Alexander P Scott
email: a.p.scott@cefas.co.uk

Matthew Sanders

Grant D Stentiford

R Allan Reese

Ioanna Katsiadaki

Evidence for estrogenic endocrine disruption in an offshore flatfish, the dab (Limanda limanda L.)

Keywords: dab, Limanda, cod, endocrine disruption, vitellogenin, VTG, oestrogen, estrogen, marine

Dab (Limanda limanda) caught in UK offshore waters show evidence of being exposed to estrogenic endocrine disrupters at a relatively low level. Two of 449 males caught between June and July 2005 had markedly elevated levels of vitellogenin (VTG; 21 and 750 µg/ml) and the remainder ranged from <0.01 to 8.6 µg/ml. Omitting the two outliers, there was a very significant positive relationship with the mass of individual males (a feature noted in previous studies on cod). Mean VTG concentrations in males differed significantly between sites. The site with the highest mean (1.6 µg/ml) was North East of the Dogger Bank and the site with the lowest (0.04 µg/ml) was in Cardigan Bay. Mean VTG concentrations in all North Sea fish were significantly higher than English Channel and Irish Sea fish, but this difference disappeared when fish mass was taken into account. VTG concentrations showed no relationship to water depth, stage of sexual maturity or age of the males.

Sixty selected male plasmas were assayed for 17β-estradiol but only two had measurable amounts (assay limit 0.04 ng/ml). Despite being the start of summer, the gonads of many of the males and females (especially those caught in the North Sea) showed signs of sexual maturity (presence of sperm in males and vitellogenic eggs in females). Many females had high VTG concentrations (up to 14 mg/ml) and 78 out of 80 had measurable concentrations of 17β-estradiol. The cause of elevated VTG levels in male dab is unknown. As seen in cod, the presence of affected males does not appear to be linked to proximity to land or to known point sources of endocrine disrupters. However, our data, showing that larger fish are more likely to have elevated VTG concentrations, suggests a gradual accumulation by marine fish, ACCEPTED MANUSCRIPT 2 probably through feeding, of persistent (probably relatively weak) estrogenic compounds.

Introduction

Contamination of UK freshwaters by estrogenic compounds was first identified through research carried out at the Ministry of Agriculture, Fisheries and Food (now Cefas) and Brunel University in the 1980s. This demonstrated elevated concentrations of the female egg-yolk protein precursor vitellogenin (VTG) in the blood plasma of caged male trout (Oncorhynchus mykiss) that were placed in UK rivers -especially near sewage treatment works [START_REF] Purdom | Estrogenic effects of effluents from sewage treatment works[END_REF]. The finding was extended by Cefas to male flounders (Platichthys flesus) in UK estuaries in the 1990s (Allen, Matthiessen, Scott, Haworth, Feist, & Thain, 1999a;[START_REF] Allen | Survey of estrogenic activity in United Kingdom estuarine and coastal waters and its effects on gonadal development of the flounder Platichthys flesus[END_REF] and to cod (Gadus morhua) in the open sea within the last five years (Scott, Katsiadaki, Witthames, Hylland, Davies, McIntosh et al., 2006b).

Significantly elevated concentrations of VTG were found in plasma of male cod from three areas of the North East Atlantic. Furthermore, the larger the fish, the higher the concentration of VTG that was found. The causative agents are not yet known. However, it is highly unlikely that the main culprits from sewage treatment works (viz. the human female hormones, 17β-estradiol (E 2 ) and estrone, and the synthetic hormone ethinylestradiol -the main ingredient of 'the contraceptive 'pill') would be able to reach fish so far offshore. Large adult cod are top predators and it is more likely to be due to the gradual accumulation of one (or even a mixture) of the many stable man-made pollutants (some of which are known to be weakly estrogenic) that are widespread in the marine food chain (e.g. DDT, PCB). This hypothesis is supported by abnormal VTG production being found in males of two other top predators, tuna (Thunnus thynnus) and swordfish (Xiphias gladius), that inhabit pelagic environments even further from shore [START_REF] Fossi | Do endocrine disrupting chemicals threaten Mediterranean swordfish? Preliminary results of vitellogenin and zona radiata proteins in Xiphias gladius[END_REF][START_REF] Fossi | Biomarkers for endocrine disruption in three species of Mediterranean large pelagic fish[END_REF][START_REF] De Metrio | Evidence of a high percentage of intersex in the Mediterranean swordfish (Xiphias gladius L.)[END_REF][START_REF] Fossi | Evaluation of ecotoxicological effects of endocrine disrupters during a fouryear survey of the Mediterranean population of swordfish (Xiphias gladius)[END_REF][START_REF] Desantis | Immunohistochemical localization of CYP1A, vitellogenin and zona radiata proteins in the liver of swordfish (Xiphias gladius L.) taken from the Mediterranean Sea, South Atlantic, South Western Indian and Central North Pacific Oceans[END_REF]. Furthermore, in tuna, it also appears that the larger fish have higher VTG production [START_REF] Barucca | Analysis of vitellogenin gene induction as a valuable biomarker of estrogenic exposure in various Mediterannean fish species[END_REF].

If, as we hypothesise, the food chain is contaminated with estrogenic endocrine disrupters, we would expect to find males of other marine fish in the waters around the UK with elevated concentrations of VTG in their blood plasma. Indeed, some early work in our laboratory showed that males of a flatfish, the plaice (Pleuronectes platessa), caught in the North Sea, all had elevated concentrations of VTG [START_REF] Scott | 17β-oestradiol in male flatfish[END_REF]. However, these fish were caught in the middle of the spawning season and mostly had readily-detectable E 2 in their plasma, so we were unable to conclude whether it was a natural or xenobiotic-induced effect.

As part of an entirely separate study of fish disease in UK waters, scientists at CEFAS Weymouth carry out annual sampling of another flatfish, the dab (Limanda limanda), a species that is widely dispersed in the North East Atlantic and, incidentally, a favourite prey item of large cod in the North Sea [START_REF] Daan | A quantitative analysis of the food intake of North Sea cod, Gadus morhua[END_REF]. The samples come from twenty-six stations in the Irish Sea, North Sea and English Channel. The dab is the main offshore sentinel species of the Clean Seas Environmental Monitoring Programme (CSEMP; formerly the National Marine Monitoring Programme), so each sample is accompanied by physical measurements of the site and biometric measurements of the fish.

For the present study, frozen plasma samples were collected from over 1000 male and female dabs during June and July 2005. A further 40 samples of male dab plasma came from a Research Vessel cruise in January, 1997. An antiserum to the egg yolk protein lipovitellin (LV) -which is formed from VTG when it is laid down in the oocytes -plus freshly purified and then freeze-dried dab LV, were used to set up an ELISA. This ELISA was tested for parallelism against purified dab VTG and female plasma samples and then used to quantify VTG in all the male plasma samples collected in 2005 (n=449) and ca. 270 female plasma samples. E 2 assays were carried out on a sub-set of the plasma samples.

Materials and methods

Purification of dab lipovitellin

Ovaries were removed from reproductively-mature female dabs and stored frozen at -20 o C. Prior to extraction, 8 g was placed in liquid nitrogen and ground to a powder in a pestle and mortar. The powder was mixed with 40 ml of ice cold 0.5 M NaCl. After gentle shaking for 30 min, the mixture was centrifuged for 30 min at 1000 G and the supernatant added to 2 l of ice-cold deionised water. The precipitate (crude LV) was allowed to settle overnight at 4 o C and concentrated by centrifuging.

Approximately 1 g of the resultant slurry was mixed first with 1 ml 2 M NaCl and then with 50 ml of Tris-HCL buffer (0.05 M; pH 7.8). Undissolved material was removed first by centrifuging and then by double filtration (0.45 and 0.25 µm pore sizes). The supernatant was loaded on a HiPrep 16/10 QXL ion exchange column (column volume 10 ml; Amersham Biosciences) at a flow rate of 2 ml min -1 . Two buffers were made up: A, 0.05 M Tris-HCl pH 7.6; B, 0.05M Tris-HCl pH 7.6 containing 1 M NaCl. The column was equilibrated with 95% A:5% B. The column was developed with a linear gradient from 5% B to 100% B over four column volumes. The effluent was monitored for UV absorption (280 nm) and fractions were collected. There were two major UV peaks (Fig. 1). The fractions corresponding to these peaks were desalted on PD-10 columns with deionised water and freeze-dried.

The sharp second peak from the ion-exchange column (B) yielded very little powder (< 5 mg). The other peak (A), however, yielded about 80 mg powder. Both powders were redissolved in carbonate buffer as described for the flounder VTG ELISA [START_REF] Allen | Survey of estrogenic activity in United Kingdom estuarine and coastal waters and its effects on gonadal development of the flounder Platichthys flesus[END_REF] and used to coat 96-well polystyrene plates at eight different concentrations (all wells in rows A to H). After washing, the plates were incubated overnight with antiserum to dab LV at 12 different concentrations (all wells in columns 1 to 12). Again, following already-published procedures [START_REF] Katsiadaki | Detection of environmental androgens: a novel method based on enzyme-linked immunosorbent assay of spiggin, the stickleback (Gasterosteus aculeatus) glue protein[END_REF], the plates were incubated with 'second antibody' labelled with alkaline phosphatase and developed. Peak B showed no cross-reaction with the antiserum and was discarded. Peak A, however, cross-reacted strongly with dab LV antiserum. The antiserum was produced in the MAFF, Lowestoft Fisheries Laboratory in 1977 by injection of a rabbit with dab LV purified using previously-published procedures [START_REF] Plack | Egg proteins in cod serum. Natural occurrence and induction by injections of oestradiol 3-benzoate[END_REF].

