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Immune modulation in the blue mussel Mytilus edulis exposed to North Sea produced water
Environmental Pollution 157 (2009) 1939–1944
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Immune modulation in the blue mussel Mytilus edulis exposed to North Sea produced water M.L. Hannam a, *, S.D. Bamber b, R.C. Sundt b, T.S. Galloway c a
Ecotoxicology & Stress Biology Research Centre, School of Biological Sciences, University of Plymouth, Drake Circus, Plymouth, Devon, PL4 8AA, UK IRIS – Biomiljø, Mekjarvik 12, 4070 Randaberg, Norway c School of Biosciences, University of Exeter, Hatherly Laboratories, Prince of Wales Road, Exeter, EX4 4PS, UK b
Exposure to produced water alters immune function in the sentinel species Mytilus edulis.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 24 October 2008 Received in revised form 12 December 2008 Accepted 18 December 2008
The discharge of oil well produced water (PW) provides a constant source of contaminants to the marine environment including polycyclic aromatic hydrocarbons, alkylated phenols, metals and production chemicals. High concentrations of PW cause adverse effects to exposed biota, including reduced survival, growth and reproduction. Here we explore the effects of PW on immune function in the blue mussel, Mytilus edulis. Mussels were exposed for 21 days to sublethal PW concentrations (0.125–0.5%) and cellular parameters were measured. Cell viability, phagocytosis and cytotoxicity were inhibited after exposure to 0.25% and 0.5% PW, whilst the 0.125% PW treatment produced significant increases in these biomarker responses. This biphasic response was only observed after 7 days exposure; longer exposure periods led to a reduction in immune parameters. Results indicate that PW concentrations close to the discharge point cause modulation to cellular immunity. The implications for longer-term disease resistance are discussed. Ó 2009 Elsevier Ltd. All rights reserved.
Keywords: Produced water Polycyclic aromatic hydrocarbons Mytilus edulis Immunotoxicity Phagocytosis
1. Introduction The marine environment is a major sink for many potentially hazardous pollutants discharged from industrial sources, such as produced water (PW). Produced water is formed as a by-product of oil extraction, with the amount of PW increasing as an oil field becomes depleted (Strømgren et al., 1995). Most PW has oil concentrations below the regulatory limit with PW discharges from North Sea installations recording an average oil content of 17 mg l1 in 2006 (OGP, 2006). In addition to oil, phenols, alkylphenols and metals also occur in PW along with chemicals (biocides, corrosion inhibitors, scale inhibitors, coagulants, detergents and surfactants) used in the process system (Neff, 2002). With 395 106 l day1 of PW entering the North Sea (OLF, 2006), these discharges represent a chronic source of pollution to marine systems with the potential to cause adverse effects to exposed biota (Neff, 2002). Adverse effects of PW may result from acute toxicity or from chronic sublethal alterations in homeostatic mechanisms, such as the immune system (Pipe et al., 1999). Immune systems of invertebrates do not operate in isolation, but are highly integrated with
* Corresponding author. Tel.: þ44 1752 232930; fax: þ44 1752 232970. E-mail address: [email protected] (M.L. Hannam). 0269-7491/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2008.12.031
other physiological systems making immune responses particularly sensitive to environmental contaminants (Galloway and Goven, 2006; Luengen et al., 2004). The immune system of invertebrates is relatively simple in contrast to that of vertebrates, providing accessible means for monitoring immune function (Galloway and Depledge, 2001). Unlike vertebrates, invertebrates lack immune specificity and memory, characteristic of the adaptive immunity observed in mammalian subjects (Anderson, 1988). Invertebrates have evolved innate defence mechanisms that can identify and protect against non-self material, with invertebrate immune response centred largely on the multifunctional haemocytes (Cheng, 1981; Galloway and Goven, 2006). The various parameters that contribute to invertebrate host defence act to provide a multifaceted immune response. It has therefore been suggested that an integrated suite of biomarkers be used for assessing immunocompetence of invertebrates (Livingstone et al., 2000; Pipe and Coles, 1995; Weeks et al., 1992). Advances in environmental risk assessment have highlighted the value of biomarkers to allow rapid evaluation of contaminant exposure and effect (Galloway et al., 2002, 2006). These biomarkers may be molecular, cellular or physiological endpoints that indicate exposure to and/or damage incurred by environmental pollutants providing early-warning indicators of contaminant-induced stress
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(Depledge and Fossi, 1994). The use of biomarkers has been recognised as a cost-effective technique in identifying the effects of pollutants on biota (Galloway et al., 2002). The majority of biomarker studies to date have been based on the mussel Mytilus edulis, and the American oyster Crassostrea virginica (Wootton et al., 2003). Bivalves exhibit worldwide distribution and sedentary, filter feeding behaviour. They have an open circulatory system that is continually exposed to fluctuating environmental conditions and pollutants (Pipe et al., 1999). In addition, a typical low metabolic transformation rate and a high bioaccumulation factor of 105 make bivalves an ideal biomonitoring organism, and as such they are now commonly used sentinels in ecotoxicology (Auffret et al., 2006; Galloway et al., 2002; Widdows and Donkin, 1992). In this study we tested the hypothesis that exposure of the blue mussel to produced water would cause modulation of immune function. 2. Method 2.1. Experimental design Mussels, M. edulis (5–7 cm) were collected from a reference site at Tysvær, Norway (59 190 400 N, 4 280 500 E), an area not subjected to large industrial or agricultural activity, and held in 600 l flow through (7.5 l min1) tanks containing filtered seawater at 8.5 C (1 C). Mussels were acclimated for 7 days prior to transfer to the exposure system. Test mussels were divided into eight exposure tanks (45 per 600 l tank with a flow rate of 7.5 l min1) maintained at 8.5 C (1 C) and a salinity of 34.5 PSU. PW was tapped on site from the Ekofisk oil field and stored at 20 C. PW treatments of 0.5, 0.25 and 0.125% and a seawater control were set up in duplicate tanks in a continuous flow system (CFS) fed from a header tank containing PW, described in detail by Sundt (2004). Immune response assays were carried out on 12 mussels from each treatment group after exposure periods of 0, 7, 14 and 21 days. 2.2. Chemical analysis Water from the PW header tank was sampled for PAH analysis on day 6 of the exposure. This was collected in 2 l amber bottles containing hydrochloric acid to ensure low acidity (pH < 2) and analysed for PAHs within 2 days of collection; the 16 PAHs on the US EPA priority pollutant list were measured along with alkyl homologues of naphthalene, chrysene, dibenzophiothene and phenanthrene/ anthracene. Eight deuterated PAHs were added as quantitative internal standards and mixed on a magnetic stirrer before triplicate extraction with cyclohexane (3 50 ml). Each extraction was performed by stirring the sample and solvent for 30 min on a magnetic stirrer and then pouring the entire bottle contents into a 2 l separating funnel. The water phase was drained back into the sampling flask and extracted a further two times. Combined organic extracts were dried with anhydrous NaSO4, filtered and concentrated to 1 ml using TurboVap 500. The samples were then transferred to vials and analysed by gas chromatography–mass spectrometry (GC-MS). PAHs were expressed as observed concentration in produced water (mg l1) and also toxic equivalents (mg l1). This latter case expresses individual PAHs as equivalents of benzo[a]pyrene, based upon their relative toxicity using toxic equivalency factors (TEFs) developed by Nisbet and LaGoy (1992). It should be noted that TEF values are based on the carcinogenic potential of PAHs in vertebrates and are not directly related to invertebrate models; it does, however, provide a method of comparison for the relative potential toxicities of PAHs present in PW. 2.3. Biological analyses Immune function assays were conducted in triplicate on haemolymph samples extracted into an equal volume of physiological saline (0.02 M HEPES, 0.4 M NaCl2, 0.1 M MgSO4, 0.01 M KCl, 0.01 M CaCl2; pH 7.4), from the posterior adductor muscle using a 21 gauge needle. Samples were transferred to siliconised EppendorfÒ tubes and stored on ice until analysis. Total haemocyte counts were carried out using an improved Neubauer haemocytometer. Haemocyte viability was assessed by measuring retention of Neutral Red (NR) dye (Babich and Borenfreund, 1992) as described in detail by Canty et al. (2007). Briefly, haemolymph samples were pipetted onto a microplate. After 45 min incubation (4 C), non-adhered cells were removed before adding aliquots of 0.004% NR solution. After 3 h incubation at 4 C excess NR solution was removed and acidified ethanol was added to resolubilise the dye and the optical density was measured at 540 nm. The optical density of Neutral Red retention was determined as a function of protein content, determined using the BCA protein assay (Pierce 23223/4).
