include anti-fouling applications on concrete marine pilings. Creosote can be a component of roofing pitch, fuel oil, and lamp black and a lubricant for die moulds. Other uses reported include animal and bird repellent, insecticide, animal dip, and fungicide.

Creosote production in the USA falls into two categories: distillate (100%) creosote and creosote in coal tar solution. Distillate production in 1992 was 240 000 tonnes; production of creosote in coal tar solution was 110 000 tonnes. The production of creosote in the EU has been estimated to be approximately 60 000–100 000 tonnes per year.

During pressure impregnation of wood products, excess creosote may be released from the treated materials. Leaching of spilled wastes from these application sites has been common. Creosote is also released to the environment from facilities through air emissions.

The environmental transport and distribution of creosote are complex processes, depending on the physicochemical properties of creosote constituents and their interaction with matrix properties, as well as environmental conditions. Generally, creosote is distributed within all environmental compartments (air, water, sediment, soil, biota). However, the major environmental sinks of creosote components are sediment, soil, and groundwater.

Generally, phenolic compounds, low-molecular-weight PAHs, and some heterocycles tend to be predominantly in the gaseous phase. Creosote constituents may also occur in the atmosphere as particulate matter.

Volatilization of creosote from water surfaces is not considered to be a significant process.

The movement of creosote within aquatic systems is dependent upon the aqueous solubility, affinity to organic phases, and sorptive capacity of the components. Generally, the highly soluble fraction includes phenolic and heterocyclic compounds and low-molecular-weight PAHs. The high-molecular-weight aromatic compounds, with relatively low solubilities and high adsorptive capacities, dominate the associated sediments. However, movement of high-molecular-weight compounds may occur by co-transport of colloid-sorbed contaminants.

Field observations and laboratory leaching experiments have shown losses of creosote components from wooden creosoted constructions during water immersion. Leachability of creosote components was higher in fresh water than in seawater. The rate of migration increased with increasing temperature and decreased with the age of the pilings. Nitrogen-containing heterocycles leached faster than PAHs and dibenzofuran.

The rate of vertical or horizontal transport of creosote components in soil is dependent upon their physicochemical properties as well as the soil properties and environmental conditions. Laboratory model and field experiments (simulating creosote spills) showed a high retardation of transport of high-molecular-weight compounds coupled with a fast downward migration of low-molecular-weight compounds. Some of the creosote compounds released from treated wood products into surrounding soil may persist for decades.

Creosote PAHs are taken up to a small degree by terrestrial plants and animals. No quantitative data on uptake of creosote compounds are available for farm animals. A number of aquatic invertebrates and fish monitored in field and laboratory studies showed significant uptake of creosote-derived PAHs. Transfer to the human food supply is possible via contaminated seafood.

The biodegradability of creosote constituents is variable. Generally, the efficacy of aerobic degradation is greater than that of anaerobic degradation. Phenolic compounds are relatively easily degraded. Within PAHs, degradability appears to be inversely related to the number of aromatic rings. Some heteroaromatic compounds are quickly removed, whereas others are recalcitrant. Biotransformation of creosote components appears to dominate over mineralization. In some cases, the intermediates formed can be more persistent, mobile, or toxic than their parent compounds.

Besides structural features of the chemicals, a number of other factors, such as bioavailability, microbial adaptation, oxygen supply, and nutrient availability, influence their degradation or transformation in situ.

Although little examined, fish appear to metabolize creosote PAHs more rapidly than aquatic invertebrates.

Photochemical transformation seems to be the most important abiotic mechanism by which creosote constituents, such as PAHs and heterocyclic and phenolic compounds, are transformed in the atmosphere and, to a lesser extent, in water and soil. Photo-oxidation prevails over direct photolysis. A study performing irradiation of selected PAHs separately or of the same PAHs present in a creosote mixture showed that there was a trend of decreased photoreactivity in the mixture compared with the individual tests.

Aquatic invertebrates and fish bioaccumulate creosote components, as has been demonstrated mainly for PAHs by field monitoring studies at creosote-contaminated sites, relocation experiments, and laboratory or microcosm studies. Generally, PAH profiles in insects and crayfish were close to that found in sediments, whereas fish had greatly altered ratios for low/high-molecular-weight PAHs. Bioconcentration factors (BCFs) in connection with creosote exposure have rarely been reported. However, BCFs for PAH components from creosote-contaminated sediments have been estimated to range from 0.3 to 73 000.

A number of remediation strategies have been developed, mainly for creosote-contaminated groundwater and soils. Most of the treatments achieved significant reductions for certain substances, but were not or only partially successful in reducing the toxic potential of the treated matrices.

Creosote-treated wood does not decay in the environment, and therefore its disposal is problematic. Creosote-treated wood should not be incinerated under uncontrolled conditions, as toxicants such as PAHs and halogenated dioxins and furans may be produced.

The very few data available for ambient air concentrations refer to concentrations of selected PAHs in the vicinity of creosote facilities. A maximum concentration of 90 ng/m³ has been reported for naphthalene at a distance of 2000 m. Concentrations decreased with increasing distances from creosoting plants: for example, from 64 ng/m³ at 500 m to 1.6 ng/m³ at 5000 m for fluoranthene or from 5 ng/m³ at 100 m to 0.6 ng/m³ at 2000 m for BaP.

Groundwater samples near creosote waste sites in several countries have been found to contain creosote-related PAHs and phenolic, heterocyclic, and BTEX (benzene, toluene, ethylbenzene, and xylene) compounds. Monitoring data from 44 Danish creosote sites showed concentrations (90th percentiles) of 30 µg/litre for BaP and 50 µg/litre for chrysene. Highest concentrations of several individual heterocyclic, phenolic, or BTEX compounds detected in the vicinity of several creosote waste sites were in the range of 10–80 mg/litre.

Concentrations in the mg/litre range have been found for some individual PAHs in river water affected by a creosote spill 10 years earlier. Twelve individual PAHs were monitored in water samples of a drainage stream near a creosote works. Maximum concentrations ranged from 0.02 µg/litre (benzo[b]- and benzo[k]fluoranthene) to 153 µg/litre (naphthalene), with BaP concentrations of up to 0.05 µg/litre.

Elevated PAH concentrations have also been observed in small waterways, where banks were protected with creosoted wood constructions, or in railway ditches, where creosote-treated power or telecommunication line poles were erected. The maximum BaP concentration measured was 2.5 µg/litre. The mean total PAH concentration in the ditches was about 600 µg/litre.

In the vicinity of wood-preserving facilities, maxima for total PAHs in sediments amounted to about 20 000–30 000 mg/kg dry weight; maxima for total nitrogen heterocycles were in the order of 1000 mg/kg dry weight. BaP concentrations as high as several hundred mg/kg dry weight have been measured. The most abundant heterocycle was carbazole (18 mg/kg dry weight). Sediments near creosoted wooden constructions (pilings, bank protection, poles/sleepers) showed total PAH concentrations of up to 1200 mg/kg dry weight, with mean BaP concentrations of about 2 mg/kg dry weight.

Elevated concentrations of creosote-derived compounds have been documented in soils near abandoned creosote-producing/using facilities in several countries, with maximum concentrations of several thousand mg/kg dry weight for total PAHs and of nearly 100 mg/kg for total phenols. Concentrations of "creosote oil contents" up to 90 000 mg/kg dry weight have been reported around creosote-treated poles. Soil from a storage area for impregnated railway ties and playground sand from sandboxes made of old impregnated railway ties contained total PAHs at concentrations up to 20 mg/kg and up to about 2 mg/kg dry weight, respectively. BaP concentrations found in soils near wood treatment/storage sites reached a maximum of 390 mg/kg dry weight, those near creosoted posts, 6 mg/kg, and those from playground sand, 0.2 mg/kg.

Creosoted wood products can contain high concentrations of PAHs, even after several decades of use; phenolic and heterocyclic compounds may also be present. For example, mean concentrations (mg/kg wood) ranging from 1510 (quinoline) to 11 990 (phenanthrene) have been found to occur in creosoted wood. Wooden sleepers installed in playgrounds showed BaP concentrations of up to 1570 mg/kg shavings.

Edible fish and seafood captured from creosote-contaminated areas or held in creosoted cages have been found to contain increased concentrations of PAHs and PAH metabolites. The mean concentration of BaP in tail meat of commercial market lobsters increased from 0.6 to 79 µg/kg wet weight after about 3 months of impoundment.

Creosote-derived PAHs have been detected at concentrations significantly over background levels in several classes of aquatic fauna, including insects, molluscs, crustaceans, and fish collected at various creosote-contaminated sites of freshwater or estuarine/marine environments. In general, concentrations were highest in invertebrates (up to several hundred mg/kg dry weight). Concentrations of total PAHs in liver of fish living on creosote-contaminated sediment and in their invertebrate food organisms were as high as 1 and 84 mg/kg dry weight, respectively (compared with 0.1 and 0.5 mg/kg dry weight in controls). Heterocyclic compounds in snails (Thais haemastoma) from a bay near a wood-preserving facility were found to be present at concentrations up to about 10 µg/kg wet weight, and PAHs were present at concentrations up to about 200 µg/kg wet weight.

The general population can be exposed to creosote or creosote components by handling creosote or products containing creosote and by contact with creosote-contaminated air, water, soil, or food. Routes of exposure include inhalation, drinking/ingestion, and skin contact.

Due to the complexity of creosote and the many different exposure situations, exposure may vary both qualitatively and quantitatively. Nevertheless, some estimations using BaP as a marker substance and based on several assumptions have been performed for two important exposure scenarios. As a result, a daily exposure through skin contact of about 2 ng BaP/kg body weight has been assessed for children playing on creosoted playground equipment. The daily intake of BaP from consumption of vegetables and fruits from gardens in the vicinity of creosoting plants has been estimated to range from 1.4 to 71.4 µg/kg body weight.

There is one study providing internal monitoring data for people living in the vicinity of a creosote impregnation plant. Excretion values of 1- and 2-naphthol were significantly higher in the exposed residents than in controls. For example, the mean concentrations of 1-naphthol in morning urine samples were 2.5 µmol/mol creatinine for the exposed and 1.2 µmol/mol creatinine for the non-exposed group. The 1-pyrenol excretion did not differ significantly.

Occupational exposure to creosote may occur during manufacture, use, transport, or disposal of creosote or creosoted wood products. Most data are available for wood-preserving workers.

Creosote aerosol concentrations monitored as the CTPV by similar methods in wood impregnation plants reached maxima of up to 9700 µg/m³. Total time-weighted average (TWA) concentrations of creosote vapours ranged from 0.5 to 9.1 mg/m³, with peaks up to 71 mg/m³, at wood impregnation plants and from 0.1 to 11 mg/m³ at workplaces where creosoted wood was handled. The mean concentrations of particulate PAHs ranged from 0.2 to 106 µg/m³ in the impregnation plants and from 0.8 to 46 µg/m³ in the handling of impregnated wood. The proportion of particulate-bound PAHs relative to total PAHs appeared to be less than 4%.

Prevailing compounds of the vapour phase of wood impregnation plants were naphthalene, methylnaphthalenes, indene, acenaphthene, and fluorene; the main PAHs of the particulate phase included fluorene, phenanthrene, anthracene, and pyrene. Maximum concentrations of the marker substances naphthalene and BaP (the latter mainly particle-bound) were as high as 41 mg/m³ and 1 µg/m³, respectively. An abundant heterocyclic PAH was benzothiophene, showing concentrations of up to 2800 µg/m³. Concentrations of phenol, biphenyl, and methyl styrenes did not exceed 2000, 1000, and 3000 µg/m³, respectively. Air monitoring during cleanup operations of highly creosote-contaminated soil revealed exposure concentrations of up to 0.9 mg/m³ for volatile PAHs, 0.2 mg/m³ for particulate PAHs, and <0.002 mg/m³ for BaP.