Validation of the enzyme-linked immunosorbent assay (ELISA) for dab vitellogenin

Disposables, equipment and basic procedures for ELISA were the same as those used for the assay of stickleback spiggin (Gasterosteus aculeatus) [START_REF] Katsiadaki | Detection of environmental androgens: a novel method based on enzyme-linked immunosorbent assay of spiggin, the stickleback (Gasterosteus aculeatus) glue protein[END_REF]. For coating plates, 1 mg heat-treated LV was dissolved in 500 µl distilled water followed by 500 µl of carbonate buffer. A 400 µl aliquot was added to 100 ml of coating buffer and 100 µl of this dilute solution was then added to all wells. The antiserum was used at a final dilution of 1:42,000 (v/v). Tests were carried out for repeatability of observations between plates (within an assay); repeatability between assays; parallelism between standard LV, standard VTG and diluted plasma samples; and comparability between plasma samples assayed first with LV as the plate-coating material and then VTG as the plate-coating material.

Collection of dab

Starting on 17 June 2005, male and female dab were captured from twenty-six sites in the Irish Sea, English Channel and North Sea using 30 min tows of a standard Granton Trawl. Sex, size (total length and weight) and external signs of disease were recorded. Blood samples were taken from the caudal vein using heparinised syringes, centrifuged for 10 min and the plasma snap frozen in liquid nitrogen and stored -20 o C. The fish were sacrificed by a blow to the head, followed immediately by severing of the spinal cord. The otolith was removed for age determination. Gonads (females only) and livers were weighed in order to calculate the Gonadosomatic Index (GSI) and Hepatosomatic Index (HSI), respectively (= 100 x [organ weight/bodyweight]).

Samples of liver, kidney, spleen and gonad were fixed in neutral-buffered formalin for 24 h followed by transfer to industrial methylated spirit. Fixed samples were processed to wax in a vacuum-infiltration processor using standard protocols. Sections were cut at 3 to 5 µm on a rotary microtome and mounted onto glass slides before staining with haematoxylin and eosin. Stained sections of the gonads were examined by light microscopy to a) check the original visual analysis of sex and b) determine the stage of reproductive maturity of the fish. Testes that showed no signs of sexual maturation were given a score of 1; and those with sperm or spermatocytes present were given a score of 2. Ovaries with only primary oocytes or oogonia present (immature) were scored as 1; those with cortical alveoli stage oocytes as 2; those with oocytes undergoing secondary vitellogenesis (VTG incorporation) as 3; and those with atretic yolky oocytes as 4. Mean scores for each sex at each site gave maturity indices. About 10 fish showed discrepancies between sex as recorded initially and as determined from histology. These fish were omitted from statistical analyses.

Statistical analysis

Data were analysed using Stata version 9.2 (2006 StataCorp, College Station, TX, USA). After scanning using a variety of exploratory data analysis techniques, log(VTG) in males was modelled as a response variable with discrete and continuous predictors with and without the two highest values. Logs were used to make the empirical distribution symmetrical. VTG values below the level of quantification were allocated a censored value of 0.01 µg/ml, prompting the use of Tobit regression to obtain unbiased estimators of left-censored data. The model was simplified by excluding or grouping factor values whose effects were not significant (based on ttests of ML estimators). An emphasis was put on producing a biologically meaningful interpretation rather than an arbitrary mathematically "best" model.

Results

Validation of the ELISA

Plasma samples from four female dabs gave dilution curves that were parallel to both LV and VTG over the steepest, most accurate, part of the curve (Fig. 2). In all assays, a pool of female plasma was included in every plate. Three assays were carried out.

The first assay yielded a within-assay (essentially between-plate) estimate (mean ± SE) for the plasma pool of 36.1 ± 2.4 µg/ml (n= 11). The second assay gave an estimate of 38.9 ± 0.9 (n= 3) and the third 35.72 ± 4.2 (n= 8). A pool of male plasma in the first assay gave an estimate of 4.8 ± 0.2, n= 11. All male plasma samples were assayed on the same date (with a single batch of standard and reagents) in order to minimise inter-assay variation.

VTG in males

Weight was strongly related to length for males; over the limited size range, linear and cubic functions gave identical goodness of fit (r 2 = 0.86). Slopes were similar, but intercepts were significantly different in each sea area. Males in the North Sea were on average larger than those in the other two sea areas. The mean + SE values were: North Sea, length 21.3 ± 0.1, weight 96.8 ± 1.8, n = 182; Irish Sea, length 19.5 ± 0.2, weight 76.8 ± 2.0, n = 182; English Channel, length 20.0 ± 0.2, weight 90.0 ± 2.4, n = 85.

Mean male VTG concentrations for each site are shown in map form in Fig. 3.

Locations of the sites are listed in Table 1 and summary statistics in Table 2. Two males with VTG concentrations higher than 10 µg/ml were treated as outliers. These data are shown separately (Table 3). Of the 1120 fish collected during 2005, 449 were male. The distribution of VTG concentrations in these males was: <0.01 µg/ml, n = 50; 0.02 to 0.1 µg/ml, n = 132; 0.1 to 1 µg/ml, n = 222; 1 to 10 µg/ml, n = 43; >10 µg/ml, n = 2.

Mean VTG concentrations in male dab in the North Sea (0.53 ± 0.08, n = 181) were marginally, but nevertheless significantly, higher than those in the Irish Sea (0.34 ± 0.07, n = 181) and English Channel (0.25 ± 0.05, n = 85). With fish weight as a covariate, the differences disappeared. In essence, the difference between the seas was driven by log(VTG) being positively related to fish weight: all fish (Fig. 4), r = 0.290, n = 447, p<0.0001; North Sea, r = 0.204, n = 182, p<0.005; Irish Sea, r = 0.358, n = 182, p<0.0001); English Channel, not significant.

Mean VTG concentrations were clearly very different between the twenty-six sites (Fig. 3). Statistically, with or without fish weight as a variate, there were at least eight separate overlapping groupings of sites by post-hoc statistical analysis.

However, sites with very low mean VTG concentrations and those with mean VTG concentrations that were up to 30 times higher were scattered across the sea areas.

Nevertheless, the map shows clusters of sites within which the values were very similar to each other (e.g. Cardigan Bay, the eastern English Channel and Liverpool Bay).

Based on summary statistics (Table 2), there was no relationship between mean VTG concentrations and water depth in males.

The mean concentration of VTG in fish with completely immature testes (0.47 ± 0.14, n = 164; stage 1) did not differ significantly from that in fish with sperm in their testes (0.43 ± 0.04, n = 283; stage 2). Mean male VTG concentrations at the different sites also showed no relationship to the stage of maturity of the females or to the date of capture. The forty male dab collected from the North Sea at the end of January 1997 had a mean length of 19 cm, a mean weight of 74 g and a mean VTG concentration of 0.62 ± 0.07 µg/ml (not dissimilar to 0.53 ± 0.08 for North Sea males in July, 2005).

There was a significant positive relationship between male log(VTG) and HSI (r = 0.276, n = 357, p<0.0001). Mean HSI values at some sites were twice as high as those at other sites (Table 2), but this was not related to mean VTG levels. There was a very significant positive correlation between male and female mean HSIs (results of analysis not shown). This suggests that whatever causes between-site differences in HSI, it affects both sexes equally.

Fifteen fish in each of the following categories were chosen for E 2 assay: immature testes and mean VTG = 0.08 µg/ml; immature testes and mean VTG = 3.3 µg/ml; mature testes and mean VTG = 0.05 µg/ml; mature testes and VTG = 2.7 µg/ml. E 2 could not be detected (<0.04 ng/ml) in any but two males. One of these males, with a VTG concentration of 5 µg/ml had an E 2 concentration of 1 ng/ml. The other male, with <0.01 µg/ml VTG, had an E 2 concentrations of 0.13 ng/ml. Age data (from otoliths) were available only for males caught on the Dogger Bank (sites 18 to 21). Some of the males were as much as 12 years old. However, there was no relationship between log(VTG) concentrations and age (Fig. 5). The positive relationship of log(VTG) to fish size, on the other hand, was particularly high (r = 0.468, n = 88, p<0.0001) for this group of fish (Fig. 6).

VTG in Females

The first ten females caught at each site were analysed for VTG. There were highly significant differences in mean VTG concentrations in females between the North Sea (2091.2 ± 219.8, n =104), the Irish Sea (409.4 ± 56.1, n = 107) and the English Channel (190.8 ± 43.2,n = 58). The site differences are very obvious in the summary statistics (Table 4).

As with males, there was a significant positive relationship between log(VTG) concentrations and fish weight (r = 0.241, n = 269, p<0.0001).

Females with ovaries that contained atretic eggs (stage 4; 2452.5 ± 433.3, n = 38) or oocytes in the early stages of vitellogenesis (stage 3; 2578.6 ± 415.6, n = 30) had significantly higher VTG concentrations than those with ovaries that contained Considering that females in the North Sea had higher mean VTG concentrations than those in the other two sea areas and that sampling was spread over 21 days starting in the Irish Sea and ending in the North Sea, it was not surprising that a significant relationship could also be found between log(VTG) concentrations and date of capture (r = 0.291, n = 269, p<0.0001). Unlike the males, there was no significant correlation between log(VTG) concentrations and HSI in females.

