Towards a scheme of toxic equivalency factors (TEFs) for the acute toxicity of PAHs in sediment

Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

Contents lists available at ScienceDirect

Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv

Towards a scheme of toxic equivalency factors (TEFs) for the acute toxicity of PAHs in sediment Tom T. Fisher a,n, Robin J. Law a, Heather S. Rumney a, Mark F. Kirby a, Carole Kelly b a b

The Centre for Environment, Fisheries and Aquaculture Science, Cefas Lowestoft Laboratory, Pakefield Road, Lowestoft, Suffolk NR33 0HT, UK Department for Environment, Food and Rural Affairs Nobel House, 17 Smith Square, London SW1P 3JR, UK

a r t i c l e i n f o

abstract

Article history: Received 18 May 2011 Received in revised form 13 July 2011 Accepted 23 July 2011 Available online 31 August 2011

Toxic equivalency factors/quotients (TEF/TEQs) express the toxicity of complex mixtures. For PAHs, TEF values are available for assessing their carcinogenic potential and are expressed as benzo[a]pyrene equivalents. This study develops a similar approach for their acute toxicity in sediments. Acute toxicity (10 day EC50) values were generated using the marine amphipod Corophium volutator bioassay for twelve low molecular weight PAHs. The results ranged from 24 to 4 1000 mg/Kg sediment dry weight for 4-methyldibenzothiophene and anthracene, respectively. Phenanthrene was used as the reference compound (TEF ¼ 1) and so the TEQ values derived are expressed as phenanthrene equivalents. In order to illustrate the applicability of this approach to the development of marine indicators we plotted TEQ values for acute toxicity to UK environmental monitoring data. Further work is required to validate the TEF values produced and to extend the TEQ approach to include a wider range of low molecular weight PAHs. & 2011 Elsevier Inc. All rights reserved.

Keywords: Polycyclic aromatic hydrocarbons Acute toxicity Toxic equivalency factors/quotients (TEF/TEQ) Sediment Indicators Corophium volutator

1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are important and ubiquitous contaminants in UK waters (Kirby et al., 2000; Woodhead et al., 1999). They are present in fossil fuels and are typically formed by the incomplete combustion of organic material. PAHs enter the marine environment from domestic sewage, industrial discharges, combustion followed by atmospheric transport and the spillage or disposal of oil and petroleum products (Kennish, 1998; Meador et al., 1995). The fact that there are many sources of PAHs and that they are readily adsorbed onto sedimentary particles in the aquatic environment has lead to high concentrations being recorded in marine sediments, particularly inshore and in historically industrialised estuaries (Woodhead et al., 1999). Worldwide, concentrations of PAH vary widely. For example, in a study of 99 sediment samples taken around the UK between 1993 and 1996, concentrations of phenanthrene, naphthalene and fluorene reached a maximum of 6.2, 2.4 and 2.3 mg/Kg dry weight, respectively (Woodhead et al., 1999). Naphthalene concentrations as high as 120 mg/Kg have been detected at highly contaminated sites, such as in estuaries and harbours and around drilling platforms

n

Corresponding author. E-mail address: tom.fi[email protected] (T.T. Fisher).

0147-6513/$ - see front matter & 2011 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2011.07.023

(Anderson, 1982; Moore and Ramamoorthy, 1984; Witt, 1995; P Woodhead et al., 1999). It is also important to consider the PAH concentrations that have been recorded in environmental samples, in particular at dredged material disposal sites. In the UK alone there are over 150 disposal sites including the Rame Head site, which serves the marine disposal requirements of the River Tamar and Plymouth Sound. The site lies in water depths of 18–38 m and is located approximately 2 km west of Rame Head and 6 km west of the entrance to Plymouth Sound. At the Rame Head disposal site, P the PAH concentrations ranged from 0.4 to 2.6 mg/Kg dry weight (Allen, 2005) and 410 mg/Kg in the Tyne and Tees estuaries (Woodhead et al., 1999). In contrast a study in Sydney harbour, P Australia, showed the maximum PAH concentrations to be as high as 380 mg/Kg dry weight (McCreedy et al., 2000). In order to ascertain the toxicity of PAHs to marine organisms a variety of ecotoxicological studies have been conducted on a number of species. However there are major data gaps for acute toxicity for low molecular weight (low MW) compounds in sediment. Few data that do exist have been focussed almost exclusively on the main parent compounds (naphthalene, anthracene and phenanthrene) with little information (actual or predictive) readily available for the methyl- and dimethyl-associated compounds or for the dibenzothiophenes. As these alkylated PAHs predominate in oil and petroleum products, this is a major gap. The limited literature information includes median lethal (LC50) sediment toxicity data for two individual low MW PAHs