Phagocytic activity of haemocytes was assessed by measuring the uptake of Neutral Red stained zymosan particles (from Saccharomyces cerevisiae) as described by Parry and Pipe (2004). Briefly, 50 ml haemolymph samples were pipetted onto a microplate, and adhered cells were incubated with 50 ml of dyed zymosan suspension (50 107 particles ml1) for 30 min at room temperature. Uptake of these dyed particles was determined spectrophotometrically at 550 nm against a standard curve prepared using zymosan suspensions (50–1.56 107 particles ml1) and expressed as a function of protein content. The ability of haemolymph to lyse sheep erythrocytes was used as a measurement of cytotoxic capability (Galloway et al., 2002). Haemolymph samples were diluted to a concentration of 1 104 cells ml1 in Tris-buffered saline (10 mM Tris– HCl, 150 mM NaCl, 10 mM CaCl2, pH 7.4). One millilitre of sheep erythrocytes (TCS Biosciences Ltd) was centrifuged at 200 g for 5 min before re-suspending in phosphate-buffered saline. The suspension was centrifuged again and a cell pack volume of 125 ml was re-suspended in 1 ml of Tris-buffered saline (TBS), this was then further diluted with 15 ml of TBS. Aliquots of diluted haemolymph (100 ml) were pipetted onto a round-bottomed microplate, along with duplicate controls of 100 ml TBS (for spontaneous release) and 100 ml TBS with 2 ml of 2% Triton-X100 (for positive control). One-hundred microlitres of the erythrocyte suspension was added and the plate was incubated at 25 C for 1 h. After centrifugation (200 g, 10 min) the supernatant was transferred to a flat-bottomed microplate, and the lysis determined through the optical density at 405 nm of the haemoglobin released into the supernatant. The cytotoxicity was expressed as percent lysis relative to the maximum release observed in the positive control. 2.4. Statistical analyses Biological endpoints were measured in 12 individuals from each treatment group at each time point, with results expressed as mean values 1 standard error. Data sets were analysed using two-way ANOVAs; tests were performed on each immune parameter to determine significant effects due to interactions (treatment time) or main factors (treatment and time). Post-hoc pairwise comparisons were conducted (Fisher’s LSD) to identify where significant differences occurred at the 95% confidence level (associated probability < 0.05).
3. Results Ekofisk PW contained 14 PAHs on the US EPA priority pollutant list (Table 1) with a total PAH content of 166 mg l1. Highly toxic PAHs, based on their toxic equivalency factors (TEF) were only present at low concentrations; levels of benzo[a]pyrene and dibenzo[a,h]anthracene, which have the highest TEF of 1.0 were present at 0.001 and 0.008 mg l1 respectively. PAH composition of Ekofisk PW is dominated by naphthalene and its C1-C3 alkyl homologues (Table 1), with these two-ringed PAHs accounting for 86% of the total PAH content. Despite the low TEF of naphthalene, its high levels make it the largest contributor to the concentration of BAP equivalents (BAPe) at 0.049 mg l1 (Table 1). Barium is the largest contributor to the metal content in Ekofisk produced water (Utvik, 1999), with an estimated 20 mg l1 in the 0.5% PW treatment (Table 2). In addition, cadmium, mercury, lead and nickel, have also been reported in PW from the Ekofisk field. However, the concentrations of these EU priority pollutants are estimated to be much lower, with levels ranging from <0.0025 mg l1 mercury to 0.965 mg l1 nickel in the 0.5% PW treatment (Table 2). Haemocyte numbers were significantly elevated following 7, 14 and 21 days PW exposure (F6,132 ¼ 6.40, P < 0.001), with cell counts demonstrating a general dose-dependent relationship with increasing PW concentrations (Fig. 1a). The maximum mean cell count of 27.1 105 cells ml1 was recorded in mussels exposed for 14 days to 0.5% PW, with more than a 100% increase in circulating haemocytes compared to the control mussels (Fig. 1a). PW exposure resulted in a general reduction in cellular viability in M. edulis with NRR in animals from the 0.5% treatment reduced to just 37% of the control values after 14 days (Fig. 1b). Cell viability indicated a significant interaction between treatment group and exposure time (F6,132 ¼ 8.34, P < 0.001) with 0.25% and 0.5% PW treatments significantly reducing NRR compared to the controls across all exposure periods (Fig. 1b). In contrast, mussels sampled on day 7 from the 0.125% PW treatment demonstrated significantly enhanced dye retention.