Twenty plasma samples were selected for E 2 assay from each of the four stages of maturity of females. Within each stage, ten samples were taken from the bottom of the distribution of VTG concentrations and ten from the top. There were no significant differences in E 2 concentrations between the different groups, apart from those females in which oocytes were undergoing vitellogenesis and had high VTG concentrations (Fig. 7). This is consistent with what is known about the major role of E 2 in females (i.e. stimulating vitellogenesis in the liver). Only two (out of 80) females had E 2 concentrations below the limit of the assay. One fish at the primary oocyte stage and with only 5 µg/ml VTG had an apparently very large E 2 concentration of > 20 ng/ml. This was omitted from the statistical comparison between VTG and E 2 concentrations.

Discussion

This study set out to determine whether dab caught in offshore waters of the UK show any evidence of exposure to estrogenic endocrine disrupters. The answer is in the affirmative. Of the 449 males caught in June/July 2005, one had a markedly elevated (750 µg/ml) and another a moderately elevated (21 µg/ml) concentration of VTG and the remainder ranged from <0.02 to 8.6 µg/ml. Without the two outliers (and even including the lower one), there was a very significant positive relationship with the size of the males. This relationship has already been noted in previous studies on cod and tuna.

Mean VTG concentrations differed between sites, but there was no clear pattern. Sites that might have been expected to have a large proportion of affected males -for example, close to the Tees estuary, the sediments of which are known to be contaminated with estrogenic compounds [START_REF] Thomas | Potency and characterization of estrogen-receptor agonists in United Kingdom estuarine sediments[END_REF]) -showed no evidence of contamination at all.

As in male cod (Scott et al., 2006b) and male flounders (Scott, Katsiadaki, Kirby, & Thain, 2006a), there was no evidence that elevated VTG concentrations in male dabs were related to their stage of reproductive maturity or to plasma E 2 concentrations.

Comparison with previous studies on estrogenic EDs on marine fish

In addition to the findings on cod and dabs, there is now a large body of evidence for the existence of estrogenic endocrine disruption in the marine environment. Male flounders (Platichthys flesus) caught in industrialised estuaries of the UK and the Netherlands have been found with elevated concentrations of VTG in their plasma [START_REF] Lye | Abnormalities in the reproductive health of flounder Platichthys flesus exposed to effluent from a sewage treatment works[END_REF][START_REF] Lye | Seasonal reproductive health of flounder Platicthys flesus exposed to sewage effluent[END_REF]Allen et al., 1999a;[START_REF] Allen | Survey of estrogenic activity in United Kingdom estuarine and coastal waters and its effects on gonadal development of the flounder Platichthys flesus[END_REF][START_REF] Matthiessen | The impact of oestrogenic and androgenic contamination on marine organisms in the United Kingdom -summary of the EDMAR programme[END_REF][START_REF] Vethaak | Estrogenic effects in fish in The Netherlands: some preliminary results[END_REF][START_REF] Kirby | Surveys of plasma vitellogenin and intersex in male flounder (Platichthys flesus) as measures of endocrine disruption by estrogenic contamination in United Kingdom estuaries: temporal trends, 1996 to 2001[END_REF][START_REF] Kleinkauf | Abnormally elevated VTG concentrations in flounder (Platichthys flesus) from the Mersey Estuary (UK) -a continuing problem[END_REF][START_REF] Kirby | Differential sensitivity of flounder (Platichthys flesus) in response to oestrogenic chemical exposure: an issue for design and interpretation of monitoring and research programmes[END_REF]. The range of VTG concentrations was great (from <0.01 µg/ml to >50 x caught in the open sea (Allen et al., 1999a), but these were hypothesised to be fish that had recently emigrated from a contaminated estuary. In estuarine and coastal areas of the USA [START_REF] Mills | Relationship between reproductive success and male plasma vitellogenin concentrations in cunner, Tautogolabrus adspersus[END_REF][START_REF] Roy | The relationship of biochemical endpoints to histopathology and population metrics in feral flatfish species collected near the municipal wastewater outfall of Orange County, California, USA[END_REF] and Japan [START_REF] Hashimoto | Elevated serum vitellogenin levels and gonadal abnormalities in wild male flounder (Pleuronectes yokohamae) from Tokyo Bay, Japan[END_REF][START_REF] Hara | Endocrine and sexual disruption in wild grey mullet[END_REF][START_REF] Ohkubo | Estrogenic activity in coastal areas around Japan evaluated by measuring male serum vitellogenins in Japanese common goby Acanthogobius flavimanus[END_REF], many male fish have also been found with high concentrations of VTG in their plasma.

In all the above studies, fish were caught either just offshore or in estuaries with a known high level of contamination. In these situations, it is difficult to dismiss the effects of sewage treatment works, fish farming and nearby industries. Evidence of estrogenic effects 'in the open sea' that can be directly compared to that in cod (and dab in the present study), comes from studies on swordfish and tuna. VTG production has been demonstrated in male swordfish from the Mediterranean and from the waters around South Africa [START_REF] Desantis | Immunohistochemical localization of CYP1A, vitellogenin and zona radiata proteins in the liver of swordfish (Xiphias gladius L.) taken from the Mediterranean Sea, South Atlantic, South Western Indian and Central North Pacific Oceans[END_REF]. Similar to the findings in dab and cod, there also appears to be a positive relationship between VTG levels and fish size in male tuna [START_REF] Barucca | Analysis of vitellogenin gene induction as a valuable biomarker of estrogenic exposure in various Mediterannean fish species[END_REF].

What is the most likely route of exposure?

If the causative agents were absorbed directly from the water, then one would expect, firstly, that all fish would be affected equally regardless of size and, secondly, that there would not be marked differences in mean VTG concentrations between sites within a sea area. Both arguments are admittedly weak as fish may get more sensitive to estrogens as they grow and currents and tidal flows also ensure that sea water is not homogeneous. More convincing is the fact that, in cod, dab and tuna, VTG concentrations are positively related to the size of the fish. The most important determinant of size is the amount of food that a fish has eaten, so it seems plausible that the main route of entry is via the food. This is consistent with VTG concentrations in dab being typically 10 to 20 times lower than those we found in cod.

Large cod (the ones in which one is most likely to find elevated VTG concentrations) are top predators (i.e. they prey on other fish, including dabs) and are thus one level up the food chain.. It has been well documented that biomagnification of lipophilic pollutants in aquatic food chains is positively correlated to both trophic level [START_REF] Burreau | Biomagnification of PBDEs and PCBs in food webs from the Baltic Sea and the northern Atlantic Ocean[END_REF] and fish size [START_REF] Olsson | Concentrations of organochlorine substances in relation to fish size and trophic position: a study on perch (Perca fluviatilis L.)[END_REF].

There are few endocrine disrupter studies in the freshwater literature to support our hypothesis. This is because, although cumulative and persistent weak estrogenic chemicals are almost certainly present in the freshwater environment, their effect would be very difficult to demonstrate in an environment dominated by shortlived but nevertheless far more potent compounds emanating from sewage treatment works and factories. However, one study (Jobling, William, Johnson, Taylor, Gross-Sorokin, Nolan et al., 2006) shows a strong positive relationship between the frequency of intersex and the age of roach (Rutilus rutilus) living in estrogencontaminated UK rivers (i.e. there is evidence of a cumulative estrogenic effect).

Another study [START_REF] Vine | Endocrine (sexual) disruption is not a prominent feature in the pike (Esox lucius), a top predator, living in English waters[END_REF], that looked for effects in a top predator, the pike (Esox lucius) preying on estrogen-contaminated roach, found that, although the bile of male pike caught downstream had significantly more estrogenic activity than that of male pike caught upstream, there was no significant difference in VTG concentrations. Only two out of fifty-two male pike had detectable VTG in their plasma. Unfortunately, fish size was not taken into account in that study.

What is the likely identity of the causative compounds?

The fact that large fish are more affected than small fish is not consistent with populations being exposed to the main 'culprits' -E 2 , estrone and 17α-ethinylestradiol [START_REF] Sumpter | Lessons from endocrine disruption and their application to other issues concerning trace organics in the aquatic environment[END_REF]. Furthermore, these compounds are relatively unstable [START_REF] Johnson | Removal of endocrine-disrupting chemicals in activated sludge treatment works[END_REF]) and likely to be found in the coastal marine environment only in the neighbourhood of sewage treatment works and run-off water [START_REF] Atkinson | Estrogens from sewage in coastal marine enviornments[END_REF]. The causative agents are more likely to be persistent, probably relatively weak, estrogenic compounds. Many likely candidates have been identified by laboratory testing [START_REF] Sumpter | Lessons from endocrine disruption and their application to other issues concerning trace organics in the aquatic environment[END_REF], so it would be futile to approach this problem by trying to second-guess what the compounds might be. A better approach is to attempt to extract them from the tissues, bile or fat of affected fish, let them signal their presence with an in vitro estrogen screen, and then purify and identify them. This approach has been used successfully with freshwater fish [START_REF] Legler | Biological measurement of estrogenic activity in urine and bile conjugates with the in vitro ER-Calux reporter gene assay[END_REF][START_REF] Gibson | Analytical methodology for the identification of estrogenic contaminants in fish bile[END_REF].

The problem of the 'outliers'.

The two male dabs with VTG concentrations above 10 µg/ml were reassayed to confirm that no mistake had been made when the plasma samples were assayed.

The gonad sections were rechecked to make sure that the sex had been correctly ascribed. The plasma with the higher value is particularly extreme and difficult to explain. It is not without the bounds of possibility that, for this particular sample, a mistake was made in tube labelling or pipetting on board ship (since 1100 samples were collected). In favour of this being a case of mistaken identity is that liver sections of this fish could not be convincingly immunostained with the dab LV antibody (unpublished data). However, immunohistology is highly insensitive compared to ELISA.

