Polynuclear aromatic hydrocarbons in the United Kingdom environment: A preliminary source inventory and budget

Environmental Pollution 88 (1995) 91-108 0 1995 Elsexier Science Limited

Printed in Great Britain. All rights reserved 0269-7491/95/$09.50 ELSEVIER

POLYNUCLEAR AROMATIC HYDROCARBONS IN THE UNITED KINGDOM ENVIRONMENT: A PRELIMINARY SOURCE INVENTORY AND BUDGET Simon R. Wild Consultants in Environmental Sciences (CES) Ltd, Kensington House, 136 Suffolk Street Queensway, Birmingham, UK, Bl ILN

Kevin C. Jones* Institute of Environmental and Biological Sciences, Lancaster University, Lancaster, UK, LA1 4 YQ (Received 6 September 1993; accepted 8 February

uncertainties identtped by data on this budget are: (I) the lack of PAH concentrations in some environmental matrices; (2) the possible importance of contaminated soils as a major repository and source of PAHs; (3) the lack of emission data (especially vapour phase releases) for some PAH sources; (4) the importance of biodegradation and volatilisation as loss mechanisms for low molecular weight PAHs in soils; and (5) the importance of creosote use in the PAH cycle.

Abstract This paper presents the first attempt to quanttfy the production, cycling, storage and loss of PAHs in the UK environment. Over 53000 tonnes of CPAHs (sum of 12 individual compounds) are estimated to reside in the contemporary UK environment, with soil being the major repository. If soils at contaminated sites are included, this estimate increases dramatically. Emission of PAHs to the UK atmosphere from primary combustion sources are estimated to be greater than 1000 tonnes CPAHs per annum, with over 95% coming from domestic coal combustion, unregulatedfires and vehicle emissions. It is estimated that approximately 210 tonnes of CPAH are delivered to terrestrial surfaces each year via atmospheric deposition. Therefore, inputs of PAHs to the UK atmosphere outweigh the outputs by a factor of over 4. This may be explained by enhanced particulate deposition near point sources, PAH degradation in the atmosphere and transport away from the UK with prevailing winds. Disposal of waste residues is estimated to contribute a further 1000 tonnes of CPAH per year to the terrestrial environment. It is illustrated that the use of creosote has the potential to release considerable quantities oj‘ PAHs to the UK environment. Temporal trends in PAH cycling are then considered. There is good evidence to suggest that air concentrations and fluxes to the UK surface are now lower than at any time throughout this century. Nonetheless, the UK CPAH burden is still increasing at the present time, principally through retention b.y soils. However, there are marked dtflerences in the behaviour of individual compounds: there is evidence, for example, that phenanthrene concentrations in soils have declined since the 1960s although soil concentrations of benzo[a]pyrene and other heavier PAHs have continued to increase through this century. Volatilisation of low molecular weight PAHs accumulated in soils over previous decades may be making an important contribution to the current atmospheric burden. The major * To whom correspondence

1994)

Keywords: PAHs, sources, environmental emissions, budgets.

distribution,

INTRODUCTION Polynuclear aromatic hydrocarbons (PAHs) are organic chemicals composed of fused benzene rings, whose environmental behaviour has been investigated for more than 20 years. PAHs have been the subject of detailed research due to their toxicity, environmental persistence and prevalence. The widespread occurrence of PAHs is largely due to their formation and release during the incomplete combustion of coal, oil, petrol and wood. Man’s long history of burning such materials has inevitably resulted in substantial releases of PAHs worldwide. Several PAHs have been shown to be acutely toxic. However, health concerns regarding PAHs focus on their metabolic transformation by aquatic and terrestrial organisms into mutagenic, carcinogenic and teratogenic agents such as dihydrodiol epoxides. These metabolites bind to and disrupt DNA and RNA, which is the basis for tumour formation. The most potent carcinogens among the PAHs are the benzofluoranthenes, benzo[a]pyrene, benz[a]anthracene, dibenzo[a,h]anthracene and indeno[ 1,2,3cdlpyrene (IARC, 1983). In order to exhibit their latent carcinogenic properties, PAHs require metabolic conversion and activation. PAHs are listed by the United States Environmental Protection Agency (USEPA) and the European Community as priority pollutants.

should be addressed. 91

92

S. R. Wild, K. C. Jones

This paper uses the existing data base on PAH behaviour and environmental levels to estimate the burden of PAHs in the UK environment. The sources of PAHs to, and losses from, the UK environment are then considered, together with likely future trends. A similar investigation by Suess (1976) conducted for the USA and the world is now dated and only focused on particulate benzo[a]pyrene. While benzo[a]pyrene is one of the most potent PAH carcinogens, it only accounts for a minor proportion of the CPAH burden in the environment. ENVIRONMENTAL

BEHAVIOUR

OF PAHs

Individual PAHs differ substantially in their physical and chemical properties. The PAHs considered in this paper (see Table 1) have been selected to reflect this range in physico-chemical properties, although some PAHs such as dibenz[a,h]anthracene have not been included due to their low environmental abundance. However, it is worth noting that there are many other compounds in the PAH group other than those selected, which may be emitted into the, UK environment where they may reside as contaminants. Overall, the low molecular weight PAHs are more volatile, water soluble and less lipophilic than their high molecular weight relatives. These physico-chemical properties largely determine the environmental behaviour of PAHs and indicate that transfer and turnover of low molecular weight PAHs will be more rapid than the other group members (Wild, 1991). The more volatile low molecular weight compounds dominate the atmospheric CPAH burden, existing primarily in the vapour phase (Jones et al., 1992). PAHs are susceptible to degradation in the atmosphere, such that their atmospheric residence times can be restricted to a few days. Following deposition, the lighter PAHs are also more susceptible to biotic and abiotic degradation (see

Wilson & Jones, 1993). Generally, the greater the number of benzene rings in the PAH molecule, the greater the resistance to degradation (Bossert & Bartha, 1986). In soils, most PAHs are strongly adsorbed on to soil organic matter, thus rendering them unavailable for plant uptake or leaching to groundwater. If plant material is found to contain PAHs, these are generally considered to be either a result of soil contamination or atmospheric deposition (Jones et al., 1989a; 1992; Wild et al., 1992a). If PAHs are incorporated into water, they are rapidly transferred into the sediments, although concentrations in the water column are likely to reflect water solubility. Following ingestion by mammals and other life forms, PAHs do not biomagnify in the same manner as some other organic chemicals (e.g. PCBs), since they tend to be metabolised at the site of entry into the body. THE BURDEN OF PAHs IN THE CONTEMPORARY UK ENVIRONMENT This section has been compiled using literature on PAH concentrations in different environmental media. Where possible, UK data have been used, but where this is not available, results from other industrialised countries have been substituted. For the purposes of this paper, the UK environment is represented by soil, inland water, sediments, air, vegetation, waste residues and biota (including man). No account has been taken of the marine environment or geological strata below soil horizons. It is important to mention that there are several different extraction and analytical methods available, while some environmental matrices are more problematic to analyse than others. Therefore, some of the differences in matrix PAH burden and emission fluxes may be, in part, due to analytical factors. soil

Table 1. Physico-chemical properties of PAHs considered in this papeP

Compoundb

Log Kow

Naphthalene (2) Acenaphthene (3) Fluorene (3) Phenanthrene (3) Anthracene (3) Fluoranthene (4) Pyrene (4) Benz[a]anthracene (4) Chrysene (4) Benzo[b]fluoranthene (5) Benzolalpyrene (5) Benzo[ghi]perylene (6)

3.5 3.95

4.28 5.61 5.33 5.33 5.32 5.61 5.61 6.57 6,3 7.23

Henry’s constant’

Water solubility (mg litre-‘)

5.0 E-2 2.6 E-3 2.6 E-3 9.58 E-4 7.92 E-4 3.51 E-4 3.74 E-4 4.83 E-5 4.38 E-5 4.96 E-4 6.46 E-5 2.23 E-6

3.20 E+l 5.30 E+O

1.85 E+O 1.24 E+O 6.40 E-2 2.45 E-l 1.32 E-l 140 E-2 1.80 E-3 l@OE-3 1.60 E-3 2.65 E-4

“Taken from Wild (1991). bThe number of benzene rings in each compound is shown in parentheses. ’Dimensionless Henry’s constants (Hc) equal to the concentration of the compound in the gas phase to its concentration in the liquid phase.