M.L. Hannam et al. / Environmental Pollution 157 (2009) 1939–1944 Table 1 PAH composition of Ekofisk produced water sampled from the header tank on day 6 of the exposure PAH
mg l1
TEF (Nisbet and LaGoy, 1992)
Naphthalene* C1-Naphthalene C2-Naphthalene C3-Naphthalene Fluorene* Phenanthrene* C1-Phen/anthr C2-Phen/anthr Acenaphthylene* Acenaphthene* Anthracene* Dibenzophiothene C1-Dibenzophiothene C2-Dibenzophiothene Chrysene* C1-Chrysene C2-Chrysene Fluoranthene* Pyrene* Benzo[a]anthracene* Benzo[b,j]fluoranthene* Benzo[k]fluoranthene* Benzo[b,j,k]fluoranthene Benzo[a]pyrene* Indeno[1,2,3-cd]pyrene* Benzo[g,h,i]perylene* Dibenzo[a,h]anthracene*
49.4 42.5 30.0 21.6 1.997 3.561 5.458 5.231 0.097 0.219 0.023 0.490 1.500 1.729 0.304 0.625 0.878 0.044 0.140 0.030 0.033 ND 0.032 0.001 ND 0.020 0.008
0.001 N/A N/A N/A 0.001 0.001 N/A N/A 0.001 0.001 0.01 N/A N/A N/A 0.01 N/A N/A 0.001 0.001 0.1 0.1 0.1 N/A 1 0.1 0.01 1
BAPe (mg l1) 0.049
0.002 0.004
<0.001 <0.001 <0.001
0.003
<0.001 <0.001 0.003 0.003 ND 0.001 ND <0.001 0.008
ND, not detected. Recorded concentration (mg l1), toxic equivalency factor (TEF) and benzo[a]pyrene equivalent concentrations (BAPe mg l1) are detailed. US EPA priority pollutants are indicated by an asterisk (*).
Phagocytic activity in haemocytes was affected by both PW concentration and exposure time with a significant interaction between these two factors (F6,132 ¼ 17.99, P < 0.001). Seven days exposure to 0.125% PW resulted in a 24% increase in phagocytic activity, with the ingestion of 12.0 107 zymosan particles mg1 protein compared to 9.7 107 particles ingested by M. edulis in the control treatment. However, after longer exposure periods, all tested PW concentrations reduced the phagocytic capability, with activity reduced to <50% of the control level after 14 and 21 days exposure to 0.25 and 0.5% PW (Fig. 1c). The cytotoxic capability of the haemolymph demonstrated a significant interaction between PW treatment and exposure period (F6,132 ¼ 6.54, P < 0.001). Seven days exposure to PW led to significant variations in observed cytotoxicity; 0.125% PW resulted in an increase in spontaneous cytotoxicity to 54%, significantly higher than the control (42%). In contrast cytotoxicity was significantly reduced following 0.25 and 0.5% PW treatments (Fig. 1d). After 14 and 21 days a significant dose-dependent reduction in cytotoxicity was observed following PW exposure.
Table 2 Estimated metal composition of the 0.5% produced water treatment, based on concentrations reported by Utvik (1999) for produced water from the Ekofisk field Metal
Concentration (mg l1)
Barium Lead Cadmium Iron Copper Mercury Nickel Zinc
20 0.002 0.002 12.5 0.05 <0.0025 0.965 <0.25
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4. Discussion There is considerable evidence suggesting that immune function in bivalves is vulnerable to modulation by environmental contaminants (Auffret et al., 2002; Brown et al., 2004; Coles et al., 1994; Dyrynda et al., 2000; Frouin et al., 2007; Pipe et al., 1999). The results obtained from this study indicate that PW exposure results in a measurable reduction in the immune fitness of M. edulis following laboratory exposure. Haemocytes are an important component of the non-specific immune system in M. edulis (Cheng, 1981). Variability in haemocyte number following sublethal exposure to PW is demonstrated by a significant increase in circulating cells at PW concentrations as low as 0.125%. Early effects of physiological alterations are often seen as changes in haemocyte counts, with elevated cell counts a common response to environmental stress (Auffret et al., 2006). Exposure to fluoranthene (Coles et al., 1994) and cadmium (Auffret et al., 2002; Coles et al., 1995) both present at low levels in PW (Tables 1 and 2), produce an increase in the number of circulating haemocytes in M. edulis. The significant reduction in haemocytes at the highest concentration (0.5% PW) may be a result of haemocyte lysis. Haemocytes contain a large number of lysosomes and it has been reported that PAHs can cause cytolysis in lysosome-enriched cells (McCormick-Ray, 1987), with haemocytes from Mercenaria mercenaria exhibiting such lysis as a result of phenol exposure (Fries and Tripp, 1980). Both PW concentration and exposure period were significant factors influencing haemocyte counts in M. edulis. This is concordant with other work that concluded crude oil concentration and exposure period significantly affect the number of circulating haemocytes in this bivalve (McCormick-Ray, 1987). Exposure of M. edulis to concentrations in excess of 0.25% PW resulted in inhibited NRR, reflecting reduced viability of the haemocytes. Lysosomes within the haemocytes of M. edulis sequester, accumulate and metabolise a range of organic xenobiotics and metals (Moore, 1985). However, this detoxification process is not effective if the storage capacity of the lysosomes is exceeded and this may lead to cell damage (Moore et al., 1984). Progressive overloading of lysosomal capacity could account for the continuing decrease in cell viability observed over the exposure period. Alternatively, contaminants can cause direct damage to cell membranes. The lipophilic nature of naphthalene and its metabolites means they can incorporate into biological membranes, impacting membrane fluidity and function (Vijayavel and Balasubramanian, 2006). Reactive oxygen species (ROS) are actively generated within M. edulis haemocytes and biotransformation reactions of PAHs such as naphthalene produce redox-cycling intermediates that proliferate ROS production (Winston et al., 1996). Elevated ROS formation can cause oxidative damage to cell membranes resulting in lipid peroxidation (Go´mez-Mendikute and Cajaraville, 2003). The consistent xenobiotic challenge posed by exposure to PW may cause oxidative stress that exceeds the normal antioxidant defences leading to membrane disturbance and reduced dye retention. Many other contaminants, including polychlorinated biphenyls, tributyltinoxide and cadmium, have also been reported to induce alterations in cell membranes leading to destabilisation and reduced cell viability (Galloway et al., 2002; Hagger et al., 2005; Werner et al., 2004). It should also be noted that a reduction in cellular viability was also observed in the control treatment over the 21 days, which may suggest a level of stress in organisms held for prolonged periods under laboratory exposure conditions. Phagocytosis by haemocytes is an essential cellular mechanism of internal defence in bivalves and contaminant-induced inhibition of this process may suppress immunocompetence within an organism (Nicholson, 2003). Results from the present study
M.L. Hannam et al. / Environmental Pollution 157 (2009) 1939–1944
a
35
THC (x105 ml-1)
30 25
7 day 14 day 21 day
+ +
*
+ +
*
b
+ +
*
20 15 10 5 0
0.125
0
0.25
OD550nm mg protein-1
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2.5 2.0
16 14
7 day 14 day 21 day
+ +
*
+ +
*
*
+ +
*
0.5
0
0.125
0.25
0.5
% produced water
d
+ +
80
* % cytotoxicity
zymosan particles phagocytosed (x107) mg protein-1
18
+ +
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% produced water
c
+ +
1.5
0.0
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7 day 14 day 21 day
12 10 8 6 4
60
7 day 14 day 21 day
+ +
*
+ +
*
+ +
*
40
20
2 0
0
0.125
0.25
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% produced water
0
0
0.125
0.25
0.5
% produced water
Fig. 1. (a) Total haemocyte counts (THC), (b) Neutral Red retention (optical density at 550 nm), (c) phagocytic activity and (d) spontaneous cytotoxicity in M. edulis exposed to produced water for 7, 14 and 21 days. Values displayed are means 1 SE. Statistically significant differences, at the 95% confidence level are shown; (z) indicates a significant difference from the 7 day control group (0% PW), () indicates a significant difference from 14 day control, and (*) indicates a significant difference from 21 day control.