An important route of occupational exposure to creosote is via skin. It has been estimated that over 90% of pyrene and 50–70% of naphthalene enters via the skin. A mean total pyrene contamination on the skin of creosote impregnation workers was approximately 1 mg/day in workers without protective clothing. Protective clothing reduced the pyrene contamination on the workers’ skin by about 35%, on average.

Concentrations of two PAH metabolites, 1-naphthol and 1-pyrenol, have been monitored as internal markers of creosote exposure. For example, the mean urinary concentrations of 1-naphthol in Finnish wood impregnation plant workers and in assemblers handling treated wood were 1350 and 1370 µmol/mol creatinine, respectively. The mean urinary concentration of 1-pyrenol was about 10 times higher in these wood impregnators (64 µmol/mol creatinine) than in the assemblers. An increase in urinary 1-pyrenol values during the workshift has also been observed in workers involved in the production of creosote or the cleanup of creosote-contaminated soil. The 1-pyrenol concentrations correlated well with differences in pyrene skin contamination, but poorly with differences in pyrene breathing-zone air concentrations.

Exposure calculations on the basis of excreted metabolites (plus air and/or skin monitoring data) suggested a total daily uptake of 15 or 16 mg/worker (assembler or impregnator) for naphthalene. Estimations for pyrene did not exceed 5 mg/worker per day.

There are no laboratory animal or human studies measuring the specific rate and extent of coal tar creosote absorption following oral, inhalation, or dermal exposure. However, evidence for a significant absorption of creosote components comes from detection of creosote PAH metabolites in urine of creosote-exposed workers or volunteers and from detection of PAH–DNA adducts in animal or human tissues following creosote exposure. Indirect evidence also comes from the toxic effects elicited by creosote in laboratory animals or humans. Additionally, single-component studies show a significant absorption potential of individual PAHs, although their predictive value for the quantitative absorption kinetics after exposure to the mixture is limited.

Specific distribution studies on coal tar creosote have not been performed.

In accordance with principal PAH metabolic pathways, hydroxy metabolites of PAHs such as 1-naphthol and 1-pyrenol have been measured in urine of creosote-exposed humans.

In general, PAHs (metabolized or unmetabolized) can be excreted into bile, faeces, and urine as well as into breast milk, regardless of the route of absorption. However, specific studies on the elimination and excretion of coal tar creosote are confined to the determination of PAH metabolites in human urine. Elevated urinary levels of 1-naphthol and 1-pyrenol have been found in workers of several wood creosoting plants and in assemblers handling creosote-impregnated wood. Comparisons between the estimated daily uptake of naphthalene/pyrene by inhalation and the urinary excretion of 1-naphthol/1-pyrenol indicated a remarkable contribution of non-inhalation routes of uptake, especially for pyrene. The relevance of dermal uptake for 1-pyrenol excretion has also been demonstrated in workers using protective clothing, which resulted in a significant reduction of skin contamination and 1-pyrenol excretion. Topical treatment of volunteers with a single dose of creosote significantly enhanced the basal excretion of 1-pyrenol.

Elimination half-lives for 1-naphthol and 1-pyrenol were in the range of hours or days.

Most studies on interactions of creosote with cellular components refer to interactions of creosote PAHs with nucleic acids. PAH–DNA adducts have been detected in mice, rats, and fish after experimental or environmental exposure to creosote.

Based on limited studies, creosotes are of low to moderate acute toxicity in experimental animals. The lowest LD₅₀ value, 433 mg/kg body weight, was reported for mice after oral exposure. There is little reliable information on effects of creosotes after short-term exposure. Body weight losses have been observed in rats, sheep, and calves following oral creosote doses.

Some earlier limited studies with mice indicated a carcinogenic activity of creosotes after topical application. Types of tumours included not only skin carcinomas and papillomas, but also lung carcinomas. A more recent epicutaneous mouse study performed with two different coal tar creosote preparations (CTP1: BaP content of 10 mg/kg; CTP2: BaP content of 275 mg/kg) confirmed the carcinogenic potential of creosotes with respect to induction of skin tumours. There was a linear dose–effect relationship between tumour incidence and BaP content of both creosotes. The creosotes were about 5 times more potent than expected from pure BaP treatments. Non-neoplastic effects observed in this long-term (78 weeks) study included skin ulcerations and decreases in life span.

Several creosotes have been shown to be skin irritants in animals. Data on eye irritancy are conflicting.

There are no adequate animal studies on the reproductive or developmental toxicity of creosotes. However, creosote has been shown to be able to elicit estrogen-mediated activities in vitro, indicating some potential for endocrine disruption. Adverse reproductive effects have also been reported in fish exposed to creosote.

A number of in vitro tests based upon bacterial and mammalian systems have shown creosote to be genotoxic. The pattern of genotoxicity observed was similar to that found in PAHs. Creosote was also genotoxic in an in vivo micronucleus test in mice.

Tests with fish cells in culture showed that the cytotoxicity of creosote is enhanced by irradiation with ultraviolet (UV) light. This is consistent with the known phototoxic potential of some PAHs.

Creosote has been shown to be a hepatic microsomal enzyme inducer in laboratory mammals.

1.7.1 General population

Information on the effects of coal tar creosote in the general population is scarce.

Creosote has been involved in incidental or accidental poisoning incidents, mainly due to its use as a pesticide. Deaths occurred following ingestion of about 1–2 g (children) or about 7 g (adults). Symptoms included salivation, vomiting, respiratory difficulties, vertigo, headache, loss of pupillary reflexes, hypothermia, cyanosis, convulsion, etc., accompanied by oropharyngeal, intestinal, pericardial, liver, and kidney damage.

Increased occurrence of skin rashes in people residing in or near an abandoned wood creosoting plant in the USA has been suggested.

Evidence of cancer incidence following environmental exposure is limited to a report on breast and gastrointestinal cancer in females of a population exposed to a creosote-contaminated water supply in the USA. However, it could not clearly be demonstrated whether creosote or confounding risk factors were responsible.

1.7.2 Occupational exposure

Most reports on the effects of coal tar creosote on humans refer to occupational exposure, resulting mainly from dermal and/or inhalational contact with creosote or creosoted wood.

The most apparent effects included irritations or lesions of skin and eyes, including phototoxic or photoallergic reactions, sometimes accompanied by general symptoms such as depression, weakness, headache, slight confusion, vertigo, nausea, increased salivation, or vomiting. Photosensitization (sensitization of the skin to UV light by creosote) has been observed in workers exposed to creosote.

Increased risks of developing lip and skin cancers have been observed in cohort studies of Swedish and Norwegian wood impregnators and in Finnish round timber workers. The possible interaction with sunlight exposure has not been adequately addressed. The mortality for cancer of the scrotum was elevated among brickmakers exposed to creosote.

Single epidemiological studies suggested a possible risk for bladder cancer, multiple myeloma, and lung cancer due to exposure to creosote. Two case–control studies suggested an increased risk of brain tumours and neuroblastoma among offspring of male workers with possible creosote exposure.

All of the epidemiological studies were based on qualitative estimations of exposure rather than on measurements.

EC₅₀ values (15 min) determined using the Microtox test (inhibition of bioluminescence from Photobacterium phosphoreum or Vibrio fischeri) by different coal tar creosotes (in acetone solutions) ranged from 0.38 to 0.63 mg/litre. Significant decreases in bioluminescence compared with controls have also been found for several creosote-contaminated environmental samples, such as sediments (including their elutriates and pore waters) and groundwater. Furthermore, a strong inhibition of nitrification by creosote-contaminated leachate has been observed.

Creosote induced signs of stress and abnormal growth in experimentally exposed aquatic plants. Visual changes in Myriophyllum spicatum could be seen at nominal creosote concentrations as low as 1.5 mg/litre. EC₅₀ values for a decrease in node production, shoot lengths, and dry weight were calculated to be 86, 55, and 33 mg/litre, respectively. Additionally, membrane ion leakage was significantly and dose-dependently increased at creosote concentrations ranging from 0.1 to 92 mg/litre. The phototoxic potential of creosote has been demonstrated in Lemna gibba: EC₅₀ values (nominal) for reduction in growth rate decreased from 54 mg/litre (under laboratory visible light) to 12 mg/litre under simulated solar radiation.

Creosote EC₅₀/LC₅₀ values for aquatic invertebrates have been measured in the range of 0.02–4.3 mg/litre. Larval stages proved to be more sensitive than adult stages. Lifetime exposure of Daphnia pulex to water-soluble fractions of creosote resulted in decreased growth rates and reproductive impairment.

An increase in susceptibility to infections has been observed in eastern oysters (Crassostrea virginica) exposed to 15% and 30% dilutions of creosote-contaminated sediment. Increased mortalities have been noted in many crustacean species exposed in the laboratory to matrices environmentally contaminated by creosote. Sublethal effects, such as decreases in dry weight gain and in proportion of gravid females, have been recorded in Mysidopsis bahia (crustacean); the 7-day EC₅₀ for these more subtle effects was 0.015 µg total identified aromatic hydrocarbons/litre.

Acetone extracts from creosote-contaminated sediments showed an acute toxicity to Nitocra spinipes (crustacean) comparable to that of creosote.

Creosote is acutely toxic to fish, with the lowest LC₅₀ reported to be 0.7 mg/litre.

Creosote-contaminated groundwater, water, or sediments (including associated waters) have been shown to cause adverse reproductive and developmental effects in fish. The LC₅₀ for hatching success was calculated to be 0.05 mg creosote/litre. LC₅₀ values determined in spot (Leiostomus xanthurus) decreased with increasing duration of exposure to creosote-contaminated sediment during 7–28 days of exposure.

Data on the effects of creosote exposure on terrestrial organisms are limited. A root elongation test of different creosotes with Allium cepa resulted in 96-h EC₅₀ values (for reduction of root length) ranging from 18 to 34 mg/litre. Earthworms (Eisenia foetida) exposed to creosote-contaminated soil (e.g., about 1000 mg total PAHs/kg dry weight) died within a few days.

In the vicinity of creosote sources, adverse effects on aquatic microorganisms, aquatic invertebrates, and fish have been observed, similar to those inducible by creosote in the laboratory. Fish from heavily creosote-contaminated sites (sediments) showed a high prevalence of hepatic and extrahepatic neoplasms, an impaired immune status (reduced macrophage activities), and reproductive impairment.

In a series of outdoor aquatic microcosm studies in which creosote was added, there was a rapid concentration-dependent reduction in zooplankton abundance and number of taxa, with an EC₅₀ (at 5 days) of 45 µg creosote/litre (nominal). In contrast, there was no direct adverse effect on the phytoplankton community. In another test, rainbow trout (Oncorhynchus mykiss) exposed to 100 µl creosote/litre (nominal) died within 3 days. At lower concentrations, immunological alterations developed within 28 days (lowest-observed-effect concentration [LOEC]: 17 µl creosote/litre [nominal]). The creosote-induced immunomodulation was reversible during continual exposure. Concentration-dependent eye damage and an elevated hepatic ethoxyresorufin-O-deethylase (EROD) activity were seen at 3 and 10 µl creosote/litre (nominal).

Field observations on terrestrial organisms refer to fatal cases of suspected creosote poisoning in wildlife (black rhinoceros Diceros bicornis) and farm animals that had access mainly to freshly creosoted wood or creosote containers.

Creosote is a genotoxic carcinogen for which a threshold has not been identified. There is consistent evidence from human studies that creosote causes skin cancer, but the studies do not allow dose–response analysis.

A study examining skin carcinogenicity of two samples of coal tar creosote, with different BaP contents, and of BaP alone in mice showed a significant increase in the rate of formation of papillomas and squamous cell carcinomas at the site of application. However, other organs were not examined. A linear relationship was observed between tumour rate and the dose of BaP in the creosote solution applied to the skin. There was no evidence of a threshold for carcinogenic effects. Analysis of the dose–response relationship resulted in a slope factor of 4.9 × 10^–3 tumours/animal for a total dose of 1 µg BaP. In this study, based on its BaP content, creosote appeared to be about 5 times more carcinogenic than a solution of BaP alone.

