Skip to main content

Multilevel responses of adult zebrafish to crude and chemically dispersed oil exposure



The application of chemical dispersants is a common remediation strategy when accidental oil spills occur in aquatic environments. Breaking down the oil slick into small droplets, dispersants facilitate the increase of particulate and dissolved oil compounds, enhancing the bioavailability of toxic oil constituents. The aim of the present work was to explore the effects of water accommodated fractions (WAF) of a naphthenic North Sea crude oil produced with and without the addition of the chemical dispersant FINASOL OSR 52 to adult zebrafish exposed for 3 and 21 d. Fish were exposed to environmentally relevant concentrations of 5% and 25% WAFOIL (1:200) and to 5% WAFOIL+D (dispersant–oil ratio 1:10) in a semi-static exposure setup.


The chemically dispersed WAF presented a 20-fold increase of target polycyclic aromatic hydrocarbons (PAHs) in the water phase compared to the corresponding treatment without dispersant and was the only treatment resulting in markedly bioaccumulation of PAHs in carcass after 21 d compared to the control. Furthermore, only 5% WAFOIL+D caused fish mortality. In general, the undispersed oil treatments did not lead to significant effects compared to control, while the dispersed oil induced significant alterations at gene transcription and enzyme activity levels. Significant up-regulation of biotransformation and oxidative stress response genes (cyp1a, gstp1, sod1 and gpx1a) was recorded in the livers. For the same group, a significant increment in EROD activity was detected in liver along with significant increased GST and CAT activities in gills. The addition of the chemical dispersant also reduced brain AChE activity and showed a potential genotoxic effect as indicated by the increased frequency of micronuclei in erythrocytes after 21 d of exposure.


The results demonstrate that the addition of chemical dispersants accentuates the effect of toxic compounds present in oil as it increases PAH bioavailability resulting in diverse alterations on different levels of biological organization in zebrafish. Furthermore, the study emphasizes the importance to combine multilevel endpoints for a reliable risk assessment due to high variable biomarker responses. The present results of dispersant impact on oil toxicity can support decision making for oil spill response strategies.


Oil spills represent one of the most important potential sources of pollution in aquatic environments [140]. Anthropogenic activities such as offshore oil exploitation and transportation, or other contaminations caused by shipwrecks and industrial discharges, pose a high risk of an oil spill that could result in devastating consequences to aquatic organisms, such as fish [19, 52, 55]. Crude oils are complex mixtures of different toxic compounds that volatilize, dissolve, emulsify or disseminate through the water column [19, 115, 128]. Oil composition and environmental factors such as temperature and salinity are key elements determining the magnitude of the biological impact of an oil spill which—along with natural weathering and the selected remediation strategies—may influence the adverse aftermath of these accidents [18, 69, 106]. Polycyclic aromatic hydrocarbons (PAHs), for example, represent up to 60% of crude oil constituents present in water accommodated fractions (WAF) [40]. Their capacity to provoke diverse toxic effects in aquatic organisms is well known. These effects range from cardiotoxicity and malformations at early developmental stages up to behavioural alterations and carcinogenesis in adults [27, 54, 78, 129].

The use of chemical dispersants is a common practice applied to limit the spreading of the oil slick, thus preventing it from reaching coastal areas and mitigating its possible toxicological impacts on mainly birds and mammals populations [19, 44, 70, 90]. Chemical dispersants are composed of hydrocarbon solvents with anionic and non-ionic surfactants, displaying both hydrophilic and lipophilic properties that enable a rupture in the interfacial tension between oil and water [32, 44, 70, 90]. Thus, they are capable of breaking down the oil slick into smaller droplets which have been suggested to be more easily degraded by bacteria or photo-oxidation [137]. However, through a massive surface increase and direct ingestion of the droplets, chemical dispersion simultaneously enhances the bioavailability of toxic oil constituents, such as PAHs, shifting the ecotoxicological damage towards pelagic species [19]. The potential adverse outcomes of the application of chemical dispersant remains a subject of major concern [17, 44, 46], which is why their effectiveness, benefits and impacts resulting from their use need to be further evaluated.