indicate exposure of M. edulis to concentrations of 0.25% PW inhibit phagocytosis. This is concordant with previous work that has demonstrated depressed phagocytic responses in marine bivalves following exposure to oil emulsion (McCormick-Ray, 1987), PAHcontaminated sediments (Sami et al., 1992), anthracene, fluoranthene and phenanthrene (Grundy et al., 1996) and oilcontaminated field sites (Auffret et al., 2004). At the lowest tested concentration of PW, stimulation in phagocytic activity was observed after 7 days exposure. Low-dose phagocytic stimulation has also been observed in response to exposure to other contaminants including copper (Pipe et al., 1999), cadmium (Coles et al., 1995), zinc and mercury (Sauve´ et al., 2002), all present at low levels in PW (Table 2). However, stimulation of immune function following contaminant exposure often changes toward suppression at higher contaminant concentrations and increased exposure periods (Cheng and Sullivan, 1984). The PW dose response for phagocytosis suggests a possible concentration threshold, below which phagocytosis is stimulated and above which inhibition of phagocytosis occurs. Low-dose stimulation may involve high energetic costs to the organism (Pipe et al., 1999) that cannot be maintained for extended periods such as 14 and 21 days exposure to PW. Changes in proportions of haemocyte types may alter phagocytosis activity, with granulocytes reported to be the main phagocytic cells in many bivalves (Chang et al., 2005, Matozzo et al., 2007; Zhang et al., 2006). However previous studies have concluded that PAH exposure does not significantly alter the relative proportions of cell types (Coles et al., 1994). Phagocytic processes are dependent on membrane properties of the haemocytes therefore any alterations in the cell membrane may impact phagocytosis activity (Grundy et al., 1996). In the present study, inhibition of phagocytosis is coupled with a reduction in cell viability following PW
exposure, suggesting that membrane disruption may impact this essential component of invertebrate immunity. Phagocytosis in M. edulis has previously been used as an immunological endpoint in water column monitoring of PW (Sundt et al., unpublished). There was no clear relationship between phagocytic activity and proximity to the PW source, though the method used produced large within group variability. However, results did indicate a reduction in phagocytosis at sites with high levels of organic contamination. The absence of a correlation between phagocytic activity and distance from the PW discharge, despite the phagocytic reduction observed at sites of higher organic pollution, may reflect variations in the spatial distribution of PW around its source due to tidal and current influences. Changes in the cytotoxic capability of M. edulis haemolymph followed similar patterns recorded in cell viability and phagocytic response. There was biphasic low-dose stimulation, high-dose inhibition in spontaneous cytotoxicity after 7 days exposure, whilst longer exposure periods resulted in a concentration-dependent reduction in cytotoxicity. Spontaneous cytotoxicity response relies on direct cell-to-cell contact between mussel haemocytes and target cells, and upon release of lysing factors (Raftos and Hutchinson, 1995). As such, any disturbance in the cell membrane is likely to alter the cytotoxic reaction (Galloway et al., 2002). This may explain the similar response patterns of spontaneous cytotoxicity and cell viability, which also demonstrated an increased response at low PW concentrations during short exposure periods. A reduction in both cytotoxic ability and phagocytosis was observed following PW exposure of 0.25% and 0.5%, despite a significant increase in the number of total haemocytes. This may be a result of increased cell proliferation, producing immature cells with a reduced capacity to ingest particles or lyse target erythrocytes. Alternatively, an increase in cell counts may suggest stimulation of
M.L. Hannam et al. / Environmental Pollution 157 (2009) 1939–1944
cell migration from the tissues (Pipe et al., 1999), with inhibited phagocytic and cytotoxic activity a direct consequence of the reduced cellular viability. The changes in immune function observed in this study, indicate the potential of these biological endpoints for use as sensitive measures in biomonitoring programmes. However, since invertebrate immunity consists of a multifaceted defence system, the use of an integrated measurement of immunocompetence would increase the ecological relevance of immunotoxic observations. Auffret et al. (2004) proposed the use of an immunotoxicological index based on the cumulative variation of each parameter from a reference value. This requires the collection of species-specific baseline data over long-term surveys, something that was beyond the scope of this study. To date, an immunotoxic index has been used in several field studies with promising results (Auffret et al., 2004, 2006), and may be a useful component in biomonitoring programmes, where responses may be affected by numerous biotic and abiotic factors. 5. Conclusion The results from this study serve to demonstrate the modulating effects that PW has on the immune parameters of M. edulis. The biomarkers tested reveal that sufficiently high levels of PW cause immune modulation in numerous aspects of the cellular immune response. Concentration-dependent increases in haemocyte count were observed. Cell viability, phagocytosis and cytotoxicity were enhanced following short-term, low level PW exposure, whilst longer exposure periods and higher PW concentrations resulted in a move towards immune inhibition. It is important to take into account this type of biphasic response when interpreting immunological data and considering the use of immune endpoints in biomonitoring studies. Whilst PW contains a variety of substances that have been reported to have immunotoxic effects, such as PAHs and metals, their concentrations in the PW exposures tested here are lower than the concentrations reported in other studies to cause effects on immune function. It is therefore possible that the contaminants present in PW are exerting an additive or synergistic immunotoxic effect. The immune system of M. edulis is sensitive to modulation by environmental contaminants, with concentrations of PW, likely to be found close to the discharge point, resulting in a general reduction in the immune function of M. edulis. Environmental concentrations of PW around discharge points vary both spatially and temporally due to tidal and current influences and heterogeneous PW composition. Concentration and dispersion of PW is also site-specific with the Ekofisk field having a circumnavigational current around the discharge point, restricting movement of PW away from the rig. Therefore, PW could have consequences for the host resistance of exposed biota close to the discharge area. Acknowledgements This study was funded by Total E&P Norge AS and ConocoPhillips Norge through the Joint Industry Project ‘Produced Water Effects: Ekofisk Case Study’, University of Plymouth (UK) and IRIS-Akvamiljø (Norway). References Anderson, R.S., 1988. Effects of anthropogenic agents on bivalve cellular and humoral defense mechanisms. American Fisheries Society 18, 238–242. Auffret, M., Duchemin, M., Rousseau, S., Boutet, I., Tanguy, A., Moraga, D., Marhic, A., 2004. Monitoring of immunotoxic responses in oysters reared in areas contaminated by the ‘Erika’ oil spill. Aquatic Living Resources 17, 297–302.