2246

T.T. Fisher et al. / Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

by Swartz et al. (1997), who conducted 10 day spiked sediment tests with the marine amphipod Rhepoxynius abronius. They related their toxicity data to levels of organic carbon (OC) within the sediment and state 10 day LC50s of 2.22 and 3.31 mg/g OC for phenanthrene and fluoranthrene, respectively. DeWitt et al. (1992) produced 10 day LC50interstitial water (iw) values for the same two PAHs using another amphipod, Leptocheirus plumulosus, but related their figures to calculated interstitial water loadings based on the partition coefficient characteristics of the substances. LC50iw figures of 1496 and 309 mgL  1 were quoted for acenaphthene and phenanthrene, respectively. Unpublished data from a similar study (again using the interstitial water loading approach) using yet another marine amphipod, Eohaustorius estuarius, is quoted by Swartz et al. (1995). For this species, 10 day LC50iw values are quoted as 708 and 158 mg/L for acenaphthene and phenanthrene, respectively. Lotufo and Fleeger (1997) conducted investigations into the impacts of sediment associated phenanthrene on the meiobenthic harpacticoid copepods, Schizopera knabeni and Nitocra lacustris. The resultant 10 day LC50s ranged from 71 (nauplius stage) to 105 (adult) mg/Kg dry weight for N. lacustris and 84 (nauplius stage) to 349 (adult) mg/Kg dry weight for S. knabeni. The only other significant study pertaining to this class of compound in sediments was by the same team (Lotufo and Fleeger., 1996), using the freshwater tubificid oligochaete, Limnodrilus hoffmeisteri, to investigate the toxicity of phenanthrene. After a 10 day exposure, LC50 values for this species were quoted as 298 mg/Kg, with sublethal impacts detected at concentrations as low as 25 mg/Kg. Other toxicity studies on low MW PAHs have tended to use water borne exposures to a variety of species including barnacle larvae (Donahue et al., 1977), unicellular algae (Djomo et al., 2004) and bioluminescent bacteria (Steevens et al., 1999). These studies each included an investigation into the toxicity of phenanthrene, naphthalene and anthracene but produced no common toxicity order for these compounds, suggesting that these water borne assays are species and technique dependant. Due to the lack of consistently expressed and reliable data sets, our study has generated toxicity data for a group of 12 low MW PAHs (physical characteristics are summarised in Table 1) using a marine amphipod. Amphipods have been shown to be very sensitive to PAH and oil polluted sediments, and are among the first benthic groups to decrease in abundance or disappear from contaminated areas (Baden, 1990; Dauvin, 1998; Rutt et al., 1998; Jewett et al., 1999). The species Corophium volutator was chosen for this initial study because it is particularly sensitive to common sediment contaminants, covers a broad geographical range and is recommended for use in sediment toxicity tests in European waters, especially in the UK (Chapman, 1992; Peters and Ahlf, 2005). The low MW PAHs selected were chosen because they are

common constituents of oil, are routinely determined in sediments within the UK Clean Seas Environment Monitoring Programme and are available in suitable purity from chemical suppliers. PAHs exert both acute and chronic toxic effects on marine biota. In general terms, it is the low MW PAH (e.g. 2- and 3-ring compounds), which exert acute toxicity, and some of the higher molecular weight PAHs (with 5-rings or more), which have carcinogenic potential, exert chronic toxicity (Harvey, 1991). PAHs themselves are not direct carcinogens, but are converted into carcinogenic derivatives when metabolised. Toxic equivalency factors (TEFs) have been developed for a number of these potentially carcinogenic PAHs, the factor for each of the PAHs expressing its potency relative to benzo[a]pyrene, which has a TEF of unity. The concentration of each of the individual PAH compounds is multiplied by its TEF, and these values are summed to yield benzo[a]pyrene equivalent concentrations. By this means, the concentrations of a suite of PAHs can be represented by a single concentration, which reflects the overall carcinogenic potential of the PAHs within the sample for which TEFs have been assigned. In the past, this technique has been applied successfully to shellfish monitoring studies, both following oil spill incidents and in wider monitoring programmes (Law et al., 2002). As an aid to the assessment of data for PAH in sediments from a variety of studies, a similar system for the acute toxicity of low MW PAHs was developed using the generated acute ecotoxicity data for C. volutator. The TEF approach is adopted because PAH contamination rarely consists of single compounds, but rather of mixtures of compounds that can affect the environment (Engraff et al., 2011). Typically risk assessment has focused on single compounds (Altenburger and Greco, 2009), which can lead to an underestimation of the actual risk, since the detrimental effects of PAH mixtures can be higher compared to the case where the assessment of environmental effects is based on single compounds only. By testing the toxicity of individual compounds and utilising the TEF approach we can begin to predict the toxicity of the combined PAHs present at a dredge disposal site and decide where to focus resources and subsequent ecotoxicity testing.

2. Materials and methods 2.1. Collection and storage of test organisms and sediment: Corophium volutator were purchased from Aqua Logistix, Dumfries, Scotland. The animals were acclimatised for at least a week prior to testing and were held in glass aquaria containing reference sediment and seawater at 157 2 1C. Reference sediment used for testing was collected from Shoebury Sands, Essex, UK. Only the upper layer of aerobic sediment (usually the top 5–10 cm) was retained for testing. Once in the laboratory the sediment was sieved through

Table 1 Physical characteristics of the target PAHs including water solubility at 25 1C (mg/L) and molecular weight (Daltons). PAH

Purity (%)

Molecular weight

log kow

Water solubility

4-methyldibenzothiophene Dibenzothiophene Phenanthrene 1-methylnaphthalene Acenaphthylene Acenaphthene Fluorene 2,3-Dimethylnaphthalene 2-methylnaphthalene Naphthalene 2-methylanthracene Anthracene

96 99 90 97 480 496 98 98 97 498 97 Z 99

198.29 184.26 178.22 142.20 152.20 154.21 166.22 156.22 142.21 128.16 192.26 178.22