The EU Horizon 2020 funded project GRACE focused on a holistic approach to investigate the environmental effects of oil spills and response measures [65]. As a part of this, the present study focused on adult zebrafish due to multiple advantages as an (eco)toxicological model including small size, fast development, low maintenance cost, ease of reproduction and large spawn. Furthermore, the zebrafish is a well-established animal model for the assessment of oil spill toxicity [34, 66, 126] and response strategies, such as chemical dispersion [75, 101]. Biomarker changes in gene transcription, development, metabolism, or even behaviour can be efficient tools to provide evidence of an exposure to pollutants [10, 31]. Measuring biomarkers on molecular and enzymatic level can indicate the exposure to toxic substances before an evident tissue-level damage occurs or even when pollutant concentrations in the media are below detection limits, supplying valuable information for environmental health evaluations [106, 121].

One of the sensitive biomarkers selected for the current study, which is linked to crude oil exposure, is the induction of phase I metabolism [10, 29, 121]. Alterations of ethoxyresorufin-O-deethylase (EROD) activity manifest changes on the expression of cytochrome P450 (e.g., the CYP1A subfamily), an essential mechanism along with the likewise investigated glutathione S-transferase (GST) activity in the phase II biotransformation reactions of xenobiotic detoxification [48, 78]. Furthermore, when these two phases of metabolism are not well coupled, the reactive metabolites produced in phase I may generate DNA adducts, which might lead to genotoxic effects [54, 94, 138]. The potential of oil constituents such as PAHs to induce DNA damage has been demonstrated previously, making genotoxicity also a relevant endpoint for crude oil toxicity assessment [12, 79, 112]. Since DNA adducts might intercalate into the DNA inducing strand breaks, genotoxicity assessment in the current study includes the investigation of micronuclei formation in peripheral blood samples in addition to selected molecular marker genes related to cell cycle control and apoptosis (tp53, casp3a). An imbalance between the appearance of reactive oxygen species (ROS) from phase I biotransformation and enzymatic antioxidant defence mechanisms, such as glutathione peroxidase (GPX) or catalase (CAT) activities [116, 121], can lead to an oxidative damage of cells and tissues, which, ultimately, could be translated into necrosis, apoptosis or carcinogenesis [53, 78, 100, 116, 121]. The assessment of the redox status or health status in general through biochemical or molecular techniques is frequently addressed in gills as the first barrier and contact with crude oil compounds present in the WAF, or in liver as a main detoxifying organ [42, 121] and hence was selected in the current study as well. Finally, the neurotoxic potential of crude oil constituents was investigated by measuring acetylcholinesterase (AChE) inhibition in brain. Though not commonly investigated in oil toxicity studies, recent studies indicate that petroleum hydrocarbons can impact neuronal development and induce neurotoxicity via transcriptional alterations, neurotransmitter regulations or behaviour analyses [45, 125, 133].

Comprehensive studies addressing different levels of biological organization in parallel and offering conclusions about the mechanisms associated with crude oil and the response strategy’s toxicity are rather scarce. Therefore, the aim of the present study was to explore the transcriptomic, biochemical and genotoxic effects in adult zebrafish (Danio rerio) exposed to different WAF dilutions of a naphthenic North Sea (NNS) crude oil, produced with and without the addition of the chemical dispersant FINASOL OSR 52. Furthermore, by means of a comprehensive chemical analysis of target PAHs from the exposure water phase and fish tissue, the biological responses were aimed to be linked to a potential bioaccumulation of oil constituents. Sensitive effect-based methods and biomarkers could further support the need for the implementation of biological monitoring for water quality assessment in the Marine Strategy Framework Directive (MSFD) [130].

Materials and methods

Zebrafish maintenance

Wild-type zebrafish of AB Salk strain from the facilities of the University of Basque Country (UPV/EHU) were used in the present study. Hatched larvae were grown until 5–6 months of age prior to the experiment. Fish were maintained in a temperature-controlled room at 28 ± 1 °C with a constant light–dark rhythm (12:12) in 100 L tanks equipped with mechanical and biological filters. Water was in continuous movement triggered by the action of an aeration siphon. Water was previously conditioned by passage through an osmosis membrane and then buffered to a pH of 7.2 with Sera pH plus (Sera, Heinsberg, Germany) and remineralized to a conductivity of 600 μs cm−1 with commercial marine salt (Sera). Fish were fed twice a day with live brine shrimp larvae (Artemia sp, INVE Aquaculture, Salt Lake City, USA) and Vipagran baby (Sera).

