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Immune effects of HFO on European sea bass, Dicentrarchus labrax, and Pacificoyster, Crassostrea gigas

Anne Bado-Nilles a,b,�, Claire Quentel c, Michel Auffret d, Stephane Le Floch b, Beatrice Gagnaire e,Tristan Renault e, Helene Thomas-Guyon a

a LIENSs Littoral Environnement et Societes UMR 6250 CNRS Universite de La Rochelle, 2 rue Olympe de Gouges, 17 000 La Rochelle, Franceb Cedre, Centre de Documentation de Recherche et d’Experimentations sur les Pollutions Accidentelles des Eaux, 715 rue Alain Colas, CS 41836, 29 218 Brest Cedex 2, Francec Afssa site de Ploufragan-Plouzane, Agence Franc-aise de Securite Sanitaire des Aliments, Technopole Brest-Iroise, 29 280 Plouzane, Franced LEMAR-UMR CNRS 6539, Institut Universitaire Europeen de la Mer, Technopole Brest-Iroise, 29 280 Plouzane, Francee Ifremer La Tremblade, Laboratoire de Genetique et Pathologie (LGP), Ronce-les-Bains, 17 390 La Tremblade, France

a r t i c l e i n f o

Article history:

Received 21 July 2008

Received in revised form

31 March 2009

Accepted 2 April 2009Available online 29 April 2009

Keywords:

Dicentrarchus labrax

Crassostrea gigas

Heavy fuel oil

Immune parameters

Complement activity

Phenoloxidase activity

Bioaccumulation

a b s t r a c t

The European sea bass, Dicentrarchus labrax, and the Pacific oyster, Crassostrea gigas, were exposed to a

soluble fraction of heavy fuel oil for 5 and 9 days, respectively. The organisms were then transferred

to non-contaminated seawater for 1 month. The bioaccumulation and elimination of PAHs in

contaminated tissues were dissimilar between species. In fish, acenaphthene and naphthalene were

detected and naphthalene was still detectable 30 days after the beginning of the recovery period. In

oysters, on the other hand, pyrene and phenanthrene were bioaccumulated and 14 days after exposure

no more PAHs were detected. Concerning innate immune parameters, the increase of haemolytic

activity of the alternative complement pathway in fish and the reduction of phenoloxidase activity in

oysters endured, respectively, 1 and 2 weeks in contaminated organisms. This indicates that these two

enzymatic cascades could be quite useful for monitoring pollution by oil.

& 2009 Elsevier Inc. All rights reserved.

1. Introduction

Recently, many studies have been directed toward elucidatingthe relationship between environmental pollutants and theoccurrence of stress-related and disease conditions in aquaticanimals. The estuarine environment is a major source of potentialchemical pollutants emitted from human activities includingpolycyclic aromatic hydrocarbons (PAHs) and polychlorinatedbiphenyls (PCBs). The PAHs may be dissolved throughout thewater column, become adsorbed onto particles and/or enter intothe sediments where they may persist, only undergoing very slowdegradation (Meador et al., 1995). These pollutants are wellknown as environmental pollutants at low concentrations andappear in the United States Environmental Protection Agency(US-EPA, 1998) priority pollutant list due to their mutagenic andcarcinogenic properties. Moreover, PAHs are phototoxic (Shiaris,

1989), reprotoxic (Diamond et al., 1995) and immunotoxic(Yamaguchi et al., 1996).

Marine organisms are at risk of being continually exposed toxenobiotics including PAHs (Gagnaire et al., 2003). The vulner-ability of aquatic species to chemical pollution depends onpollutant properties, pollutant concentrations entering ecosys-tems and the capacity of ecosystems to defend themselvesespecially by biodegradation (Fochtman, 2000). However, bioac-cumulation constitutes a phenomenon, which may be theconsequence of direct contamination through water or indirectcontamination through the food chain (Amiard-Triquet, 1989).When bioaccumulation occurs, pollutants can directly interactwith tissues and cells such as immune cells.

For immunotoxicological risk assessment studies on mammals,innate immune responses, which was the first line of immunesystem defence of the organisms acting against pathogenwithout prior exposure to any particular microorganism, may besuppress by xenobiotics (Luster et al., 1988). Innate defencemechanisms may prevent infections by cellular reactions such asphagocytosis (Glinski and Jarosz, 1997). Intruders may also beeliminated by humoral components formed by the comple-ment system or prophenoloxidase/phenoloxidase (proPO/PO)system, lysozyme activity and lectins (Glinski and Jarosz, 1997).

ARTICLE IN PRESS

Contents lists available at ScienceDirect

journal homepage: www.elsevier.com/locate/ecoenv

Ecotoxicology and Environmental Safety

0147-6513/$ - see front matter & 2009 Elsevier Inc. All rights reserved.

doi:10.1016/j.ecoenv.2009.04.001

� Corresponding author at: LIENSs Littoral Environnement et Societes UMR 6250

CNRS Universite de La Rochelle, 2 rue Olympe de Gouges, 17 000 La Rochelle,

France. Fax: +33 298 449 138/+33 298 055165.

E-mail addresses: Anne.Bado.Nilles@cedre.fra,

a.bado-nilles@afssa.fr (A. Bado-Nilles).

Ecotoxicology and Environmental Safety 72 (2009) 1446–1454

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Xenobiotic-mediated suppression of these innate immune re-sponses seems to impact on pathogen resistance. Nevertheless,little is known about the mechanisms by which PAHs induceimmunotoxicity in organisms and more especially in fish(Reynaud and Deschaux, 2006) and marine bivalves (Coles et al.,1994).