We also examined our data to determine whether there was any possibility that any of the male samples might have been accidentally contaminated by an immediately preceding female sample (e.g. through accidental re-use of a syringe). However, we found no evidence of this (data not shown).

What can be considered a 'basal' VTG concentration?

Despite the lack of difference in VTG concentrations between the Irish Sea and the North Sea (after correcting for fish size) there were significant differences between individual sites and between clusters of sites. Such differences suggest that the fish either experience a different degree of exposure to environmental estrogens depending on where they live or that they belong to populations with different basal concentrations of VTG. In relation to the latter possibility, mean VTG concentrations from all sites plotted in rank order form a smooth distribution with no clear breaks, suggesting a single underlying continuous variable and arguing against there being 'separate populations'. Also, the five sites with the lowest mean VTG concentrations (< 0.1 µg/ml) were scattered between the three sea areas (three in the Irish Sea, one in the English Channel and one in the North Sea), suggesting that the 'basal concentration' is the same (and at some value less than 0.1 µg/ml) wherever the fish come from.

It is impossible to define the basal concentration exactly (i.e. the concentration below which a male could be assumed not to have been exposed to environmental estrogens) because everywhere on earth is now affected to some extent by pollution.

Hence, it is impossible to determine what is 'normal' [START_REF] Sumpter | Lessons from endocrine disruption and their application to other issues concerning trace organics in the aquatic environment[END_REF].

We can only draw the conclusion that some dabs have been exposed to exogenous estrogens (albeit with very low levels of induction compared to cod and flounder) on the basis of relative differences between the sites and the fact that VTG concentrations are linked to fish size.

Other possible explanations for differences in VTG concentrations between sites include: the fish are all exposed to estrogens but fish at some sites respond more sensitively; the fish are all exposed to estrogens but that fish at some sites are also exposed to anti-estrogens; or that fish at some sites have more natural phytoestrogens in their diet than at others. The first two possibilities have recently been examined experimentally in the flounder and found to have little or no effect on VTG concentrations [START_REF] Kirby | Differential sensitivity of flounder (Platichthys flesus) in response to oestrogenic chemical exposure: an issue for design and interpretation of monitoring and research programmes[END_REF]Kirby, Smith, Neall, Rooke, Scott, & Katsiadaki, in press). The third possibility cannot be easily dismissed. Our only argument against it is that, if phytoestrogens were the causative agents, it is unlikely that they would be cumulative or persistent and also unlikely that they would only be available to the larger fish.

Relationship of VTG concentrations to hepatosomatic index in males

The finding that HSI was positively correlated to VTG concentrations has been noted before in male flounders from the estuary of the River Mersey (Kleinkauf, Connor, Swarbreck, Levene, Walker, Johnson, & Leah, 2004a). One would fully expect to find this relationship in females, as the liver is the organ where the VTG is made and it would be expected to increase in weight as it expands protein production and takes on board the necessary precursors. However, no such correlation was found in the female dabs in the present study and since plasma VTG concentrations in male dabs (with the exception of the high outlier) were well below those in many of the females -and in fact all fell within the bottom 0.1 percentile of their full possible range (i.e. the rate of production of VTG in even the most stimulated fish was likely to have been miniscule) -this seems an unlikely explanation. Further examination of our data revealed a correlation between HSI and fish weight (r = 0.22, n = 357, p<0.0001 for the males and r = 0.37, n= 264 , p<0.0001 for females), suggesting that the relationship between VTG and HSI in males is probably just an 'association' (i.e. there is no direct causal link), although this does not mean that the increase in HSI with increase in fish size might not also be the result of the accumulation of persistent pollutants through the diet. However, it should be noted that, in the above-quoted study, HSI and body length were positively correlated in male flounders not only from the estuary of the River Mersey (heavy industrial contamination) but also from the River Dee (very low industrial contamination).

We have no good explanation for why, at three of the sites, mean HSI values were markedly higher than those found at the other sites -other than it might have reflected food availability. This was not assessed, however.

Evidence against the hypothesis that the presence of VTG in males is a natural ageing phenomenon.

The question most often asked at conferences where the results of the cod work (Scott et al., 2006b) have been presented is 'Could the appearance of VTG in the blood of large male fish be just a natural ageing phenomenon?'. One of the arguments against this from the cod study was that the fact that many of the affected male cod that were caught in the North Sea were actually quite young (only five years at the most) and were younger than completely unaffected male cod (six years old) caught off the Norwegian coast. The crucial difference was that the North Sea cod were much larger than the Norwegian cod. In the present study as well, the evidence points entirely to size and not age as the cause of raised VTG concentrations in dabs.

The significant site differences in mean VTG concentrations (as discussed above) also suggest that the appearance of VTG in male dabs is not something that happens as part of an endogenous process. There were highly significant site differences in male swordfish, with no affected fish being found in the Pacific Ocean [START_REF] Desantis | Immunohistochemical localization of CYP1A, vitellogenin and zona radiata proteins in the liver of swordfish (Xiphias gladius L.) taken from the Mediterranean Sea, South Atlantic, South Western Indian and Central North Pacific Oceans[END_REF]. Similarly, male tuna caught in the Mediterranean had VTG in their plasma [START_REF] Fossi | Biomarkers for endocrine disruption in three species of Mediterranean large pelagic fish[END_REF] while tuna (Thunnus obesus) caught in the Pacific Ocean [START_REF] Hashimoto | Gonadal histology and serum vitellogenin levels of bigeye tuna Thunnus obesus from the northern Pacific ocean -absence of endocrine disruption bio-indicators[END_REF] did not. This dependence on where the fish are living reinforces the argument against the hypothesis that the appearance of VTG is something that happens as a matter of course to male fish as they grow older. Yet one further fact that can be used against this hypothesis is that there was a strong correlation between VTG induction and CYP1A induction (an enzyme involved in the deactivation of planar xenobiotics) in male swordfish, such that CYP1A was only present in affected males from the Atlantic Ocean and Mediterranean and absent from the unaffected males in the Pacific Ocean [START_REF] Desantis | Immunohistochemical localization of CYP1A, vitellogenin and zona radiata proteins in the liver of swordfish (Xiphias gladius L.) taken from the Mediterranean Sea, South Atlantic, South Western Indian and Central North Pacific Oceans[END_REF]. This finding is the strongest evidence yet that elevated VTG concentrations in male fish can be directly attributed to pollution.

Although fish could hypothetically make VTG without any stimulation, such a mechanism has not yet been demonstrated and in all studies to date, some form of hormonal induction, operating through a receptor, is required. The main natural endogenous signal in fish is E 2 . Although E 2 could be measured in male cod, its levels were no different from those in immature females and were not correlated with VTG concentrations (Scott et al., 2006b). In the dab, E 2 was undetectable in most of the males that were examined.

VTG in females

Examination of ovaries, plus the results of the VTG assays, indicated that a high proportion of females (especially in the North Sea) had either not ended their previous reproductive cycle or were just beginning the next one. This concurs with observations (Htun-Han, 1978a, 1978b) that the reproductive cycle of the dab tends to extend throughout the year (i.e. there is possibly never a time when all individuals in the population are reproductively quiescent). In this situation, we feel it would be impossible to decide whether the presence of VTG in females, even those that appear to be immature (i.e. primary oocytes only), has an endogenous or exogenous cause.

The reason for measuring VTG in female plasmas in future studies would be, as in the present study, to validate the assay. VTG differences in females were consistent with its role as the precursor of the egg yolk protein -being low in immature females and much higher in those showing signs of VTG incorporation or atresia in their ovaries.

VTG was also highly correlated with maturity index and GSI in females. As our females were more mature in the North Sea than in the Irish Sea, VTG concentrations were also significantly higher in the former area.

Would the dab be useful as a monitor of biological effects of estrogens?

In 2006, a report was submitted to the Working Group on Concentrations, Trends and Effects of Substances in the Marine Environment [START_REF] Robinson | Draft OSPAR Background Document for Biological Effects Monitoring Techniques: Fish Vitellogenin as a Biomarker of Exposure to Xenoestrogens[END_REF] for consideration in relation to the OSPAR Convention for the Protection of the Marine Environment of the North East Atlantic. The main recommendations of the report were that: 1) in order to assess the occurrence, distribution, bioavailability and effects of estrogenic endocrine-disrupting chemicals in the OSPAR region, it is recommended that the determination of VTG in the blood plasma of male fish be added to the list of Contaminant-specific Biological Effects included in the OSPAR CEMP; 2) the recommended sentinel species are flounder (estuaries) and cod (offshore), which should be sampled in January/February. The committee reported back and accepted the first recommendation -plus the use of the flounder for monitoring estuaries. However, doubt was expressed about the choice of cod for offshore monitoring, because of its relative scarcity in some areas (especially of large male fish). The dab may be a better choice -with the advantage of demonstrated differences between sites. However, dab is lower in the food chain than adult cod, which, because the biomagnification factor of contaminants (and thus relative induction of VTG) is much lower, is a major disadvantage. Despite the successful outcome of the present study, if the dab were chosen as the sentinel species, it would be necessary to develop a more sensitive assay for VTG. Too many samples in the present study came close to or below the limits of the assay. This would (we suggest) be unacceptable for long-term trend data. The development of a more sensitive type of assay is not an insuperable problem [START_REF] Fukada | Development and validation of chemiluminescent immunoassay for vitellogenin in five salmonid species[END_REF]. It does require, however, considerably more effort and a higher degree of sophistication to develop than the type of ELISA described in this study.

What are the implications of estrogen exposure for the health of the fish?