4.84 4.38 4.46 3.87 3.94 3.92 4.18 4.40 3.86 3.37 5.00 4.54

0.26 0.21 1.30 25.80 2.49 2.53 1.68 12.44 24.60 31.70 0.20 0.04

T.T. Fisher et al. / Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

2247

a 500 mm sieve to remove the larger benthic organisms that may predate on the test organisms. In order to verify that the sediment did not contain significant background levels of PAH a sample was taken and analysed. The results showed no detectable levels of PAH (detection limit 0.1 mg/Kg). The reference sediment had a median grain size of 156.19 mm comprising of 4.66% fines (o 63 mm). The organic carbon (OC) content was 0.12%. The C. volutator were not fed during these studies as it is generally considered that feeding during sediment toxicity testing should be avoided (Hill et al., 1993).

control mortality remained within validity criteria i.e. o 20% (Roddie and Thain, 2002). In addition, two studies were conducted in order to assess the sensitivity of the C. volutator; one using the reference toxicant ammonium chloride and the other using ivermectin. The subsequent LC50 values produced were within acceptable limits (Roddie and Thain, 2002).

2.2. Toxicity testing

In order to chemically verify the stock concentration (highest test concentration used to prepare other test concentrations), samples were taken for analysis at days 0 and 10. The samples were taken by transferring approximately 25 g of test sediment to hexane-rinsed 60 ml glass jars and the samples were then immediately frozen at  20 1C prior to analysis. The determination of PAH concentrations in spiked test sediments was carried out using established methods that use GC/MS in full-scan, electron impact mode. The method is fully described elsewhere (Kelly et al., 2000). Briefly, sediments are subjected to alkaline saponification using methanolic KOH and the digests solvent extracted with pentane to recover PAH. Extracts are cleaned-up using alumina column chromatography, reduced in volume and transferred to an injection vial. Deuterated analogues of the parent PAH determined are used as surrogate standards. All samples are run under an analytical quality control (AQC) protocol, which involves analysing procedural blanks and reference materials within each sample batch. In this study we used a marine sediment reference material (HS-6) from the Canadian National Research Council, produced within its marine analytical chemistry standards programme. A detection limit of 0.1 mg/Kg dry weight in sediment for each compound or class is one, which we have found readily achievable in all samples, and that provides an adequate detection capability for interpretation of the data.

The Corophium volutator tests were conducted in accordance with internationally recognised guidelines (Roddie and Thain, 2002). 2.2.1. Spiking sediment with PAH Test solutions of PAH were prepared on the day the sediment was spiked. This was achieved by dissolving a specific quantity of the required PAH in 50 ml of acetone (Fisher Scientific, 98%). In order to allow for any potential toxic effects of acetone an additional control was set up using an identical amount of acetone; this was termed the ‘solvent control’. Once dissolved the solution was added to an appropriate amount of ‘dry sediment’, which achieved a stock concentration of PAH to dry sediment (mg/Kg dry weight). This was then left for one hour to allow the solvent to evaporate. Appropriate quantities of the stock sediment were added to ‘wet sediment’ in order to create a concentration series; each treatment had a total dry weight of 1 kg. To calculate the amount of ‘wet sediment’ required the dry weight of the reference sediment was calculated using the method described in Roddie and Thain (2002). The spiked sediment was added to filtered (0.2 mm) seawater (2:1 ratio) and was shaken for 3 h on an orbital shaker to ensure thorough mixing. Once removed from the shaker the sediment slurry was added to 1 L glass beakers (3 replicates of approximately 300 g); this was then left overnight and then reference seawater was added to the 850 ml graduation mark. The beakers were aerated throughout the study.

2.3. Chemical analysis

3. Results 2.2.2. Corophium volutator survival assay Corophium volutator were collected from the stock tank using a 500 mm sieve and individuals between 4–6 mm long were selected for testing. A total of ten organisms were added to each beaker; 3 replicates were set up for each of a minimum of 7 test concentrations and the concentration separation factor did not exceed 3.2. After 10 days exposure, the C. volutator were sieved from the sediment and the number of surviving organisms recorded. Missing animals were presumed dead. The physical parameters (salinity, dissolved oxygen and temperature) were monitored at least four times during the study (typically on days 0, 2, 4, 7 and 10) and, when necessary, reverse osmosis water was added to adjust the salinity to within guideline limits (307 5%). As the toxicity of each PAH to Corophium volutator was unknown, a shortened version of the test was conducted initially using the same procedure but with larger concentration ranges. 2.2.3. Statistics The 50% lethal concentration (LC50), no observed effect (NOEC) and lowest observed effect (LOEC) concentrations were calculated using the Toxcalc statistical programme version 5 (Tidepool Scientific Software, USA). 2.2.4. Test validity and quality assurance Physical parameters remained within the following guideline limits: salinity (307 5%), dissolved oxygen, 485%; temperature, 15 7 2 1C; pH, 87 0.5. The mean