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Immune modulation in the blue mussel Mytilus edulis exposed to North Sea produced water M.L. Hannam a, *, S.D. Bamber b, R.C. Sundt b, T.S. Galloway c a
Ecotoxicology & Stress Biology Research Centre, School of Biological Sciences, University of Plymouth, Drake Circus, Plymouth, Devon, PL4 8AA, UK IRIS – Biomiljø, Mekjarvik 12, 4070 Randaberg, Norway c School of Biosciences, University of Exeter, Hatherly Laboratories, Prince of Wales Road, Exeter, EX4 4PS, UK b
Exposure to produced water alters immune function in the sentinel species Mytilus edulis.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 24 October 2008 Received in revised form 12 December 2008 Accepted 18 December 2008
The discharge of oil well produced water (PW) provides a constant source of contaminants to the marine environment including polycyclic aromatic hydrocarbons, alkylated phenols, metals and production chemicals. High concentrations of PW cause adverse effects to exposed biota, including reduced survival, growth and reproduction. Here we explore the effects of PW on immune function in the blue mussel, Mytilus edulis. Mussels were exposed for 21 days to sublethal PW concentrations (0.125–0.5%) and cellular parameters were measured. Cell viability, phagocytosis and cytotoxicity were inhibited after exposure to 0.25% and 0.5% PW, whilst the 0.125% PW treatment produced significant increases in these biomarker responses. This biphasic response was only observed after 7 days exposure; longer exposure periods led to a reduction in immune parameters. Results indicate that PW concentrations close to the discharge point cause modulation to cellular immunity. The implications for longer-term disease resistance are discussed. Ó 2009 Elsevier Ltd. All rights reserved.
Keywords: Produced water Polycyclic aromatic hydrocarbons Mytilus edulis Immunotoxicity Phagocytosis
1. Introduction The marine environment is a major sink for many potentially hazardous pollutants discharged from industrial sources, such as produced water (PW). Produced water is formed as a by-product of oil extraction, with the amount of PW increasing as an oil field becomes depleted (Strømgren et al., 1995). Most PW has oil concentrations below the regulatory limit with PW discharges from North Sea installations recording an average oil content of 17 mg l1 in 2006 (OGP, 2006). In addition to oil, phenols, alkylphenols and metals also occur in PW along with chemicals (biocides, corrosion inhibitors, scale inhibitors, coagulants, detergents and surfactants) used in the process system (Neff, 2002). With 395 106 l day1 of PW entering the North Sea (OLF, 2006), these discharges represent a chronic source of pollution to marine systems with the potential to cause adverse effects to exposed biota (Neff, 2002). Adverse effects of PW may result from acute toxicity or from chronic sublethal alterations in homeostatic mechanisms, such as the immune system (Pipe et al., 1999). Immune systems of invertebrates do not operate in isolation, but are highly integrated with
* Corresponding author. Tel.: þ44 1752 232930; fax: þ44 1752 232970. E-mail address: [email protected] (M.L. Hannam). 0269-7491/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2008.12.031
other physiological systems making immune responses particularly sensitive to environmental contaminants (Galloway and Goven, 2006; Luengen et al., 2004). The immune system of invertebrates is relatively simple in contrast to that of vertebrates, providing accessible means for monitoring immune function (Galloway and Depledge, 2001). Unlike vertebrates, invertebrates lack immune specificity and memory, characteristic of the adaptive immunity observed in mammalian subjects (Anderson, 1988). Invertebrates have evolved innate defence mechanisms that can identify and protect against non-self material, with invertebrate immune response centred largely on the multifunctional haemocytes (Cheng, 1981; Galloway and Goven, 2006). The various parameters that contribute to invertebrate host defence act to provide a multifaceted immune response. It has therefore been suggested that an integrated suite of biomarkers be used for assessing immunocompetence of invertebrates (Livingstone et al., 2000; Pipe and Coles, 1995; Weeks et al., 1992). Advances in environmental risk assessment have highlighted the value of biomarkers to allow rapid evaluation of contaminant exposure and effect (Galloway et al., 2002, 2006). These biomarkers may be molecular, cellular or physiological endpoints that indicate exposure to and/or damage incurred by environmental pollutants providing early-warning indicators of contaminant-induced stress
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(Depledge and Fossi, 1994). The use of biomarkers has been recognised as a cost-effective technique in identifying the effects of pollutants on biota (Galloway et al., 2002). The majority of biomarker studies to date have been based on the mussel Mytilus edulis, and the American oyster Crassostrea virginica (Wootton et al., 2003). Bivalves exhibit worldwide distribution and sedentary, filter feeding behaviour. They have an open circulatory system that is continually exposed to fluctuating environmental conditions and pollutants (Pipe et al., 1999). In addition, a typical low metabolic transformation rate and a high bioaccumulation factor of 105 make bivalves an ideal biomonitoring organism, and as such they are now commonly used sentinels in ecotoxicology (Auffret et al., 2006; Galloway et al., 2002; Widdows and Donkin, 1992). In this study we tested the hypothesis that exposure of the blue mussel to produced water would cause modulation of immune function. 2. Method 2.1. Experimental design Mussels, M. edulis (5–7 cm) were collected from a reference site at Tysvær, Norway (59 190 400 N, 4 280 500 E), an area not subjected to large industrial or agricultural activity, and held in 600 l flow through (7.5 l min1) tanks containing filtered seawater at 8.5 C (1 C). Mussels were acclimated for 7 days prior to transfer to the exposure system. Test mussels were divided into eight exposure tanks (45 per 600 l tank with a flow rate of 7.5 l min1) maintained at 8.5 C (1 C) and a salinity of 34.5 PSU. PW was tapped on site from the Ekofisk oil field and stored at 20 C. PW treatments of 0.5, 0.25 and 0.125% and a seawater control were set up in duplicate tanks in a continuous flow system (CFS) fed from a header tank containing PW, described in detail by Sundt (2004). Immune response assays were carried out on 12 mussels from each treatment group after exposure periods of 0, 7, 14 and 21 days. 2.2. Chemical analysis Water from the PW header tank was sampled for PAH analysis on day 6 of the exposure. This was collected in 2 l amber bottles containing hydrochloric acid to ensure low acidity (pH < 2) and analysed for PAHs within 2 days of collection; the 16 PAHs on the US EPA priority pollutant list were measured along with alkyl homologues of naphthalene, chrysene, dibenzophiothene and phenanthrene/ anthracene. Eight deuterated PAHs were added as quantitative internal standards and mixed on a magnetic stirrer before triplicate extraction with cyclohexane (3 50 ml). Each extraction was performed by stirring the sample and solvent for 30 min on a magnetic stirrer and then pouring the entire bottle contents into a 2 l separating funnel. The water phase was drained back into the sampling flask and extracted a further two times. Combined organic extracts were dried with anhydrous NaSO4, filtered and concentrated to 1 ml using TurboVap 500. The samples were then transferred to vials and analysed by gas chromatography–mass spectrometry (GC-MS). PAHs were expressed as observed concentration in produced water (mg l1) and also toxic equivalents (mg l1). This latter case expresses individual PAHs as equivalents of benzo[a]pyrene, based upon their relative toxicity using toxic equivalency factors (TEFs) developed by Nisbet and LaGoy (1992). It should be noted that TEF values are based on the carcinogenic potential of PAHs in vertebrates and are not directly related to invertebrate models; it does, however, provide a method of comparison for the relative potential toxicities of PAHs present in PW. 2.3. Biological analyses Immune function assays were conducted in triplicate on haemolymph samples extracted into an equal volume of physiological saline (0.02 M HEPES, 0.4 M NaCl2, 0.1 M MgSO4, 0.01 M KCl, 0.01 M CaCl2; pH 7.4), from the posterior adductor muscle using a 21 gauge needle. Samples were transferred to siliconised EppendorfÒ tubes and stored on ice until analysis. Total haemocyte counts were carried out using an improved Neubauer haemocytometer. Haemocyte viability was assessed by measuring retention of Neutral Red (NR) dye (Babich and Borenfreund, 1992) as described in detail by Canty et al. (2007). Briefly, haemolymph samples were pipetted onto a microplate. After 45 min incubation (4 C), non-adhered cells were removed before adding aliquots of 0.004% NR solution. After 3 h incubation at 4 C excess NR solution was removed and acidified ethanol was added to resolubilise the dye and the optical density was measured at 540 nm. The optical density of Neutral Red retention was determined as a function of protein content, determined using the BCA protein assay (Pierce 23223/4).