The human monitoring data concerning this type of exposure are limited; therefore, a sample risk assessment was not included here.

Creosote has been measured in air, water, soil, sediment, and biota. The fate of creosote components is largely dependent on the physicochemical properties of the components, matrix properties, the presence of degrading or accumulating organisms, and environmental conditions. Creosote may pose a significant risk to biota encountering spills or loading events. Laboratory studies have shown toxicity of creosote to aquatic and terrestrial organisms, while field studies have also demonstrated adverse effects following exposure to creosote. To date, it is not clear which creosote components may serve as indicators of environmental creosote contamination and toxicity.

This document is on coal tar creosote. Wood creosote is a different product that is used mainly in pharmaceutical preparations and is not covered in this document.

Coal tar creosote is a brownish-black/yellowish-dark green oily liquid with a characteristic sharp odour, obtained by the fractional distillation of crude coal tars (see Figure 1). The approximate distillation range is 200–400 °C (ITC, 1990). Table 1 provides some of the physical properties of creosote.

Table 1: Physical properties of creosote.^a

^a From ITC (1990); von Burg & Stout (1992).

Fig. 1: The formation of creosote from coal tar distillation.

The chemical composition of creosotes is influenced by the origin of the coal and also by the nature of the distilling process; as a result, the creosote components are rarely consistent in their type and concentration. Therefore, the trade names and/or the manufacturers of the creosotes mentioned in this document have been specified, wherever possible.

The legislation concerning the composition of creosote varies from country to country. Creosotes used in wood preservation are classified according to national/ international standards in terms of specifications — e.g., American Wood-Preservers’ Association (AWPA) standards P1 and P2 and the Western European Institute for Wood Preservation creosote grades A, B, and C (see Table 2). For example, before 1994, creosote could contain up to 20% phenolic compounds; in 1994, however, this was limited to 3% (EC, 1994). Further, in recent years, legislation in many countries has required that the BaP content of creosote be reduced. The EU (European Committee for Standardization, 2000) has recently finalized a new standard on classification and methods of testing for creosotes. European industry uses only creosote grades B and C with a BaP content lower than 50 mg/kg (0.005 weight %) and, for grade C, also lower volatile compounds (European Committee for Standardization, 2000).

Table 2: Specifications for creosote (according to the Western European Institute for Wood Preservation).^a

Creosote is a mixture of several hundred, probably a thousand, chemicals, but only a limited number of them — less than 20% — are present in amounts greater than 1% (Lorenz & Gjovik, 1972; Nylund et al., 1992). The chemical structures of some of these constituents are given in Figure 2.

Fig. 2: Chemical structures of some creosote constituents.

There are six major classes of compounds in creosote (Willeitner & Dieter, 1984; US EPA, 1987) (see Table 3):

• aromatic hydrocarbons, including PAHs, alkylated PAHs (non-heterocyclic PAHs can constitute up to 90% of creosote by weight), and BTEX;
• tar acids / phenolics, including phenols, cresols, xylenols, and naphthols (tar acids, 1–3 weight %; phenolics, 2–17 weight %; Bedient et al., 1984);
• tar bases / nitrogen-containing heterocycles, including pyridines, quinolines, benzoquinolines, acridines, indolines, and carbazoles (tar bases, 1–3 weight %; nitrogen-containing heterocycles, 4.4–8.2 weight %; Heikkilä, 2001);
• aromatic amines, such as aniline, aminonaphthalenes, diphenyl amines, aminofluorenes, and aminophenanthrenes (Wright et al., 1985), as well as cyano-PAHs, benzacridine, and its methyl-substituted congeners (Motohashi et al., 1991);
• sulfur-containing heterocycles, including benzothiophenes and their derivatives (1–3 weight %); and
• oxygen-containing heterocycles, including dibenzofurans (5–7.5 weight %).

Table 3: Reported chemical analyses of some coal tar creosotes.^a,b

Nylund et al. (1992) analysed four different creosotes from Poland, Germany, Denmark, and the former Soviet Union. They could identify about 85 components (96–98% of the total amount). The main components in all four creosotes were naphthalene and its alkyl derivatives, phenanthrene, fluorene, acenaphthene, alkylphenols, and dibenzofuran. The sum concentration of PAHs with 2–3 aromatic rings was about the same in all four analysed creosotes, but there were considerable differences (1.3–8.6%) in the content of PAHs with >4 aromatic rings (Nylund et al., 1992). An analysis of the PAHs in a mixture of German and Polish creosotes showed that phenanthrene was by far the predominant PAH, followed by naphthalene (Lehto et al., 2000).

Table 3 gives the results of some coal tar creosote analyses, and Table 4 lists the PAH content of some coal tar creosotes used in regulatory monitoring and in some environmental/toxicological studies. Many of the earlier studies on creosote concentrated on PAHs, because they are known carcinogens and represent the largest chemical group in creosote itself. However, PAHs are not very soluble and have a high adsorption to particulate matter. More recent studies (e.g., Arvin & Flyvbjerg, 1992; Mueller et al., 1993; Middaugh et al., 1994a,b) have concentrated on the other compounds in creosote — e.g., BTEX, nitrogen-containing heterocycles, sulfur-containing heterocycles, or phenolics — that are more soluble in water (see Table 5) and are found at a much higher percentage in leachate, contaminated water, soil, and sediment (see section 5). Heterocyclic compounds constitute about 5% of creosote compounds; however, due to their relatively high solubility and weak sorption, they can amount to 35–40% of the water-soluble fraction of creosote and are therefore potential groundwater contaminants (Licht et al., 1996). The mixture profiles of creosote-associated chemicals found in the environment (see section 5) are quite different from those in Table 3.

Table 4: Priority PAHs and concentration of PAHs in some creosotes used in environmental/toxicological studies.

Table 5: Physical properties of some components of creosote.

Creosote formulations can contain, for example, petroleum oils (Fowler et al., 1994). For some wood preservation uses, creosote is mixed 1:1 with fuel oil (Hoffman & Hrudey, 1990). In order to increase anti-microbial efficacy, creosote has been mixed with "topped" coal tar (i.e., CTPV) (Todd & Timbie, 1983).

2.2.1 Vapour pressure

The vapour pressure of creosote is variable because of the number of compounds involved and is difficult to characterize. Vapour pressures for individual components range from, for example, 12 700 Pa for benzene to 2.0 × 10^–10 Pa for the PAH dibenzo[a,h]anthracene (see Table 5).

2.2.2 Solubility and K_ow values

Creosote itself is given as immiscible with water (US EPA, 1984a) or slightly soluble (von Burg & Stout, 1992). The individual components of creosote have very differing solubilities (see Table 5). The aqueous solubility and the mobility of PAHs in, for example, groundwater systems decrease as the molecular mass increases. The PAHs with three or more aromatic rings have a solubility of less than 1 mg/litre, whereas the solubilities of BTEX, phenols, and nitrogen-, sulfur-, and oxygen-containing heterocycles (NSO compounds) are orders of magnitude higher.

The aqueous solubilities given for individual chemicals are usually given as solid solubilities if the chemicals are a solid at ambient temperatures. In creosote, however, these compounds are present in liquid form. Liquid solubilities are always greater than solid solubilities, the differences increasing in proportion to their boiling points; for the compounds found in creosote, their liquid solubilities are 3–240 times greater than their solid solubilities (Raven & Beck, 1992). The few data found on liquid solubilities are given in Table 6. When liquids are present in a mixture, the properties of an individual component in the mixture vary from those of the pure component. Furthermore, as dissolution proceeds, the composition of the non-aqueous phase will change (Mackay et al., 1991). The term "effective solubility" is used to describe the solubility of a particular component in a complex mixture. As dissolution of creosote proceeds, the more soluble components will be rapidly lost, causing the mole fraction and therefore the effective solubilities of the other constituents to increase. Effective solubilities of the 10 US EPA priority PAHs in creosote are given in Table 6.

Table 6: Differences in aqueous solid and liquid solubilities for 10 US EPA priority PAHs in creosote where data were available, together with their effective solubilities.^a

Log octanol/water partition coefficient (K_ow) values for PAHs range from 3 to about 7. Other components of creosote have widely varying log K_ow values, from 0.65 for pyridine (Leo et al., 1971) to 4 for biphenyl and dibenzofuran (see Table 5). The log organic carbon sorption coefficient (K_oc) values for PAHs range from 2.4 to 7.0 (IPCS, 1998). Some experimental log tar/water partition coefficient (K_tw) values derived from partitioning studies of coal tar constituents were found to be reasonably comparable with the respective log K_ow cited from other studies (Rostad et al., 1985; see Table 5).

2.2.3 Other physical/chemical properties

Creosoted timber has a low electrical conductivity, which is recognized in the use of creosote-impregnated poles for electrical power transmission and for sleepers (railroad ties) where track signalling is practised (ITC, 1990).

Corrosive effects on a range of metals were slight: for example, liquid creosote on mild steel produced a weight loss of 2.3 µg/dm² per day, whereas creosoted timber produced a weight loss of 27 µg/dm² per day. Natural rubber, neoprene, polyvinyl chloride (PVC), and polythene were significantly affected by creosote, whereas other substances, such as PTFE and polypropylene, were least affected (ITC, 1990).

The ignition temperature for creosoted timber is 50–100 °C higher than that of untreated timber (ITC, 1990).

Thermal decomposition of creosote and wood treated with creosote at 400 °C, 600 °C, and 800 °C produced in the condensate the same PAHs present in the original substance. Up to about 400 °C, the substance distilled off; between 400 °C and about 545 °C, creosote was oxidatively decomposed. In addition, at about 400 °C, treated wood showed oxidative decomposition to aldehydes, ketones, and phenolic compounds. Experiments to detect polychlorinated dibenzodioxins (PCDDs) and dibenzofurans (PCDFs) showed raised levels of these compounds compared with blanks; however, due to the small number of samples, the difference was not significant (Becker, 1997).

The analysis of creosote — a mixture of hundreds of chemicals — is very complex. The presence of creosote is confirmed by the profiling analysis of its components. Different profiles of creosote chemicals are found in the different matrices: the most volatile are found in air, the most soluble in water, and those with greater sorptive capacity in sediment/soil (see section 5; see also Hale & Aneiro, 1997). Depending upon the matrix (e.g., air, water, soil/sediment, biological materials) from which the sample is taken, suitable cleanup and extraction are necessary (see sections below). HRGC with FID, HRGC-MS, or reversed-phase HPLC with FL have been the separation and determination methods most commonly used. Thin-layer chromatography (TLC)-FID can supplement methods such as gas chromatography (GC)-FID and GC-MS through its ability to quantify the polar and high-boiling fractions (Breedveld & Karlsen, 2000).

Most analytical efforts have concentrated on the PAHs, the dominant components of coal tar and coal tar creosote. However, a number of constituents, notably oxygen- and nitrogen-containing heterocycles, which exhibit appreciable solubilities, have been identified as major contributors to the acute toxicity of creosote leachates, and recent studies have aimed at analysis of these compounds.

2.3.1 "Pure" (undiluted) creosote

First attempts to analyse creosote were by fractional distillation, but the process is tedious, and fractions overlap considerably. Lorenz & Gjovik (1972) used GC to analyse creosote both quantitatively, by a simulated fractional distillation, and qualitatively (see Table 3).

Samples of creosote can be separated into chemical class fractions (aliphatic hydrocarbons, neutral aromatic hydrocarbons, sulfur- and oxygen-containing aromatic compounds, and nitrogen- and hydroxyl group-containing aromatic compounds) using adsorption column chromatography with neutral alumina by elution with hexane, benzene, chloroform, and tetrahydrofuran/ ethanol, according to the method of Later et al. (1981).