3.1. LC50 LC50 determinations The 10 day LC50s to C. volutator of the selected PAH varied significantly, from 24 to 41000 mg/Kg based on dry weight of sediment (Table 2). These figures are based on nominal sediment concentrations. The highest toxicity (lowest LC50) was induced by 4-methyldibenzothiophene and the lowest by anthracene. The only PAH tested, which did not elicit any acute toxicity over the 10 day period, was anthracene. During the initial studies, no significant mortality occurred even at an anthracene concentration of 1000 mg/Kg. In the environment, anthracene concentrations are unlikely to reach these levels therefore it was unrealistic to test any higher concentrations. The other eleven PAHs did cause acute toxicity to C. volutator, resulting in EC50s between 24 and 861 mg/Kg. The LOEC also varied from 10 to 41000 mg/Kg. Some of the PAHs tested created a LOEC higher than the LC50 value produced; this occurred because there was no induced high mortality ( 450%) at the NOEC concentration and the next highest concentration significant mortality occurs e.g. 90%. Therefore the

Table 2 PAH results for 10 day acute sediment bioassay using the marine amphipod Corophium volutator (mg/Kg dry weight). Values are based on nominal concentrations. PAH

LC50 (95% confidence limits)

NOEC

LOEC

TEF

4-methyldibenzothiophene Dibenzothiophene Phenanthrene 1-methylnaphthalene Acenaphthylene Acenaphthene Fluorene 2,3-Dimethylnaphthalene 2-methylnaphthalene Naphthalene 2-methylanthracene Anthracene

24 (16–33) 48 (33–73) 103 (68–142) 129 (91–176) 173 (121–220) 389 (292–522) 457 (347–602) 524 (412–636) 644 (490–847) 845 (621–1150) 861 (715–1036a) 41000

10 5.6 25 10 100 40 100 50 32 56 500 1000

30 10 100 30 300 100 320 300 50 100 1000 41000

4.3 2.2 1.0 0.8 0.6 0.3 0.2 0.2 0.2 0.1 0.1 0.1

a

indicates an extrapolated value as it is above the highest concentration tested.

2248

T.T. Fisher et al. / Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

Table 3 Chemical confirmation of PAH concentrations during study, all units expressed in mg/Kg dry weight. PAH

Nominal concentration

Day 0 measured concentration

Day 10 measured concentration

Percentage recovery of nominal Day 0

Percentage recovery of nominal day 10

Percentage difference in measured concentrations over the test period

2-methylanthracene Dibenzothiophene Anthracene Phenanthrene Fluorene 4-methyldibenzothiophene Naphthalene 2-methylnaphthalene 1-methylnaphthalalene Acenaphthylene Acenaphthene 2,3-dimethylnaphthalene

1000 320 1000 1000 1320 500 1560 1560 1000 1000 1300 1500

1243 310 716 920 1172 835 1419 1501 863 734 1512 1797

1244 240 567 410 1026 371 499 782 673 349 1411 1631

124 97 72 92 89 167 91 96 86 73 116 120

124 75 57 41 78 74 32 50 67 35 109 109

0 22 21 55 12 56 65 48 22 52 7 9

LC50 is between the NOEC and LOEC. The LC50 is statistically produced where as the NOEC and LOEC depend on the chosen test concentrations.

3.1.1. Chemical analysis Chemical analysis showed that, at the initiation of the study, all measured concentrations were within 728% of the nominal concentration, except 4-methyldibenzothiophene where the recovery was very high (167% of the nominal) (Table 3). The concentration stability of the PAH spiked sediments varied widely, with measured concentrations on day 10 ranging from 7% to 65% (Table 3) remaining in comparison to the initial concentrations. The following PAHs remained within 722% of the initial measured concentration during the test period: 2-methylanthracene, dibenzothiophene, anthracene, fluorene, 1-methylnaphthalene, acenaphthene and 2,3-dimethylnaphthalene. However, the other PAHs were less stable during the test period: phenanthrene, 4-methyldibenzothiophene, naphthalene, 2-methylnaphthalene and acenaphthylene showed between 48% and 65% reduction compared to the initial measured concentration (Table 3).

3.2. Toxic equivalency factors 3.2.1. Development Using the LC50 data generated for Corophium volutator, toxic equivalency factors were calculated for all of the low MW PAHs studied. As it was centrally ranked in the toxicity range, phenanthrene was chosen as the reference compound to which the others were related (TEF ¼1). Essentially the numbers used in the equation below for calculating the acute TEQ values are the ratios of the toxicity compared to phenanthrene i.e. if the number is 4 1 that particular PAH is more toxic than phenanthrene. These numbers are the TEF values and are also displayed in Table 2. Therefore in order to assess the potential toxicity of a particular site, traditional chemistry data and the TEF values are used to create a TEQ value. For each PAH concentration determined in a particular environmental sample it is multiplied by the appropriate TEF and summed in order to ascertain the TEQ. TEQ (as phenanthrene equivalents)¼(0.6  acenaphthylene concentration)þ (0.26  acenaphthene concentration)þ(0.23  fluorene concentration)þ(0.12  naphthalene concentration)þ(0.23  C1-naphthalenes concentration)þ(0.2  C2-naphthalenes concentration)þ phenanthrene concentrationþ(0.12  C1-phenanthrenes/anthracenes concentration)þ(2.15  dibenzothiophene concentration)þ(4.3  C1-dibenzothiophenes concentration).