This is the big question. As with freshwater fish, it is difficult to answer by studying fish in the field as there many other factors influence the survival of fish populations, including fishing pressure and long-term climate change [START_REF] Rose | Reconciling overfishing and climate change with stock dynamics of Atlantic cod (Gadus morhua) over 500 years[END_REF].

The occurrence of males with intersex gonads [START_REF] Lye | Abnormalities in the reproductive health of flounder Platichthys flesus exposed to effluent from a sewage treatment works[END_REF][START_REF] Allen | Survey of estrogenic activity in United Kingdom estuarine and coastal waters and its effects on gonadal development of the flounder Platichthys flesus[END_REF][START_REF] Cho | Histological abnormalities in the gonads of konoshiro gizzard shad (Konosirus punctatus) from coastal areas of Japan[END_REF][START_REF] De Metrio | Evidence of a high percentage of intersex in the Mediterranean swordfish (Xiphias gladius L.)[END_REF] or feminised secondary sexual characteristics [START_REF] Kirby | The presence of morphologically intermediate papilla syndrome in United Kingdom pupulations of sand goby (Pomatoschistus spp.): endocrine disruption[END_REF] has been directly ascribed to estrogenic exposure. Although, two intersex dab were found in the Dogger Bank area during the CSEMP survey in 2003 [START_REF] Stentiford | First reported case of intersex (ovotestis) in the flatfish species dab Limanda limanda: Dogger Bank, North Sea[END_REF], none was found in the 2004 and 2005 surveys. This rarity of intersex in dab could be taken as evidence that the small amount of estrogen to which they are apparently being exposed is not detrimental to the health of individuals or populations. However, intersex is neither a very specific nor very sensitive outcome of estrogen exposure in most species.

As already mentioned, VTG concentrations in male dabs lie within the bottom 0.1 percentile of possible VTG concentrations (assuming a VTG concentration of 10 mg/ml in a fully mature female or an E 2 -treated male). In cod, values lie within the bottom 2.0 percentile and, in flounder, occupy, and in some cases exceed, the entire range. On the basis of such differences, it would be tempting to conclude that the degree of estrogen exposure to which the dabs are being exposed is likely to be of little or no biological significance. However, with our present state of knowledge, it is dangerous to make such judgements. If, as we hypothesise, the causative agents in dab and cod are cumulative and persistent, VTG induction might just be the 'tip of the iceberg' in terms of biological and, ultimately, population effects.

We believe that we will only make headway in answering the question of population effects if we know what sorts of compounds (and routes of exposure) we are dealing with in marine fish. This information is also needed by governments to make rational decisions on how to deal with any problems (e.g. possible concentration further up the food chain to mammals that eat fish). Such an approach would also help to resolve whether the changes that are observed in cod and dab might have a simple natural cause such as ingestion of naturally-occurring phytoestrogens. CB01706). Dab collection and histology (under the auspices of CSEMP) were funded by Marine and Environmental Division, Defra, UK (contract AE003). We thank the numerous CEFAS staff who were involved in the collection of the plasma samples, histology, fish ageing and statistical analysis. 2. Two outlying observations are excluded (Table 3). Fig. 4. Vitellogenin concentrations (µg/ml) v. fish weight (plus least-squares regression line) for all but two (outlier) male dabs caught in June/July 2005. Fig. 5. Vitellogenin concentrations (µg/ml) v. fish age for males that were caught in the Dogger Bank area of the North Sea (sites 18 to 21). Relationship not statistically significant. Fig. 6. Vitellogenin concentrations (µg/ml) v. weight (plus least-squares regression line) for males that were caught in the Dogger Bank area of the North Sea (sites 18 to 21). Same males as shown in Fig. 5. Fig. 7. Mean ± SEM concentration of 17β-estradiol (lower graph) in female dab v. stage of maturation of the ovaries and also based upon whether they had 'high' or 'low' vitellogenin concentrations (upper graph). Each group had ten fish, except the first group that had an outlier (primary stage of development; plasma vitellogenin = 5 µg/ml; E 2 > 20 ng/ml) that was excluded from the analysis. The only group in which E 2 concentrations were significantly different from the other groups is marked with an asterisk. The stages of development were

Figure captions

  primary (stage 1; 424.0 ± 76.3, n = 125) or cortical alveoli-stage (642.2 ± 91.6, n = 76) oocytes only. A positive relationship was evident between log(VTG) concentrations and log(GSI) (r = 0.583, n = 235, p<0.0001). There was also a significant correlation between mean VTG concentrations and the mean female maturity index at all sites (r = 0.70, n = 26, p<0.0001). Mean maturity index of females in the North Sea (2.68 ± 0.10, n = 104) was significantly higher than that in the Irish Sea (1.19 ± 0.05. n = 107) and English Channel (1.95 ± 0.13, n = 58).

Fig. 1 .

 1 Fig. 1. UV absorption trace following anion-exchange chromatography of crude precipitated dab lipovitellin. The thinner line shows the gradient of 1M NaCl that was applied to the column: from 5% at the beginning to 100% at the top. Only the first major peak (A: 90-130 ml) cross-reacted with the antiserum to dab lipovitellin.

Fig. 2 .

 2 Fig.2. Cross-reaction in the ELISA of five-fold dilutions of purified dab lipovitellin (solid circles), purified dab vitellogenin (solid squares) and four randomly chosen female plasma samples (open symbols) that had been diluted 1:10 before being added to the wells. In two separate experiments, there was a high correlation in concentrations of vitellogenin in samples that had been measured using either lipovitellin or vitellogenin as plate-coating material and standard (r 2 = 0.87, n=60 in first experiment and 0.90, n=40 in second experiment, p<0.0001 for both).

Fig. 3 .

 3 Fig. 3. Map of mean plasma vitellogenin concentrations (µg/ml) in male dab. The heights of the bars indicates the mean vitellogenin concentration. The highest bar represents a value of 1.6 µg/ml. Data taken from Table2. Two outlying observations are excluded (Table3).

  Fig 2 above

  Fig 3 above

Fish

  Fig 4 above

Fish

  Fig 5 above

Fish

  Fig 6 above

Table 1

 1 Location of sites of capture of dab in June/July 2005

	Site No.	Date of capture	Site name	Latitude	Longitude	Depth (m)
	1	June 17 th , 2005	Carmarthan Bay	51.55N	4.65W	40
	2	June 18 th , 2005	Inner Cardigan Bay	52.18N	4.51W	25
	3	June 18 th , 2005	South Cardigan Bay	52.31N	4.26W	24
	4	June 19 th , 2005	North Cardigan Bay	52.70N	4.54W	51
	5	June 20 th , 2005	Dundrum Bay	54.08N	5.62W	46
	6	June 21 st , 2005	St Bee' s	54.55N	3.85W	29
	7	June 21 st , 2005	SE Isle of Man	54.06N	3.82W	47
	8	June 22 nd , 2005	Morecambe Bay	53.90N	3.40W	18
	9	June 23 rd , 2005	Burbo Bight	53.47N	3.32W	19
	10	June 24 th , 2005	Liverpool Bay	53.47N	3.70W	34
	11	June 24 th , 2005	Red Wharf Bay	53.37N	4.17W	44
	12	June 25 th , 2005	West Lundy	51.19N	5.46W	81
	13	June 28 th , 2005	Lyme Bay	50.61N	2.91W	31
	14	June 28 th , 2005	Eddystone	50.11N	4.15W	73
	15	June 29 th , 2005	Newhaven	50.76N	0.11W	21
	16	June 30 th , 2005	Rye Bay	50.76N	0.75E	39
	17	June 30 th , 2005	Outer Gabbard	52.04N	2.11E	45
	18	July 2 nd , 2005	Central Dogger	54.52N	2.69E	24
	19	July 2 nd , 2005	NE Dogger	55.27N	2.90E	32
	20	July 3 rd , 2005	West Dogger	54.79N	1.28E	37
	21	July 3 rd , 2005	North Dogger	55.06N	2.06E	32
	22	July 4 th , 2005	Off Flamborough	54.25N	0.48E	64
	23	July 4 th , 2005	Off Humber	54.08N	1.81E	82
	24	July 7 th , 2005	Indefatigable Bank	53.57N	2.09E	51
	25	July 7 th , 2005	Amble	55.25N	1.26W	68