Wright et al. (1985) isolated the PAHs and nitrogen-containing polycyclic aromatic compounds (NPACs) from creosote and separated these further using adsorption column chromatography with silicic acid using hexane:benzene, benzene, and benzene:ethyl ether eluents to isolate carbazole, amino-substituted enriched, and azaarene subfractions. Comparative quantitative chemical analysis of the PAH and NPAC fractions was achieved by HRGC using GC-FID. Before analysis of the amino-substituted enriched fraction, the amino-PAHs were selectively derivatized using pentafluoroproprionic anhydride. Over 30 PAHs and over 20 NPACs were identified (Wright et al., 1985; see Table 3).

The composition of creosotes from four producers was characterized by Nylund et al. (1992). The creosote samples were fractionated according to the method of Later et al. (1981) into four chemical classes. The creosotes and chemical fractions were analysed and their components identified by HRGC with FID, with MS, or with an alkali thermoionization detector (Nylund et al., 1992). In addition to the HRGC analysis, the 13 PAHs in the creosotes and in the distilled fractions boiling above 240 ºC were analysed with HPLC with FL. About 85 components were identified, some of which are listed in Table 3.

For the analysis of PAHs in creosote, the sample was dissolved in acetone, then injected into a GC-mass selective detector (GC-MSD) (Priddle & MacQuarrie, 1994).

Benz[c]acridine in creosote has been determined in silica-alumina column chromatography enriched fractions using GC with nitrogen-specific detection and GS-MS (Motohashi et al., 1991).

2.3.2 Air monitoring

Constituents of creosote can appear both in the gaseous phase and/or on particles in air. Mass equilibrium depends, for example, on vapour pressure and adsorptive affinity of a compound for particles. Also, high airflow over the filter increases evaporation from particles to the vapour phase. High sampling velocities are generally used in environmental air monitoring in contrast to occupational exposure monitoring, potentially resulting in different ratios of compounds in the vapour phase to compounds on particles.

2.3.2.1 Creosote vapours

A sample of creosote was heated to 60 °C in a chamber, and evaporated constituents were collected simultaneously into absorption solution (toluene), on silica (desorbed with diethyl ether), on activated charcoal (desorbed with carbon disulfide), and on XAD-2 resin (desorbed with diethyl ether). The samples were analysed by means of HRGC-MS, and 28 compounds were identified. With activated charcoal, many components were lacking. Of the four sampling methods, XAD-2 was selected for further testing. The main components were analysed by GC with FID. The recovery of the tested main components ranged from 82 to 102%. The detection limit was about 1–5 µg/sample, corresponding to 0.01–0.05 mg/m³ for an air volume of 100 litres (Heikkilä et al., 1987). The 12 main components were phenol, cresols, xylenols, methyl styrene, indene, naphthalene, biphenyl, dibenzofuran, benzothiophene, quinoline, isoquinoline, and fluorene.

2.3.2.2 Occupational air monitoring

The concentration of airborne particles originating from coal tar or coal tar pitch has been monitored as CTPV. This is also known as benzene-soluble matter (BSM) or cyclohexane-soluble matter (CSM). The method has also been applied to creosote fumes (Markel et al., 1977; Todd & Timbie, 1983). However, the precision of the CTPV method with glass fibre filters was very poor when tested with creosote fumes (Todd & Timbie, 1983).

The US National Institute for Occupational Safety and Health (NIOSH) recommended that CTPV be collected on PTFE filters. Although this method has been withdrawn, NIOSH Method 5042 is fundamentally the same procedure and can be used (NIOSH, 1998). Air samples are collected by drawing a known volume of air through the sampler. The filters are analysed by extracting with benzene and gravimetrically determining the BSM. The US Occupational Safety and Health Administration (OSHA) Method 58 is also used to collect CTPV, but on glass fibre filters (OSHA, 1986). The glass fibre filters are analysed by extracting with benzene and gravimetrically determining the BSM of half of the extract. If the BSM exceeds the permissible exposure limit, the remaining extract is analysed by reversed-phase HPLC with UV-FL detection for select PAH determination. Borak et al. (2002) sampled creosote particles and vapours on closed-faced cassettes containing PTFE filters connected in a series with XAD-2 sorbent tubes. The filters and sorbent tubes were extracted with benzene to determine the BSM. The extracts were redissolved in acetonitrile, and 16 PAHs were analysed by HPLC with UV detection (Borak et al., 2002). Because Borak et al. (2002) had to evaporate the benzene to determine the BSM, some of the lower-molecular-weight PAHs may have been lost; hence, these determinations may be an underestimate of the actual exposure. The results showed that the BSM method was not sensitive enough to reliably measure low concentrations of creosote fumes.

Creosote vapours were sampled on XAD-2 and analysed with HRGC and FID, and particulate PAHs on prewashed glass fibre filters were extracted and analysed by HPLC with FL (Heikkilä et al., 1987).

Using GC with atomic emission detection, Becker et al. (1999) measured thiarene (sulfur-containing PAHs; 0.4–19 µg/m³) in the personal air space of workers in electrolysis halls of an aluminium reduction plant who were exposed to CTPV.

For details on the use of biological monitoring to monitor occupational exposure of creosote workers, see sections 2.3.8 and 5.3.

2.3.3 Water samples

The water-soluble fraction of creosote is an extremely complicated mixture of low-molecular-weight aromatic compounds. When creosote comes into contact with water, the low-molecular-weight aromatics — the water-soluble fraction — are enriched in the aqueous phase. The fraction of phenols increases to about 45%, the NSO fraction increases to 38%, and the fraction of PAHs decreases to 17%. The dominant compounds in the water-soluble fraction are phenolics (phenol, mono- and dimethylphenols), benzene and alkylated benzenes, low-molecular-weight PAHs, and heterocycles containing nitrogen, sulfur, and oxygen (Arvin & Flyvbjerg, 1992). Changing the pH of a sample, followed by extraction of the aqueous phase by liquid–liquid extraction or solid-phase extraction, is widely used. It is valuable in separating neutral (e.g., PAHs), basic (e.g., azaarenes), and acidic (e.g., phenolics) compounds (Turney & Goerlitz, 1990; Mueller et al., 1991a; Middaugh et al., 1994a; see Figure 3). Available methods for the extraction, purification, identification, and quantification of creosote-related constituents in the aquatic environment have been reviewed (Johansen et al., 1996, 1998; Hale & Aneiro, 1997).

Fig. 3: Analysis of creosote according to Middaugh et al. (1994a).

Using the scheme for extraction of groundwater samples according to Figure 3, the components were analysed (Middaugh et al., 1994a). The limit of detection for PAHs in creosote was 400 ng/ml. Creosote heterocycles in organic extracts were also detected with GC-FID, but with slightly different temperature conditions. The limit of detection for creosote heterocycles was 200 ng/ml (Middaugh et al., 1994a). Phenolic compounds were detected by GC-FID/electron capture detector (ECD). The limit of detection for creosote phenolics was 50 ng/ml.

Turney & Goerlitz (1990) used a method similar to that given in Figure 3. Water samples were first analysed with HPLC. If organic compounds were present, they were extracted at different pHs with dichloromethane (DCM) into three separate fractions: the neutral aromatics, the phenolic compounds, and the nitrogen-containing heterocycles. The three isolated fractions were then analysed qualitatively and quantitatively using GC-MS. Two different analyses on two different fused silica columns, one polar and the other non-polar, were required to resolve the complex mixtures. The GC-MS was supplemented by GC with FID where appropriate. Quinolinone and isoquinolinone are difficult to extract from water and decompose at the high temperature of GC, so they were analysed by HPLC only (Turney & Goerlitz, 1990; Bestari et al., 1998a,b).

Whereas some authors (e.g., Rostad et al., 1984; Mueller et al., 1991a,b,c; Middaugh et al., 1994a,b; Sved & Roberts, 1995) have analysed for all creosote components, others (see below) have concentrated on particular groups of compounds. BTEX compounds were analysed with a hexane liquid–liquid microextraction GC technique (Barbaro et al., 1992). Aromatic hydrocarbons and derivatized phenols were extracted from groundwater with pentane. Compounds present at concentrations higher than 30 µg/litre were identified by GC-MS and quantified by GC-FID. Detection limits were 0.01 mg/litre for aromatic hydrocarbons and 0.02–0.03 for the phenols (Flyvbjerg et al., 1993; Dyreborg & Arvin, 1994).

Analysis of PAHs in aqueous samples was carried out by extraction with DCM and GC-MSD (Priddle & MacQuarrie, 1994). Other investigators used HPLC with FL for analysis of PAHs (Schoor et al., 1991; Hattum et al., 1998; Karrow et al., 1999).

NSO compounds (thiophene, benzothiophene, benzofuran) were extracted with diethyl ether/pentane and the organic phase analysed with GC-FID (Dyreborg et al., 1996a; Licht et al., 1997). Johansen et al. (1996, 1997, 1998) used the classical liquid–liquid extraction with DCM at pH 8 followed by GC-MS using scan mode or selective ion monitoring (SIM), giving detection limits for most heterocyclic aromatic compounds containing nitrogen, sulfur, or oxygen of 0.05 µg/litre. Other creosote compounds were extracted with diethyl ether/pentane followed by GC-FID. HPLC was used to detect 2-hydroxyquinoline, 1-hydroxy-isoquinoline, and 4-methyl-2-hydroxyquinoline, giving detection limits of 10, 10, and 50 µg/litre, respectively (Johansen et al., 1996, 1997, 1998).

2.3.4 Sediment samples

Methods to extract coal tar and coal tar creosote from sediments include aqueous caustic reflux and sonification/blending. Many extraction procedures require initial drying, such as air or oven drying, lyophilization, or use of a chemical desiccant (Hale & Aneiro, 1997). Other methods include mechanical shaking with acetone followed by petroleum ether (Hattum et al., 1998), DCM, and acetone/hexane; and evaporation to dryness, the residue being subsequently dissolved in hexane (Hyötyläinen & Oikari, 1999a). The methods for the extraction and fractionation of creosote from soil and sediment are similar to those given in Figure 3. However, Mueller et al. (1991b,c) used a slightly different fractionation, whereby the oxygen- and sulfur-containing heterocycles are extracted in the organic phase with PAHs, leaving the nitrogen-containing heterocycles in a fraction of their own.

2.3.5 Soil samples

Wet soil was extracted in acetone, then centrifuged to remove solids. The supernatant was evaporated and resuspended in DCM, dried with anhydrous sodium sulfate, and then analysed by GC-FID. Phenols were quantified colorimetrically using the 4-aminoantipyrene method after extraction in alkaline methanol solution. Oil hydrocarbons were determined using infrared spectroscopy after extraction with 1,2-trichlorotrifluoroethane (Ellis et al., 1991). Breedveld & Karlsen (2000) likewise used GC-FID to determine the PAH content in dried soil samples.

Soil samples were collected and analysed for PAHs by US Environmental Protection Agency (EPA) Methods 3550 and 8310 (HPLC with UV-FL) and for total phenols using American Public Health Association (APHA) method APHA 5530 C (Guerin, 1999).

Hydrocarbons (mainly PAHs) were extracted from heavily contaminated soil using headspace solid-phase microextraction, and analysis was done by GC-MS (Eriksson et al., 2000, 2001).

2.3.6 Wood samples

Samples of creosoted wood were taken and dried at room temperature and Soxhlet-extracted with DCM, then washed with sulfuric acid and then sodium hydroxide. After drying with sodium sulfate, the solvent was changed to cyclohexane for GC-MSD determination of the PAHs (Gurprasad et al., 1995).

Creosote oil content in wood and soil samples was determined by extraction with a toluene/xylene mixture or Soxhlet extraction using toluene. Compound identification was by GC-MSD, and quantitative analysis was by GC-FID (Becker et al., 2001). Mostly PAHs and nitrogen-containing heterocycles were found. The same method was used by Bergqvist & Holmroos (1994).