Phagocytic activity of haemocytes was assessed by measuring the uptake of Neutral Red stained zymosan particles (from Saccharomyces cerevisiae) as described by Parry and Pipe (2004). Briefly, 50 ml haemolymph samples were pipetted onto a microplate, and adhered cells were incubated with 50 ml of dyed zymosan suspension (50 107 particles ml1) for 30 min at room temperature. Uptake of these dyed particles was determined spectrophotometrically at 550 nm against a standard curve prepared using zymosan suspensions (50–1.56 107 particles ml1) and expressed as a function of protein content. The ability of haemolymph to lyse sheep erythrocytes was used as a measurement of cytotoxic capability (Galloway et al., 2002). Haemolymph samples were diluted to a concentration of 1 104 cells ml1 in Tris-buffered saline (10 mM Tris– HCl, 150 mM NaCl, 10 mM CaCl2, pH 7.4). One millilitre of sheep erythrocytes (TCS Biosciences Ltd) was centrifuged at 200 g for 5 min before re-suspending in phosphate-buffered saline. The suspension was centrifuged again and a cell pack volume of 125 ml was re-suspended in 1 ml of Tris-buffered saline (TBS), this was then further diluted with 15 ml of TBS. Aliquots of diluted haemolymph (100 ml) were pipetted onto a round-bottomed microplate, along with duplicate controls of 100 ml TBS (for spontaneous release) and 100 ml TBS with 2 ml of 2% Triton-X100 (for positive control). One-hundred microlitres of the erythrocyte suspension was added and the plate was incubated at 25 C for 1 h. After centrifugation (200 g, 10 min) the supernatant was transferred to a flat-bottomed microplate, and the lysis determined through the optical density at 405 nm of the haemoglobin released into the supernatant. The cytotoxicity was expressed as percent lysis relative to the maximum release observed in the positive control. 2.4. Statistical analyses Biological endpoints were measured in 12 individuals from each treatment group at each time point, with results expressed as mean values 1 standard error. Data sets were analysed using two-way ANOVAs; tests were performed on each immune parameter to determine significant effects due to interactions (treatment time) or main factors (treatment and time). Post-hoc pairwise comparisons were conducted (Fisher’s LSD) to identify where significant differences occurred at the 95% confidence level (associated probability < 0.05).
3. Results Ekofisk PW contained 14 PAHs on the US EPA priority pollutant list (Table 1) with a total PAH content of 166 mg l1. Highly toxic PAHs, based on their toxic equivalency factors (TEF) were only present at low concentrations; levels of benzo[a]pyrene and dibenzo[a,h]anthracene, which have the highest TEF of 1.0 were present at 0.001 and 0.008 mg l1 respectively. PAH composition of Ekofisk PW is dominated by naphthalene and its C1-C3 alkyl homologues (Table 1), with these two-ringed PAHs accounting for 86% of the total PAH content. Despite the low TEF of naphthalene, its high levels make it the largest contributor to the concentration of BAP equivalents (BAPe) at 0.049 mg l1 (Table 1). Barium is the largest contributor to the metal content in Ekofisk produced water (Utvik, 1999), with an estimated 20 mg l1 in the 0.5% PW treatment (Table 2). In addition, cadmium, mercury, lead and nickel, have also been reported in PW from the Ekofisk field. However, the concentrations of these EU priority pollutants are estimated to be much lower, with levels ranging from <0.0025 mg l1 mercury to 0.965 mg l1 nickel in the 0.5% PW treatment (Table 2). Haemocyte numbers were significantly elevated following 7, 14 and 21 days PW exposure (F6,132 ¼ 6.40, P < 0.001), with cell counts demonstrating a general dose-dependent relationship with increasing PW concentrations (Fig. 1a). The maximum mean cell count of 27.1 105 cells ml1 was recorded in mussels exposed for 14 days to 0.5% PW, with more than a 100% increase in circulating haemocytes compared to the control mussels (Fig. 1a). PW exposure resulted in a general reduction in cellular viability in M. edulis with NRR in animals from the 0.5% treatment reduced to just 37% of the control values after 14 days (Fig. 1b). Cell viability indicated a significant interaction between treatment group and exposure time (F6,132 ¼ 8.34, P < 0.001) with 0.25% and 0.5% PW treatments significantly reducing NRR compared to the controls across all exposure periods (Fig. 1b). In contrast, mussels sampled on day 7 from the 0.125% PW treatment demonstrated significantly enhanced dye retention.