This approach can be refined and extended as additional data become available. 3.2.2. Application In order to demonstrate the utility of this approach for use as an indicator, two datasets were investigated. These were the PAHs in sediment data from the England and Wales component of the UK Clean Seas Environmental Monitoring Programme (CSEMP) and the dredged material disposal site monitoring programme, which Cefas operates on behalf of the Marine Management Organisation. Along with calculating the acute toxicity TEQ, the TEQ relating to carcinogenic potential was also calculated. For this purpose we used the same factors as those employed in Law et al. (2002): TEQ (benzo[a]pyrene equivalents)¼(1.05  dibenz[a,h]anthracene concentration)þbenzo[a]pyrene concentrationþ (0.25  indeno[1.2.3 -cd]pyrene concentration)þ (0.13  pyrene concentration)þ(0.11  benzo[b]fluoranthene concentration)þ(0.07  benzo[k]fluoranthene concentration)þ(0.03  benzo[ghi]perylene concentration)þ(0.02  fluoranthene concentration)þ (0.014  benz[a]anthracene concentration)þ(0.013  chrysene concentration) The derived TEQ data from 2008 were plotted together (see Fig. 1a–c). In each case, the left-hand bar shows the acute toxicity summation and the right-hand bar the carcinogenic potential summation. In Fig. 1a, the acute toxicity potential is highest in the eastern Irish Sea and off the Tyne and Humber estuaries in the North Sea. Fig. 1b and c show plots for two of the disposal sites studied, in Tees Bay and at Rame Head (outside Plymouth Sound) . As can be seen from Fig. 1c, sediments off the Tees are dominated by low MW PAH, while around Rame Head (Fig. 1b) the proportion of high MW PAHs is much greater. These simple graphical illustrations summarise the concentration data for ten individual low MW PAHs and groups and ten individual high MW PAHs in relation to their potential impacts on sediment-dwelling organisms.

4. Discussion: TEQ data (acute TEQ for low MW PAH and chronic TEQ for high MW PAH) for 13 offshore CSEMP stations (TEQ ranges indicated) and seven dredged material disposal sites, all sampled in 2008, are given in Table 4. This clearly highlights the importance of (primarily oilderived) low MW PAH in the coastal estuaries of NE England (Tyne and Tees: North Tyne, Souter Point and Tees disposal sites) and of high-MW parent (primarily combustion-derived) in the North Tyne and Tees, and at Goole in the River Ouse, in terms of

T.T. Fisher et al. / Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

2249

Fig. 1. Average toxic equivalency quotients (TEQ) for low MW PAH and high MW PAH in relation to their acute toxicity and carcinogenic potential (as mg/Kg phenanthrene and benzo[a]pyrene, respectively): (a) at stations around England and Wales sampled under the UK Clean Seas Environmental Monitoring Programme, (b) at stations around the Tees dredged material disposal site, (c) at stations around the Rame Head dredged material disposal site.

Table 4 Ranges of acute and chronic PAH TEQs found at 13 CSEMP stations and 7 dredged material disposal sites in 2008 (mg/Kg phenanthrene equivalents and mg/Kg benzo[a]pyrene equivalents, respectively). Location

Acute TEQ

Chronic TEQ

CSEMP Falmouth North Tyne Rame Head Scarborough Souter Point Tees Goole

6.1–932 57–632 1290–15,600 12–654 74–5740 1490–10,800 393–31,300 161–2940

0.2–195 28–393 221–1380 5.1–502 18–710 172–736 47–2210 63–2030

their potential to cause environmental harm during 2008. The highest TEQ values, and therefore those sites likely to exhibit the greatest acute toxicity, were around the Tees, North Tyne and Souter Point disposal sites. The lowest TEQ values were seen at the Rame Head disposal site. Looking in closer detail at the Tees and Rame Head data (Fig. 1b and c), there is a lot of variation in TEQ values around the dredged sites. Tees Bay A shows greater

low MW acute and chronic TEQ values (more potential toxicity) than Bay C. In addition, when compared to Rame Head the TEQ values are many times higher. As demonstrated these TEQ indicators can assist in visualising environmental levels of contaminants and their potential impacts. This could be used to aid monitoring programmes and to decide where resources should be concentrated during toxicity assessments and sampling. Analytical support was used during this study. The purpose of this was two-fold: (i) to provide chemical confirmation of laboratory spiked stock concentrations, and (ii) to give an insight into the stability of the individual PAH in the test system. During this study the initial measured concentrations supported the nominal concentrations within 728% for all except 4-methyldibenzothiophene, where the data suggest that the nominal LC50 may be an over estimation i.e. this compound may be of lower toxicity as the initial measured concentration was greater than the nominal. The data also suggest that the LC50 values calculated for phenanthrene, 4-methyldibenzothiophene, naphthalene, 2-methylnaphthalene and acenaphthylene may be under-estimations as the concentrations decreased substantially over the test period, resulting in the Corophium volutator being exposed to a decreasing concentration