The aims of the present work were (i) to develop an in vivo

system of experimental contamination by heavy fuel oil (HFO)representative of concentrations experienced after an oil spill and(ii) to carry out a preliminary validation of the experimentalsystem by studying pollutant effects on immune parameterssimultaneously in two marine species, the European sea bass,Dicentrarchus labrax, and the Pacific oyster, Crassostrea gigas. TheHFO used was similar to the oil that was released into the AtlanticFrench coastal waters, from Brittany to Charente-Maritime, duringthe Erika oil spill that occurred in December 1999. The capacityof organisms to recover their initial status after exposure to asoluble fraction of HFO was analysed over a 1-month period bymonitoring cell mortality, cell subpopulations (sea bass leuco-cytes and oyster haemocytes) and phagocytosis activity by flowcytometry. Some authors have described phagocytosis and cellsubpopulations as suitable biomarkers of pollution in differentfish (Carlson et al., 2004; Seeley and Weeks-Perkins, 1991) andbivalve species (Auffret, 2005; Bado-Nilles et al., 2008; Gagnaireet al., 2006). Concerning humoral innate immune parameters,haemolytic activity of the alternative complement pathway(ACH50) in sea bass, Dicentrarchus labrax (Bado-Nilles et al.,2009), and phenoloxidase (PO) activity in the Pacific oyster,Crassostrea gigas (Bado-Nilles et al., 2008; Bouilly et al., 2006),have been reported to be modified by xenobiotics. These twohumoral activities were therefore also performed.

2. Materials and methods

All experiments were conducted in accordance with the Commission

recommendation 2007/526/EC on revised guidelines for the accommodation and

care of animals used for experimental and other scientific purposes. Cedre is

authorised to conduct experimentation on animals in its capacity as a certified

establishment, according to the administrative order no. 2006-0429 dated 9 May

2006. Furthermore, the experimentation carried out as part of this study was

conducted under the responsibility and supervision of Dr. Stephane Le Floch, who

holds a certificate awarded by the National Veterinary School of Nantes entitling

him to direct scientific experimentation on animals.

2.1. Organisms

The 120 European sea bass, Dicentrarchus labrax, 144732 g, used for this

experiment were taken from one pond (aquaculture facility: Ecloserie Marines de

Gravelines) and were raised in an experimental facility at Afssa (Agence Franc-aise

de Securite Sanitaire des Aliments), Plouzane site (France).

Pacific oysters, Crassostrea gigas, 8–10 cm in shell length, were purchased from

a shellfish farm located in the Brest bay (Brittany, France).

European sea bass and Pacific oysters were acclimated for 2 weeks, together in

1200 L tanks at a density of 20 kg m�3 and with a flow of 0.3 m3 h�1 at 1271 1C

(dissolved oxygen 9674%, pH 7.670.4, salinity 3672%, free of nitrate and nitrite).

The sea bass were fed 150 g of granulates (Grower Extrude Natura 4 mm, Le

Gouessant Aquaculture) every day and the oysters with 15 L of Isochrysis galbana at

4�106 cell mL�1 every 3 days. This rate maintained a minimum concentration of

15�104 cell mL�1 per day, which was enough to ensure continuous feeding of the

oysters (Rico-Villa et al., 2006).

At the end of the acclimation period, a small notch was carved in the dorsal

shell of the oysters. Normally, oysters close their valves during short periods of

water contamination period. Carving a notch therefore affects the environmental

relativity of this study by favouring direct contact between oyster tissues and

pollutants.

2.2. Pollutants

A HFO, similar to that of the Erika, which contained 24–25% of saturated

hydrocarbons, 54–55% of aromatic hydrocarbons and 20–21% of polar compounds,

was selected to perform the exposure. The HFO contained, among others, the 16

PAHs of the United States Environmental Protection Agency list (US-EPA, 1998),

which are benzo[a]anthracene, naphtalene, acenaphtylene, acenaphtene, fluorene,

phenanthrene, anthracene, fluoranthene, pyrene, chrysene, benzo[b]fluoranthene,

benzo[k]fluoranthene, benzo[a]pyrene, indeno[1,2,3-c,d]pyrene, dibenz[a,h]anthra-

cene and benzo[g,h,i]perylene at concentrations ranging from 1936.97116.6mg g�1

for phenanthrene to 27.675.7mg g�1 for dibenz[a,h]anthracene (Table 1).

2.3. Experimental arrangement

2.3.1. Experimental system

The experimental system was constituted of four tanks, one control tank

(Fig. 1A), one exposed tank (Fig. 1B) and two tanks for recovery and acclimation

period (Fig. 1C). Each tank had a volume of 1200 L and was 120 cm deep. The tanks

were placed in a greenhouse, which is thermoregulated (T ¼ 1471 1C) and has a

total renewal of the air in 6 h.

For the four different tanks, an air stone and a degassing column were used to

maintain a level of dissolved oxygen of around 9674%. The degassing column was

constituted of a 2 m high pipe, with a diameter of 20 cm, which was filled with

40 kg of glass beads (diameter of 6 mm). In addition, the flow of seawater

(1.5 m3 h�1), which this column ensured the elimination of the dissolved carbon

dioxide (Fig. 1).

For the exposure period, a mixing tank was connected to the exposure tank

in order to generate a closed circulation of seawater with a flow of 0.5 m3 h�1

(Fig. 1B). At the water surface of the mixing tank, 3 L of HFO were released slowly.

This equipment allows seawater contaminated by only the soluble fraction of the

HFO in the exposure tank to be obtained. This system was adapted from Anderson

et al. (1974) and modified by Cedre to obtain a stable oil concentration in the

exposure tank throughout the experimentation (7337111 ng L�1). After 2 weeks of

mixture of HFO and seawater, the organisms were placed in exposure tanks.

2.3.2. Exposure conditions

After the acclimation period, 60 sea bass and 60 oysters were placed in the

control tank (Fig. 1A), while 60 other sea bass and 60 other oysters were exposed

(Fig. 1B) with the natural light/dark cycle (12 h/12 h). The fish were exposed for

5 days, due to their constant contact with pollutants by their skin, and the oysters

for 9 days, due to their capacity to close their shell. Control groups received clean

seawater during the same period. During this exposure period, control and

exposed organisms were not fed to prevent potential bioaccumulation of PAHs by

food. The seawater quality was monitored in both tanks (oxygen level,

temperature, pH, salinity, nitrate and nitrite) and, also, the concentration of the

HFO soluble fraction.