M.L. Hannam et al. / Environmental Pollution 157 (2009) 1939–1944 Table 1 PAH composition of Ekofisk produced water sampled from the header tank on day 6 of the exposure PAH
mg l1
TEF (Nisbet and LaGoy, 1992)
Naphthalene* C1-Naphthalene C2-Naphthalene C3-Naphthalene Fluorene* Phenanthrene* C1-Phen/anthr C2-Phen/anthr Acenaphthylene* Acenaphthene* Anthracene* Dibenzophiothene C1-Dibenzophiothene C2-Dibenzophiothene Chrysene* C1-Chrysene C2-Chrysene Fluoranthene* Pyrene* Benzo[a]anthracene* Benzo[b,j]fluoranthene* Benzo[k]fluoranthene* Benzo[b,j,k]fluoranthene Benzo[a]pyrene* Indeno[1,2,3-cd]pyrene* Benzo[g,h,i]perylene* Dibenzo[a,h]anthracene*
49.4 42.5 30.0 21.6 1.997 3.561 5.458 5.231 0.097 0.219 0.023 0.490 1.500 1.729 0.304 0.625 0.878 0.044 0.140 0.030 0.033 ND 0.032 0.001 ND 0.020 0.008
0.001 N/A N/A N/A 0.001 0.001 N/A N/A 0.001 0.001 0.01 N/A N/A N/A 0.01 N/A N/A 0.001 0.001 0.1 0.1 0.1 N/A 1 0.1 0.01 1
BAPe (mg l1) 0.049
0.002 0.004
<0.001 <0.001 <0.001
0.003
<0.001 <0.001 0.003 0.003 ND 0.001 ND <0.001 0.008
ND, not detected. Recorded concentration (mg l1), toxic equivalency factor (TEF) and benzo[a]pyrene equivalent concentrations (BAPe mg l1) are detailed. US EPA priority pollutants are indicated by an asterisk (*).
Phagocytic activity in haemocytes was affected by both PW concentration and exposure time with a significant interaction between these two factors (F6,132 ¼ 17.99, P < 0.001). Seven days exposure to 0.125% PW resulted in a 24% increase in phagocytic activity, with the ingestion of 12.0 107 zymosan particles mg1 protein compared to 9.7 107 particles ingested by M. edulis in the control treatment. However, after longer exposure periods, all tested PW concentrations reduced the phagocytic capability, with activity reduced to <50% of the control level after 14 and 21 days exposure to 0.25 and 0.5% PW (Fig. 1c). The cytotoxic capability of the haemolymph demonstrated a significant interaction between PW treatment and exposure period (F6,132 ¼ 6.54, P < 0.001). Seven days exposure to PW led to significant variations in observed cytotoxicity; 0.125% PW resulted in an increase in spontaneous cytotoxicity to 54%, significantly higher than the control (42%). In contrast cytotoxicity was significantly reduced following 0.25 and 0.5% PW treatments (Fig. 1d). After 14 and 21 days a significant dose-dependent reduction in cytotoxicity was observed following PW exposure.
Table 2 Estimated metal composition of the 0.5% produced water treatment, based on concentrations reported by Utvik (1999) for produced water from the Ekofisk field Metal
Concentration (mg l1)
Barium Lead Cadmium Iron Copper Mercury Nickel Zinc
20 0.002 0.002 12.5 0.05 <0.0025 0.965 <0.25
1941
4. Discussion There is considerable evidence suggesting that immune function in bivalves is vulnerable to modulation by environmental contaminants (Auffret et al., 2002; Brown et al., 2004; Coles et al., 1994; Dyrynda et al., 2000; Frouin et al., 2007; Pipe et al., 1999). The results obtained from this study indicate that PW exposure results in a measurable reduction in the immune fitness of M. edulis following laboratory exposure. Haemocytes are an important component of the non-specific immune system in M. edulis (Cheng, 1981). Variability in haemocyte number following sublethal exposure to PW is demonstrated by a significant increase in circulating cells at PW concentrations as low as 0.125%. Early effects of physiological alterations are often seen as changes in haemocyte counts, with elevated cell counts a common response to environmental stress (Auffret et al., 2006). Exposure to fluoranthene (Coles et al., 1994) and cadmium (Auffret et al., 2002; Coles et al., 1995) both present at low levels in PW (Tables 1 and 2), produce an increase in the number of circulating haemocytes in M. edulis. The significant reduction in haemocytes at the highest concentration (0.5% PW) may be a result of haemocyte lysis. Haemocytes contain a large number of lysosomes and it has been reported that PAHs can cause cytolysis in lysosome-enriched cells (McCormick-Ray, 1987), with haemocytes from Mercenaria mercenaria exhibiting such lysis as a result of phenol exposure (Fries and Tripp, 1980). Both PW concentration and exposure period were significant factors influencing haemocyte counts in M. edulis. This is concordant with other work that concluded crude oil concentration and exposure period significantly affect the number of circulating haemocytes in this bivalve (McCormick-Ray, 1987). Exposure of M. edulis to concentrations in excess of 0.25% PW resulted in inhibited NRR, reflecting reduced viability of the haemocytes. Lysosomes within the haemocytes of M. edulis sequester, accumulate and metabolise a range of organic xenobiotics and metals (Moore, 1985). However, this detoxification process is not effective if the storage capacity of the lysosomes is exceeded and this may lead to cell damage (Moore et al., 1984). Progressive overloading of lysosomal capacity could account for the continuing decrease in cell viability observed over the exposure period. Alternatively, contaminants can cause direct damage to cell membranes. The lipophilic nature of naphthalene and its metabolites means they can incorporate into biological membranes, impacting membrane fluidity and function (Vijayavel and Balasubramanian, 2006). Reactive oxygen species (ROS) are actively generated within M. edulis haemocytes and biotransformation reactions of PAHs such as naphthalene produce redox-cycling intermediates that proliferate ROS production (Winston et al., 1996). Elevated ROS formation can cause oxidative damage to cell membranes resulting in lipid peroxidation (Go´mez-Mendikute and Cajaraville, 2003). The consistent xenobiotic challenge posed by exposure to PW may cause oxidative stress that exceeds the normal antioxidant defences leading to membrane disturbance and reduced dye retention. Many other contaminants, including polychlorinated biphenyls, tributyltinoxide and cadmium, have also been reported to induce alterations in cell membranes leading to destabilisation and reduced cell viability (Galloway et al., 2002; Hagger et al., 2005; Werner et al., 2004). It should also be noted that a reduction in cellular viability was also observed in the control treatment over the 21 days, which may suggest a level of stress in organisms held for prolonged periods under laboratory exposure conditions. Phagocytosis by haemocytes is an essential cellular mechanism of internal defence in bivalves and contaminant-induced inhibition of this process may suppress immunocompetence within an organism (Nicholson, 2003). Results from the present study
M.L. Hannam et al. / Environmental Pollution 157 (2009) 1939–1944
a
35
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30 25
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+ +
*
+ +
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20 15 10 5 0
0.125
0
0.25
OD550nm mg protein-1
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2.5 2.0
16 14
7 day 14 day 21 day
+ +
*
+ +
*
*
+ +
*
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zymosan particles phagocytosed (x107) mg protein-1
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60
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40
20
2 0
0
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0
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% produced water
Fig. 1. (a) Total haemocyte counts (THC), (b) Neutral Red retention (optical density at 550 nm), (c) phagocytic activity and (d) spontaneous cytotoxicity in M. edulis exposed to produced water for 7, 14 and 21 days. Values displayed are means 1 SE. Statistically significant differences, at the 95% confidence level are shown; (z) indicates a significant difference from the 7 day control group (0% PW), () indicates a significant difference from 14 day control, and (*) indicates a significant difference from 21 day control.