2250

T.T. Fisher et al. / Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

over time (Table 3). There are many possible reasons for the loss of PAH in the test system over time, microbial degradation is believed to be one of the major processes to remove PAHs (Quan et al., 2009) and is likely to be a significant factor in the loss of PAH during these studies. Aromatic hydrocarbons are also photosensitive ( Zepp and Cline, 1977; Zepp and Baughman, 1978; Zepp and Schlotzhauer, 1979). In addition the environmental conditions (in particular temperature and pH) will affect the rate of degradation as well as loss to biota, particulate matter and the atmosphere. When evaluating the LC50 values it is important to consider the physical characteristics of the PAH. It is well known that chemicals with a low log Kow require substantially higher concentrations to produce mortality, which is consistent with previous observations that LC50s are typically higher (lower toxicity) for hydrophilic compounds when compared to hydrophobic compounds (Landrum et al., 2003). This is due to the bioavailability being reduced. The LC50 values generated during this study however did not show any clear correlation with Log Kow, water solubility, molecular weight or percentage loss of the chemical during the test period. It is, therefore, likely that because both the molecular weight and log Kow values are relatively similar for the range of compounds tested that this did not play a significant role in the toxicity of the PAHs. It remains highly important that we are able to develop and subsequently test the validity of these approaches through the use of quality toxicity datasets as it is known that chemistry alone can often underestimate toxic potential (Hendriks et al. 1994) due to a range of factors. For example, the methods only work when assessing chemical mixtures that demonstrate toxic additivity. Recent work by Landrum et al. (2003) has used the additivity model for PAH proposed by Swartz et al. (1995) and so this approach seems appropriate for low molecular MW PAH; however in environmental samples synergistic activity may occur caused by other contaminants (Hendriks et al., 1994; Birnbaum and DeVito, 1995). The interaction of PAHs in mixtures may also cause antagonism where one PAH may repress the effects of other PAH (Cassee et al., 1998). In addition it is important to note when using these models for the prediction of sediment toxicity, the effects of sediment organic carbon on toxicity (Meador et al., 1995; Rogers, 2002) and the complex equilibrium kinetics that dictate bioavailability in sediments (O’Connor and Paul, 2000; Rogers, 2002). This latter issue has been somewhat addressed by the development of an equilibrium partitioning- toxic unit (EqP-TU) approach (Swartz et al., 1995; Rogers, 2002). A toxicity unit (TU) for a compound is the proportion of the total toxicity of a complex mix (as might be found in effluents or contaminated sediments) that can be attributed to that compound (Kirby et al., 1998). In theory the sum of the TUs for all toxicologically active compounds in a given medium should add up to 1, equivalent to the concentration giving the measured effect (e.g. mortality). However, in reality not all compounds can be known in complex mixtures and there is always the potential for synergistic effects to occur and, therefore, the sum of individual TUs is often less than 1. The TU approach has been applied to concentrated water extracts for a range of PAHs and these could only account for a maximum of 11% (TU¼0.11) (Hendriks et al., 1994) and 18% (TU¼0.18) (Kirby et al., 1998). The sediment toxicity data generated in this study could also be used to help explain the contribution to toxicity of certain low MW PAHs in contaminated sediments. Recent studies have shown how the toxic unit approach is applicable to PAH assessments in sediments from as varied locations as European high mountain lakes (Quiroz et al., 2010) and oil contaminated sediments in the Arabian Gulf (Bejarano and Michel, 2010). Another widely used approach to predict the toxicity (or other effect) of broad categories of similar compounds whose physical

structure and characteristics can be reliably quantified is the development of quantitative structure activity relationships (QSARs). This approach relates physical attributes of compounds (solubility, Log Koc, etc.) to known toxicity in order to enable a model to be developed to predict the toxicity of other similar compounds. This approach was applied by Swartz et al. (1995) to produce predicted 10 day LC50iw values in sediments for a number of low MW PAHs as follows (figures quoted as mg/L): naphthalene, 3500; acenaphthylene, 490; acenaphthene, 970; fluorene, 270; phenanthrene 240; anthracene, 180. Future work is required to develop the indicator and the next stage would be to validate with actual environmental samples. A recent study has reviewed TEQ values generated for river water and effluent samples from a sewage treatment plant in 2007 using Vibrio fischeri and it was found that all validation experiments with mixtures of reference compounds and complex environmental mixtures confirmed the applicability of the TEQ concept (Escher et al., 2008). There are many challenges to be faced when assessing the acute toxicity of PAH in sediment to marine amphipods and other species. There are an infinite number of possible mixtures, other contaminants, differing environmental parameters (including sediment types). Nevertheless, while it is right to acknowledge these limitations it is not always feasible to assess the toxic potential of large sample numbers or wide geographical areas (as is being proposed under the European Marine Strategy Framework Directive) using biological toxicity test approaches and the development of TEF factors and the subsequent TEQ indicators will be an important tool in predicting potential problem areas.

5. Conclusion This paper generated LC50 values for 12 LMW PAHs and subsequent TEF values for use as an indicator of sediment toxicity to the amphipod Corophium volutator.