2.3.3. Recovery period

After the exposure period, organisms of each tank were transferred into two

other tanks, which were similar to the acclimatized tank (Fig. 1C), with an inlet of

clean seawater with a flow rate of 0.3 m3 h�1. The organisms were put back in the

same conditions as during the acclimation period. During the 30-day recovery

period, the organisms were fed again in the same conditions as during the

acclimatized period.

ARTICLE IN PRESS

Table 1Concentration of 16 priority PAHs of the US EPA list in the heavy fuel oil used

during the exposure period.

Name of PAH compounds Molecular weight (g mol�1) Concentration

(mg g�17SE)

Naphthalene 128.2 687789

Acenaphthylene 152.2 5173

Acenaphthene 154.2 273714

Fluorene 166.2 396720

Phenanthrene 178.2 19377116

Anthracene 178.2 214728

Fluoranthene 202.3 125714

Pyrene 202.3 516757

Benz[a]anthacene 228.3 213713

Chrysene 228.3 465714

Benzo[b+k]fluoranthene 252.3 81711

Benzo[a]pyrene 252.3 16877

Benzo[g,h,i]perylene 276.3 4873

Indeno [1,2,3-c,d] pyrene 276.3 1773

Dibenz[a,h]anthracene 278.4 2875

PAH detection was performed by gas chromatography coupled with mass

spectroscopy (GC–MS). The results are expressed in mg g�1 (n ¼ 3, mean7standard

error).

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2.4. Sampling and sample preparation

2.4.1. Seawater

One litre of seawater was collected in triplicates in heat-sterilised (500 1C)

Duran glass bottles (Bioblock) on days (-8), (-4) and (-2) after the beginning of the

exposure period to characterise the seawater quality.

2.4.2. Sea bass

The fish were anaesthetised with phenoxy-2-ethanol. Peripheral blood was

then collected on the first day of the recovery period, immediately after the

disconnection of the mixing tank (day 0), and on days 1, 3, 9, 14 and 30 of the

recovery period. For each fish, 2 mL of blood was withdrawn from the caudal vein

with a lithium heparinized vacutainer (BD VacutainerTM LH 85 U.I.). The fish were

then killed, with an overdose of anaesthetic, and weighed. The blood and fish were

kept on ice until they were processed.

For flow cytometry analysis, 0.5 mL of blood collected from 10 contaminated

and 10 control fish (n ¼ 10) were immediately diluted with 10 mL of Leibovitz 15

medium (L15, Eurobio) containing 10 U heparin lithium (Sigma). Samples were

then loaded onto Ficoll gradient (Histopaques1077, Eurobio) to a density of

1.07–1.08 g cm�3. After centrifugation (400 g, 30 min, 15 1C), mononuclear cells at

the interface were collected and were washed twice (400 g, 5 min, 4 1C) with L15.

Finally, cells were resuspended in 1 mL of L15 medium.

For spectrophotometry analysis, remaining blood samples were centrifuged

(1200 g, 10 min, 4 1C). After centrifugation, the plasma of each fish were split into

two aliquots of 100mL and stored at �80 1C for further analysis.

After blood collection, about 20 g of muscle of 10 control and 10 contaminated

fish was collected and frozen at �80 1C until analysis.

2.4.3. Oyster

Haemolymph was collected on the first day of the recovery period,

immediately after the disconnection of the mixing tank (day 0), and on days 3,

14 and 30 post-exposure.

After opening the oyster shell by cutting off the adductor muscle, approxi-

mately 2 mL of haemolymph was withdrawn from the pericardial cavity using a

2 mL syringe equipped with a 0.7�30 mm needle. The haemolymph samples were

kept on ice until they were processed to reduce haemocyte aggregation (Auffret

and Oubella, 1997).

As for cytometry, 1 mL of haemolymph was collected individually from five

contaminated oysters and five control oysters (n ¼ 5). All samples were treated

immediately and separately to study cellular activity.

As for spectrophotometry, three pooled samples, each composed of haemo-

lymph from five oysters, were formed to obtain a sufficient volume of

haemolymph. The three pooled haemolymph samples were centrifuged (260g,

10 min, 4 1C). The acellular fraction (supernatant) was frozen at �80 1C for further

analysis.

2.5. Analytical methods

2.5.1. Seawater PAH concentrations

For the seawater PAH content, samples were extracted with 30 mL of

dichloromethane pestipur quality (Carlo Erba Reactifs, SDS). After separation of

the organic and aqueous phases, water was extracted two additional times with

the same volume of dichloromethane (2�30 mL). The combined extracts were

purified and treated using gas chromatography coupled with mass spectrometry

(GC–MS, Hewlett Packard HP5890 coupled with an HP5972 mass selective

detector) following published procedures (Douglas et al., 1992). Finally, the

16 US-EPA PAHs were quantified.

2.5.2. PAH concentrations

PAH levels in fish muscles were determined by GC–MS using the procedure of

Baumard et al. (1997) with some modifications. Prior to extraction, each muscle

sample was homogenized using an Ultraturax (Janke & Kunkel, IKAs-Labortech-

nik). One hundred and fifty mL of perdeuterated internal standards (CUS-7249,

Ultra Scientific, Analytical solutions) were added to 16 g of homogenized muscle

and samples were digested for 4 h under reflux in 50 mL of an ethanolic solution of

potassium hydroxide (2 mol L�1, Fisher Chemicals). After cooling, settling and

addition of 20 mL of demineralised water, the digest was extracted in a 250 mL

funnel two times with 20 mL of pentane (Carlo Erba Reactifs, SDS). The extract was

evaporated with a Turbo Vap 500 concentrator (Zyman, Hopkinton, MA, USA, at

880 mbar and 50 1C) to obtain 1 mL of concentrated extract. The purification of the

extract was performed by transfer to a silica column (5 g of silica). Hydrocarbons

ARTICLE IN PRESS

1.5 m3.h-1

Control tank

Degassing column

Mixing tank

Degassing column

Exposure tank

1.5 m3.h-1

1.5 m3.h-1

Degassing column

Pump

Fresh seawater inlet (0.3 m3.h-1

Outlet (0.3 m3.h-1

0.5 m3.h-1

Fig. 1. Experimental tanks for control (A) and contaminated (B) organisms as a closed circuit and during the recovery period (C) as an open circuit.