PAHs have been delivered to all soils by atmospheric deposition (Jones et al., 1989b), so it is expected that most combustion-derived PAHs will be restricted to the surface few centimetres. However, soil profile data show that PAHs can extend to well below the 15 cm plough layer (Jones et al., 1989~). Soil PAH concentrations have been found to vary by orders of magnitude spatially, depending on proximity to point sources and soil factors (Jones et al., 19894. PAH concentrations are generally higher near emission sources; thus, elevated soil concentrations have been found in urban soils (Jones et al., 19894 and roadside soils (Johnston & Harrison, 1984), while very high concentrations have been reported for contaminated sites such as old gas works (Bewley & Theile, 1988; Wilson & Jones, 1993). Soil PAH concentrations have also been found to correlate positively with soil organic matter content, with peaty and forest soils generally containing higher PAH concentrations than agricultural soils (Jones et al., 1989d; Wild & Jones, 1993). The UK has a surface area, not covered by freshwater, of 2.475 X 10” m2 (Geodata, 1983) and has an

Polynuclear aromatic hydrocarbons in the UK environment

93

Table 2. PAH concentrations and estimated burden In UK soils Compound _____~__. Naphthalene Acenaphtheneifluorene’ Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene’ Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene CPAH

Rural soil concentration” (cLg kg-‘)

Urban soil concentration’ (cLg kg-‘)

Forest soil concentrationb (cLgkg-‘)

UK soil burden (tonnes)

6.2 37 14 2.4 27 31 20 9.8 6.8 33

30.5 166 481 88 1256 645 1153 613 379 428

154 86 533 17 1132 573 796 492 352 685

1000 2400 4 800 550 11 200 6 400 9 100 5 100 3 400 5 900

187.2

4239.5

4809

50 000

’ Jones er al. (19896): rural soil concentration median of 30 samples; urban soil concentration ’ Wild & Jones (1993): mean of five coniferous forest soils. (‘These compounds are quantified together by the HPLC procedure used here.

soil bulk density of 1200 kg m-3 (Brady, 1991). Taking a soil depth of 15 cm, a total UK soil weight of 4.45 x 10” tonnes is obtained. Approximately 10% of UK soil is assumed to be in urban areas, while those in forests and woodlands cover 2.3 million ha (DOE, 1992). This leaves approximately 80% of UK soils being classed as ‘rural’. Using these area estimates and the PAH concentrations for rural soil (Jones et al., 19894, urban soil (Jones et al., 19894 and forest soil (Wild & Jones, 1993) it is estimated that the total UK XPAH soil burden is approximately 50 000 tonnes (see Table 2). Fluoranthene is the most abundant PAH found in UK soils, followed by pyrene, benzo[ghi]perylene, benzo[b]fluoranthene and phenanthrene. Of the PAHs considered, anthracene and naphthalene are the least abundant in the soil environment. It should be noted that this estimate takes no account of the presence of PAHs below the plough layer. The PAH burden of UK coal and oil reserves has been ignored, since this budget is for the surface environment.

assumed

Contaminated soils Soil PAH concentrations at industrially contaminated sites can be significantly higher than those detected in urban and rural areas. Elevated soil PAH concentrations may occur at sites involved with the following industrial activities: gasification/liquefaction of fossil fuels (gasworks sites), coke production, asphalt production, coal tar production, wood treatment processes, wood preservative production, fuel processing, etc. (Wilson & Jones, 1993). Estimating the burden of PAHs in such contaminated sites is difficult due to a lack of information regarding the number and size of sites, the PAH concentrations found and the depth of PAH penetration. However, the potential importance of contaminated sites can be illustrated by estimating the burden of PAHs at old gasworks sites. It has been estimated that there were 3000 + 1000 gasworks sites of sizes between 0.3 and 200 ha operat-

mean of 10 samples.

ing in the UK in the early twentieth century (DOE, 1987). However, many of these locations will since have been reclaimed. Here it is assumed that there are currently 1000 derelict gasworks sites in the UK, and that at each site there is a 1 ha area of PAH contamination. If the surface 15 cm is considered (1200 kg mm3bulk density) it is estimated that about 1.8 X 10’ kg of soil is contaminated. If this soil is considered to be contaminated to the levels recorded by Bewley et al. (1989), a UK gasworks site PAH burden of 8100 tonnes is estimated. Derivation of this figure is given in Table 3. The PAH burden given in Table 3 constitutes about 16% of the burden in ‘uncontaminated’ UK soils. However, it is important to emphasise that this estimate is critically dependent on the assumptions made. For example, if a 1 m soil depth is considered, the PAH burden in contaminated gasworks soils increases to over 53 000 tonnes, while if there are 1000 gasworks sites of 10 ha of PAH contamination to 15 cm, then the PAH burden increases to over 80 000 tonnes. From Table 3. PAH concentrations and estimated burden in UK gas works site soils Compound

Gaswork soil concentrationsa (cLgkg-‘)

UK gaswork soil burden (tonnes)

Naphthalene Acenaphthenelfhtorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluorantheneb Benzo[a]pyrene Benzo[gbi]perylene

nq 227 379 156 2 174 491 662 92 260 nq

410 680 280 4 000 890 1 200 170 470

CPAH

4441

8 100

nq, Not quantified. ’ Bewley et al. (1989). b Plus benzo[k]fluoranthene.

S. R. Wild, K. C. Jones

94

Table 4. PAH concentrations and estimated burden in UK vegetation Compound

Vegetation concentrationa (I-Lgkg-‘)

Table 5. PAH concentrations and estimated burden in tbe UK air mass

UK vegetation burden (tonnes)

Compound

Naphthalene Acenaphthenelfluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

32 300 260 21 60 10 16 33 8 63

2.1 19 17 1.3 3.9 0.63 1.0 2.1 0.51 4.0

Naphthalene Acenaphthenelfluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

CPAH

803

52.0

CPAH

‘Wild et al. (19926).

this it is apparent that the UK soil PAH burden may be dominated by contaminated sites, especially given that other industrial sites, such as wood preservation plants can be contaminated with PAHs to levels in excess of those found at gasworks sites (see Mueller et al., 1991). Until further information is available it is stressed that the estimates derived here are very tentative. Vegetation

PAHs detected in most surface vegetation are essentially derived from atmospheric deposition. Taking the above-ground vegetation PAH data reported by Wild et al. (19926) for a 1986-1989 composite herbage sample and an assumed dry aerial biomass of 260 g mm2 (Haygarth et al., 1991) a total PAH burden for UK vegetation of 52 tonnes is derived, with the lighter compounds, acenaphthene/fluorene and phenanthrene, accounting for about 70% of this load. These estimates (shown in Table 4) have been derived using data from contemporary rural/semi-urban vegetation and hence may underestimate the EPAH burden of UK vegetation, because of elevated levels nearer sources. Measurable loadings of PAHs are also associated with below-ground plant parts (Wild & Jones, 1992). Air

The UK air mass is assumed here to be that above the UK surface area to a height of 1 km. This amounts to 2.5 X lOI m3 of air. UK urban air data taken from Halsall et al. (1993) is shown in Table 5. The values shown are median air concentration in 1991 (particulate and vapour phases) for Manchester, Cardiff, Stevenage and London. Interestingly, the concentration of some of the lower molecular weight species in rural areas may actually be equivalent to, and even exceed, those observed in urban regions, probably due to soil outgassing (see later). Those PAHs with less than four benzene rings (see Table 1) exist in the atmosphere almost exclusively in the vapour rather than the particulate phase, while the opposite is true for PAHs with

Air concentrations’ (ng mm3)

UK atmospheric burden (tonnes) 2.8 16.0 12.0 0.85 2.6 1.9 0.15 0.24 0.16 0.35

11.3b 63.4 46.2 3.4 10.3 7.5 3.0 0.95 0.65 1.48 148

38

’ Halsall et al. (1993) for UK air. b Taken from Vogt et al. (1986) for Norwegian air.

more than four rings (see Jones et al., 1992). The defined UK air mass is estimated to have a PAH burden of 38 tonnes at any one time, with the low molecular weight compounds acenaphthene/fluorene and phenanthrene dominating. Freshwater

Freshwater has been assumed to cover 1% of the total UK surface area. If this freshwater has a depth of 10 m it can be estimated that the volume of freshwater in the UK totals 2.5 X 1O’Om3. In the absence of reliable ‘background’ UK freshwater PAH data, figures published by McVeety and Hites (1988) have been substituted. These relate to water from Lake Superior in an isolated location considered to be representative of ‘background’ conditions. Much higher PAH concentrations have been reported for rivers and estuaries in urban regions (Readman et al., 1982). It is assumed that 10% of freshwater is in urban regions, with PAH concentrations as those reported in this cited reference. The total UK freshwater PAH burden derived in Table 6 is approximately 260 kg. The majority of this burden is expected to be associated with suspended organic matter, although some PAHs may be adsorbed on to mineral particles or in a true dissolved state (Herbes, 1977). The burden of 260 kg in UK freshwater seems particular!y low, although it is expected that PAHs will rapid11 transfer from the water column to the sediments. It is noted that water PAH concentrations may be significantly higher at industrial outlets. However, due to a lack of information available in the literature, this has not been included here. Groundwater has not been included as this is a budget for the UK surface environment. Freshwater sediments

As indicated above, most PAHs in water are associated with the particulate phase. Sedimentation of these particulates limits the residence time of PAHs in the water column and results in an accumulation of PAHs in sed-

95

Polynuclear aromatic hydrocarbons in the UK environment Table 6. PAH concentrations aud estimated burden in UK freshwater Compound

_____ Naphthalene Acenaphthenel fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/ chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene ZPAH

Urban Rural freshwater freshwater concentration”concentrationb (ng litre-‘) (ng litre-‘)

Table 8. PAH concentrations aud estimated burden in UK sewage sludge

UK freshwater burden (kg)

nq

13.9

35

0:;4 0.075 0.48 0.25

;8 4.9 10.4 18.0

-TO 15 35 50

0.165 Onq nd

18.7 9.3 9.1 -

50 25 10 -

1.81

93.1

260

nd, Not detected. nq, Not quantified. ’ McVeety and Hites (1988). ’ Readman et al. (1982).