Preparation of water accommodated fractions

A naphthenic North Sea (NNS) crude oil (Equinor, Stavanger, Norway) and the dispersant FINASOL OSR 52 (TOTAL Special Fluids, Paris, France) were used in the present study. Water accommodated fractions (WAF) of oil (WAFOIL) and chemically dispersed oil (WAFOIL+D) were prepared in 5 L (WAFOIL+D) or 20 L (WAFOIL) glass flasks according to Singer et al. [114] with modifications. Briefly, oil or a dispersant–oil mixture (1:10, w/w) was gently applied on the surface of formulated fish water (remineralized osmosis water) at an oil-to-water ratio of 1:200 (w/w). Regardless glass flask volumes the ratio of water, oil and dispersant and headspace were kept constant. Both setups were stirred for 40 h at 20 °C in dark with low energy avoiding a vortex in the water phase. Water fractions were then carefully drained off from a stopcock at the bottom part of the flask and immediately used for exposure.

Exposure regime

Exposure concentrations were selected based on pre-tests. A pilot experiment was carried out to test the general toxicity of 25% WAFOIL and 5% WAFOIL+D on adult zebrafish exposed for 6 d with a solution renewal at day 3 resulting in a 100% survival. For the main experiment, eight glass tanks with 45 L capacity covering 1 control and 3 treatment groups (5% WAFOIL, 25% WAFOIL and 5% WAFOIL+D) in duplicate were established in a temperature-controlled room (28 °C) with 12:12 light:dark rhythm. Formulated fish water for the control group, and exposure solutions for the treatments were already filled into the tanks 3 d before fish were added in order to saturate the system. Biological and mechanical filters were not used during the experiment to avoid interference with the exposure. Tank internal water circulation was maintained using a circulatory pump. Fifty adult zebrafish (AB Salk) at the age of 5–6 months were placed in each tank at a sex ratio of approximately 1:1. Exposure medium was renewed every 3 d at a 75% exchange rate. During the experiment, fish were fed twice a day with live brine shrimp larvae. After 3 and 21 d of exposure, 50 adult zebrafish per treatment were individually anesthetized in a benzocaine solution (200 mg L−1) prepared in a 1:9 (v/v) ethanol–water stock, and immediately dissected. For both sampling timepoints one of the two replicate tanks was used. As several biomarker responses are reported to be influenced by fish gender [36, 121, 132], males and females were separated for individual biomarkers as described in detail below.

Chemical analysis of exposure media and fish tissue

Chemical analysis of target PAHs in exposure media

Chemical analysis of exposure media was performed using stir-bar (twister, Gerstel GmbH & Co. KG, Mülheim an der Ruhr, Germany) sorptive extraction (SBSE) technique according to Prieto et al. [104]. Twisters (20 mm length and 0.5 mm film thick) were placed on magnetic stirrers in the control and treatment tanks and exchanged at 6, 12, 24, 36, 48, 60 and 72 h during two complete exposure cycles of 3 d in order to evaluate the concentrations of PAHs in exposure media along the experiment.

Detection of PAHs was performed using gas chromatography–mass spectrometry (GC (Agilent 6890)–MS (Agilent 5975), Agilent Technologies, Santa Clara, USA) analysis according to the protocol described in Prieto el al., [104] with modifications. A mix standard solution of 16 PAHs (NS 9815: S-4008-100-T, Norwegian Standard supplied by Chiron, Trondheim, Norway) was used for calibration (Additional file 1: Table S1).

PAHs were desorbed from the twisters using a commercial thermal desorption TDS-2 unit connected to a CIS-4 injector (Gerstel GmbH & Co. KG, 10 min, 300 °C, flow 23 mL min−1, cryo-focusing − 50°, 7 psi). An Agilent DB-5MS + DG column (Agilent Technologies) was used to separate analytes in a helium stream (1.3 mL min−1). Transfer line, ion source and quadrupole analyser temperatures were maintained at 300, 230 and 150 °C, respectively. The following temperature program was used for target PAHs and lineal hydrocarbons: 170 °C for 5 min; ramp at 30 °C min−1 to 260 °C; ramp at 8 °C min−1 to 300 °C and hold 15 min. The mass selective detector was operated in selected ion monitoring for quantification of target compounds.