A. Bado-Nilles et al. / Ecotoxicology and Environmental Safety 72 (2009) 1446–14541448

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were eluted with 50 mL of pentane:dichloromethane (80:20, v:v, SDS) and

concentrated to 200mL by means of a Turbo Vap 500 concentrator (Zyman,

880 mbar, 50 1C). Aromatic compounds were analysed by GC–MS, with a detection

limit of 5mg kg�1of dry weight, and PAHs were quantified relative to the

perdeuterated internal standards introduced at the beginning of the sample

preparation procedure.

Whole tissues of 15 control and 15 contaminated oysters were collected,

pooled and frozen at �80 1C until further analysis. The PAHs levels were

determined by GC–MS (Laboratoire Municipal de Rouen, ETSA, Rouen, France)

(Munschy et al., 2005).

2.5.3. Blood leucocyte and haemocyte analysis by flow cytometry

Cell enumeration was performed with a Thoma’s cell haemocytometer and

adjusted at 106 cells mL�1 with L15 medium for leucocytes and with Tris buffer

saline (TBS, 1000 mOsm L�1) for haemocytes. Morphological characteristics, cell

mortality and the phagocytosis percentage were analysed with a Facscalibur flow

cytometer (Becton Dickinson) using the previously described protocols (Gagnaire

et al., 2003). For each cell sample, 10,000 events were counted. Analyses were

carried out on whole immune cells without distinguishing subpopulations, and the

results were expressed as a percentage of positive cells.

The cellular subpopulation percentage and structure (size and complexity)

were analysed using forward light scatter (FSC) and side light scatter (SSC)

photodetectors, respectively, which enable differentiation between cell types and

the establishment of the proportion in fish of lymphocytes and monocytes-

granulocytes and in oysters of hyalinocytes and granulocytes. These parameters

were monitored using 200mL of cell suspension without previous treatment.

Cell mortality was measured using FL3 (red fluorescence). Propidium iodide

(PI, 1.0 g L�1, Molecular Probes) is membrane impermeant and is excluded from

viable cells. Mortality was determined using 200mL of haemocyte suspension and

10mL of PI. Cell suspensions were incubated for 30 min at 4 1C.

The phagocytosis percentage was determined using FL1 (green fluorescence).

Fluorescent microspheres (2.7�1010 particles mL�1, Fluorospheress carboxylate-

modified microspheres, diameter 1mm, Molecular Probes) were used and the

fluorescence setting was established using a suspension of fluorescent beads in

distilled water. Only the events showing a fluorescence of at least three beads were

considered positive for phagocytic activity. Phagocytic activity of haemocyte

suspensions was analysed on 200mL of haemolymph samples and 10mL of a 1/10

dilution of fluorescent beads. Cell suspensions were incubated for 1 h at room

temperature.

2.5.4. Plasma and haemolymph analysis by spectrophotometry

2.5.4.1. Fish analysis. Determination of the alternative pathway of plasma

complement activity was carried out by haemolytic assay with rabbit red blood

cells (RRC, Biomerieux) as described by Yano (1992) and adapted to microtitration

plates. Sea bass samples, diluted to 1/64 in EGTA-Mg-GVB buffer to avoid natural

haemolytic activity, were added in increasing amounts, from 10 to 100mL wells�1.

The wells were then filled with EGTA-Mg-GVB buffer to a final volume of 100mL.

Fifty mL of 2% RRC (Biomerieux) suspension was finally added to all wells. Control

values of 0% and 100% haemolysis were obtained using: 100mL of EGTA-Mg-GVB

buffer and 100mL of non-decomplemented trout haemolytic serum at 1/50 in

ultrapure water, respectively. Samples were incubated for 1 h at 20 1C. The

microplates were centrifuged (400 g, 5 min, 4 1C, Jouan). Then, 75mL of supernatant

from each well was transferred with 75mL of phosphate buffer saline (PBS,

Biomerieux) into another 96-well microplate. The absorbance (A540) was read in a

Labsystems’iEMS analyser and the number of ACH50 units per mL of plasma was

determined by reference to the 50% haemolysis.

2.5.4.2. Oyster analysis. Haemolymph samples were centrifuged (260g, 10 min,

4 1C) and supernatants recovered. PO activity in acellular fraction samples was

detected by the measurement of L-3,4-dihydroxyphenylalanine (L-Dopa, Sigma)

transformation in dopachromes as previously described by Gagnaire et al. (2004).

Samples were distributed in 96-well microplates (Nunc, France). PO modulators

were used to confirm the specificity of the detection. The purified trypsin TPCK

(N-Tosyl-L-phenylalanine chloromethyl ketone, 1 g L�1, Sigma) was used as an

activator and the b-2-mercaptoethanol (10 mM, Sigma) was used as an inhibitor.

To determine the PO activity, 80mL of cacodylate buffer (CAC buffer: sodium

cacodylate (10 mM), trisodium citrate (100 mM), NaCl (0.45 M), CaCl2 (10 mM),

MgCl2 (26 mM), pH 7.0), 20mL of L-Dopa (3 mg mL�1) and 20mL of samples were

added to each well. To measure the PO activity modulation, 60mL of CAC buffer,

20mL of PO modulators, 20mL of L-Dopa and 20mL of samples were added to each

well. Control (120mL of CAC buffer) and negative control (100mL of CAC buffer,

20mL of L-Dopa) wells were used to determine respectively the purity of the buffer

and the autooxidation capacities of L-Dopa. Each sample was tested in nine

replicates and absorbance was measured at 490 nm after a 21 h incubation period

at room temperature.