2.3.7 Biological materials

Freshwater isopods were blotted dry, homogenized with anhydrous sodium sulfate, and extracted with n-hexane with micro-Soxhlet. Analysis of PAHs was by HPLC (Hattum et al., 1998).

Trout hepatic tissue was ground with anhydrous sodium sulfate and extracted with DCM. Gel permeation was used to remove the lipids and other hydrophobic co-extractives. The second gel permeation chromatographic fraction containing PAHs of interest was cleaned up by solid-phase extraction (SPE) with Florisil. Analysis was by GC-FID, with limits of detection of 2–7 ng/g lipid (Whyte et al., 2000).

Wet sediment or minced tissues (fish) were digested in boiling ethanol/potassium hydroxide. Following sample digestion, hydrocarbons were partitioned into cyclohexane by liquid–liquid extraction. The cyclohexane phase was concentrated and the PAH-containing fraction isolated by chromatography on Florisil. Analysis was by HPLC (Black et al., 1981).

Saponified snail tissue was extracted with isooctane/ dimethylsulfoxide (DMSO) and the resulting sample analysed for PAHs and nitrogen-, sulfur-, and oxygen-containing heterocycles using GC-MS (Rostad & Pereira, 1987).

Homogenized tissue was digested with lithium hydroxide and extracted with diethyl ether. The ether extracts were dried, concentrated, and fractionated on glass columns with activated silica gel. The saturated hydrocarbons were eluted with hexane; the aromatic hydrocarbons were eluted with DCM/hexane. Analysis was by GC-FID and GC-MS (DeLeon et al., 1988).

2.3.8 Biological monitoring

The urinary PAH metabolites 1-pyrenol (1-hydroxypyrene) and 1-naphthol (1-hydroxynaphthalene) have been used to attempt to monitor creosote PAH exposure (Heikkilä, 2001). The use of a single marker to control or monitor exposure to a mixture assumes that the components of the mixture do not act additively and that there are no toxicokinetic interactions between the members of the mixture (Viau, 2002). Experimental studies have shown that co-exposure to naphthalene does not alter the toxicokinetic profile of the urinary excretion of 1-pyrenol in the rat exposed to pyrene (Bouchard et al., 1998). In human volunteers, it was shown that dermal exposure to pure pyrene or to creosote produced the same toxicokinetic profile for the urinary excretion of 1-pyrenol (Viau & Vyskocil, 1995).

In workers of a coke oven plant, there was an excellent correlation between airborne pyrene and BaP; further, 1-pyrenol was correlated with both airborne pyrene and airborne PAHs, suggesting that 1-pyrenol could be used as a biomarker of exposure to carcinogenic PAHs (Kuljukka et al., 1996) (see also section 5.3).

2.3.8.1 1-Pyrenol

2.3.8.2 1-Naphthol

There are no natural sources of coal tar creosote.

3.2.1 Processes and production levels

3.2.1.1 Processes

3.2.1.2 Production levels

3.2.2 Uses

3.2.2.1 Wood uses

3.2.2.2 Non-wood uses

3.2.3 Release into the environment

There are a large number of creosote treatment facilities in the USA. Bennett et al. (1985) reported more than 4000, Mueller et al. (1989) reported over 700, and Fowler et al. (1994) reported 600 wood-preserving plants using almost half a million tonnes of creosote each year. Micklewright (1998) reported that there were 70 wood-preserving plants in the USA that treated wood with creosote. During pressure impregnation of wood products, excess free product may be released from the treated materials. Leaching of spilled wastes from these application sites has been common (e.g., Black, 1982; Borthwick & Patrick, 1982; Goerlitz et al., 1985; Malins et al., 1985; Merrill & Wade, 1985; Elder & Dresler, 1988; Mueller et al., 1989; Pollard et al., 1994; Pereira et al., 1987). Large companies either treat aqueous wastes in their own biological treatment plants or discharge these wastes into municipal systems that receive biological treatment.

Seventy-seven large handlers of coal tar creosote in the USA report that 97% (500 000 kg) of the creosote released to the environment from their facilities is emitted through air (TRI97, 2000).

The behaviour of creosote in the environment depends upon the physical and chemical properties of its components. Most of the information available refers to creosote PAHs, but there are some data on the heterocyclic and phenolic components.

After input to the environment, creosote is "weathered," a multifactor process involving evaporation, dissolution, adsorption to particulate matter, and photo-oxidation. These processes affect the various constituents to different degrees, depending on their physicochemical properties (and possible mutual interactions). The extent of reactions is influenced by weather conditions and other environmental factors.

Characteristic of creosote-impregnated wood products is "bleeding" or exudation of creosote. The exudate may evaporate, remain liquid, or harden into a semisolid state on the surface of the treated wood. Bleeding continues for many years and is enhanced on hot and sunny days (see ITC, 1990).

4.1.1 Air

Creosote constituents occur in the atmosphere in the gaseous and the particulate phases. The distribution between the two phases depends strongly on the vapour pressures of the different creosote constituents. According to Eisenreich et al. (1981), compounds with vapour pressures of >10^–5 kPa should exist predominantly in the vapour phase, and those having vapour pressures of <10^–9 kPa should exist predominantly in the particulate phase, although, in reality, most high-molecular-weight organics lie between these extremes.

Vapour pressures of individual components detected in creosote range from, for example, 12 700 Pa for benzene to 2.0 × 10^–10 Pa for dibenzo[a,h]anthracene (see Table 5). Generally, low-molecular-weight PAHs (e.g., naphthalene, anthracene, and phenanthrene) are mainly in the gas phase, and high-molecular-weight PAHs are mainly bound to particles. Phenolic compounds, including cresols, as well as the heterocyclic fraction tend to be in the vapour state (see section 2). However, it is not clear how the specific composition of creosote may modify the distribution behaviour of the individual components.

PAHs can be transported over long distances without significant degradation (IPCS, 1998), but this is not the case for phenolic compounds, due to rapid photochemical attack and rain scavenging (IPCS, 1995). There is no comprehensive evaluation of the atmospheric transport of heterocyclic compounds.

The volatilization of five PAHs (acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene) from creosote-treated wood has been determined using chambers in the laboratory. The total exposed surface area of wood (yellow pine, Pinus strobus) was 0.118 m² and was painted with approximately 120 ml of creosote (density: 0.9 g/ml). Rates of desorption followed first-order kinetics and were higher at 30 °C than at 4 °C. Mean sum PAH fluxes varied from 2.6 (±1.5) mg/m² treated wood per day at 4 °C to 29.5 (±6.1) mg/m² treated wood per day at 30 °C. Desorption half-lives ranged from 0.7 to 31 years at 4 °C and from 0.3 to 1 year at 30 °C for fluoranthene and acenaphthene, respectively (Gevao & Jones, 1998). Emissions of creosote compounds from treated wooden ties used in the Swiss railway network were calculated to be 1710 tonnes per year, corresponding to an emission factor of 208 mg/m² per day. Emission factors for volatile PAHs and phenolic compounds were calculated to be 20.3 and 0.58 mg/m² per day, respectively (Kohler et al., 2000).

4.1.2 Water and associated sediments

4.1.2.1 Volatilization from water

4.1.2.2 Distribution within aquatic systems

4.1.3 Soil

The same principal physicochemical properties as discussed in section 4.1.2 for the aquatic environment regulate the mobility of creosote (components) in soil, resulting in partitioning between soil organic matter, solid surfaces, soil water, residual oily phases, and gaseous phases.

4.1.3.1 Volatilization from soil

Loss of creosote components from soil by volatilization may be significant only for some low-molecular-weight compounds (with relatively high vapour pressures; see section 2) and for highly contaminated soils. Quantitative data obtained with creosote-contaminated soil as the source were not available.

4.1.3.2 Transport within soil

Depending on soil type, hydrogeology, amount of creosote released, etc., transport processes such as advection (movement with the bulk fluid), dispersion, adsorption, or decay contribute to a different degree to the movement of creosote in soil. Creosote components often persist for long times near their sources; on the other hand, some migration to groundwater and surface water has occurred. Some data have been summarized for wood products in service and for creosote waste sites, as follows.

1) Wooden creosoted constructions

Creosote components are slowly released from treated wood products (poles, railroad ties, etc.) by oil exudation, rainwater (or irrigation) leaching, and volatilization of the lighter fractions (Petrowitz & Becker, 1964, 1965; Bosshard, 1965; Henningsson, 1983; Nurmi, 1990; Behr & Baecker, 1994; Gurprasad et al., 1995; Gevao & Jones, 1998). Accordingly, soil samples around and under impregnated poles showed high contents of creosote (see section 5) after 10 years (up to 90 g/kg dry weight; Nurmi, 1990) or 40 years (up to 1.5 g/kg dry weight) (Bergqvist & Holmroos, 1994) of service. Similarly, PAHs have entered surrounding soil, ditches, and groundwater in a storage and use area of old railway ties; migration of PAHs in soil appeared to be slow (Sandell & Tuominen, 1996).

In a terrestrial microcosm study with wooden (pine) posts impregnated with creosote spiked with ¹⁴C-labelled phenanthrene and acenaphthene, the mass balance of the radioactive material within compartments was investigated 2.5 months after introduction of the posts in the experimental ecosystem; the major part of labelled phenanthrene/acenaphthene remained in the posts (95%/93.5%), and a small portion distributed to soil (2.7%/4.3%), air (1.4%/1%), and biota (0.9%/1.2%). The highest concentrations were found in areas immediately surrounding the posts. No radioactivity was detected below the first 10-cm soil layer or in leachate or groundwater (Gile et al., 1982).

2) Waste creosote sites

Waste creosote in soil can occur as lighter- and heavier-than-water fractions, or even as a free liquid pool. The light fraction (moving with fluctuating water levels) contains mainly the water-soluble components. The heavy fraction tends to travel downwards to impervious soil layers (which it can flow along), in this way sometimes reaching groundwater and surface waters (CEPA, 1993; see also section 4.1.2). Near old wood-preserving facilities, creosote residues have been found to persist in soil for many years (e.g., Sundström et al., 1986; Borazjani et al., 1990; Acharya & Ives, 1994; see also section 5). At deeper soil layers (about 0.5–1.5 m), the composition of residues was sometimes similar to that of the original creosote product, whereas in the surface layer (0–20 cm), many low-molecular-weight compounds were missing (KEMI, 1995). At greater depths (4–5 m), mainly two- and three-ring PAHs could be observed (Breedveld & Karlsen, 2000).

In a laboratory model experiment (simulating a creosote spill), the leaching of six aromatic compounds from a saturated sand column contaminated artificially with creosote (116 g mixed with 2.56 kg sand) was studied over 36 days. The compounds studied (benzene, toluene, o-xylene, phenol, o-cresol, and naphthalene) accounted for 21.8% (w/w) of the creosote (no further specification). Phenol and o-cresol were totally leached from the contaminated sand within the first 5 days, followed by benzene within 10 days and toluene after 36 days. Naphthalene and o-xylene were not totally leached when the experiment ended. Some loss by evaporation was indicated by a typical creosote smell before the experiment started (Dyreborg & Arvin, 1994). Another soil column leaching test with soil (heterogenic composition) from a creosote-contaminated area showed that about 9% of the creosote (no specification) could be leached out (Ellis et al., 1991). Broholm and co-workers studied the transport of 25 organic compounds typical of creosote (monocyclic aromatic hydrocarbons [MAHs], PAHs, heterocyclic aromatic compounds [HACs], phenolic compounds) in a large (0.5 m in height, 0.5 m in diameter) macroporous clayey till column (139 days, biodegradation prevented by sodium azide). Results showed that the transport of low-molecular-weight compounds was not retarded relative to bromide, whereas the transport of the high-molecular-weight organic compounds was retarded significantly. The following order (with increasing retardation) was observed: benzene, pyrrole, toluene, o-xylene, p-xylene, ethylbenzene, phenol, benzothiophene, benzofuran < naphthalene < 1-methylpyrrole < 1-methylnaphthalene, indole, o-cresol, quinoline < 3,5-dimethylphenol, 2,4-dimethylphenol < acridine < carbazole < 2-methylquinoline < fluorene < dibenzofuran < phenanthrene, dibenzothiophene. This order was unexpected based on the octanol/water partition coefficients of the organic compounds (Broholm et al., 1999a). Transport through a column of fractured clayey till resulted in a comparable order of retardation (Broholm et al., 1999b). A fast downward migration of selected compounds (naphthalene, 1-methylnaphthalene, toluene, phenol, dimethylphenols, o-cresol, benzothiophene, quinoline) — coupled with some attenuation, probably due to biodegradation — was also observed in a clayey till field experiment (Broholm et al., 2000).