Table 2

 2 Summary statistics of male dab (Mean ± S.E.)

	Site No.	Site	n	Length (cm)	Weight (g)	HSI	VTG (µg/ml) Maturity Index
	1	Carmarthan Bay	15	21.9 ± 0.6	111.7 ± 9.6	2.0 ± 0.2	0.828 ± 0.383	1.0
	2	Inner Cardigan Bay	12	17.8 ± 0.3	51.2 ± 3.2	1.3 ± 0.1	0.044 ± 0.014	1.1
	3	South Cardigan Bay	10	18.0 ± 0.4	52.7 ± 2.8	1.1 ± 0.1	0.062 ± 0.019	1.0
	4	North Cardigan Bay	9	16.9 ± 0.7	45.1 ± 5.4	1.4 ± 0.1	0.086 ± 0.022	1.0
	5	Dundrum Bay	6	17.3 ± 0.5	55.5 ± 4.0	1.7 ± 0.1	0.329 ± 0.178	1.3
	6	St Bee' s	20	18.7 ± 0.3	66.6 ± 3.6	1.4 ± 0.1	0.137 ± 0.021	1.8
	7	SE Isle of Man	35	20.4 ± 0.2	88.0 ± 3.1	1.4 ± 0.0	0.266 ± 0.091	1.0
	8	Morecambe Bay	15	19.9 ± 0.4	80.2 ± 5.1	1.6 ± 0.1	0.196 ± 0.057	1.0
	9	Burbo Bight	23	18.8 ± 0.2	64.2 ± 2.1	1.7 ± 0.1	0.501 ± 0.356	1.8
	10	Liverpool Bay	21	20.7 ± 0.5	93.1 ± 5.9	1.8 ± 0.1	0.583 ± 0.259	2.0
	11	Red Wharf Bay	15	19.3 ± 0.4	84.0 ± 6.3	2.4 ± 0.3	0.407 ± 0.129	1.9
	12	West Lundy	10	20.3 ± 0.5	87.9 ± 6.4	1.8 ± 0.1	0.232 ± 0.066	1.9
	13	Lyme Bay	20	19.8 ± 0.3	93.1 ± 4.6	2.6 ± 0.1	0.193 ± 0.040	2.0
	14	Eddystone	21	19.9 ± 0.3	94.8 ± 3.8	2.6 ± 0.1	0.565 ± 0.178	1.9
	15	Newhaven	16	20.2 ± 0.6	87.8 ± 7.2	1.5 ± 0.1	0.118 ± 0.021	2.0
	16	Rye Bay	18	20.2 ± 0.4	84.2 ± 5.6	1.8 ± 0.3	0.069 ± 0.017	1.9
	17	Outer Gabbard	10	22.5 ± 0.5	115.2 ± 6.7	1.5 ± 0.1	0.150 ± 0.058	1.8
	18	Central Dogger	22	20.6 ± 0.2	86.5 ± 2.9	1.7 ± 0.1	0.378 ± 0.143	1.7
	19	NE Dogger	24	23.0 ± 0.3	122.0 ± 5.3	1.2 ± 0.1	1.600 ± 0.450	1.6
	20	West Dogger	22	20.9 ± 0.3	87.9 ± 4.7	1.3 ± 0.1	0.277 ± 0.052	1.4
	21	North Dogger	22	21.3 ± 0.4	98.4 ± 5.3	1.6 ± 0.1	0.377 ± 0.154	1.8
	22	Off Flamborough	23	20.2 ± 0.2	79.8 ± 2.9	1.9 ± 0.1	1.353 ± 0.934	1.7
	23	Off Humber	16	20.9 ± 0.4	91.4 ± 4.3	1.8 ± 0.1	0.548 ± 0.120	2.0
	24	Indefatigable Bank	11	20.5 ± 0.4	90.0 ± 6.0	1.7 ± 0.1	0.172 ± 0.091	2.0
	25	Amble	17	21.0 ± 0.3	89.4 ± 4.6	1.8 ± 0.1	0.678 ± 0.231	1.8
	26	Tees Bay	15	22.3 ± 0.3	114.1 ± 4.2	1.8 ± 0.1	0.067 ± 0.012	1.7
	Two outliers were excluded (data in Table 3)

Table 3

 3 Sites and date of capture of two male dab outliers (high VTG).

	Site No.	Date of capture	Site name	Length (cm) Weight (g) Liver weight (g) E2 (ng/ml) VTG (µg/ml)	Maturity index
	22	July 4 th , 2005	Off Flamborough	21	84	1.3	<0.04	21.6	1
	10	June 24 th , 2005	Liverpool Bay	22	116	3.9	<0.04	755.8	2

Table 4

 4 Summary statistics of female dab (Mean ± S.E.)

	Site No. Date of capture	Site name	n	Length (cm) Weight (g)	HSI	GSI	VTG (µg/ml)	Maturity index
	1	June 17 th , 2005	Carmarthan Bay	12	24.3 ± 0.7 161.3 ± 14.6 2.3 ± 0.3	0.8 ±0.1	673.2 ± 229.6	1.0
	2	June 18 th , 2005	Inner Cardigan Bay	10	21.3 ± 0.4	89.5 ± 7.0	1.5 ± 0.1 0.8 ± 0.1	107.6 ± 30.2	1.0
	3	June 18 th , 2005 South Cardigan Bay	10	19.0 ± 0.5	62.8 ± 3.4	1.1 ± 0.0 0.8 ± 0.1	225.2 ± 77.7	1.0
	4	June 19 th , 2005 North Cardigan Bay	10	22.2 ± 0.3	97.2 ± 3.2	1.5 ± 0.2 0.8 ± 0.1	84.9 ± 62.0	1.0
	5	June 20 th , 2005	Dundrum Bay	5	18.4 ± 0.9	63.0 ± 9.6	1.5 ± 0.2 0.7 ± 0.1	117.1 ± 76.4	1.0
	6	June 21 st , 2005	St Bee' s	10	22.0 ± 0.8 114.8 ± 10.0 1.5 ± 0.2	0.9± 0.1	771.4 ± 216.3	1.3
	7	June 21 st , 2005	SE Isle of Man	10	23.3 ± 0.9 129.0 ± 17.0 1.5 ± 0.2 0.9 ± 0.1	574.6 ± 267.8	1.3
	8	June 22 nd , 2005	Morecambe Bay	10	23.7 ± 0.6 145.6 ± 12.1 1.8 ± 0.3 0.9 ± 0.1	562.2 ± 106.2	1.2
	9	June 23 rd , 2005	Burbo Bight	10	21.8 ± 0.7 105.7 ± 9.5 2.0 ± 0.2 0.8 ± 0.1	457.2 ± 203.2	1.1
	10	June 24 th , 2005	Liverpool Bay	10	25.6 ± 0.7 168.6 ± 14.5 1.8 ± 0.2 1.1 ± 0.3	584.5 ± 244.4	1.4
	11	June 24 th , 2005	Red Wharf Bay	10	23.7 ± 1.1 158.7 ± 24.5 2.2 ± 0.2 0.9 ± 0.1	146.5 ± 69.9	1.7
	12	June 25 th , 2005	West Lundy	10	20.2 ± 0.2	83.1 ± 3.3	1.4 ± 0.0 0.6 ± 0.0	242.1 ± 91.3	1.8
	13	June 28 th , 2005	Lyme Bay	20	22.6 ± 0.9 149.2 ± 20.8 2.4 ± 0.1 0.5 ± 0.0	126.1 ± 68.4	2.2
	14	June 28 th , 2005	Eddystone	9	19.9 ± 0.4	87.8 ± 4.7	2.4 ± 0.2 0.5 ± 0.0	94.1 ± 53.3	2.1
	15	June 29 th , 2005	Newhaven	10	22.2 ± 0.6 124.4 ± 12.0 1.8 ± 0.2 0.6 ± 0.0	214.5 ± 121.5	1.6
	16	June 30 th , 2005	Rye Bay	9	23.1 ± 0.5 121.8 ± 9.0 1.2 ± 0.2 1.1 ± 0.0	348.2 ± 152.7	1.8
	17	June 30 th , 2005	Outer Gabbard	10	23.2 ± 0.4 127.8 ± 8.1 1.3 ± 0.1 0.8 ± 0.1	1369.7 ± 599.9	1.9
	18	July 2 nd , 2005	Central Dogger	10	24.2 ± 0.6 143.1 ± 11.2 1.4 ± 0.1 1.2 ± 0.1	2776.2 ± 745.3	2.6
	19	July 2 nd , 2005	NE Dogger	11	24.1 ± 0.4 130.7 ± 7.9 2.2 ± 0.1 1.9 ± 0.6	1675.8 ± 561.6	2.6
	20	July 3 rd , 2005	West Dogger	10	21.7 ± 0.5 105.7 ± 8.1 1.3 ± 0.1 1.2 ± 0.2	2537.4 ± 541.9	2.9
	21	July 3 rd , 2005	North Dogger	10	23.2 ± 0.7 130.1 ± 9.0 1.3 ± 0.1 2.0 ± 0.5	3936.9 ± 924.5	3.1
	22	July 4 th , 2005	Off Flamborough	11	23.5 ± 0.5 121.5 ± 8.0 1.3 ± 0.1 1.1 ± 0.1	3185.9 ± 1138.9	3.0
	23	July 4 th , 2005	Off Humber	10	23.3 ± 0.6 123.6 ± 9.4 1.5 ± 0.1 1.3 ± 0.1	2341.6 ± 456.0	3.2
	24	July 7 th , 2005	Indefatigable Bank	10	24.3 ± 0.7 144.0 ± 11.6 1.9 ± 0.1 0.9 ± 0.1	765.7 ± 282.0	2.3
	25	July 7 th , 2005	Amble	10	25.3 ± 0.7 167.8 ± 15.5	-	-	1653.2 ± 515.9	2.5
	26	July 8 th , 2005	Tees Bay	11	22.5 ± 0.5 120.5 ± 6.7 1.8 ± 0.1 1.0 ± 0.1	890.2 ± 208.8	2.7

Age analysis was only carried out for females caught on the Dogger Bank (sites 18 to 21). The oldest female was 8 years old. VTG was not related to either age or water depth in females.

based upon the most advanced stage of oocytes present in the ovaries: primary; cortical alveoli; vitellogenic (essentially the new season's growing oocytes) and atretic (essentially the previous season's unused oocytes).