Compound

Sewage sludge concentration’ (mg kg-‘)

UK sewage sludge burden (tonnes)

Naphthalene Acenaphthene/fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

0.86 2.40 3.00 0.33 1.50 2.00 1.70 0.99 0.82 2.5

o-95 2.7 3.3 0.37 1.7 2.2 1.9 1.10 0.91 2.8

CPAH

16.1

18.0

’Wild & Jones (1992). in UK sediments. It should be noted that this estimate does not consider contaminated sediments.

found

Waste residues

iment (Cranwell & Koul, 1989). Sanders et al. (1993) detected PAHs in a semi-rural freshwater UK lake (Esthwaite Water, NE England) to a depth of at least 0.68 m. However, the majority of PAHs (except the biogenically formed perylene) were mainly restricted to the surface O-5 m. UK inland water sediments are assumed to cover an area of 2.5 X 10’ m2 (1% of UK surface area). Taking a sediment density of 0.13 g dry weight cm~3 (based on a sediment moisture content of 87% and a solid density of 1 g cm-“) and averaged PAH concentrations for the entire 0.5 m core depth, the UK freshwater sediment CPAH burden is calculated to be about 2800 tonnes (see Table 7). Note that the naphthalene concentration came from the study by Readman et al. (1982) on the Tamar Estuary (UK). Pyrene and fluoranthene are the most abundant individual PAHs Table 7. PAH concentrations aud estimated burden in UK freshwater sediments ____ Compound Mean sediment UK freshwater dry weight sediment concentrationa burden (tonnes) (pg kg-‘) __-24sb 40 Naphthalene Acenaphthenelfluorene ?5 130 Phenanthrene 200 30 Anthracene 3 450 560 Fluoranthene 3 520 570 Pyrene 340 Benz[a]anthracene/chrysene 2 082 380 Benzo[b]fluoranthene 2 355 260 Benzo[a]pyrene 1 582 470 Benzo[ghi]perylene 2 992 ZPAH ’ Sanders et al. (1993). ’ Readman et al. (1982) for UK surface sediment. nq, Not quantified.

2800

Wastes such as sewage sludge, municipal waste, industrial wastes, dredgings, excavated material and ash residues may all contain PAHs. These wastes may become incorporated into the environment. There is very little information regarding the PAH content of most of these wastes. However, there are data on the PAH content of UK sewage sludges and municipal fly-ashes. Using the PAH data reported by Wild and Jones (1992) for contemporary UK sewage sludge and a total UK sewage sludge production figure of 1.1 million tonnes (dry weight) (DOE, 1993a) it can be estimated that UK sludge contains about 18 tonnes of CPAH (see Table 8). The UK currently produces some 140 million tonnes (wet weight) of controlled waste (DOE, 1992). This includes about 30 million tonnes of municipal waste (MSW), which has been shown to contain PAHs (Hagenmaier et al., 1986). Taking the limited MSW PAH concentrations given by Hagenmaier et al. (1986) for Germany and correcting for 65% organic matter, it is estimated that about 82 tonnes of CPAH reside in UK MSW (42.4 tonnes fluoranthene, 21.2 tonnes benzo[b]fluoranthene, 9.1 tonnes benzo[a]anthracene/ chrysene, 9.1 tonnes benzo[ghi]peqlene). Inclusion of the other PAHs probably increases this estimate to over 100 tonnes while the CPAH burden associated with the remaining 110 million tonnes of controlled waste is considered to be anywhere from 500 to 1000 tonnes. General comments

The above sections estimate that about 50000 tonnes of ZPAH reside in ‘uncontaminated’ soil, while 52 tonnes are associated with vegetation, 38 tonnes in air, 260 kg in freshwater and 2800 tonnes in freshwater sediments. Thus, the total UK natural environment contains approximately 53 000 tonnes of PAHs (excluding highly contaminated soils/sediments and waste

96

S. R. Wild, K. C. Jones

Table 9. Smmary of the contemporaryUK natural envirommd Compound

Soil

PAH burden(toums) (excludingcouhmhated sites and waste matrices) Air

Vegetation

Naphthalene Acenaphthenelfluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

1 000 (96.15) 2400 4 800 (96.00) 550 (94.83) 11 200 (94.92) 6 400 (91.43) 9 100 (96.81) 5 100 (92.73) 3 400 (91.89) 5900

2.1 19 17 1.3 3.9 0.63 1.0 2.1 0.51 4.0

CPAH

50 000

52.0

(0.20) (0.34) (0.22) (0.03) (0.01) (0.01) (0.04) (0.01)

2.8 16.0 12.0 0.85 2.6 1.9 0.75 0.24 0.16 0.35

Freshwater

(0.27)

0.035 (0.000 3)

(0.24) (0.15) (0.02) (0.03) (0X)08) (0.004) (0.004)

&40 0.015 0.035 0.050 0.050 0.025 0.010 -

38.0

0.26

(0.000 8) (0.003) (0.000 3) (0.000 7) (0.000 5) (0.000 5) (0.000 3)

Freshwater sediment 40

(3.85)

1; 30.5 560 570 340 380 260 470

(2.60) (5.17) (4.75) (8.14) (3.62) (6.91) (7.03)

2 800

Approximate totals 1040 2400 5000 580 118 000 7000 9400 5500 3700 6400 53 000

a Figures in brackets are percentages of the total.

Greater than 90% of the total UK burden of CPAH resides in the surface soil, while the bulk of the remainder is associated with freshwater sediments (53%). The burdens and relative proportions of individual PAH compounds in these environmental compartments is summarised in Table 9 (excludes waste matrices and contaminated soils). The burden of PAHs in contaminated soils is potentially very large. Under the assumptions detailed earlier it is likely that over 8000 tonnes of CPAHs reside in contaminated soils at old gasworks sites. It is possible that the total burden of PAHs at contaminated sites is actually greater than the total UK environmental burden given above. However, until more information is available regarding soil PAH concentrations at such sites, the number and size of sites, and the depth of PAH penetration, the burden cannot be assessed very accurately. The above estimates are considered to be reasonably reliable, but could be refined subsequently. The assumptions regarding soil and sediment depths clearly have a profound effect on the overall budget. In some locations, it is likely that the bulk of the PAHs are found within zones deeper and shallower than assumed. The above calculations have not considered the contribution of biota to the UK PAH budget. However, as indicated earlier, PAHs are metabolised and not bioaccumulated to the extent of some other persistent organic pollutants. Thus, it is expected that higher animals constitute a minor environmental compartment of PAHs. For example, if we assume that human tissue contains the same PAH concentrations as the meat reported by Dennis et al. (1983), the human body burden of benzo[a]pyrene is only 3 pg (assuming a 60 kg human). Thus, the UK human population only contains 165 g of benzo[a]pyrene, considerably less than 1% of the total UK burden. The same is true for most other individual PAHs, although some of the low molecular weight compounds may be relatively enriched in biotic tissues. Overall, it is considered that biota (e.g. humans, livestock and wildlife), as an environment compartment, do not constitute a major PAH sink (e.g. 400 kg). matrices).

PRIMARY SOURCES OF PAHs TO THE UK ATMOSPHERE Introduction

PAHs are formed principally by incomplete combustion of organic materials. Biogenic formation has been suggested for some PAHs, but its importance remains uncertain. However, perylene is one PAH whose biogenie production in anaerobic sediments is well characterised (Sanders et al., 1993). For the purposes of this study, biogenic sources for the PAHs of interest have been ignored, given the overwhelming importance of production during combustion. This section quantifies sources of PAHs to the UK atmosphere. Combustion can be both naturally or anthropogenically driven. Natural sources of PAHs, such as forest fires (McMahon & Tsoukalas, 1978) and volcanic activity, whilst important in some countries, are unlikely to be significant for the UK. Anthropogenic PAH releases are considered to be from two main combustion categories. The first is combustion of materials for energy supply (e.g. coal, oil, gas, wood, etc.). The second is combustion for waste minimalisation (e.g. waste incineration) (WHO, 1987). Both categories use materials which contain PAHs, but it is thought that most of these are destroyed during the combustion process. New PAHs are formed due to incomplete pyrolysis and are emitted in vapour and particulate forms, predominantly to the atmosphere, although solid waste residues also contain detectable quantities of PAHs (e.g. Wild et al., 1992~). Energy supply

Combustion of organic based materials for energy supply can be from mobile or stationary sources. Vehicles are the principle mobile sources, while stationary sources can produce energy for small scale direct use, such as domestic heating, or for large_$cale energy production and industrial applications. The amount of PAHs produced and released to the atmosphere from all these sources varies greatly and is dependent on factors such as fuel type, combustion conditions, emission

97

Polynuclear aromatic hydrocarbons in the UK environment Table 10. Total UK emissions of gas and particulate pbase PAHs by vehicles per anuum Gas phase emission” (pg km-‘)

Compound

Leaded Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/ chrysene Benzol[b]fluorantheneb Benzo[a]pyrene Benzo[ghi]perylene CPAH

particulate phase Emission” (pg km-“)

Unleaded

Leaded

Unleaded

11.7 17.3 35 128 12.3 13

23 42 91 28 18 17

nq

“4 “4

tfp30 0.14 1.9 3.2

5.3 nq nq nq

1.6 nq nq nq

107

221

Total car/taxi emission (tonnes)

Lorries/ coaches emission (tonnes)

Trains/ aircraft emission (tonnes)

Total (tonnes)

nd nd 2.5 4.6

5.5 9.4 20 6.5 5.6 6.1

1.8 3.2 7.0 2.1 1.6 1.6

0.09 0.16 0.38 0.11 0.09 0.10

7.4 12.8 27.4 8.7 7.3 7.8

3.6 1.9 3.3 2.9

6.7 4.8 6.6 5.5

2.8 1.0 1.6 1.3

0.6 0.4 0.5 0.4

0.04 0.02 0.03 0.02

3.4 1.4 2.1 1.7

17.3

30.7

19.2

l-04

80.2

60

’ Westerholm et al. (1988). * Plus benzo[k]fluoranthene. nq, Not quantified. nd, Not detected.

control measures, etc. (Hasanan

et al., 1986; Mitra et

al., 1987).