Total protein concentration was measured using the Bradford method (Micro

BCA Protein Assay Kit, Pierce, Rockford, IL). The bovine albumin serum (BSA, from

0.0 to 1.0 g L�1, Pierce, Rockford, IL) was used as a standard. Ten mL of samples or of

the standard were distributed in 96-well microplates with 90mL of milliQ water

and 100mL of reagent kit. The microplates were then incubated for 2 h at 37 1C and

the absorbance was read at 570 nm (A570). Specific PO activity (g L�1) was

determined with PO activity (A490) per total protein concentration (A570).

Enzymatic activity was analysed in triplicates.

2.5. Statistical analysis

Statistical tests were carried out using XLStat Pro 7.5.3. Verification of

normality and of homogeneity of covariance matrices (homescedasticity) were

conducted using respectively the Anderson–Darling test and the Bartlett test. For

normal values and homogeneous variance, an F-test was applied to analyse HFO

effects. p-Values lower than 0.05 were used to identify significant differences.

3. Results

No sea bass or oyster mortality was observed during the courseof the experiment. The physico-chemical parameters of theseawater were the same as during the acclimation periodthroughout the entire experimentation (dissolved oxygen9674%, pH 7.670.4, salinity 3672%, temperature 1271 1C, freeof nitrate and nitrite).

ARTICLE IN PRESS

Table 2Concentration of 16 US-EPA PAHs in contaminated tanks throughout the in vivo experiment.

Name of PAH compounds Concentration at day (-8) Concentration at day (-4) Concentration at day (-2) Mean concentration (ng L�17SE)

Naphthalene 88 60 48 66712

Acenaphthylene 4 18 2 875

Acenaphthene 15 22 18 1972

Fluorene 7 7 7 770

Phenanthrene 482 300 614 466791

Anthracene 37 33 31 3472

Fluoranthene 37 48 64 5078

Pyrene 4 5 7 571

Chrysene 7 9 10 971

Benzo[a]anthracene 65 32 13 37715

Benzo[b]fluoranthene 7 15 32 1877

Benzo[k]fluoranthene 5 11 23 1375

Benzo[a]pyrene 4 4 5 571

Indeno[1,2,3-c,d]pyrene n.d. n.d. n.d. n.d.

Benzo[g,h,i]perylene n.d. n.d. n.d. n.d.

Dibenz[a,h]anthracene n.d. n.d. n.d. n.d.

Sum of the 16 PAHs 762 564 874 7337157

The results are expressed in ng L�1 in seawater (n ¼ mean values establish from three analysis at day (-8), day (-4) and day (-2) after the beginning of the exposure period,

mean7standard error, n.d. ¼ not detected).

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3.1. PAH concentrations in seawater

In the control tank, PAHs were not detected. In the exposuretank, 13 out of 16 PAHs were detected with a total meanconcentration of 7337111 ng L�1 and with mean concentrationsranging from 571 ng L�1 for benzo[a]pyrene to 466791 ng L�1 forphenanthrene. Concerning the diaromatic compounds, theirconcentrations decreased during the exposure period (e.g.naphthalene, from 88 to 48 ng L�1). For compounds with morebenzenic cycles, kinetics of dissolution were inverted: on day (-8)concentrations were globally lower than on day (-2) (e.g.benzo[b]fluoranthene, from 7 to 32 ng L�1). By the end, thethree heaviest (indeno[1,2,3-c,d]pyrene, dibenz[a,h]anthraceneand benzo[g,h,i]perylene) were undetected (Table 2).

3.2. PAH concentrations in fish muscles and oyster tissues

PAHs were neither detected in the muscles of control fish norin the tissues of the control oyster (lower than 5mg kg�1ofdry weight, which corresponds to the detection limit of GC–MS,Table 3).

For the fish, after 5 days of exposure, two PAHs were detectedin the muscles of contaminated sea bass, with 56.8mg kg�1 of dryweight composed of about 79.6% of naphthalene and 20.4% ofacenaphthene. After 1 day of recovery period, the fish hadeliminated 4.4% of naphthalene and 21.6% of acenaphthene. Byday 3, 34.3% of naphthalene and 29.3% of acenaphthene had beeneliminated. By day 9, 39.8% of naphthalene and 55.2% of

acenaphthene had been eliminated. From the 14th day of therecovery period, only naphthalene was detected in fish musclewith 6.4mg kg�1 of dry weight (85.5% elimination) on day 14 and5.7mg kg�1 of dry weight (87.4% elimination) on day 30.

Concerning oysters, after 9 days of exposure, the contaminatedoyster tissues had bioaccumulated only two PAHs, with 22.8mgkg�1 of wet weight composed of about 25.5% of phenanthrene and74.5% of pyrene. By day 3 post-contamination, the oysters hadeliminated 100% of the phenanthrene and 55.3% of the pyrenebioaccumulated in their tissues. By the 14th day of the recoveryperiod, PAHs were no longer detected in oyster tissues.

3.3. PAH effects on cellular parameters

Concerning fish results, leucocyte subpopulations were repre-sented by mean values of 14.673.9% for granulocytes–monocytesand 70.776.4% for lymphocytes. Leucocyte subpopulation per-centages and phagocytic activity were not significantly differentin control and contaminated fish at the end of the contaminationperiod and during the recovery period. Nevertheless, cellmortality had significantly increased in the contaminated fish(1.7%) in comparison to the control fish (0.8%) before transfer intorecovery tanks (Table 4).

In oysters, haemocyte subpopulations were represented bymean values of 33.277.5% of granulocytes and 66.877.5% ofhyalinocytes. Haemocyte subpopulation percentages and cellmortality were not significantly different in the control andcontaminated oysters during the recovery period. By the end of

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Table 3Concentration of the 16 US-EPA PAHs in sea bass muscles and oyster tissues.