4.1.4 Biota

Individual creosote components are bioavailable to widely varying degrees, as seen for PAHs (e.g., Hattum et al., 1998; IPCS, 1998), heterocyclic compounds (e.g., Southworth et al., 1978, 1980; Eastmond et al., 1984), and phenolic compounds (e.g., IPCS, 1995).

Most data in connection with creosote exposure refer to PAHs.

A limited uptake of PAHs in plants (ryegrass, Lolium perenne L.) has been observed in a terrestrial microcosm study with creosote-treated wood as initial creosote source. After 2.5 months, 0.1% of the applied ¹⁴C-labelled phenanthrene and 0.04% of the acenaphthene were detected in plant tissue (Gile et al., 1982).

Measurements of the possible adsorption of PAHs to or deposition of PAHs onto roots or leaves of plants near creosote sources are not available. However, in general, PAHs are significantly subjected to both wet and dry deposition onto plant surfaces, with little transport to inner plant tissues possible (e.g., Kipopoulou et al., 1999; Howsam et al., 2000, 2001). Adsorption of PAHs to roots of terrestrial plants occurs, but translocation to shoots seems to be limited (e.g., Binet et al., 2000). Terrestrial soil PAHs are probably less available to terrestrial root systems than are aquatic sediment PAHs to aquatic rooted plants (e.g., McGlynn & Livingston, 1997).

Animals (earthworms: Lumbricus spp., pill bugs: Armadillarium and Porcellia spp., mealworm larvae: Tenebrio molitor, gray crickets: Acheta domesticus, garden snails: Helix pomata, and a vole: Microtus canicaudus) kept in the terrestrial microcosm mentioned above were found to have taken up 0.8% and 1.2% of the labelled phenanthrene and acenaphthene, respectively, applied to wood. Soil-dwelling and litter-feeding species (pill bugs and earthworms) showed higher concentrations than crickets or snails residing above-ground and feeding on vegetation. The vole had high concentrations of ¹⁴C in the gastrointestinal tract, but also in the brain, suggesting some systemic intake (Gile et al., 1982; see also section 4.3.2).

Field observations also indicated uptake of creosote-derived PAHs by aquatic organisms. Elevated levels (see section 5.1.6) have been detected, for example, in molluscs and crustaceans taken from creosote-treated pilings (Shimkin et al., 1951; Dunn & Stich, 1976) and in (mostly sediment-associated) invertebrates and fish captured from creosote-contaminated areas in fresh waters (Black et al., 1981; DeLeon et al., 1988; Pastorok et al., 1994) and marine environments (Zitko, 1975; Malins et al., 1985; Rostad & Pereira, 1987; Elder & Dresler, 1988).

Relocation experiments with oysters and clams and laboratory studies with fish confirmed the uptake (and accumulation) of PAHs following creosote exposure (see section 4.3.1). An increase in PAH bile metabolites has been detected in fish (rainbow trout, Oncorhynchus mykiss) from a 28-day creosote microcosm study, indicating some uptake (and metabolism) (Karrow et al., 1999). Extracts of oysters (Crassostrea virginica) exposed to the water-soluble fraction of creosote-contaminated sediment and the corresponding sediment extracts showed similar profiles of organic compounds (Hale & Aneiro, 1997).

A transfer primarily of the lipophilic creosote constituents (or their metabolites) to the human food supply is possible and has been documented in some cases (e.g., for shellfish and fish; see sections 5.1.4 and 5.1.6). Some transfer to livestock is also likely via contact of farm animals with creosote-treated farm buildings or fences or when creosote is used as a general disinfectant (Oehme & Barrett, 1986); however, measurements of tissue burdens of animals affected (see section 9) were not available.

4.2.1 Biodegradation/biotransformation

Compared with the many studies with individual components of creosote, relatively little is known of the biotransformation and biodegradation of creosote constituents when they are present in the creosote mixture.

4.2.1.1 Microbial organisms

Creosote is not easily degraded by microorganisms (see Tables 7 and 8), consistent with its use as a wood preservative (see section 3) and with field monitoring results (see section 5).

Table 7: Survey on laboratory investigations of aerobic degradation of creosote (under simulated conditions as expected in situ).^a

Table 8: Survey on laboratory investigations of anaerobic degradation of creosote (under simulated conditions as expected in situ).^a

Table 9: Groundwater contamination by creosote: sum concentrations.^a

Increased concentrations of methane found in some creosote-contaminated aquifers suggested that some anaerobic degradation was occurring (e.g., Ehrlich et al., 1982; Goerlitz et al., 1985). Products of aerobic degradation have also been identified at some creosote-contaminated sites (e.g., Pereira et al., 1988; Johansen et al., 1998).

The extent of microbial degradation of creosote is difficult to determine because of the large number of chemicals present in the original or environmentally fractionated mixtures and the large variability in concentration ratios.

The individual creosote constituents together cover a wide range of microbial degradability or recalcitrance. For details, see reviews or other publications on MAHs (e.g., Barker et al., 1987; Barbaro et al., 1992; Rippen, 1999), PAHs (e.g., Cerniglia & Heitkamp, 1989; Cerniglia, 1992; Mueller et al., 1996; IPCS, 1998; Juhasz & Naidu, 2000; Kanaly & Harayama, 2000), HACs (e.g., Grbic-Galic, 1989; Kuhn & Suflita, 1989; Dyreborg et al., 1996a; Licht et al., 1996; Bianchi et al., 1997; Bressler et al., 1998; Fetzner, 1998), or phenolic compounds (e.g., Arvin et al., 1991; Nielsen & Christensen, 1994; IPCS, 1995).

However, extrapolations from studies with single substances are of limited value, because many interactions between the individual components are possible. Most interactions observed were degradation inhibiting, but there were also promoting effects on the degradation of co-substances in a few cases (Arvin et al., 1988, 1989; Bouchez et al., 1995; Millette et al., 1995, 1998; Dyreborg et al., 1996b,c; Lantz et al., 1997; Broholm et al., 1999b; Lotfabad & Gray, 2002).

Therefore, in this context, mainly studies using simple or complex mixtures of creosote constituents as test material (i.e., artificial, environmental, original mixtures) have been considered. A survey of the results of bacterial aerobic (Keck et al., 1989; Thomas et al., 1989; Mueller et al., 1991a; Evanshen et al., 1992; Dyreborg et al., 1997; Mohammed et al., 1998; Broholm et al., 1999b; Lehto et al., 2000) or anaerobic (Godsy et al., 1992; Flyvbjerg et al., 1993; Dyreborg et al., 1997; Sharak Genthner et al., 1997) degradation studies modelling (more or less sophisticated) natural in situ conditions is given in Tables 7 and 8, respectively. Despite the great complexity of creosote degradation processes, some general trends can be observed. The majority of components are not completely degradable under simulated natural conditions, even with (in situ adapted) inocula from creosote-contaminated sites. Aerobic degradation proceeds more quickly than anaerobic degradation (e.g., Dyreborg et al., 1997). Phenolic compounds are relatively easily degraded. Within PAHs, degradability appears to be inversely related to the number of aromatic rings. Some HACs were quickly degraded or disappeared (e.g., quinoline), whereas others (e.g., pyrrole) were rather recalcitrant. Most studies monitored only the disappearance of the compounds, so it is often not clear if there was biotransformation rather than complete mineralization.

Besides structural features of the chemicals, a series of other factors influence the degradation/transformation of creosote constituents in situ — for example, bioavailability (e.g., with respect to sorption phenomena, trapping in pore water, etc.), initial concentration, and nutrient or oxygen supply (Fetzner, 1998; Johansen et al., 1998; Broholm et al., 1999b; Breedveld & Sparrevik, 2000; Juhasz et al., 2000a). It is assumed that in typical creosote-contaminated groundwater compartments, the oxygen concentration will often not be sufficient for complete biodegradation (Lee & Ward, 1985; Wilson et al., 1986; Broholm et al., 1999b). Biodegradation rates of some PAHs have been found to be enhanced transiently by pre-irradiation (Lehto et al., 2000). Microbial adaptation to some creosote compounds also occurred (Wilson et al., 1986).

There is a great variety in microbial metabolic pathways of creosote compounds, but they all involve the incorporation of oxygen into the ring structure, ring cleavage, and the production of intermediates with specific breakdown patterns (Gibson & Subramanian, 1984; Pereira et al., 1987, 1988; Arvin et al., 1988; Miller & Comalander, 1988; Wilson & Jones, 1993; Chapman et al., 1995; Mueller et al., 1996; Licht et al., 1997; Fetzner, 1998; Johansen et al., 1998; Bressler & Fedorak, 2000). In some cases, the intermediates formed can be more persistent (mobile or toxic) than their parent compounds, as has been found, for example, for quinoline/quinolinone (Fetzner, 1998), acenaphthene / several acenaphthene oxidation products (Selifonov et al., 1998), and some other PAH metabolites (Singleton, 1994).

Because of the numerous sites with creosote contamination (mainly soils, groundwater), many efforts have been undertaken to develop useful remediation strategies. Generally, there are three basic approaches. One method involves the removal of contaminated soil and its treatment on-site (prepared bed, etc.) or in a slurry reactor under conditions optimal for microbial growth; inocula used can be native or specifically enriched (Mueller et al., 1989, 1991b,c; Borazjani et al., 1990; Ellis et al., 1991; Davis et al., 1993; Otte et al., 1994; Glaser & Lamar, 1995; Brooks et al., 1998; Guerin, 1999; Eriksson et al., 2000). In another technique used for treating groundwater, the contaminated water is pumped to the surface, where it can be treated aerobically and thereafter recycled (to nearby surface waters) (Mahaffey et al., 1989; Mueller et al., 1993; Middaugh et al., 1994b). The third method involves procedures to promote in situ biodegradation — for example, by adding nutrients, electron acceptors, adapted microorganisms, and sometimes surfactants or other materials, such as manure, straw, compost, or sewage sludge, to the soil (e.g., Ellis et al., 1991; Evanshen et al., 1992). Creosote-contaminated groundwaters have been treated in similar ways (Dust & Thompson, 1973). Frequently, a combination of several strategies is used for remediation — including physical or chemical methods such as encapsulating, washing (plus surfactants), and sorption (Tobia et al., 1994; Zapf-Gilje et al., 2001; Bates et al., 2002; Rasmussen et al., 2002).

In many cases, it has been possible to achieve significant reductions for certain substances. However, advantages seem to be limited. Primarily the high-molecular-weight PAHs continued to be recalcitrant in treated soil (e.g., Mueller et al., 1991b,c; Davis et al., 1993; Glaser & Lamar, 1995; Breedveld & Karlsen, 2000; Breedveld & Sparrevik, 2000). Despite the removal of a majority of creosote contaminants from groundwater through biotreatment, only a slight decrease in toxicity and teratogenicity (see section 9) of biotreated groundwater was observed (Mueller et al., 1991a). A study in which creosote-contaminated groundwater was treated with a sorption/bio-barrier (peat/sand barrier material under aerobic conditions) found trimethylphenols to be the most difficult to remove (Rasmussen et al., 2002). Some constraints connected with bioremediation of creosote-contaminated soils have been reviewed in detail (Wilson & Jones, 1993; Pollard et al., 1994; Alexander, 2000; Juhasz et al., 2000b; Reid et al., 2000). A recent study (Brooks et al., 1998) demonstrated that several treatments have been successful in reducing, for example, the total PAH concentration in contaminated soils, but actually increased the microbial mutagenicity of these soils; an analysis indicated that the mutagenic fraction contained azaarenes, a substance class that was often not included in monitoring programmes. Other experiments showed accumulations of PAH metabolites/ degradation products, such as 9H-fluorenone, 4-hydroxy-9H-fluorenone, 9,10-phenanthrenedione, and 4H-cyclopenta[def]phenanthrenone (Eriksson et al., 2000).