July 8 th , 2005 Tees Bay 54.77N 1.14W

Acknowledgements

This work was primarily funded by the Chemicals and Nanotechnologies Division, Department for the Environment, Food and Rural Affairs (Defra), UK (contract


Gail V Ashton

Kate J Willis

Michael T Burrows

Elizabeth J Cook

Environmental tolerance of Caprella mutica: implications for its distribution as a marine non-native species

Keywords: Biological invasion, caprellid amphipod, non-native distribution, physiological tolerance, salinity tolerance, temperature tolerance

Physiological tolerances limit the distribution of marine species, with geographical ranges being set by environmental factors, such as temperature and salinity, which affect the rates of vital processes and survival of marine ectotherms. The physiological tolerances of the non-native marine amphipod Caprella mutica were investigated in laboratory experiments. Adult C. mutica were collected from a fish farm on the west coast of Scotland and exposed to a range of temperatures and salinities for 48 h. C. mutica were tolerant of a broad range of temperature and salinity conditions, with 100% mortality at 30 °C (48 h LT50, 28.3 ± 0.4 °C), and salinities lower than 16 (48 h LC50, 18.7 ± 0.2). Although lethargic at low temperatures (2 °C), no mortality was observed, and the species is known to survive at temperatures as low as -1.8 °C. The upper LC 50 was greater than the highest salinity tested (40), thus it is unlikely that salinity will limit the distribution of C. mutica in open coastal waters.

However, the species will be excluded from brackish water environments such as the heads of sea lochs or estuaries. The physiological tolerances of C. mutica are beyond the physical conditions experienced in its native or introduced range and are thus unlikely to be the primary factors limiting its present distribution and future spread.

Introduction

The marine amphipod Caprella mutica, a native of north-east Asia, is now widely distributed in the northern hemisphere between latitudes 35˚N and 70˚N, representing a considerable extension to its native latitudinal range (Ashton et al., in press). The non-native populations discovered to date are restricted to temperate oceanic coasts that experience temperatures ranging from 0 to 22 ˚C, similar to that of the native habitat, which has an annual seawater temperature range of -1.8 to 25 °C [START_REF] Schevchenko | Seasonal dynamics of the diatoms of the genus Chaetoceros Ehrenberg in Amursky Bay (Sea of Japan)[END_REF].

Environmental factors are likely to play an important role in influencing the present and future global distribution of C. mutica, as physiological tolerances result in ranges being set by physical characteristics that affect the rates of vital processes and survival [START_REF] Vermeij | An agenda for invasion biology[END_REF]. Temperature and salinity are often considered to be the principle variables affecting growth, reproduction and survival of marine ectotherms (Kinne, 1970a;[START_REF] Kinne | Temperature: Animals: Invertebrates[END_REF], and can act singly or in combination to influence distribution patterns [START_REF] Wiederholm | Distribution of Potamoschistus minutus and P. microps (Gobiidae, Pisces) in the Bothnian Sea: importance of salinity and temperature[END_REF][START_REF] Stauffer | Effect of Salinity on the Temperature Preference and Tolerance of Age-0 Mayan Cichlids[END_REF]. Previous work has shown that while temperature does not limit the latitudinal distribution of C. mutica on the west coast of Scotland (Ashton et al., in press), it does influence global distribution patterns of caprellid amphipods [START_REF] Laubitz | Studies on the Caprellidae (Crustacea, Amphipoda) of the American North Pacific[END_REF][START_REF] Thiel | The distribution of littoral caprellids (Crustacea: Amphipoda: Caprellidea) along the Pacific coast of continental Chile[END_REF]. For example, temperature is the primary factor limiting the distribution of caprellids in Australasia and Southern Chile [START_REF] Thiel | The distribution of littoral caprellids (Crustacea: Amphipoda: Caprellidea) along the Pacific coast of continental Chile[END_REF], and in the North Pacific [START_REF] Laubitz | Studies on the Caprellidae (Crustacea, Amphipoda) of the American North Pacific[END_REF], with most species occurring in the 10 to 15 °C temperature range, and only a few in the sub-antarctic 5 to 10 °C range. The relatively disjunctive non-native distribution of C. mutica within the region experiencing temperatures between 0 to 22 °C, suggests that either range expansion is still underway, or factors other than temperature are affecting its establishment at a global scale.

A wide salinity tolerance is typical of many amphipod species (Kinne, 1970a), with several caprellid species displaying tolerance to salinities lower than 15 [START_REF] Takeuchi | Gill ultrastructure and salinity tolerance of Caprella spp. (Crustacea: Amphipoda: Caprellidae) inhabiting the Sargassum community[END_REF]. Upper salinity tolerance limits are unlikely to be a limiting factor, as illustrated by [START_REF] Takeuchi | Gill ultrastructure and salinity tolerance of Caprella spp. (Crustacea: Amphipoda: Caprellidae) inhabiting the Sargassum community[END_REF], who found the salinity tolerance of four Caprella spp. to be greater than 34. In the native habitat of C. mutica, salinities range from 11 to 35 [START_REF] Schevchenko | Seasonal dynamics of the diatoms of the genus Chaetoceros Ehrenberg in Amursky Bay (Sea of Japan)[END_REF]. The presence of C. mutica in Scottish sea lochs that, at times, receive considerable freshwater inputs (Ashton et al., in press), suggests that this caprellid is relatively tolerant of low salinities.

Tolerance of a wide range of environmental conditions is suggested as a characteristic of successful non-native species [START_REF] Van Der Velde | Ecological impact of crustacean invaders: General considerations and examples from the Rhine River[END_REF][START_REF] Strasser | Mya arenaria -an ancient invader of the North Sea coast[END_REF]. However, it is currently unknown whether the environmental conditions experienced by C. mutica throughout its distribution range correspond with its environmental tolerances. The caprellid may be limited to its current range by its physiological tolerance limits to temperature and salinity, or other factors may be limiting its biogeographic distribution. Here, we present the results of a study to investigate the lethal salinity and temperature tolerance limits of the non-native amphipod C. mutica on the west coast of Scotland. This study aimed to establish the physiological tolerances of the species to these two key environmental variables that are likely to be important in limiting the caprellids global distribution.

Methods

Caprellid collection and maintenance

Caprella mutica were collected from filamentous red algae attached to pontoon structures at Dunstaffnage fish farm (56° 27.090' N 05° 27.733' W) in June 2005 and 2006. The animals were transported in sea water, still attached to the algae, in 10 L buckets. Animals were acclimatised to laboratory conditions for 48 h at a constant temperature (14 ±1 °C, ambient temperature) at a photoperiod of 8 D:16 L, the photoperiod regime was maintained throughout the tolerance experiments. The buckets received constant aeration and no additional food was given.

Temperature tolerance experiments

Fifteen containers (1.5 L) were filled with filtered sea water (100 m; salinity ~35) and five pieces of 5 mm mesh (150 cm 2 ) were added to each container. Larger, healthy-looking animals were carefully isolated from the algae using forceps and transferred to the mesh in the experimental containers. A maximum of 34 individuals, including at least five male and five female C. mutica were added to each container.

Starting numbers in each container were recorded. Experimental containers were then transferred into water baths (60 x 30 cm) maintained at five temperatures, with three replicate containers at each temperature. The containers reached test temperatures within 10 minutes of the start of the experiment. Temperatures were maintained using either TropicMarin heaters, or a Grant chiller, and were monitored throughout the experiment with temperature data loggers (Tinytag Aquatic, TG-3100). After 48 h, the containers were removed from the water baths and animals were immediately inspected for mortality, defined as a lack of response to mechanical stimulus (gentle prodding with a blunt probe).

A preliminary range-finding test using temperatures of 2, 10, 22, 31 and 40°C was undertaken to define approximately the lethal tolerance limits of C. mutica. These temperatures were selected pragmatically in an effort to include mortality rates ranging from 0% to 100%. In the range finding test, caprellids exposed to 40 o C died within 5 minutes, while at 2 o C they were immobile after 48 h, and lethargic when stimulated. The second definitive experiment used a narrower range of temperatures (21, 24, 26, 28 and 30 o C) encompassing the upper lethal limit, which was estimated to lie between 22 and 31 o C in the range finding test.

Salinity tolerance experiments

Experimental procedures and conditions were as described for the temperature tolerance tests, with the exception that the salinity tolerance experiments were maintained at a temperature of 14 ±1 °C in a constant temperature room.

Large healthy-looking caprellids were transferred to exposure tanks (1.5 L) containing filtered sea water (100 m) at five different salinities. Salinities were achieved by addition of artificial sea salt (Tropical Marin) to water treated by reverse osmosis. Salinity was confirmed using a portable refractometer (Atago, S/MILL) at the beginning and end of the 48 h exposure periods. After 48 h, individuals were inspected for mortality by gentle prodding with a blunt probe.

As above, a preliminary range-finding test was done using salinities of 0, 10, 20, 30 and 40 in an attempt to incorporate concentrations causing 0% and 100% mortality. In the range finding test, caprellids exposed to a salinity of 0 were immobile within 5 minutes, with 100% mortality at salinities of 0 and 10 after 48 h. The upper lethal limit was higher than 40 (<10% mortality at salinities of 30 and 40). The second definitive experiment employed a narrower range of salinities (9, 13, 15, 18 and 20) encompassing the lower lethal limit, which was estimated to lie between 10 and 20 in the range finding test.

Statistical Analysis

Median lethal LT 50 and LC 50 values and associated 95% confidence limits were calculated using probit [START_REF] Finney | Probit Analysis. 3rd edn[END_REF] or trimmed Spearman-Karber analysis (TSK; [START_REF] Hamilton | Trimmed Spearman-Karber method for estimating median lethal concentrations in toxicity bioassays[END_REF]; U.S.E.P.A., 1990) using MINITAB. MINITAB uses a Chisquared analysis ( 2 ) to test whether the data fits the Probit model. When 2 rejected the Probit model (P > 0.05), a trimmed Spearman-Karber analysis was used to calculate a non-parametric estimate of the lethal value. A disadvantage of the TSK analysis is that it only yields the median lethal response (LC 50 ) and 95% confidence limits. When the test results are not unimodal or symmetrical about the median lethal response, the confidence intervals (CI) can be unreliable. A pairwise students' t-test was used to compare differences in mortality of male and female C. mutica at each temperature.