Sampling dates Fish muscles (mg kg�1 of dry weight) Oyster tissues (mg kg�1 of dry weight)

Control Contaminated Control Contaminated

D0 n.d. 56.8 (naphthalene: 45.2, acenaphthene: 11.6) n.d. 22.8 (phenanthrene: 5.8, pyrene: 17.00)

D1 n.d. 52.3 (naphthalene: 43.2, acenaphthene: 9.1) – –

D3 n.d. 37.9 (naphthalene: 29.7, acenaphthene: 8.2) n.d. 7.6 (pyrene)

D9 n.d. 32.4 (naphthalene: 27.2, acenaphthene: 5.2) – –

D14 n.d. 6.4 (naphthalene: 6.4) n.d. n.d.

D30 n.d. 5.7 (naphthalene: 5.7) n.d. n.d.

The results are expressed in mg kg�1 of dry weight (n ¼ 10 for fish and n ¼ 15 for oysters, n.d. ¼ not detected). The experiment was performed by gas chromatography

coupled with mass spectroscopy (GC–MS).

Table 4Cellular subpopulation, cell mortality and phagocytosis percentage monitored by flow cytometry after in vivo exposure of sea bass and oysters to heavy fuel oil (HFO).

Values7SE Experimental conditions

Recovery period

Day 0 Day 1 Day 3 Day 9 Day 14 Day 30

Control HFO Control HFO Control HFO Control HFO Control HFO Control HFO

European sea bass

Lymphocytes (%) 72.279.1 74.973.1 72.874.2 65.272.7 71.976.1 64.478.2 75.177.4 66.975.6 77.776.1 78.575.3 71.579.3 57.179.9

Monocytes/granulocytes (%) 15.474.9 14.071.2 19.673.7 23.872.9 18.274.8 20.975.0 16.275.6 20.373.1 12.873.6 13.473.9 16.773.9 22.374.4

Mortality (%) 0.870.2 1.770.3* 1.170.3 0.570.1 0.770.3 0.770.2 1.670.6 1.270.3 0.770.1 1.170.2 0.770.2 0.770.2

Phagocytosis (%) 35.377.9 26.976.5 44.179.8 36.9715.4 27.3714.0 26.479.9 16.973.9 12.772.0 3.570.5 3.371.0 20.678.7 11.375.9

Pacific oyster

Granulocytes (%) 47.276.0 37.872.1 – – 38.374.9 31.274.3 – – 32.272.8 26.571.4 26.672.7 25.474.0

Hyalinocytes (%) 52.876.0 62.272.1 – – 61.774.9 68.874.3 – – 67.872.8 73.571.4 73.472.7 74.674.0

Mortality (%) 10.971.2 11.771.6 – – 4.970.9 7.870.9 – – 6.671.6 6.171.1 5.771.0 6.572.3

Phagocytosis (%) 35.674.2 24.570.5* – – 27.572.9 23.772.5 – – 24.874.0 23.175.7 23.674.4 28.573.0

Days 0, 1, 3, 9, 14 and 30 concerned the recovery period. No effect was detected during the entire recovery period and for all values (n ¼ 10 for fish and n ¼ 15 for oysters,

mean7standard error). * ¼ statistical difference for pp0.05.

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the contamination period, the phagocytosis activity had signifi-cantly decreased in the contaminated oysters (24%) comparedwith controls (36%) (Table 4).

3.4. PAH effects on extracellular parameters

In fish, the ACH50 significantly increased in contaminated fishcompared with controls with 40.772.5 ACH50 units mL�1 forcontaminated fish and 31.972.2 ACH50 units mL�1 for control fishon day 3 and with 36.172.9 ACH50 units mL�1 for contaminatedfish and 27.073.0 ACH50 units mL�1 for control fish on day 9. Ateach other sampling date of the recovery period, no significantdifference was observed between the contaminated fish and thecontrol fish (Fig. 2).

Specific PO activity in oysters significantly decreased incontaminated oysters compared with controls from days 0 to 14of the recovery period. This difference decreased from 0.9 on day0, to 0.8 on day 3 and to 0.7 on day 14, and this difference was nolonger significant by day 30 (o0.1) (Fig. 2).

4. Discussion

4.1. Chemical analysis: validation of the experimental system

Pollutant detection and exposure in soluble fractions dependon the dissolution technique used. In this study, the Andersonmethod (Anderson et al., 1974) was chosen and modified by Cedre

to expose organisms only to the soluble fraction of HFO. Duringthe entire exposure period of the experiment, the overallconcentration of the soluble fraction of HFO (733 ng L�1) remainedconstant, validating the experimental system, which moreoverensures a homogenous exposure in the whole tank. Thisconcentration was within the classic range often observed afteran oil spill. In fact, much wider ranges of concentrations aredescribed in the literature covering the concentrations of PAHs inseawater after different oil spills. After the Exxon Valdez wreckage,Boehm et al. (2007) reported seawater PAH concentrationsranging up to 600 ng L�1. Law (1978) also reported high concen-trations of PAHs in seawater ranging up to 1700 ng L�1 after theEkofisk accident. Thirteen out of 16 PAHs present in the HFO weredetected in the soluble fraction, obtained after the HFO has beenin contact with seawater for 2 weeks. No detectable concentrationof indeno(1,2,3-c,d)pyrene, benzo(g,h,i)perylene and dibenz(a,h)anthracene was found. The absence of these three compounds

may be explained by their low content in HFO and their highmolecular weight that makes them less soluble. The low solubilityof pyrene (202.25 g mol�1), related to its high molecular weight,probably explains the small quantity of pyrene (5.0 ng L�1) inseawater, in spite of its high concentration in HFO (516mg g�1). Onthe contrary, the high concentration of phenanthrene in the HFO(1936mg g�1), and its low molecular weight (178.23 g mol�1),explain its high concentration in seawater. So, the composition inPAHs of the soluble fraction depends on the initial concentrationof the different elements in HFO, of their molecular weight and oftheir number of benzenic cycles, which determine also theirsolubilisation kinetics (Meador et al., 1995). Moreover, seawatercontamination by HFO is an equilibrium between all compoundsand their kinetics of dissolution, which explains why the totalmean concentration of PAHs stays stable during the exposureperiod. These results, close to those observed during thebehaviour of an oil slick at sea, validate the experimental system.Thus, organisms contaminated by the soluble fraction of HFOwere exposed to the 13 PAHs during different lab times, 5 days forfish and 9 days for oysters.