The amount of PAHs generated from petrol-driven vehicles is controlled by engine factors (e.g. age, maintenance, capacity, etc.), the vehicle model and whether the car vehicle has warmed up. However, an indication of the quantity of PAH released into the UK atmosphere from petrol driven vehicles can be estimated from the data provided by Westerholm et al. (1988) combined with the fact that UK cars and taxis consume 24 million tonnes of petrol per year (45% unleaded) resulting in over 330 billion vehicle kilometers (DOE, 1992). Table 10 shows the PAH concentrations in leaded and unleaded petrol car emissions (gas and particulate phases) and illustrates that at least 60 tonnes of PAHs are emitted to the atmosphere each year from this source in the UK. PAH emissions are higher from combustion of unleaded petrol, due to the higher aromatic nature of the fuel (Baek et al., 1992). The above estimate assumes that PAH emissions from diesel engines are the same as leaded petrol. This is conservative in the light of the data given by Hagemann et al. (1981) which showed that PAH diesel emissions were equal to or greater than those from combustion of lead-free fuel. A further 76 billion vehicle kilometres are travelled by motor cycles, buses, coaches, vans and heavy goods vehicles. All these vehicles are assumed to run on diesel fuel and emit PAHs at the same rate as emissions from unleaded driven cars. Under this scenario a further 19.2 tonnes of CPAHs are emitted to the UK atmosphere each year (see Table 10). PAH emissions from trains and civil aircraft have not been evaluated in the literature. However, a crude estimate of PAH emissions from these sources can be obtained if it is assumed that the ratio of PAH emissions from railways/aircraft to road traffic is the same as volatile organic compound (VOC) emission (e.g. 1 :

81) (DOE, 1992). Using this approach it is estimated that a further 1 tonne of CPAHs are emitted to the UK atmosphere from railways and civil aircraft (only emissions associated with ground movement and take off and landing cycles up to 1 km). Therefore, overall from all UK road, rail and air traffic vehicles over 80 tonnes of CPAH are emitted each year to the atmosphere. Emission fluxes for some vapour phase PAHs are absent because of a lack of measured data. This is an important omission, since it is clear that total vehicular emissions could be substantially higher than indicated (Benner et al., 1989). Overall, phenanthrene emissions from vehicles are higher than any other PAH, while about 70% of total emissions are composed of low molecular weight compounds with less than four benzene rings. Energy production by stationary sources includes most coal consumption, combustion of fuel oil and natural gas, as well as wood utilisation for domestic heating. In 1991, 96 million tonnes of coal were used in the UK, 78% for power generation 9% for coke manufacture, 1% for production of smokeless fuel and 11% for domestic heating and industry (DOE, 1992). A further 5 million tonnes of solid smokeless fuel were used for domestic and industrial purposes. Using the PAH emission data measured by Masclet et al. (1987) for French coal-powered power stations (vapour and particulate phase PAHs), it is possible to calculate that at least 3.1 tonnes of CPAHs are emitted from UK power stations each year (see Table 11). This estimate excludes naphthalene, anthracene and benzo[ghilperylene. Pyrene and phenanthrene emissions are significantly higher than those of other individual PAHs, accounting for approximately 70% of the CPAH emission flux from this source. Applying the same emission rates for coal combustion at power stations to coal used to manufacture coke and smokeless fuel, and to coal used for industrial purposes it is esti-

S. R. Wild, K. C. Jones

98

Table 11. PAH emissions from UK coal-&d power stations

Compound

Table 13. PAH emissions from UK oil-firedpower statious

Emission rate” Total emission (cLgkg-‘) (kg) 130 950

Compound

Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzolblfluoranthene Benzo[a]pyrene Benza[ghi]perylene

nq 1.7 12.6 nq 7.5 158 4.1 0.072 0.077 nd

560 1 200 300 5.4 58 -

Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzolblfluoranthene Benzo[a]pyrene Benzo[ghi]perylene

CPAH

41.9

3 140

ZPAH

Emission rate” (pg kg-’ fuel)

Total emission (tonnes)

1;\ 88.0

o& 1.5

1;P, 18.5 16.6 nd 0.88 nd

oY4 0.31 0.28

157

0.015 2.7

’Masclet et al. (1987). nd, Not detected. nq, Not quantified.

’Masclet et al. (1987). nq, Not quantified. nd, Not detected.

mated that a further 400 kg of CPAHs is emitted to the UK atmosphere each year. Domestic combustion of coal occurs under less controlled and less efficient conditions than coal use at power stations. Combustion temperatures are also significantly lower. Thus, PAH emissions are significantly higher as a result. Grimmer et al. (1983, 1985) presented PAH emissions from domestic pyrolysis of hard coal briquets in Germany. Table 12 illustrates that using the average emission condensate rates given in these papers, and the fact that approximately 10.5 million tonnes of coal/solid smokeless fuel and other solid fuels (excluding wood) are used for domestic heating, it is estimated that 600 tonnes of CPAH may be released to the UK atmosphere from domestic appliances per annum. Fluoranthene has the highest emission flux from this source, followed closely by pyrene, although naphthalene emissions have not been determined. Each year 17 million tonnes of fuel oil and 21758 million therms of natural gas and 2934 million therms of other gas forms are combusted for energy supply (1 million therms are equivalent to 4000 tonnes of coal or

2400 tonnes of petrol) (DOE, 1990, 1992). PAH emissions from oil-fired power stations and industrial oil users are estimated to be in the region of 2.7 tonnes using data derived by Masclet et al. (1987) (see Table 13). Phenanthrene has the highest individual emission flux from this source, although emission rates for some PAHs have not been quantified. No information could be found to assess PAH emissions from domestic oil heating systems, while there are few PAH emission figures for any form of natural gas combustion. Total PAH emissions from gas combustion are likely to be low, although the data to support this statement are lacking. PAH emissions by industry can be derived from either the fuel being used or the raw materials being processed. Emissions from industrial fuel use have been incorporated into the paragraphs above. However, it is important to note that PAHs may be present in emissions from plants involved with aluminium production, petroleum processing, production of carbon black for tyres and air blowing of asphalt. All these industries use raw materials which contain PAHs which can feasibly be lost to the atmosphere during processing. However, no data regarding PAH emissions from these sources can be found in the literature; thus, an accurate assessment of the significance of these sources cannot be made. It is apparent that most of these industries are covered by Integrated Pollution Control measures and as such it is likely that PAH emissions to the atmosphere will not be excessively high due to emission abatement systems. Industries and refineries produce about 3.2 times more black smoke than power stations (DOE, 1992). If this ratio is assumed to apply to PAH emissions it is estimated that about 18 tonnes of PAHs are emitted to the UK atmosphere each year. Clearly this is only a very crude estimate which requires future clarification. The final atmospheric PAH combustion source from heat/energy production is wood burning. The amount of wood used as a domestic heat source in the UK is currently unquantified. As such, the contribution of

Table 12. PAH emissions from UK domestic coal combustion

Compound

Emission rate” (mg kg-’ fuel)

Total emission (tonnes)

Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluorantheneb Benzo[a]pyrene Benzo[ghi]perylene

o”P,65 3.8 1.3 16.8 15.1 10.3 4.1 2.7 2.1

0.69 40 14 180 160 110 43 28 22

ZPAH

56.4

600

“Grimmer et al. (1983, 1985). ’Benzolblfluoranthene + benzoljlfluoranthene + benzo[k]fluoranthene. nq, Not quantified.

99

Polynuclear aromatic hydrocarbons in the UK environment

Table 14. Estimated PAH particulate emissions from domestic UK wood combustion

Table 15. PAH emissions from UK incinerators Compound

Compound

Particulate concentrationsa (CLg8-V

Particulate emission (tonnes)

Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

nq nq nq nq 88 102 102 30 21 5.8

0.92 1.1 1.1 0.32 0.29 0.06

CPAH

355

3.8

-

’Freeman and Cattell (1990). nq, Not quantified. this source to atmospheric releases of PAHs cannot be fully assessed. However, Dasch (1982) illustrates that on average 10 g of particles is emitted per kg of wood cornbusted in fireplaces, while Freeman and Cattell (1990) report PAH concentrations in particulate matter from wood burning. Using these figures and a conservative estimate of 1 million tonnes of wood burnt per annum, an estimate of approximately 4 tonnes CPAH is derived (excluding naphthalene, acenaphthene/ fluorene, phenanthrene, anthracene) (see Table 14). This also excludes vapour phase emissions, which are likely to be significantly higher than particulate emissions, especially with respect to the lower molecular weight PAH compounds. It is apparent that although this source cannot currently be quantified accurately, it is potentially more important than power station emissions. Waste disposal and minimalisation Incineration represents a major waste disposal and waste minimalisation process. Each year in the UK, 3 million tonnes of municipal waste, 61 000 tonnes of chemical waste, 170 000 tonnes of clinical waste and 77 000 tonnes (dry weight) of sewage sludge are incinerated. As indicated in previous sections, incineration results in the formation of solid waste ash residues containing detectable quantities of some PAHs. These constitute a source of PAHs to the UK environment on disposal. However, as for material combustion for energy generation, incineration of waste also results in emissions of PAHs from the stacks in vapour and particulate phases. These emissions will be highly variable, depending on the feedstock, the incinerator design and combustion conditions. Hence, the figures derived here should be regarded as very tentative. If the stack PAH concentrations given by Davies et al. (1976), Williams (1990) and Colmsjii et al. (1986) are used in combination, an estimate of total UK PAH emissions from incinerators (MSW, sewage sludge, clinical waste, chemical waste) can be derived. However, the estimate detailed below assumes that 655 m3

Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene CPAH

Emission rate (pg m-‘1
Annual emission (kg) 0.‘14 4.3 0.14 0.14 8.2 22 10 4.5 0.28 6.0 56

Emission data given by Davies et al. (1976) except “Colmsjij et al. (1986) and b Williams (1990). Assumes 655 m3 min-’ emission rate, 12 h operation burning -9.14 tonnes of waste per hour. of stack gases are released per minute, while the incinerator operates for 12 h per day, burning 9.14 tonnes of refuse per hour (Davies et al., 1976). It also assumes that all refuse produces stack gases of the same PAH composition and burden. While these assumptions will not be applicable for all incinerators, this was the most practical method to derive crude estimates. Table 15 suggests that an estimated 56 kg of ZPAH is emitted annually from UK incinerators. The annual ZPAH release would be substantially higher if the stack gas PAH concentrations reported by Colmsja et al. (1986) found during incinerator startup were used, as under ‘cold startup’ PAH stack gas concentrations are significantly enhanced due to poorer combustion conditions. This total PAH emission estimate is substantially higher than the value of -2 kg ZPAH obtained by considering only particulate fly ash PAH concentrations (reported by Wild et al., 1992~) and the ash emission rate of 9.8 X lo6 kg derived by Harrad et al. (1991), thus illustrating the importance of vapour phase releases. Accurate estimates for PAW emissions from UK incinerators can only be made with accurate stack gas PAH concentrations. This study highlights the lack of such data. It is also worth noting that future PAH emissions from UK incinerators will probably decrease with improvements in incineration technology and pollution abatement systems. Miscellaneous sources The previous sections estimated the annual PAH emissions to the UK atmosphere from major combustion sources. There are, however, other sources which should be mentioned. One source which is potentially extremely important is PAH emissions from unregulated fires. Unregulated fires includes forest fires, heathland fires, recreational fires, factory/house fires and the unregulated burning of tyre, waste wood, etc. While there are no available estimates of how much unregulated burning takes place in the UK, it is clear that during such fires a considerable

S. R. Wild, K. C. Jones

100

Table 16. PAH particulate emissions during UK stubble burning Particulate emission (tonnes)

Particulate concentration” (/-% g-l)

Compound

Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthraeene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

:: 102 102 30.2 27.3 5.8

6 1.8 1.8 0.54 0.49 0.10

CPAH

355.2

6.4

’ Freeman & Cattell nq, Not quantified.

nq nq nq

-

(1990).

quantity of PAHs is likely to be emitted. Therefore, while no estimates can be given, it is considered that the quantity is likely to be of the order of tonnes rather than kilograms. In the past, straw and stubble burning has been widely used as an agricultural practice. In 1984 an estimated 1.26 million ha of land was burnt, resulting in emissions of about 18 000 tonnes of black smoke (Larkin et al., 1986). Since then, the practice has reduced in importance-indeed, stubble burning has been banned in the UK since 29 June 1993 (DOE, 19933). However, as an indication of its recent significance if the above figure is used, combined with the particulate woodsmoke concentrations given by Freeman and Cattell (1990), an estimated 6.3 tonnes of CPAH would be emitted in particulate form (see Table 16). Vapour phase releases are likely to increase this flux substantially, as is the inclusion of the low molecular weight PAH compounds. PAH emissions from cigarette smoking can also be

considered. PAH concentrations in cigarette smoke have been found to be significantly higher than background atmospheric concentrations (Tuominen, 1990). The importance of this source on environmental PAH levels is currently unquantified, although if 10 million people smoke 10 cigarettes each day for a year, over 0.26 kg of benzo[a]pyrene would be emitted, using the figures of Tuominen (1990). If benzo[a]pyrene emissions account for about 4% of the ZPAH emissions from this source, over 6 kg of CPAH could be emitted by cigarettes. Perhaps surprisingly, these crude estimates suggest that cigarettes could be a more important source of PAHs to the UK atmosphere than waste incineration. PAH emissions from this source are also likely to be important with respect to human health, given the direct physical link between PAH formation, emission and inhalation. Indeed, benzo[a]pyrene has been suggested as a causative agent in the increased incidence of lung cancer in smokers. In a later section, the importance of PAH volatilisation from soils, vegetation and other surfaces will be discussed. Here it is sufficient to state that we consider this process to constitute a potentially very significant source of some low molecular weight PAHs to the UK atmosphere (see Table 1). The UK air mass is not a static body. Generally, air passing over the UK moves from the southwest, and leaves to the east. This air movement clearly results in the delivery and loss of PAHs. Since to the west of the UK there are no major industrialised countries producing PAHs, air entering the UK is expected to be cleaner with respect to PAH load, than air leaving the UK. The UK is therefore likely to be a net exporter of PAHs. Nevertheless, the influx of air to the UK may constitute a source of volatile PAHs which may be released from the marine environment. Other miscellaneous PAH atmospheric sources includes crematoria, gas leakages, volatilisation from fuel spills, etc.

Table 17. Summary of annual PAH emissions to the UK atmosphere (tonnes) Compound

Vehicles

Naphthalene Accnaphthenelfluorene Phenanthrene Authraccne Fluoranthene Pyrene Benz[a]anthracene/ chrysene Beuzoblfluoranthene Benzo[a]pyrene Bcnzo[ghi]perylene

7.4 12.8 27.4 8.7 7.3 7.8

CPAI-I

80.2

LIOnly particulate phase.

3.4 1.4 2.1 1.7

Coal-fired power stations

Cokemanufacture/ smokeless fuel production and industrial coal use

OL 0.95 0.56 1.20

0.32 0.24 0.144 O-30

0.7 40 14 180 160

0.32 1.5 0.24 0.31

0.30 0.005 4 0.005 8 -

0.078 0@014 OGOl 5

110 43 28 22

0.28

3.1

0.8

600

Domestic coal combustion

Oil-fire power stations and industrial oil users

Domestic MSW/etc. incineration wood combustion”

Stubble burning’

0.92 1.1

0JxKl 14 0.004 3 0.000 14 oxlOO 14 OflO 2 0.022

1.6 1.8

0,015 -

1.1 0.32 0.29 0.061

0.010 0.004 5 oxlOO3 0.006 0

1.8 0.5 0.5 0.1

2.7

3.8

0.056

6.3

1 -

Industrial processing

L39 7.74

-

2.50 4.77

I

1.83 0.017 0.066 18.30

Total

7.4 15 78 23 193 177 119 45 31 24 712

Polynuclear aromatic hydrocarbons in the UK environment

Summary Table 17 presents a summary of the estimated PAH annual primary emissions made to the UK atmosphere. The total estimated input of over 710 tonnes ZPAH per year is subject to many variations and is only as accurate as the assumptions and raw data used. The estimate excludes all inputs from unregulated fires and any inputs related to volatilisation from soil and other surfaces. From the table it is also clear that for many of the PAHs (especially the low molecular weight compounds) accurate emission estimates cannot be made. Vapour phase release estimates are absent for several important sources. These omissions are likely to exert a profound effect on the total emission flux estimate. To illustrate this point, the likely missing compound emission rates can be crudely estimated using the known ratios between compounds from other emission sources. When this is done the total CPAH annual emission flux increases to over 900 tonnes. Of this, about 67% comes from domestic coal combustion, 22% from domestic wood combustion and stubble burning, while 9% comes from vehicle emissions. In contrast, coal- and oil-fired power stations and waste incineration account for less than 0.5% of annual PAH releases to the UK atmosphere. If other unregulated fire and volatilisation sources were to be included it is considered that the estimate would rise to over 1000 tonnes CPAH, potentially substantially over this figure. Overall, these figures indicate that uncontrolled, unregulated fuel and waste combustion, predominantly from numerous small relatively uncontrolled sources, are responsible for the majority of contemporary PAH releases. These figures compare favourably with the few PAH emission estimates for other European industrial nations. PAH emissions from Sweden in 1985 and 1988 have been estimated to be 250 and 200 tonnes, respectively (OECD, 1993). However, there is no indication of the number of PAHs included, nor whether the vapour and particulate phases are included. In 1985 PAH emissions from The Netherlands and Germany have been estimated to be 1116 and 8218 tonnes, respectively (OECD, 1993). In The Netherlands 20% of emissions were considered to be from transport and 56% from non-industrial sources. The German emissions estimate of 8218 tonnes per year appears much higher than the UK estimate. However, on an individual compound basis the estimates are very similar except for naphthalene and phenanthrene. German emission estimates of benzo[ghi]perylene, benzo[a]pyrene, fluoranthene, benz[a]anthracene/chrysene and anthracene are given as 41, 29, 168, 372 and 152 tonnes per year and are thus in the same order as the UK estimates given in this paper. The estimates for naphthalene and phenanthrene are given as 5900 and 1484 tonnes per year, respectively. Given that most emission fluxes for naphthalene were absent during the derivation of the UK estimate, it is clearly possible that the UK naphthalene emission figure is similar to that given for Germany. However, it is stressed that the source data for the German estimates are not given, nor are the emission

101

flux estimates. The fact that the UK phenanthrene estimate has been derived with quoted fluxes, and is still an order of magnitude lower than the German figure, indicates the risk in simply substituting German figures for the UK. PAH LOSSES AND RECYCLING ENVIRONMENT