Bacteria involved in degradation of creosote components (some PAHs, HACs, phenolic compounds) have been isolated and identified as belonging mostly to the genera of Pseudomonas or Sphingomonas (Drisko & O’Neill, 1966; Ehrlich et al., 1983; Bennett et al., 1985; Rothenburger & Atlas, 1990; Chapman et al., 1995; Grifoll et al., 1995; Lantz et al., 1997; Selifonov et al., 1998; Eriksson et al., 2000; Leblond et al., 2001) and Mycobacterium (Grosser et al., 1995). Lignin-degrading fungi such as Phanerochaete sordida, P. chrysosporium, and Pleurotus ostreatus have also been found to be capable of transforming several creosote PAHs (Glaser, 1990; Davis et al., 1993; Glaser & Lamar, 1995; Eggen & Majcherczyk, 1998).

4.2.1.2 Other organisms

Little is known about the transformation of creosote by other organisms. Generally, the most striking feature seems to be that PAH components are transformed more rapidly in fish than in several invertebrate species (NRCC, 1983; IPCS, 1998). Mostly, these PAH metabolites are not routinely detected (Meador et al., 1995). The occurrence of PAH metabolites in fish after controlled creosote exposure has been reported only rarely (Karrow et al., 1999). Formation of PAH–DNA adducts in fish is addressed in section 6.6. Transformation results for mammals are available only from studies with laboratory mammals (see section 6).

4.2.2 Abiotic degradation

4.2.2.1 Photodegradation

Photochemical transformation seems to be the most important abiotic mechanism by which creosote constituents, such as PAHs, HACs, and phenolic compounds, are transformed in the atmosphere and, to a lesser extent, in water and soil. Indirect photolysis (photo-oxidation, involving peroxyl, hydroxyl, and other radicals) appears to prevail over direct photolysis. Half-lives measured (mostly in single-component studies with and without reference to creosote) varied widely (0.2 h –550 days), depending on test conditions and compounds (e.g., NRCC, 1983; IPCS, 1995 [review on cresols], 1998 [review on PAHs and HACs]). There are few data on transformation products. Several oxygenated compounds have been observed, as well as nitroPAHs and nitro-cresols (e.g., Andersson & Bobinger, 1992; Kochany & Maguire, 1994; IPCS, 1995, 1998).

Little information is available on photodegradation of creosote components when present in the creosote mixture. Irradiation (arc xenon lamp, aqueous media) of six PAHs separately or of these PAHs in creosote (initial PAH concentrations / duration: 0.61–3.1 µmol/litre / 5–30 min or 0.55–4.4 µmol/litre / 10 min, respectively) resulted in the following photodegradation rates (% separately / in mixture): naphthalene (57/47.6), acenaphthene (47/50), fluorene (48.4/29.3), phenanthrene (91.7/6.82), anthracene (83.6/29.1), and pyrene (38.3/8.64). With one exception (acenaphthene), there was a trend of decreased photoreactivity in the mixture compared with the individual tests; this was explained by the competition of light absorption in the presence of co-occurring compounds. The photochemical products detected by GC-MS seemed to be quinone derivatives of the original compounds (Lehto et al., 2000).

For the purpose of remediation of creosote- and PCP-contaminated waters, laboratory-scale experiments using the photo-Fenton reaction, which employs ferric ion (Fe³⁺), hydrogen peroxide, and UV light, have been conducted. Saturated aqueous solutions of creosote/PCP (American creosote-P2) were treated by the photo-assisted Fenton reaction (1 mmol Fe³⁺/litre; 10 mmol hydrogen peroxide/litre; black lamp UV light; pH: 2.75; 25 °C), and the disappearance of 36 identified creosote compounds and PCP was monitored during a 180-min reaction period. The following order of reactivity was found: two-ring PAHs > heterocyclics > phenolics (including PCP) > three-ring PAHs > four- and five-ring PAHs. Within 5 min, the concentrations of 18 of the 37 compounds declined to values at or below their detection limits; 13 compounds were at least 90% transformed, but six PAHs (phenanthrene, fluoranthene, 2,3-benzo[b]fluorene, chrysene, benzo[b]fluoranthene, and BaP) resisted to a greater extent. By 180 min, a more extensive transformation was observed, except for chrysene and BaP (showing 70–80% transformation). No new peaks were observed in MS chromatograms, and added ¹⁴C-phenanthrene and ¹⁴C-pyrene were mineralized by 93% and 35%, respectively; about 33% of the organic nitrogen was converted to inorganic nitrogen-containing compounds, accompanied by an undetermined yield of sulfate. Concomitantly, the acute toxicity of the treated solution to fish and water flea (see section 9) was reduced (Engwall et al., 1999). Another method using photocatalysis (near-UV irradiation in the presence of titanium dioxide) appeared to achieve total mineralization of creosote in water (separate phase; 100 or 360 mg creosote/litre, Armor Coat, commercial domestic grade, Canada), as demonstrated by changes in absorbance or reflectance and/or carbon dioxide evolution; intermediate products were not monitored (Serpone et al., 1994). A single-component photolysis study with BaP (using hydrogen peroxide and UV light and additional analytical methods) detected a series of polar BaP photoproducts, including methoxy, hydroxy, and dihydroxy isomers of BaP and even more polar compounds (Miller et al., 1988).

4.2.2.2 Hydrolysis

Abiotic hydrolysis is considered not to be a significant environmental degradation process for PAHs (IPCS, 1998) and phenolic compounds (IPCS, 1995). Similar conclusions may be valid for HACs.

4.3.1 Aquatic organisms

Field monitoring studies showed that aquatic organisms such as invertebrates (Shimkin et al., 1951; Zitko, 1975; Dunn & Stich, 1976; Black et al., 1981; Malins et al., 1985; Rostad & Pereira, 1987; DeLeon et al., 1988; Elder & Dresler, 1988) and fish (Black et al., 1981; Malins et al., 1985; Pastorok et al., 1994) living at creosote-contaminated sites have absorbed PAHs (and heterocyclic compounds) typical of creosote at concentrations above reference values (see section 5.1.4).

Characteristic differences between invertebrate and vertebrate species have been pointed out by the field study of Black et al. (1981). Insects and crayfish (Procambarus sp.) had much higher levels of phenanthrene, 1,2-benzanthracene, and BaP than most of the fish (brown trout, Salmo trutta; white sucker, Catostomus commersoni). An exception was lamprey (Lampetra sp.), which appeared to accumulate high levels of phenanthrene (a 3.5-fold increase compared with sediments). Generally, PAH profiles in insects and crayfish were close to that found in the sediments, whereas fish had greatly altered ratios for low/high-molecular-weight PAHs (see also section 5.1.4).

Relocation experiments with mollusc and crustacean species also indicated accumulation of PAHs. Oysters (Crassostrea virginica; total n = about 60) sampled from an industrially non-impacted site (Piatatank River, USA) were exposed in situ to PAH-contaminated sediments near a creosote facility on the Elizabeth River (Virginia, USA). Within 3 days of exposure, the concentrations of several measured PAHs (benz[a]anthracene/chrysene, benzofluoranthenes, BaP, fluoranthene, phenanthrene, pyrene) increased from non-detectable to as much as 11.7 mg/kg dry weight; they then stabilized during the 15-day observation period (Pittinger et al., 1985). Similarly, oysters (n = 5 per group) introduced for a 6-week period into an estuarine environment near a creosote contamination site (Pensacola, Florida, USA) showed a significant increase of phenanthrene, fluorene, and pyrene (diagram only) in soft tissues, whereas naphthalene was not accumulated. Minimum BCFs relative to sediments have been estimated to be in the range of 0.3–1.0 for phenanthrene and fluoranthene (Elder & Dresler, 1988). Clams (Rangia cuneata; n = 3–4 per group) translocated from a relatively pristine site to a bayou (Bayou Bonfouca, USA) flowing over a creosote spill site showed gradual increases of several PAHs in their shucked bodies over a 4-week period; the most pronounced increase was seen for benzopyrenes, from pre-exposure (background) levels of 87 µg/kg wet weight to 132 µg/kg wet weight after 2 weeks and 600 µg/kg wet weight after 4 weeks (DeLeon et al., 1988). Accumulation of PAHs in newly moulted and intermoult crustaceans (blue crab, Callinectes sapidus) has been examined in the Elizabeth River (USA) near a creosote spill site. Paired premoult and intermoult crabs (n = 9–12 per group) were placed in baskets for approximately 3 days (until ecdysis was completed). The mean total PAH concentrations (cyclopenta[def]phenanthrene, fluoranthene, pyrene) in hepatopancreas and muscle increased from non-detectable levels to significant burdens in both stages, with higher concentrations in newly moulted crabs (hepatopancreas/muscle: 9560/1380 µg/kg dry weight) than in intermoult crabs (3360/498 µg/kg dry weight). Low-molecular-weight compounds such as acenaphthene, dibenzofuran, fluorene, and phenanthrene were present in both control and exposed crabs (Mothershead & Hale, 1992).

Adult lobsters (Homarus americanus; n = n.sp.) exposed to creosote (no specification) at concentrations ranging from 0.3 to 2.5 mg/litre during a laboratory lethality test showed considerably higher values of "creosote" in their hepatopancreas than the control lobsters (3220–47 500 mg/kg lipid versus 670 mg/kg lipid; as indicated by fluorescence measurements). The residues increased with the test concentrations (0.3, 1.3, 2.5 mg/litre) as well as with exposure time (up to 120 h) at the 0.3 mg/litre exposure, which was the non-lethal concentration (McLeese & Metcalfe, 1979). Guppies (Poecilia reticulata) kept in the laboratory in aquaria that contained a sediment layer taken from a creosote-contaminated small drainage stream had significant residues of several PAHs (acenaphthene, anthracene, benz[a]anthracene, benzo[b]fluoranthene, benz[k]fluoranthene, BaP, chrysene, fluoranthene, phenanthrene, pyrene) in their tissues (composite samples of carcasses without liver; n = n.sp.) 43 days after exposure. Naphthalene and fluorene were not detected in the fish tissue, although they were present along with the other PAHs in sediment and water (Schoor et al., 1991). In a microcosm study, livers of rainbow trout (Oncorhynchus mykiss; n = 10) contained 4 of 16 individual PAHs scanned for, but there was no relationship to creosote dose after 28 days of exposure. This lack was thought to be partly due to rapid metabolism in fish prior to analytical detection (Whyte et al., 2000).