4.2. Chemical analysis: PAH bioaccumulation

In the present study, the contamination period induced PAHbioaccumulation in fish muscles (naphthalene; acenaphthene)and in oyster tissues (pyrene; phenanthrene) confirming theefficacy of the experimental system. Several factors may be at theorigin of this difference between the uptake and accumulation ofPAHs in fish and oysters: (i) physical and chemical properties ofPAHs (i.e. molecular weight, solubility, half-life period, bioavail-ability); (ii) time of exposure and environmental parameters suchas water oxygenation and temperature; (iii) intraspecific factorssuch as living conditions, pelagic and carnivorous for certainspecies, sessile and filter-feeding for the others, or the physiolo-gical status of the animals (e.g. ventilatory rate, reproductivecondition, age, capacities to uptake and metabolise PAHs, rate ofexcretion) (Meador et al., 1995; Varanasi et al., 1985). In fact, thefish possess very rapid rates of uptake, which may be related inpart to high ventilatory rates; they can metabolise PAHs muchmore efficiently and hence acquire faster elimination rates(Meador et al., 1995). In this work, comparable conditions for fishand oysters were provided by the different experimental proce-dures: similar exposure to the soluble fraction of hydrocarbons ledto an analogous route of contaminant uptake by diffusion acrossgills and intestinal or cutaneous teguments for each species. Thus,

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0

10

20

30

40

50

60

70

D0

Sea bass sampling dates

AC

H50

(U

nity

.mL

-1)

Control

HFO

0

2

4

6

DO

Oyster sampling dates

Spec

ific

PO

act

ivity

(g.

L-1

) Control

HFO

* *

* *

*

D1 D3 D9 D14 D30D3 D14 D30

Fig. 2. Haemolytic activity of alternative complement pathway (ACH50, A) in fish and specific phenoloxidase activity (PO, B) in oysters measured by spectrophotometry

during the recovery period following exposure to heavy fuel oil (HFO). n ¼ 10 for fish and n ¼ 15 for oysters. The bars represent the standard error. * ¼ statistical difference

for pp0.05.

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the principal differences between the two species were theexposure time, 5 days for fish and 9 days for oysters, and thedifferential accumulations of PAHs according to molecular weightand the number of rings. Van der Oost et al. (1994) showed thatthe eel, Anguilla anguilla, presents a higher proportion of 2- and3-ring PAHs and less of the 4- and 5-ring compounds. Theelimination of higher lipophilic PAHs (3-, 4- and 5-rings) was dueto fish metabolism (Varanasi et al., 1985) for which the typicaleffect is the formation of hydrophobic metabolites that are rapidlyexcreted (Van der Oost et al., 1994). These differential accumula-tions of PAHs shown here were in concordance with theaccumulation of naphthalene and acenaphthene, both 2-ringmolecules, in fish muscle. The oysters, on the other hand,accumulated the highest mean weight percentage of PAHs, whichwas produced by 3- and 4-ring compounds (Wade et al., 1988),like the pyrene and the phenanthrene bioaccumulated in thisstudy. In contrast with fish, the accumulation of higher molecularweight compounds in oyster tissues seems to be due to the lowcapacity of bivalves to metabolise PAHs (Varanasi et al., 1985),whereas the 2- and 3-ring molecules were apparently releaseddirectly through the skin and gills without being metabolised(Varanasi et al., 1985). This metabolism discrepancy explains whythe fish bioaccumulated naphthalene and acenaphthene, 2-ringcompounds, whereas the oysters preferentially bioaccumulatedphenanthrene and pyrene, respectively 3- and 4-rings. Fish, whichefficiently metabolise higher PAHs, may have excreted PAHmetabolites at a similar rate to the rate of uptake of parentcompounds. Nevertheless, since accumulation of PAHs in fish wasdirectly proportional to the ability of the organisms to metabolisePAHs, the absence of metabolisation of lower molecular weightcompounds seems to induce an increase in accumulation.

Metabolism, excretion and diffusive loss are processes that candecrease tissue concentrations of parent PAHs. The principaldiscrepancy between bivalves and fish was the type of elimina-tion. In bivalves, diffusive loss prevails, the decrease in tissueburden being caused by simple diffusion, whereas in fish activeexcretion is the predominant method, a physiological process thateliminates parent compounds and metabolites through bile orurine (Meador et al., 1995). In addition to excretion type,molecular weight and PAH hydrophobicity induced different ratesof elimination (Neff, 1979). All these factors could explain thedifference between the time taken to eliminate the molecules inoyster and fish tissues. Indeed, lower molecular weight PAHs candiffuse easily out of organisms through their gills, whereas a PAHwith higher molecular weight is not appreciably eliminatedthrough this route (Meador et al., 1995). This indicates that seabass could easily diffuse naphthalene and acenaphthene throughtheir gills; nevertheless, a part of these PAHs could be metabolisedand excreted later. In the same way, since oysters presentintensive filtration by their gills, phenanthrene and pyrene couldbe easily eliminated by important diffusive loss. Therefore, arelationship between the type of excretion and the bioaccumu-lated PAH decline could explain the difference observed betweenoysters and fish elimination duration.