IN THE

Introduction From the previous sections it was tentatively estimated that at least 900 tonnes of CPAH are emitted into the UK atmosphere each year. Earlier in the paper it was indicated that at any one time the UK air mass only contains -40 tonnes of CPAH. From this it is very clear that the atmosphere is not a repository and collector of PAHs, but is more likely to be a transporter, dilutor and reactor. The following paragraphs briefly detail the principal means by which emitted PAHs are lost from the atmosphere. Atmospheric deposition One mechanism of PAH loss from the atmosphere is deposition. Since many PAHs released into the atmosphere are associated with particulates, they are open to gravitational settling and scavenging by precipitation/water vapour. Such processes will clearly have less effect on the vapour phase PAHs and those associated with fine aerosols. Using the mean deposition fluxes derived from two UK urban sites over 2.5 years (Halsall, C. & Jones, K. C., unpublished data) it is estimated that each year over 210 tonnes of CPAH are delivered to the ground surface by wet/dry deposition (see Table 18). There is an apparent major discrepancy between atmospheric inputs, storage and loss. By the estimates given here, inputs of PAHs to the UK atmosphere (at least 900 tonnes) outweigh the outputs (210 tonnes) by a factor of over 4. This is interesting, as this situation is in marked contrast to the mass balance calculations derived for another group of combustion related compounds, the polychlorinated dibenzo-p-dioxins and furans (PCDD/Fs) (Harrad & Jones, 1992). The primary Table 18. Annual UK PAH deposition fluxes

Compounds Naphthalene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene CPAH

Flux deposition” (pg rn. - day- ‘) nq 397 556 22.1 403 243 131 187 201 169

36 51 2.0 37 22.2 12.0 17 18 15

2 310

’ Halsall, C. & Jones, K. C. (unpublished nq, Not quantified.

Annual flux (tonnes)

210 data).

102

S. R. Wild, K. C. Jones

sources of PCDD/Fs were estimated to account for only about 10% of the estimated annual deposition flux (Travis & Hattemer-Frey, 1991; Harrad & Jones, 1992). The imbalance between atmospheric PAH inputs and loss can be explained by three main processes: (i) PAH losses within the atmosphere due to reactions/breakdown; (ii) PAH transport away from the UK with prevailing winds; and (iii) greater PAH deposition to terrestrial and aquatic surfaces (e.g. Gardner et al., 1992) near point sources. PAHs are subject to complex physico-chemical reactions and transformations in the atmosphere. Atmospheric PAHs are susceptible to both chemical oxidation and photochemical alterations (Baek et al., 1991). PAHs have been found to react with atmospheric ozone (0,) (Alebic’-Juretic’ et al., 1990), with NOx to produce nitro-PAHs which are potentially more toxic than the PAH precursors, and with SOx and -OH radials (Nikolaou et al., 1984; Valerio et al., 1984). The importance of such reactions is difficult to assess, since many of the investigations conducted to date have been laboratory based. Thus, there is the potential for creating unrealistic, controlled reaction conditions which may not occur in reality. Loss of PAHs by photochemical transformations has been cited as being one of the most important decomposition mechanisms (Masclet et al., 1988). Such processes are expected to affect PAHs in association with particulates less than those in the vapour phase (Butler & Crossby, 1981). At present it is not possible to give accurate halflives for atmospheric PAHs under realistic conditions, although it is accepted that half-lives can be anything from hours to weeks. It also appears that such transformations are less effective if the PAHs are adsorbed on particulates such as soot or fly-ash. However, as stated earlier, such particles are susceptible to settlement and/or scavenging, thus removing them from the atmosphere. Movement of air masses away from the UK also constitutes a loss mechanism of PAHs. It is likely that PAHs are transported from the UK eastwards towards Scandinavia and mainland Europe where they may be deposited. The amount of PAHs transported by this method is very difficult to assess. However, there is the potential for many tonnes of PAHs to be transported. For example, if the UK air mass was replaced every 2.5 days (assumes a wind air speed of 8 km h-‘) over 5000 tonnes of associated PAHs could be transported from the UK every year. Of course, in reality, under certain prevailing winds the UK could receive PAHs emitted elsewhere. Nonetheless, long-range transport can clearly contribute to the PAH input/output atmospheric imbalance. Atmospheric deposition of PAHs near point sources is another mechanism which can also be invoked to explain the apparent imbalance between release and deposition. The overall UK PAH wet/dry deposition figure of 210 tonnes represents average national deposition away from immediate sources. However, the release figure of at least 900 tonnes includes all

quantifiable releases, except unregulated fires and volatilisation from soil and other surfaces. Several studies have shown that for some PAHs, especially the particulate phase compounds, deposition in the immediate vicinity of the emission site occurs. This explains the elevated soil PAH concentrations at source points such as road-sides. In fact, Hewitt and Rashed (1990) found that up to -30% of some PAHs are deposited within 50 m of major motorways. Deposition close to source is also likely to be significant near open burning sources such as coal/wood fires, straw burning, etc. Thus, the PAH deposition figure is likely to substantially underestimate the actual PAH loss from the atmosphere. Taking into account the enhanced PAH deposition fluxes near point sources and the other atmospheric loss processes, it is possible that inputs and outputs balance. However, it is possible that as for other nonvolatile organic compounds such as PCBs and PCDD/Fs, volatilisation from the soil surface with subsequent transport within, and redeposition from, the atmosphere will be occurring (Harrad & Jones, 1992; Harrad et al., 1993). This flux is considered in the following section. INPUTS AND OUTPUTS OF PAHs TO/FROM THE TERRESTRIAL ENVIRONMENT Introduction This section quantifies the inputs of PAHs to the terrestrial system. The terrestrial environment can be viewed as a very important long-term sink for PAHs. There are two main inputs, disposal of waste material and atmospheric deposition. In addition to these there are other known sources such as creosote use, fuel spills and industrial wastewaters. Waste disposal Disposal of sewage sludge, municipal wastes and other controlled waste, etc., while themselves being sinks for PAHs (see earlier), can also be viewed as constituting a source to the terrestrial environment. Current estimates show that 42% of UK sewage sludge is applied to agricultural land (DOE, 1993~). However, about 55% of all sludge is disposed of to land (includes agriculture, dedicated sites, disposal within curtilage, etc.), while -8% is landfilled, -30% sea disposed and the remainder incinerated. Using the sewage sludge PAH concentrations given in Table 8, it is apparent that -10 tonnes of ZPAH per annum are applied to land and 1.4 tonnes of ZPAHs are disposed of to landfill in the form of sewage sludge every year. Other material disposed of to landfill include MSW and other controlled wastes. Taking the limited PAH data reported by Hagenmaier et al. (1986) and correcting for 65% per annum MSW organic matter content, it can be calculated that about 74 tonnes of CPAH (fluoranthane, benzo[b]fluoranthene, benzo[a]pyrene and benzo[ghi]perylene) are landfilled associated with MSW (assuming about 90% of MSW is landfilled).

Polynuclear aromatic hydrocarbons in the UK environment

As stated in the section of PAH production, during combustion ash residues are produced which may contain PAHs. Incineration of MSW causes a 90% reduction in the initial volume (Woodfield, 1987). The remaining 10% is composed of ash which constitutes about 2.4% of the initial waste weight (Woodfield, 1990). Thus, for every tonne of MSW incinerated, 24 kg of ash are formed. This ash residue is generally disposed of to landfill. If it is assumed that MSW incineration practices produce 72 000 tonnes of ash per year which contains PAH concentrations similar to those measured by Wild et al. (1992~) it can be calculated that 13.6 kg of CPAH are associated with MSW ash disposed of to landfill each year. Similarly, an estimated 1848 tonnes of ash are produced by the incineration of 77 000 tonnes (dry weight) of sewage sludge each year. This assumes that 1 tonne of dry sludge produces 24 kg of dry ash as opposed to 1 tonne of MSW (fresh weight). Taking the incinerated sewage sludge ash PAH concentrations given by Wild et al. (1992c) it is calculated that about 0.3 kg of CPAH are disposed of to landfill each year associated with sludge ash. About 61000 tonnes of chemical waste and 170 000 tonnes of clinical waste are also incinerated each year in the UK. Again assuming the same proportions of ash production and PAH concentrations similar to MSW ash (both very crude assumptions) it is estimated that about 1 kg of CPAH is landfilled. It is estimated that approximately 100 tonnes of other controlled wastes (commercial, industrial, demolition, special, imported) are also landfilled in addition to the waste materials already considered. If it is assumed that this material has a dry weight content of 65% and has a PAH content similar to that of urban soil, a CPAH load of 340 tonnes is derived. Clearly this is subject to substantial errors, but it serves to illustrate the potential contribution from this source. It is worth noting that about 107 million tonnes of mining and quarrying waste and about 80 million tonnes of agricultural waste (animal excreta) are also produced each year in the UK (not controlled waste as defined by the Control of Pollution Act, 1974). Most of this material is either stored or disposed of to land (not landfill). Again if we assume a dry weight of 65% and a CPAH content of urban soil, then about 640 tonnes of XPAH are associated with these wastes (57% mining/ quarry waste, 43% agricultural waste). In summary, it is estimated that in excess of 400 tonnes of ZPAH are disposed of to UK landfill sites each year, with the majority being associated with untreated bulky controlled waste. Disposal of non-controlled waste and imported hazardous waste will add to this load. CPAH additions to other land uses from waste residues are tentatively estimated to be in the order of 650 tonnes, with the bulk coming from disposal of agricultural and quarrying/mining waste. In total it is estimated that disposal of waste arising constitutes the addition of over 1000 tonnes of CPAHs to land, although much will not be part of the general environmentally mobile ‘pool’ of PAHs, subject to recycling.