indicate exposure of M. edulis to concentrations of 0.25% PW inhibit phagocytosis. This is concordant with previous work that has demonstrated depressed phagocytic responses in marine bivalves following exposure to oil emulsion (McCormick-Ray, 1987), PAHcontaminated sediments (Sami et al., 1992), anthracene, fluoranthene and phenanthrene (Grundy et al., 1996) and oilcontaminated field sites (Auffret et al., 2004). At the lowest tested concentration of PW, stimulation in phagocytic activity was observed after 7 days exposure. Low-dose phagocytic stimulation has also been observed in response to exposure to other contaminants including copper (Pipe et al., 1999), cadmium (Coles et al., 1995), zinc and mercury (Sauve´ et al., 2002), all present at low levels in PW (Table 2). However, stimulation of immune function following contaminant exposure often changes toward suppression at higher contaminant concentrations and increased exposure periods (Cheng and Sullivan, 1984). The PW dose response for phagocytosis suggests a possible concentration threshold, below which phagocytosis is stimulated and above which inhibition of phagocytosis occurs. Low-dose stimulation may involve high energetic costs to the organism (Pipe et al., 1999) that cannot be maintained for extended periods such as 14 and 21 days exposure to PW. Changes in proportions of haemocyte types may alter phagocytosis activity, with granulocytes reported to be the main phagocytic cells in many bivalves (Chang et al., 2005, Matozzo et al., 2007; Zhang et al., 2006). However previous studies have concluded that PAH exposure does not significantly alter the relative proportions of cell types (Coles et al., 1994). Phagocytic processes are dependent on membrane properties of the haemocytes therefore any alterations in the cell membrane may impact phagocytosis activity (Grundy et al., 1996). In the present study, inhibition of phagocytosis is coupled with a reduction in cell viability following PW
exposure, suggesting that membrane disruption may impact this essential component of invertebrate immunity. Phagocytosis in M. edulis has previously been used as an immunological endpoint in water column monitoring of PW (Sundt et al., unpublished). There was no clear relationship between phagocytic activity and proximity to the PW source, though the method used produced large within group variability. However, results did indicate a reduction in phagocytosis at sites with high levels of organic contamination. The absence of a correlation between phagocytic activity and distance from the PW discharge, despite the phagocytic reduction observed at sites of higher organic pollution, may reflect variations in the spatial distribution of PW around its source due to tidal and current influences. Changes in the cytotoxic capability of M. edulis haemolymph followed similar patterns recorded in cell viability and phagocytic response. There was biphasic low-dose stimulation, high-dose inhibition in spontaneous cytotoxicity after 7 days exposure, whilst longer exposure periods resulted in a concentration-dependent reduction in cytotoxicity. Spontaneous cytotoxicity response relies on direct cell-to-cell contact between mussel haemocytes and target cells, and upon release of lysing factors (Raftos and Hutchinson, 1995). As such, any disturbance in the cell membrane is likely to alter the cytotoxic reaction (Galloway et al., 2002). This may explain the similar response patterns of spontaneous cytotoxicity and cell viability, which also demonstrated an increased response at low PW concentrations during short exposure periods. A reduction in both cytotoxic ability and phagocytosis was observed following PW exposure of 0.25% and 0.5%, despite a significant increase in the number of total haemocytes. This may be a result of increased cell proliferation, producing immature cells with a reduced capacity to ingest particles or lyse target erythrocytes. Alternatively, an increase in cell counts may suggest stimulation of
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cell migration from the tissues (Pipe et al., 1999), with inhibited phagocytic and cytotoxic activity a direct consequence of the reduced cellular viability. The changes in immune function observed in this study, indicate the potential of these biological endpoints for use as sensitive measures in biomonitoring programmes. However, since invertebrate immunity consists of a multifaceted defence system, the use of an integrated measurement of immunocompetence would increase the ecological relevance of immunotoxic observations. Auffret et al. (2004) proposed the use of an immunotoxicological index based on the cumulative variation of each parameter from a reference value. This requires the collection of species-specific baseline data over long-term surveys, something that was beyond the scope of this study. To date, an immunotoxic index has been used in several field studies with promising results (Auffret et al., 2004, 2006), and may be a useful component in biomonitoring programmes, where responses may be affected by numerous biotic and abiotic factors. 5. Conclusion The results from this study serve to demonstrate the modulating effects that PW has on the immune parameters of M. edulis. The biomarkers tested reveal that sufficiently high levels of PW cause immune modulation in numerous aspects of the cellular immune response. Concentration-dependent increases in haemocyte count were observed. Cell viability, phagocytosis and cytotoxicity were enhanced following short-term, low level PW exposure, whilst longer exposure periods and higher PW concentrations resulted in a move towards immune inhibition. It is important to take into account this type of biphasic response when interpreting immunological data and considering the use of immune endpoints in biomonitoring studies. Whilst PW contains a variety of substances that have been reported to have immunotoxic effects, such as PAHs and metals, their concentrations in the PW exposures tested here are lower than the concentrations reported in other studies to cause effects on immune function. It is therefore possible that the contaminants present in PW are exerting an additive or synergistic immunotoxic effect. The immune system of M. edulis is sensitive to modulation by environmental contaminants, with concentrations of PW, likely to be found close to the discharge point, resulting in a general reduction in the immune function of M. edulis. Environmental concentrations of PW around discharge points vary both spatially and temporally due to tidal and current influences and heterogeneous PW composition. Concentration and dispersion of PW is also site-specific with the Ekofisk field having a circumnavigational current around the discharge point, restricting movement of PW away from the rig. Therefore, PW could have consequences for the host resistance of exposed biota close to the discharge area. Acknowledgements This study was funded by Total E&P Norge AS and ConocoPhillips Norge through the Joint Industry Project ‘Produced Water Effects: Ekofisk Case Study’, University of Plymouth (UK) and IRIS-Akvamiljø (Norway). References Anderson, R.S., 1988. Effects of anthropogenic agents on bivalve cellular and humoral defense mechanisms. American Fisheries Society 18, 238–242. Auffret, M., Duchemin, M., Rousseau, S., Boutet, I., Tanguy, A., Moraga, D., Marhic, A., 2004. Monitoring of immunotoxic responses in oysters reared in areas contaminated by the ‘Erika’ oil spill. Aquatic Living Resources 17, 297–302.
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