Acknowledgments This work was funded by the Department for Environment, Food and Rural Affairs under R&D Contract ME4118. The author would like to thank Angela Ward for her contribution to this work. References Allen, Y., 2005. Environmental Impacts Resulting from Disposal of Dredged Material at the Rame Head Disposal Site, S.W. England: An Analysis of Existing Data and Implications for Environmental Management. Cefas Contract report BA004, June 2005. Altenburger, R., Greco, W.R., 2009. Extrapolation concepts for dealing with multiple contamination in environmental risk assessment. Integr. Environ. Assess. Manage. 5, 62–68. Anderson, J.W., 1982. The transport of petroleum hydrocarbons from sediments to benthos and the potential effects. In: Mayer, G. (Ed.), Ecological Stress and the New York Bight: Science and Management. Estuarine Research Federation, Columbia, pp. 165–179. Baden, S.P., 1990. The cryptofauna of Zostera marina (L.): abundance, biomass and population dynamics. Neth. J. Sea Res. 27, 81–92. Bejarano, A.C., Michel, J., 2010. Large-scale risk assessment of polycyclic aromatic hydrocarbons in shoreline sediments from Saudi Arabia: environmental legacy after twelve years of the Gulf war oil spill. Environ. Pollut. 158, 1561–1569. Birnbaum, L.S., DeVito, M.J., 1995. Use of toxic equivalency factors for risk assessment for dioxins and related compounds. Toxicology 105, 391–401. Chapman, P.M., 1992. Pollution status of North Sea sediments—an international integrative study. Mar. Ecol. Prog. Ser. 91, 313–322. Cassee, F.R., Groten, J.P., Bladeren, P.J., Feron, V.J., 1998. Toxicological evaluation and risk assessment of chemical mixtures. Crit. Rev. Toxicol. 28, 73–101. Dauvin, J.-C., 1998. The fine sand Abra alba community of the Bay of Morlaix twenty years after the Amoco Cadiz oil spill. Mar. Pollut. Bull. 36, 669–676.

T.T. Fisher et al. / Ecotoxicology and Environmental Safety 74 (2011) 2245–2251

DeWitt, T.H., Redmond, M.S., Sewall, J.E., Swartz, R.C., 1992. Development of a chronic sediment toxicity test for marine benthic amphipods. EPA CBP/TRS/89/93. US Environmental Protection Agency, Newport, Oregon, 252 pp. Donahue, W.H., Wang, R.T., Welch, M., Nicol, J.A.C., 1977. Effects of water-soluble components of petroleum oils and aromatic hydrocarbons on Barnacle larvae. Environ. Pollut. 13, 187–202. Djomo, J.E., Dauta, A., Ferrier, V., Narbonne, J.F., Monkiedje, A., Nijine, T., Garrigues, P., 2004. Toxic effects of some major polyaromatic hydrocarbons found in crude oil and aquatic sediments on Scenedesmus subspicatus. Water Res. 38, 1817–1821. Engraff, M., Solere, C., Smith, K., Mayer, P., Dahllof, I., 2011. Aquatic toxicity of PAHs and PAH mixtures at saturation to benthic amphipods: linking toxic effects to chemical activity. Aquat. Toxicol. 102, 142–149. Escher, B.I., Bramaz, N., Mueller, J.F., Quayle, P., Rutishauser, S., Vermeirssen, E.L.M., 2008. Toxic equivalent concentrations (TEQs) for baseline toxicity and specific modes of action as a tool to improve interpretation of ecotoxicity testing of environmental samples. J. Environ. Monit. 10, 612–621. Harvey, R.G., 1991. Polycyclic Aromatic Hydrocarbons: Chemistry and Carcinogenicity. Cambridge University Press, Cambridge. Hendriks, A.J., Maas-Diepeveen, J.L., Noordsij, A., Van der Gaag, M.A., 1994. Monitoring response of XAD-concentrated water in the Rhine delta: a major part of the toxic compounds remains unidentified. Water Res. 28, 581–598. Hill, I.R., Matthiesse, P., Heimbach, F., 1993. Guidance Document on Sediment Toxicity Tests and Bioassays for Freshwater and Marine Environments—Workshop on Sediment Toxicity Assessment, Renesse, The Netherlands, 8–10 November, SETAC, 105 pp. Jewett, S.C., Dean, T.A., Smith, R.O., Blanchard, A., 1999. ‘Exxon Valdez’ oil spill: impacts and recovery in the soft-bottom benthic community in and adjacent to eelgrass beds. Mar. Ecol. Prog. Ser. 185, 59–83. Kelly, C.A., Law, R.J., Emerson, H.S., 2000. Methods of analysing hydrocarbons and polycyclic aromatic hydrocarbons (PAH) in marine samples. Cefas Science Series, Aquatic Environment Protection: Analytical Methods No. 12. Centre for Environment, Fisheries and Aquaculture Science, Lowestoft, 18 pp. Kennish, M.J., 1998. Pollution Impacts on Marine Biotic Communities. Marine Science Series. CRC press, Boca Raton. Kirby, M.F., Blackburn, M.A., Thain, J.E., Waldock, M.J., 1998. Assessment of water quality in estuarine and coastal waters of England and Wales using a contaminant concentration technique. Mar. Pollut. Bull. 36, 631–642. Kirby, M.F., Lyons, B.P., Waldock, M.J., Woodhead, R.J., Goodsir, F., Law, R.J., Matthiessen, P., Neall, P., Stewart, C., Thain, J.T., Tylor, T., Feist, S.W., 2000. Biomarkers of Polycyclic Aromatic Hydrocarbons (PAH) Exposure in Fish and Their Application in Marine Monitoring. Cefas Science Series, Technical Report No. 110. Centre for Environment, Fisheries and Aquaculture Science, Lowestoft, 30 pp. Landrum, P.F., Lotufo, G.R., Gossiaux, D.C., Gedeon, M.L., Lee, J.-H., 2003. Bioaccumulation and critical body residue of PAHs in the amphipod, Diporeia spp.: additional evidence to support toxicity additivity for PAH mixtures. Chemosphere 51, 481–489. Law, R.J., Kelly, C., Baker, K., Jones, J., McIntosh, A.D., Moffat, C.F., 2002. Toxic equivalency factors for PAH and their applicability in shellfish pollution monitoring studies. J. Environ. Monit. 4, 383–388. Lotufo, G.R., Fleeger, J.W., 1996. Toxicity of sediment-associated pyrene and phenanthrene to Limnodrilus hoffmeisteri (Oligochaeta: Tubificidae). Environ. Toxicol. Chem. 15, 1508–1516.