103

Atmospheric deposition

Atmospheric deposition was considered as a mechanism of PAH loss from the atmosphere. This process clearly results in the delivery and incorporation of PAHs into the terrestrial environment. It was estimated earlier that the UK atmosphere loses approximately 210 t CPAH per year due to wet/dry deposition each year. Thus, this process provides the terrestrial environment with approximately 20% of the amount of PAHs delivered by waste disposal: In fact the quantity of PAHs delivered by deposition is likely to be considerably higher than 210 tonnes due to greater deposition fluxes near to point sources. This is especially the case near point sources such as domestic coal/wood fires where the uncontrolled nature of the process results in significant PAH formation and production of particulate material. Soil PAH concentrations have been found to increase with time, as a response to atmospheric deposition of PAHs (Jones et al., 1989b). However, the longterm response of soil PAH concentrations to atmospheric deposition fluxes also depends on the effectiveness of PAH loss processes within the soil environment. Miscellaneous sources

This section details several methods by which PAHs may be introduced into the terrestrial environment. Some of these methods are not quantified, but are not generally considered to represent major sources. However, on a local scale they may dominate PAH inputs. Fuel spills and leakages are one method by which PAHs are uncontrollably released into the environment. This can take the form of major oil spillages to minor leakages of petrol from fuel storage tanks. Since fuels are relatively enriched with low molecular PAHs, fuel spills may be important in applying these compounds to some locations. PAHs may also be released into the environment via industrial wastewater. This source is currently unquantified, although it can be estimated using the figures of Davies et al. (1976) that MSW incinerators release approximately 1.44 kg of CPAH each year, associated with wastewater. Other minor sources include biogenic production, soil conditioners, natural fire ash, car tyre shredding, etc. Creosote is a distillate of coal tar made by high temperature carbonisation of bituminous coal (Merrill & Wade, 1985). Creosote has been used widely as a wood preserver both domestically and industrially for many years. It has been estimated that approximately 40 000 tonnes of creosote are manufactured each year in the UK, 25% of which is exported, 25% is used by industry while 50% is used for immersion treatment and retail domestic use (DOE, 1988). Merrill and Wade (1985) identified over 30 individual PAH compounds in a sample of US creosote. Typically, creosote consists of 85% PAHs, 10% phenols and 5% nitrogen and sulphur heterocycles (DOE, 1988). From this it is apparent that creosote treatment of wood potentially delivers to the UK over 25 000 tonnes of CPAH each year-20 times more than waste

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disposal practices and atmospheric deposition combined. It is also apparent that creosote-treated wood is potentially the largest sink in the UK environment. This is especially the case given that after 30 years of use, only a small decrease in CPAH concentrations was found in railway sleepers originally vacuum-highpressure treated with creosote (Rotard 8z Mailahn, 1987). Given the potential overwhelming importance of creosote as a source and sink of environmental PAHs, more investigations are warranted. Such investigations should attempt to assess the importance of creosote on the PAH cycle, since PAH emissions during combustion of creosote-treated wood may be elevated, while disposal of waste creosote-treated wood may be an important source of PAHs to landfill sites, etc. Losses of PAHs from the terrestrial environment In the previous sections of this paper the burden and sources of PAHs to the terrestrial environment have been estimated. It is important to emphasise that the environment is a dynamic system through which chemicals suc’h as PAHs are produced, transported, deposited and transformed. Thus, PAHs once incorporated into terrestrial compartments are not static and recalcitrant. Losses of PAHs can be via several processes; the major mechanisms include biodegradation, chemical transformation, photolysis, volatilisation, leaching, metabolism and erosion. Erosion is often disregarded as a loss mechanism. However, it is an important means of material removal from land to water. Without estimates of the total amount of sediment washed into the sea, the amount of PAHs lost from land cannot be quantified. PAHs in the dissolved phase are lost to the sea with river-water, while sea disposal of sewage sludge and marine outfall of sewage and treated wastewater can be mechanisms of PAH loss from the terrestrial system. Metabolism by higher animals is not likely to be important for the mass balance, but is the mechanism by which PAHs exhibit their carcinogenic potential. At present, the amount of PAH lost by metabolism within living tissue cannot be estimated. However, average livestock soil consumption rates are estimated to be about 8 kg per day (5% soil ingestion). Assuming a soil dry matter content of 60%, a vegetation dry matter content of 10% and average (vegetation and) rural soil CPAH concentrations given earlier (see Table 2), it can be estimated that about 14 tonnes of CPAH are consumed by the 63.4 million cattle, sheep and pigs in the UK each year. PAH metabolism usually occurs at the site of contact with the body, so presumably only a small proportion of this figure will be transferred across the gastro-intestinal tract intact. Leaching of PAHs from soil surface horizons to deeper aquifers may be viewed as a loss mechanism. However, most PAHs are either too well adsorbed on to soil organic matter or degradable that total leaching losses are not expected to be an important loss mechanism. Leaching of PAHs may be more significant where there is ground contamination with solvents.

For the purposes of this discussion, PAH losses by biodegradation and chemical transformation are considered together. PAH susceptibility to degradation is now well characterised as being structure dependent. Several studies have revealed that degradation potential decreases as the number of benzene rings in the PAH molecule increases (Sims & Overcash, 1983; Bossert & Bartha, 1986; Heitkamp & Cerniglia, 1987; Park et al., 1990; Wild & Jones, 1993). Half-lives for PAHs in the soil environment applied in the form of sewage sludge have been quantified both in the laboratory (Wild & Jones, 1993) and in the field (Wild et al., 1990; 1991a,b). PAHs applied to soil in sewage sludge are likely to be significantly more susceptible to degradation than PAHs already present in soil, since sludge provides a readily biodegradable substrate which stimulates bacterial and fungal activity. However, it is also possible that the sludge organic matter acts as a competitive carbon source which is degraded in preference to the less readily degradable PAHs. Indigenous soil PAHs may be strongly adsorbed on to soil organic matter and can be viewed as a relatively recalcitrant residual fraction that has been left in the soil after the readily degradable fractions have gone. Thus, PAHs present in most soils may be significantly more persistent than the sludge-to-land experiments cited above indicate. PAH degradation is also dependent on other factors such as previous contamination of the soil with PAHs, soil PAH concentrations, soil structure, environmental conditions, presence of vegetation, etc. However, if it is assumed that PAHs degrade in soils at rates similar to those obtained at the Luddington field experiment (Wild et al., 1991b), the soil CPAH content may fall by about 10% each year-ignoring fresh inputs (see Table 19). Clearly these estimates are speculative. In reality, losses by degradation are probably considerably lower than indicated. However, there is evidence to show that the high molecular weight PAHs exhibit a higher propensity for long-term enrichment in agricultural soils than low molecular compounds, due to their Table 19. Reductionin soil PAH burdendue to degradation Compound

Naphthalene

Soil burden (tonnes) 1 000

Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benz[a]anthracene/ chrysene Benzo[b]fluoranthene Benzo[a]pyrene Benzo[ghi]perylene

2 400 4 800 550 11 200 6400

ZPAH

50 000

Wild et al. (1991b).

9 5 3 5

100 100 400 900

Half-life* Soil concentration after (years) 1 year (tonnes) 2.1 3.2 5.7 7.9 7.8 8.5

760 2000 4400 515 10 500 6 000

8.1 9.0 8.2 9.1

8 500 4 800 3200 5600 46 000

Polynuclear aromatic hydrocarbons in the UK environment greater resistance to degradation and volatilisation (Jones et al., 19896). PAHs may also be lost by biodegradation and chemical transformations in water and sediments, while processes such as photolysis may be an important loss mechanism for some compounds in watercourses (Sanders et al., 1993). It is considered that there is insufficient information available in the literature to estimate the quantity of PAHs lost by these processes each year, although the low molecular weight PAHs will again be most susceptible. Soil-to-air recycling of PAHs is a potentially important, but poorly understood (and unquantified) flux mechanism. Volatilisation is influenced by several physicochemical and environmental factors. Table 1 indicates that the volatility (expressed as Henry’s constants) decreases as molecular weight increases. Unfortunately there is no satisfactory method that enables flux estimates of semivolatile organic chemicals, incorporated evenly through the soil, to be made. Measured volatilisation rates are also extremely sparse. Park et al. (1990) found that 3O?hof naphthalene added to soil was lost by volatilisation, while for other PAHs volatilisation was negligible. The importance of volatilisation losses for the low molecular weight PAHs has also been implicated in sludge-amended soils (see Wild & Jones, 1993). Empirical investigations have also indicated that some loss by volatilisation is possible, even for the high molecular weight PAHs (Jury et al., 1987). As an illustration it can be estimated that if only about 1% of the naphthalene in ‘uncontaminated’ soils volatilised, it would result in the release of over 8 tonnes to the atmosphere. This release is significantly higher than the atmospheric load derived earlier and equivalent to combustion related releases. Atmospheric concentrations of phenanthrene at a rural site (Halsall, C. & Jones, K. C., unpublished data) were significantly higher than those in urban areas mentioned earlier (Halsall et al., 1993). This observation may be explained, at least in part, by volatilisation from soil. If the observations at the rural site are apparent elsewhere, the atmospheric burden of phenanthrene (and possibly other low molecular weight PAHs) may have been significantly underestimated. It is also apparent that volatilisation of some PAHs from other sources such as sewage sludge, creosote-treated wood may also make important contributions to the atmospheric load of PAHs. Because of the potential importance of volatilisation from soil and the lack of reliable data, this process warrants detailed research. Summary Under the scenarios described in the preceeding sections it is estimated that the UK terrestrial environment currently receives over 1200 tonnes CPAH each year from waste disposal practices and atmospheric deposition. When other sources are considered, as well as enhanced deposition near combustion sources, this figure is likely to be significantly higher. In particular, the use of creosote to treat wood may release over 25 000 tonnes of CPAH per annum.