Results

Temperature tolerance

Following 48 h exposure to temperatures between 21 o C and 30 o C, there was a marked increase in mortality of C. mutica at temperatures greater than 26 o C, with 100% mortality at 30 ºC (Figure 1). At 24 o C, mortality was increased (~25%) relative to that at 26 o C, and similar to mortality at 28 o C. The reason for this is unknown, but may be a response to other uncontrolled experimental conditions (e.g. pH or oxygen).

Mortality was <10% at temperatures of 21 o C and 26 o C. The 48 h LT 50 (± 95% C.L.) was 28.3 ± 0.4 ºC. There was no significant difference in mortality between male and female C. mutica. Salinity tolerance 100% mortality was observed following 48 h exposure of C. mutica to salinities lower than 15 (Figure 2). At a salinity of 18, individuals were lethargic and less responsive to stimulation than those at 20. The lower 48 h LC 50 (± 95% C.L.) was 18.7 ± 0.2. The lower LC 50 for females (18.0 ± 0.3) was significantly higher (P<0.05) than that for males (19.3 ± 0.4).

Discussion

Adult C. mutica collected on the west coast of Scotland were tolerant of a wide range of temperature and salinity conditions in 48 h laboratory exposures. The upper thermal tolerance limit of C. mutica (48 h LT 50 , 28.3 °C) was higher than maximum temperatures experienced in its non-native Scottish habitat (16 o C; Ashton et al., in press), or in its native habitat (25 °C, [START_REF] Schevchenko | Seasonal dynamics of the diatoms of the genus Chaetoceros Ehrenberg in Amursky Bay (Sea of Japan)[END_REF]. The lower lethal temperature was not determined as there was no mortality at the lowest temperature tested (2 °C), although the caprellids were immobile after 48 h at 2 °C and were lethargic when stimulated. After 48 h at a salinity of 18, C. mutica were lethargic, and at salinities lower than 16, 100% mortality was observed. These salinity values are also higher than those recorded in the native and non-native habitat of the caprellid (11 and 10.5 respectively; [START_REF] Schevchenko | Seasonal dynamics of the diatoms of the genus Chaetoceros Ehrenberg in Amursky Bay (Sea of Japan)[END_REF]Ashton et al., in press).

The current experiments were run for a period of 48 hours, and therefore only give an indication of the short-term tolerance of C. mutica. However, the thermal tolerance limits recorded in this study suggest that C. mutica could survive introduction to the tropics or the poles, extending its current distribution. In the northern hemisphere, C. mutica has been introduced to both coasts of North America, from 37.5 to 48.5 o N and to the Atlantic coasts of Europe from 49.5 to 69.1 o N. In the southern hemisphere, C. mutica has only been described from New Zealand (Ashton et al., in press). Sea surface temperatures at the equator are relatively stable at 26 to 27 o C throughout the year [START_REF] Tait | Elements of Marine Ecology[END_REF], slightly lower than the 48 h upper lethal limit of 28 o C. However, the ability of C. mutica to survive prolonged exposure to high temperatures is unknown, and it is also possible that earlier life stages may not be as tolerant to high temperatures as the adults.

In its native habitat, populations of C. mutica survive at temperatures as low as -1.8 °C, and non-native populations have established in areas that freeze at the surface during winter, such as Passammaquoddy Bay, Canada (Ashton et al., in press). Given the lethargic state of the caprellids after 48 h at 2 o C in the present study, it is likely that growth and reproduction are reduced, or cease, at such low winter temperatures.

The salinity tolerance of adult C. mutica is similar to or slightly less than that of other Caprella species. The lower 48 h LC 50 (18.7 ± 0.2) is higher than that of C. scaura and C. equilibria, with 24 h LC 50 values of 7.5 and 11.5 respectively [START_REF] Cockman | Caprellidae of the Swan River Estuary[END_REF], noting that these values are for 24 h exposures, not 48 h. For other Caprella species, C. danilevski, C. subinermis, C. penantis and C. verrucosa, 48 h LC 50 values ranged from 12 to 17 [START_REF] Takeuchi | Gill ultrastructure and salinity tolerance of Caprella spp. (Crustacea: Amphipoda: Caprellidae) inhabiting the Sargassum community[END_REF]. The increased salinity tolerance of male C. mutica in the present study may reflect the sexual dimorphism of the species (the males are considerably larger) or differences between the sexes in resource allocation to reproduction [START_REF] Hoback | Lethal limits and sublethal effects of hypoxia on the amphipod Gammarus pseudolimnaeus[END_REF]. It is surprising, however, that this trend was not observed with temperature, which might be expected to produce a greater physiological response [START_REF] Davenport | Effects of shore height, wave exposure and geographical distance on thermal niche width of intertidal fauna[END_REF].

In the native habitat, C. mutica individuals experience salinities as low as 11 [START_REF] Schevchenko | Seasonal dynamics of the diatoms of the genus Chaetoceros Ehrenberg in Amursky Bay (Sea of Japan)[END_REF], considerably lower than the 48 h LC 50 in this study. Thus, the current realized distribution of C. mutica with respect to salinity is beyond its fundamental niche described here. This is contrary to the expectation for marine ectotherms [START_REF] Sprague | Resistance of four freshwater crustaceans to lethal high temperature and low oxygen[END_REF][START_REF] Tait | Elements of Marine Ecology[END_REF]. In coastal habitats, C. mutica are probably able to survive fluctuating salinity conditions during the tidal cycle and short periods of exposure to low salinity during periods of heavy rainfall. However, prolonged exposure to salinity lower than 18 is likely to adversely affect the caprellid, and may explain why populations are absent from the upper reaches of Scottish sea lochs that regularly experience low salinities (Ashton pers. obs.).

The broad physiological tolerance of adult C. mutica would suggest that changing temperature and salinity conditions experienced during transport through equatorial or polar regions would be unlikely to adversely affect individuals in ballast water [START_REF] Witt | Echinogammarus ischnus: another crustacean invader in the Laurentian Great Lakes basin[END_REF] or on ships' hulls [START_REF] Minchin | Fouling and ships' hulls: how changing circumstances and spawning events may result in the spread of exotic species[END_REF]. Species distributions that are limited by thermal tolerances are expected to change in response to climate change (e.g. marine mammals, [START_REF] Tynan | Observations and predictions of Arctic climate change: potential effects on marine mammals[END_REF]coral reefs, Hoegh-Guldberg, 1999;pelagic ecosystems, Boyd & Doney, 2002). Predicted rates of range expansion in terrestrial environments include 6.1 km per decade [START_REF] Parmesan | A globally coherent fingerprint of climate change impacts across natural systems[END_REF] and 12.5 -24.8 km per decade [START_REF] Hickling | The distributions of a wide range of taxonomic groups are expanding polewards[END_REF]. In particular, evidence suggests that many invasive species possess traits that will enable them to capitalize on the various elements of global change, exacerbating the impacts of invasions on stressed ecosystems [START_REF] Dukes | Does global change increase the success of biological invaders?[END_REF]. Climate change will be important in determining the future distribution of C. mutica, which can be expected to extend polewards.

In addition to environmental conditions, the success of non-native species is also influenced by the frequency of introduction events, which are generally more common in temperate waters [START_REF] Minchin | Fouling and ships' hulls: how changing circumstances and spawning events may result in the spread of exotic species[END_REF][START_REF] Drake | Global hot spots of biological invasions: evaluating options for ballast-water management[END_REF] compared to tropical or polar regions, propagule pressure (the frequency and size of introduction events), and biotic interactions in the recipient region such as competition, facilitation and predation [START_REF] Bruno | Insights into Biotic Interactions from Studies of Species Invasions[END_REF]. Thus, the current distribution of C. mutica is likely to reflect a combination of introduction opportunities, as well as favourable abiotic and biotic conditions in these areas.

Broad physiological tolerance and plasticity are commonly cited as characteristics of successful invading species [START_REF] Van Der Velde | Ecological impact of crustacean invaders: General considerations and examples from the Rhine River[END_REF], with temperature and salinity tolerances often combining to limit the potential distributions of marine species [START_REF] Wiederholm | Distribution of Potamoschistus minutus and P. microps (Gobiidae, Pisces) in the Bothnian Sea: importance of salinity and temperature[END_REF][START_REF] Stauffer | Effect of Salinity on the Temperature Preference and Tolerance of Age-0 Mayan Cichlids[END_REF]. Given the broad temperature and salinity tolerance of adult C. mutica observed in this study, the caprellid is likely to be limited to marine habitats experiencing a temperate climate (not exceeding 26 °C), and regions where salinity does not drop below 18 for extended periods of time. Evidence suggests that dispersal via human vectors is increasing the distribution of C. mutica at a faster rate than climate change [START_REF] Ashton | Distribution and dispersal of the non-native caprellid amphipod, Caprella mutica Schurin 1935[END_REF]. Therefore, at present human dispersal will be the most important factor in determining the range expansion of C. mutica.

Figure 1 .

 1 Figure 1. Mortality of C. mutica following 48 h exposure to temperatures between 21

Figure 2 .

 2 Figure 2. Mortality of C. mutica following 48 h exposure to salinities between 9 and

  Figure 1.

Acknowledgements

We thank A. Keay for help with the experimental setup and H. Brown and R. Shucksmith for assistance with field collection of C. mutica. This work was funded through a PhD studentship by the UK Natural Environmental Research Council (NER/S/A/2003/11899).