2251

Lotufo, G.R., Fleeger, J.W., 1997. Effects of sediment associated phenanthrene on survival, development and reproduction of 2 species of meiobenthic copepods. Mar. Ecol. Prog. Ser. 151, 91–102. McCreedy, S., Slee, D.J., Birch, G.F., Taylor, S.E., 2000. The distribution of Polycyclic Aromatic Hydrocarbons in surficial sediments of Sydney Harbour, Australia. Mar. Pollut. Bull. 40, 999–1006. Meador, J.P., Stein, J.E., Reichert, W.L., Varanasi, U., 1995. Bioaccumulation of polycyclic aromatic hydrocarbons by marine organisms. Rev. Environ. Contam. Toxicol. 143, 79–165. Moore, J.W., Ramamoorthy, S., 1984. Organic Chemicals in Natural Waters: Applied Monitoring and Impact Assessment. Springer-Verlag, New York. O’Connor, T.P., Paul, J.F., 2000. Misfit between sediment toxicity and chemistry. Mar. Pollut. Bull. 40, 59–64. Peters, C., Ahlf, W., 2005. Reproduction of the estuarine and marine amphipod Corophium volutator (Pallas) in laboratory for toxicity testing. Chemosphere 59, 525–536. Quan, X., Tang, Q., He, M., Yang, Z., Lin, C., Guo, W., 2009. Biodegradation of polycyclic aromatic hydrocarbons in sediments from the Daliao River watershed, China. J. Environ. Sci. 21, 865–871. Quiroz, R., Grimalt, J.O., Fernandez, P., 2010. Toxicity assessment of polycyclic aromatic hydrocarbons in sediments from European high mountain lakes. Ecotoxicol. Environ. Saf. 73, 559–564. Roddie, B.D., Thain, J.E., 2002. Biological Effects of Contaminants: Corophium sp. Sediment Bioassay and Toxicity Test. Techniques in Marine Environmental Science, No. 28. ICES, Copenhagen, 22pp. Rogers, H.R., 2002. Assessment of PAH contamination in estuarine sediments using the equilibrium partitioning-toxic unit approach. Sci. Total Environ. 290, 139–155. Rutt, G.P., Levell, D., Hobbs, G., Rostron, D.M., Bullimore, B., Law, R.J., Robinson, A.W., 1998. The effect on the marine benthos. In: Edwards, R., Sime, H. (Eds.), The Sea Empress Oil Spill: Proceedings of the International Conference held in Cardiff, CIWEM, London, 11–13 February 1998, pp. 189–206. Steevens, J.A., Slattery, M., Schlenk, D., Aryl, A., Benson, W.H., 1999. Effects of ultraviolet-B light and polyaromatic hydrocarbon exposure on sea urchin development and bacterial bioluminescence. Mar. Environ. Res. 48, 439–457. Swartz, R.C., Schults, D.W., Ozretich, R.J., P Lamberson, J.O., Cole, F.A., DeWitt, T.H., Redmond, M.S., Ferraro, S., 1995. PAH: a model to predict the toxicity of polynuclear aromatic hydrocarbon mixtures in field-collected sediments. Environ. Toxicol. Chem. 14, 1977–1987. Swartz, R.C., Ferraro, S.P., Lamberson, J.O., Cole, F.A., Ozretich, R.J., Boese, B.L., Schults, D.W., Behrenfeld, M., Ankley, G.T., 1997. Photoactivation and toxicity of mixtures of polycyclic aromatic hydrocarbon compounds in marine sediment. Environ. Toxicol. Chem. 16, 2151–2157. Witt, G., 1995. Polycyclic aromatic hydrocarbons in water and sediment of the Baltic Sea. Mar. Pollut. Bull. 31, 237–248. Woodhead, R.J., Law, R.J., Matthiessen, P., 1999. Polycyclic aromatic hydrocarbons in surface sediments around England and Wales, and their possible biological significance. Mar. Pollut. Bull. 38, 773–790. Zepp, G.G., Schlotzhauer, P.F., 1979. Photoreactivity of selected aromatic hydrocarbons in water. In: Jones, P.W., Leber, P. (Eds.), Polynuclear Aromatic Hydrocarbons. Ann Arbor Science. Publishers, Ann Arbor, pp. 141–158. Zepp, R.G., Baughman, G.L., 1978. Prediction of photochemical transformation of pollutants in the aquatic environment. In: Hutzinger, O. (Ed.), Aquatic Pollutants. Pergamon, Oxford, pp. 237–263. Zepp, R.G., Cline, D.M., 1977. Rates of direct photolysis in aquatic environment. Environ. Sci. Technol. 11, 359–366.