105

It is estimated that about 4000 tonnes of CPAH are lost from the terrestrial environment per annum due to degradation. Volatilisation from soil and other surfaces may substantially increase the total lost each year. Jones et al. (19893) and Wild et al. (1990, 1991b) have measured increasing soil CPAH concentrations in rural areas in southeast England due to atmospheric deposition up to the late 1980s. This imbalance is probably explained by an overestimation of the importance of soil degradation of PAHs above. The balance between terrestrial PAH inputs and outputs is currently undefined due to lack of information on transformation rates of indigenous soil PAHs. However, it seems likely that the high molecular weight PAHs are continuing to increase in concentration to the present, while phenanthrene deposited to soil in the past may be being rereleased to air (Jones et al., 1989b; Wild et al., 1990, 1991b) by outgassing now that ambient concentrations have declined (see next section). PAST AND FUTURE TRENDS IN PAH EMISSIONS AND ENVIRONMENTAL BURDEN PAH emissions and environmental levels are clearly linked and cannot be considered independently. Several studies have quantified PAH concentrations in archived (or dated) environmental samples to obtain evidence of changing levels over time. The environmental matrices used include soils (Jones et al., 1989b; Wild et al., 1990, 1991b), sediments (Sanders et al., 1993) and vegetation (Wild et al., 19926; Jones et al., 1992). These studies have consistently shown that from the 1930s through to the 195Os/196Os CPAH deposition fluxes were significantly higher than those observed today. Since the 1960s CPAH deposition fluxes have declined, although contemporary fluxes are higher than those apparent in pre-industrial times. There is evidence that not all individual PAHs behave the same. For example, phenanthrene soil concentrations appear to have declined since the 1960s (probably due to degradation and volatilisation), whereas benzo[a]pyrene and other high molecular weight PAHs appear to be accumulating in soils (Jones et al., 1989~). As this paper has shown, PAHs are formed by incomplete pyrolysis and as such the trend in PAH atmospheric deposition and environmental levels must be related to the changing combustion inputs. Coal combustion was the principal means of power generation and domestic heating in the UK up to the 1960s. Inefficient and uncontrolled residential burning of coal is likely to be the principal reason for high deposition fluxes, observed until the 196Os, although power generation to that date would have emitted more PAHs than power stations today due to poorer combustion conditions. Since the 1960s coal utilization has fallen by almost 75%, being replaced with ‘cleaner’ fossil fuels, such as natural gas, oil and petroleum, while there have been concurrent improvements in emission control measures. Since the 1960s however, PAH emissions from vehicles will have probably increased. The use of

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in the UK appears to have decreased at a rate of about 3% per annum since the early 1970s (DOE, 1988). Contemporary atmospheric deposition PAH fluxes are probably lower than at any period during this century in the UK, as illustrated by contemporary herbage PAH concentrations (Jones et al., 1992; Wild et al., 1992b). Future trends in PAH emission are difficult to predict, although continued reduction in coal use, the cessation of stubble burning, improvements in combustion technology and pollution abatement suggest that the trend will continue downward. As a caveat to this, incineration as a waste disposal process is increasing in popularity, although this paper indicates that PAH emissions from incinerators are likely to be relatively insignificant nationally. Future national PAH emissions from vehicles are difficult to predict, given the continuing increases in (a) vehicles, (b) the use of unleaded petrol, and (c) catalytic converters. The effect of decreasing PAH deposition fluxes on soil PAH concentrations is difficult to predict. However, as stated earlier, soil concentrations of the high molecular weight PAH compounds are continuing to increase due to inputs exceeding degradational losses. In contrast, the low molecular compounds may be lost from soils by volatilisation quicker than the influx rate; thus, the soil acts as a source of PAHs to the UK atmosphere. creosote

CONCLUSIONS

AND RECOMMENDATIONS

This paper presents a detailed but preliminary inventory of PAH sources and sinks in the UK. Transformations within and fluxes between environmental compartments have also been tentatively estimated. The ZPAH burden of the UK natural environment is estimated to be -54000 tonnes, with ~90% of this associated with soil. There is the potential for this estimate to increase dramatically if contaminated soils are included. However, it is considered that at present there is not enough information available regarding the size and number of sites to obtain an accurate PAH burden. If this information becomes available, it may show that the UK burden is predominantly associated with contaminated soils. Combustion probably results in the release of over 1000 tonnes of CPAH to the UK atmosphere each year, with ‘uncontrolled’ burning the major source. Combustion of fuel by vehicles is also an important source of PAHs, emitting over 10 times more PAHs than power stations and waste incineration combined. Over 210 tonnes of CPAH are lost from the UK atmosphere by wet/dry deposition each year, while unquantified amounts are lost by deposition near point sources, transformations in the atmosphere and by long range transport. Atmospheric deposition and waste disposal result in over 1200 tonnes of PAHs entering the UK terrestrial environment each year, while the use of creosote has the potential to release over 25 000 tonnes of CPAH each year. Thus, creosote treated

wood must also be considered to be a major PAH sink, possibly containing more PAHs than the ‘uncontaminated’ soil environment. This paper has highlighted major gaps in the data required to present a detailed precise budget and source inventory. There is a lack of PAH concentration data for some important environmental matrices, while there is a serious lack of data on low molecular weight PAHs and vapour phase emissions from some combustion sources. It is clear that only reporting particulatebound emissions from combustion sources seriously underestimates inputs to the environment. Thus PAH emissions from some combustion sources, such as domestic wood combustion, stubble burning and unregulated fires, cannot currently be adequately assessed. Atmospheric residence times (half-lives) and the importance of volatilisation from soil and other surfaces also needs measurement. The potential volatilisation of PAHs from contaminated sites and creosote-treated wood requires particular attention. It is surprising, given the known carcinogenicity of some of the PAHs and their prevalence in the environment, that there are such serious gaps in our understanding. In contrast, PCDD/Fs (which are considerably more problematical and expensive to quantify) are probably better understood, even though health effects of PAHs have been more clearly demonstrated than those of the PCDD/Fs.

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Westerholm, R. N. et al. (1988). Effect of fuel polycyclic aromatic hydrocarbons and other mutagenic substances from a gasoline-fuelled automobile. Environ. Sci. Technol., 22, 925-30. Wild, S. R. (1991). The fate and behaviour of polynuclear aromatic hydrocarbons in sewage sludge amended agricultural soils and their uptake by plants. PhD thesis, Lancaster University, UK, unpublished. Wild, S. R. & Jones, K. C. (1992). Polynuclear aromatic hydrocarbon uptake by carrots grown on sludge,_amended \ soils. J. Environ. Qual., 21, 217-25. Wild, S. R. & Jones, K. C. (1993). Biological and abiotic losses of polynuclear aromatic hydrocarbons (PAHs) from soils freshly amended with sewage sludge. Environ. Toxicol. Chem., 12, 5-12. Wild, S. R., Waterhouse, K. S., McGrath, S. P. & Jones, K. C. (1990). Organic contaminants in an agricultural soil with a known history of sewage sludge amendments: Polynuclear aromatic hydrocarbons. Environ. Sci. Technol., 24, 170611. Wild, S. R., Berrow, M. L. & Jones, K. C. (1991~). The persistence of polynuclear aromatic hydrocarbons (PAHs) in sewage sludge-amended agricultural soils. Environ. Pollut., 72, 141-57. Wild, S. R., Obbard, J. P., Munn, C., Berrow, M. L. & Jones, K. C. (199lb). The long-term persistence of polynuclear aromatic hydrocarbons (PAHs) in an agricultural soil amended with metal contaminated sewage sludges. Sci. Tot. Environ., 101,235-53. Wild, S. R., Berrow, M. L., McGrath, S. P. & Jones, K. C. (1992~). Polynuclear aromatic hydrocarbons in crops from long-term field experiments amended with sewage sludge. Environ. Polk, 76, 25-32. Wild, S. R., Jones, K. C. & Johnston, A. E. (1992b). The polynuclear aromatic hydrocarbon (PAH) content of herbage from a long-term grassland experiment. Atmos. Environ., 26A, 1299-307. Wild, S. R. Mitchell, D., Yelland, C. M. & Jones, K. C. (1992~). Arrested municipal waste incinerator ash as a source of polynuclear aromatic hydrocarbons (PAHs) to the UK environment. Waste Man. Res., 10, 99-l 11. Williams, P. T. (1990). A review of pollution from waste incinerators. J. Inst. Wat. Environ. Man., 4, 2634. Wilson, S. C. 8c Jones, K. C. (1993). Bioremediation of soil contaminated with polynuclear aromatic hydrocarbons (PAHs): A review. Environ. Polk., 81, 229949. Woodfield, M. (1987). The environmental impact of refuse incineration in the UK. Warren Spring Laboratory Report. Stevenage, Hertfordshire, UK. Woodfield, M. J. (1990). Warren Spring Laboratory Reports. Stevenage, Hertfordshire, UK. 19861990. World Health Organisation (1987). Emissions of Heavy Metals and PAH Compounds from Municipal Solid Waste Incinerators: Control Technology and Health Effects, Geneva, Switzerland.