BCFs in connection with creosote exposure have rarely been determined; there is an estimate for oysters and sediment (see above: Elder & Dresler, 1988). Recently, biota–sediment accumulation factors (BSAFs) of several PAHs have been determined for a creosote-contaminated lake in Finland. Duck mussels (Anodonta anatina) were exposed to sediment for 10 months (1998–1999) by caging at four experimental sites. The BSAFs derived from concentrations of PAHs (acenaphthene, phenanthrene, anthracene, fluoranthene, pyrene, and benz[a]anthracene) in sediment (on an organic matter basis) and in duck mussel tissue (on a lipid weight basis) varied from 0.79 to 1.45. The highest BSAF (1.45) was calculated for benz[a]anthracene (Hyötyläinen et al., 2002). Values from other studies for some creosote constituents, such as PAHs and heterocyclic or phenolic compounds, have been compiled elsewhere (e.g., Lu et al., 1978; Southworth et al., 1978, 1980; Veith et al., 1980; Eastmond et al., 1984; Sundström, et al., 1986; IPCS, 1995, 1998). They vary over a wide range, depending on compound, aquatic species, and test conditions — for example, BCFs (organism/water in wet weight; IPCS, 1998) for naphthalene: 19.3–10 844 (measured in crustaceans), 2.2–320 (measured in fish); or for BaP: 458–73 000 (measured in crustaceans), 12.5–4900 (measured in fish). Generally, equilibrium concentration factors increased with increasing molecular weight (or with log K_ow) within a compound group. Some sulfur-containing heterocycles have been found to be bioconcentrated to a greater extent than their PAH counterparts — for example, naphthalene and benzo[b]thiophene showed BCFs of 50 and 750, respectively, in Daphnia magna (Eastmond et al., 1984). Accumulation may also be influenced by the presence of co-substances. Studies with anthracene showed that BCFs in fish (rainbow trout, Oncorhynchus mykiss) differed during single-compound and complex-mixture (oil shale retort water) exposures (Linder et al., 1985).

Biomagnification of creosote PAHs within aquatic food-chains appears to be limited as far as fish are involved, because vertebrates generally have a more effective metabolic capacity for these compounds than invertebrates (e.g., NRCC, 1983).

4.3.2 Terrestrial organisms

Only few data are available on bioaccumulation of creosote constituents in the terrestrial environment following creosote exposure.

In a microcosm experiment with creosote containing ¹⁴C-labelled acenaphthene and phenanthrene, an accumulation (ecological magnification) factor between a grey-tailed vole (Microtus canicaudus) and soil has been determined. The calculations (mg/kg of ¹⁴C equivalents in vole / mg/kg of ¹⁴C equivalents in soil) resulted in factors of 12 for phenanthrene and 31 for acenaphthene (Gile et al., 1982).

The ultimate fate of creosote components is largely dependent on the physicochemical properties of the components, matrix properties, the presence of degrading/accumulating organisms, and environmental conditions. Components may be distributed to the atmosphere (the more volatile fraction), leached to water and soil (compounds with high solubilities), with the potential for migration, or sorbed onto soil or sediment particles (compounds with high K_ow). Movement of sediment-sorbed creosote components may also occur through transport of colloidal material. Some creosote components are readily degradable via biotic (aerobic and anaerobic) and abiotic processes; however, many high-molecular-weight compounds are recalcitrant and may persist in the environment for decades. Degradation of creosote components often leads to the formation of transformation products (i.e., the compounds are not mineralized), which may be more toxic and mobile than the parent compounds. There is also the potential for marine and terrestrial organisms to bioaccumulate creosote components; however, this is dependent on the bioavailability of the compounds, the organisms’ mode of feed, and metabolism.

For high-molecular-weight PAHs, which are the most persistent creosote components, sediments and soils are the major environmental sinks (IPCS, 1998), consistent with creosote-related findings discussed previously (sections 4.1–4.3). However, the possibility for redistribution processes should be noted. Additionally, groundwater is an important sink for many creosote components (see section 5).

There are only a few specific measurements of the thermal decomposition of creosote or creosoted wood (Marutzky, 1990; Becker, 1997; see also section 2). For example, combustion of creosoted wood in a small incinerator resulted in elevated emission values (carbon monoxide, nitrogen oxides, hydrocarbons) compared with untreated wood (Marutzky, 1990). Analyses for PCDDs/PCDFs in a laboratory-scale combustion experiment gave (preliminary) positive results (Becker, 1997). Because a multiplicity of undesired combustion products may be generated, it is recommended that creosoted residues be incinerated only in a licensed high-temperature incinerator (UNEP, 1995; HSDB, 1999). In particular, dioxins, which are formed during waste incineration processes in the temperature range between 250 and 650 °C, require high temperatures (~1000 °C) for destruction (e.g., Tuppurainen et al., 1998).

The reuse of creosoted railway ties or telephone poles, etc., in children’s playgrounds or other public or private places is now restricted in many countries (RPA, 2000).

The contribution of creosote to environmental contamination in highly industrialized regions is difficult to assess because there are releases of PAHs from many other sources. However, profile and concentration gradients of creosote components in relation to creosote point sources are helpful markers.

The choice of PAHs to be quantified in research and environmental policy as indicators for pollution or health risk assessment depends on the purpose of the investigation, and there is no international agreement. The US EPA (1984b) proposed 16 PAHs as "priority pollutant PAHs," and these are often taken as a reference list for analysis of various environmental matrices. Some countries have their own lists of "priority PAHs" (IPCS, 1998). For example, a group of 10 PAHs is used by the Dutch Ministry of Environment (BKH, 1995), and a group of 11 PAHs is used by the United Kingdom Health and Safety Executive (see Table 4). Sometimes only the three most abundant PAHs (pyrene, phenanthrene, and fluoranthene) are recorded.

Typically, contamination by creosote has been traced by monitoring selected PAHs, but occasionally HACs (NSO) have also been determined.

Generally, there is great variability in the concentrations of the various creosote components, depending, for example, on the varying composition of original creosotes, the amount of creosote released, distance from the source, and the degree of weathering (combined influence of dissolution/adsorption, volatilization, and degradation; see also section 4) of the creosote.

There are few data available for ambient air and surface water matrices, partly reflecting the difficulties in relating the volatile or water-soluble components of creosote to their source. Stationary matrices such as sediment and soil are the best indicators of creosote contamination due to the association of many characteristic creosote components with organic matter (see section 4). Creosote constituents frequently reach groundwater, which remains enriched with low-molecular-weight aromatics, including benzene, toluene, xylenes, naphthalene, and phenolic and heterocyclic compounds.

Metabolites from the degradation of creosote are not usually included in routine analyses of creosote-contaminated samples.

5.1.1 Air

There are few data concerning ambient atmospheric concentrations of creosote-derived compounds; these refer to PAHs from creosote point sources — for example, in the vicinity of creosote facilities.

Fluoranthene concentrations near a creosoting firm in the Netherlands were reported to be 64 ng/m³ at a distance of 500 m, which decreased to 7.2 and 1.6 ng/m³ at 2000 m and 5000 m, respectively. At 2000 m from the creosoting company, the following PAHs were present: naphthalene (90 ng/m³), phenanthrene (44.6 ng/m³), fluoranthene (7.2 ng/m³), and anthracene (2.2 ng/m³) (Slooff et al., 1989). In another study, measured and calculated BaP concentrations in air at a distance of 100, 200, and 2000 m from a creosoting plant were given (without details) as 2–5 ng/m³, 0.5–1.5 ng/m³, and 0.6 ng/m³, respectively (BKH, 1995). Occasionally, neighbouring residents of creosote facilities have complained about odour nuisance (and irritated mucous membranes) (BKH, 1995).

A survey on outdoor concentrations of selected PAHs (including BaP) from other or undefined sources is given by IPCS (1998).

Indoor air concentrations measured at workplaces are compiled in section 5.3.

5.1.2 Water

5.1.2.1 Groundwater

Creosote-related compounds have been detected in groundwater samples near former gasworks and creosoting facilities in Canada, Denmark, and the USA. They include MAHs (e.g., benzene, toluene, xylenes), PAHs (e.g., naphthalene, methylnaphthalene, phenanthrene, anthracene, fluorene, pyrene, chrysene, BaP), and phenolic and heterocyclic compounds. A survey is given in Table 9 (showing sum concentrations) and Tables 10–12 (showing individual concentrations). Highest concentrations have been found in on-site monitoring wells of a former creosote works (Pensacola, Florida, USA), with average concentrations of 1419 mg/litre for total PAHs, 178 mg/litre for total heterocycles, and 0.77 mg/litre for total phenolics (Mueller et al., 1993; see also Table 9). The average concentration for BaP was 37.6 mg/litre (Mueller et al., 1993; see also Table 10). In this study, the elevated concentrations of total PAHs, total heterocycles, and BaP may be reflective of the sample preparation methodologies used. Monitoring data from 44 Danish creosote sites showed concentrations (90th percentiles) of 30 µg/litre for BaP and 50 µg/litre for chrysene (Kiilerich & Arvin, 1996; see also Table 10). Highest concentrations of several individual heterocyclic compounds (e.g., carbazole, dibenzofuran, dibenzothiophene, quinoline/quinolinone) were in the order of 10–80 mg/litre (see Table 11). Within monocyclic aromatic and phenolic compounds, a maximum of 25 mg/litre was reported for m/p-cresol (see Table 12).

Table 10: Groundwater concentrations of non-heterocyclic PAHs detected at creosote-contaminated sites.

Table 11: Groundwater concentrations of heterocyclic compounds detected at creosote-contaminated sites.

Table 12: Groundwater concentrations of monocyclic aromatic and phenolic compounds detected at creosote-contaminated sites.

Mean total PAH concentrations in municipal water supply wells near a former creosote facility (including a distillation plant) in the USA ranged from 0.5 to 4 mg/litre; individual PAH concentrations are not given (Hickok et al., 1982; see also Table 9).

A comparison of groundwater analyses from 44 Danish creosote sites showed that the highest concentrations found for most of the creosote constituents were of the same order of magnitude as the calculated solubilities found in the literature. Exceptions were chrysene and BaP concentrations, which were 1–2 orders of magnitude higher than their solubilities. The reason for this was not given (Kiilerich & Arvin, 1996).

5.1.2.2 Surface waters

It has been estimated that 80% of the (diffuse) PAH pollution of surface waters (from small waterways) in the Netherlands originates from creosote-treated banks (BKH, 1995). Occasionally, films of creosote oil have been identified in small Dutch waterways in which creosoted wood was being placed for bank protection (BKH, 1995) or in a small river near an abandoned US wood treatment facility (Black, 1982). During the application of creosoted bank protection in a small Dutch waterway, BaP concentrations of 1.3 µg/litre were measured; concentrations of other PAHs were not reported (BKH, 1995).

None of the nine PAHs monitored was detected in river water samples taken downstream from a wood treatment facility in Pensacola, Florida, USA (detection limit not specified; probably 30 µg/litre), despite high PAH concentrations in sediments and biota (Elder & Dresler, 1988); however, it should be noted that many PAHs are difficult to detect in surface waters due to adsorption and other phenomena (see also section 4).

On the other hand, PAHs have been detected in river water (Bayou Bonfouca, Lousiana, USA) from an area that remained contaminated following a creosote spill (fire at a wood products treatment plant) 10 years previously. Eight PAHs were monitored (at one control and three "contaminated" sites) — for example, naphthalene (up to 14.1 mg/litre), fluorene (up to 12.3 mg/litre), phenanthrene (up to 155 mg/litre), and BaP (up to 6.6 mg/litre) (Catallo & Gambrell, 1987). Monitoring of 12 PAHs in water samples of a drainage stream near a creosote works (Pensacola, Florida, USA) resulted in total concentrations of up to 153 µg/litre, with BaP concentrations of up to 0.05 µg/litre (Schoor et al., 1991). Total PAH concentrations (16 PAHs) in railway ditch water flowing to salmon streams (British Columbia, Canada) did not exceed 1 µg/litre at four sampling points, but reached maxima of 122 µg/litre and 3516 µg/litre at two sites where creosote-treated power/ telecommunication line poles were erected in the ditches (mean: 606.9 µg/litre; n = 6). Maximum BaP and chrysene concentrations were 2.5 µg/litre and 441 µg/litre, respectively (Wan, 1991; see also Table 13). Data on additional creosote-derived individual PAHs detected in surface waters are listed in Table 13.

Table 13: Surface water concentrations of PAHs detected at creosote-contaminated sites.

For associated sediment concentrations, see Tables 14 and 15.

Table 14: Sediment contamination by creosote: sum concentrations.^a

Table 15: Sediment concentrations of PAHs and heterocyclic compounds detected at creosote-contaminated sites.