4.3. PAHs and immunological analysis

In this study, no organism mortality was recorded. Neitherinternal nor external lesions were observed in exposed animals,probably due to the short exposure time and the controlledexternal conditions (e.g. filtered water, constant temperature andoxygenation, etc.), which prevent or at least delay secondaryinfections often observed after immunotoxicity phenomena. Theexperiment was designed in order to detect effects of pollutantson innate immune parameters. All experimental conditions were

equal in both control and contaminated tanks except for thepresence of pollutants. If pollutants were capable of inducingsome modification of cellular parameters, significant differencesbetween both conditions (control versus contaminated) would beexpected. Similar temporal trends detected in both control andcontaminated organisms were thus out of the scope of the study.However, such results indicated that the rearing conditions maybe improved.

Variations in cellular composition of blood and haemolymph,as total and differential cell counts, are reported among the firstphysiological disturbance described in organisms exposed toenvironmental stressors (Fisher, 1988; Reynaud and Deschaux,2006). In the two species tested, exposure to 733 ng L�1 of thesoluble fraction of HFO did not significantly modulate cellularsubpopulation percentages. In fact, in sea bass, no modification tothe classic proportions of granulocytes–monocytes (20%) andlymphocytes (80%) (Scapigliati et al., 2003) was observed.Approximately 40% of oyster haemocytes were granulocytes and60% were hyalinocytes, which corresponds to classical valuesreported for Pacific oysters (Hegaret et al., 2003). The principaldiscrepancy between fish and oyster cellular composition con-cerns cellular mortality: a significant increase in sea bassleucocyte mortality was observed after the 5-day exposure toHFO (day 0), whereas no haemocyte mortality was noted in theinvertebrates. This difference was probably due to the PAHsbioaccumulated in organisms. Indeed, the naphthalene, quantifiedin fish muscle and absent in oyster tissues, is known to causedeleterious membrane damage, which induces a decrease in thenumber of immune cells (Ahmad et al., 2003). In the present work,the increase of leucocyte mortality could be explained by theliposoluble properties of naphthalene and acenaphthene: (1) thenaphthalene could induce damage to cellular membranes, such asglycoprotein and protein alteration, when they are transportedthrough the leucocyte membrane; (2) the penetration ofnaphthalene into lysosomes could induce membrane rupture,which in turn would induce pH modification in leucocyte anddefinitely cell destruction (Grundy et al., 1996a).

Moreover, this lysosomal damage could also interfere withphagocytic percentage. In the present work, fish do not presentmodifications of this immune parameter, thus naphthaleneimpact seems to induce cellular mortality by direct damage tothe leucocyte membrane. On the other hand, in oysters, thephagocytic percentage was significantly decreased by short-termexposure to 733 ng L�1 of the soluble fraction of HFO. This mightbe correlated with the bioaccumulation of PAHs in oyster tissuesat this sampling date, and in particular of phenanthrene, which isknown to modify phagocytic activity (Grundy et al., 1996a). Theseauthors suggest that PAHs, or their metabolites to a lesser degree,disrupt lysosomal integrity by altering membrane fluidity, whichprevents deformation of the membrane essential in the process ofphagocytosis (Grundy et al., 1996a). Moreover, the disruption oflysosomal membrane integrity induces an acidification of hae-mocytes and then an alteration of internal pH of haemocytes. Thisacidosis affects proteolytic enzymes that may also contribute tothe decrease in phagocytic activity (Grundy et al., 1996b). Thus foroysters, as described by Auffret et al. (2006), among functionalparameters, the phagocytic capacity of haemocytes appears as apossible tool for monitoring pollution.

Concerning innate and humoral immune parameters, thecomplement of fish is an essential part of the innate immunesystem, which shares some similarities with the proPO/PO systemin invertebrates. Concerning sea bass, a significant increase inhaemolytic activity of the alternative pathway was observed incontaminated fish from days 3 to 9 of the recovery period. Thisresult suggests that high amounts of complement componentswere secreted, probably by phagocytes, which are known to be an

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important source of these components in mammals (McPhadeand Whaley, 1993). This activation of the metabolism of these cellsmay reflect an inflammatory response, probably due to thebioaccumulation of naphthalene and/or acenaphthene whosetoxic potential due to their polycyclic aromatic nature was known.This inflammatory phenomenon, already described following PAHcontamination (Myers et al., 1998; Stentiford et al., 2003), wasassociated, when exposure time was extended, with cellulardamage favouring an increase in pathologies due to opportunistpathogens. An activation of the alternative pathway to fight theseaggressors could be considered (Sakai, 1992). In Pacific oysters, POactivity decreased after short exposure to the soluble fraction ofHFO and the recovery of this activity was obtained after 2 weeks.Some other pollutants have been known to cause to cause asimilar decrease in PO activity in vitro such as mercury in thePacific oyster (Gagnaire et al., 2004) and in vivo such as trichlorfonin prawns (Chang et al., 2006).

5. Conclusions

An in vivo system of experimental contamination by HFO wasdeveloped and a preliminary validation carried out. PAHs weredetected both in seawater and organisms’ tissues, confirming theefficacy of the developed system. In a second step, the effects ofHFO were reported on some immune parameters in the Europeansea bass, D. labrax, and in the Pacific oyster, C. gigas. In vivo

contamination by the soluble fraction of HFO caused leucocytedeath in fish and affected oyster phagocytosis. Moreover, the twoenzymatic cascades (PO activity in bivalves and ACH50 in fish)appear to be useful factors for monitoring pollution by oil.Nevertheless, a recovery period could rapidly recondition immuneparameters after this short-term exposure.

Acknowledgments

This research was supported by the European programs‘‘Emergency Response to Coastal Oil, Chemical and Inert Pollutionfrom Shipping’’ (EROCIPS INTERREG IIIB) and Total. The authorsthank to Sally Ferguson (Alba Traduction) for her reading thisdocument.

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A. Bado-Nilles et al. / Ecotoxicology and Environmental Safety 72 (2009) 1446–14541454