Chemical analysis of target PAHs in fish tissue

After 21 d of exposure, fish carcasses from the biological endpoints (described below) were preserved at − 20 °C for the analysis of the bioaccumulation of PAHs. The analysis of PAHs in the fish samples was carried out according to Navarro et al. [93]. Briefly, the fish tissue samples were pooled for each group before analysis. Tissue sample (0.5 g) was accurately weighed in the extraction vessel and 5 mL of acetone were added. The PAHs extraction was performed by focused ultrasound solid–liquid extraction using a SONOPULS HD 2070 (Bandelin electronic GMBH & Co. KG, Berlin, Germany) provided with a 3 mm titanium microtip at 45% of ultrasound amplitude and 0 °C during 120 s. The extracts were filtered by Millex® HV PVDF 0.45 μm (Millipore, Carrigtwohill, Ireland) and concentrated to ~ 0.5 mL under a nitrogen stream (TurboVap LV, Zymark, Barcelona, Spain) after the addition of 1 mL of iso-octane. The concentrated extracts were cleaned with 5 g Florisil cartridges previously conditioned with n-hexane, and eluted with 25 mL of n-hexane:toluene 75:25 mixture. Subsequently, they were concentrated to dryness, redissolved in iso-octane and kept at − 20 °C in the dark until the GC–MS analysis (previously described).

Gene transcription analysis

Livers of 15 male zebrafish per experimental group and sampling time were dissected, transferred individually to cryovials prefilled with RNA later (Sigma-Aldrich, Merck KGaA, Darmstadt, Germany) and immediately frozen in liquid nitrogen. Samples were preserved at − 80 °C until analysis. We prioritised the use of males for the gene transcription levels based on the results obtained in previous studies [71, 72, 83].

Total RNA was extracted from a pool of 3 male fish livers with 5 replicates for each of the 4 experimental groups and for each sampling timepoint, except in the case of 5% WAFoil+D group sampled at 21 d with 4 replicates due to the mortality recorded in this treatment. Tissue was homogenized in 300 µL of Trizol using an electric disperser (Pellet Pestle® Cordless Motor, Kimble Kontes, Merck KGaA). RNA was extracted using the Trizol reagent method (Invitrogen, Thermo Fisher Scientific, Waltham, USA) following the manufacturer’s instructions with minor modifications. Samples were transferred to Phase Lock Gel Heavy pre-filled tubes (Quantabio, Beverly, USA) prior to the first centrifugation at 12,000 g and 4 °C for 15 min. RNA was diluted in 50 µL RNAse- and DNAse-free distilled water (Invitrogen). Sample concentration and quality control were evaluated through an Agilent 2100 Bioanalyzer System (Agilent Technologies) using the Agilent RNA 6000 Nano Kit and the corresponding 2100 Expert Software.

First-strand cDNA was synthesized from 2 µg of RNA using the AffinityScript QPCR cDNA Synthesis Kit (Agilent Technologies) following the manufacturer’s protocol. The conditions for cDNA synthesis were: primer annealing at 25 °C (5 min), cDNA synthesis at 42 °C (15 min) and reaction termination at 95 °C (5 min). cDNA was stored at − 20 °C. The Quant-iT OliGreen ssDNA Assay Kit (Thermo Fisher Scientific) was used for quantification of cDNA samples according to manufacturer’s instructions using a Synergy HT Multi-Mode Microplate Reader with Gen5 Microplate Reader and Imager Software (BioTek Instruments, Agilent Technologies, Santa Clara, USA).

Quantitative real-time PCR was performed using TaqMan Gene Expression Assays (Applied Biosystems, Thermo Fisher Scientific, reaction efficiency reported from manufacturer 100 ± 10%) inventoried for cyp1a (Dr03112441_m1), gstp1 (Dr03118992_g1), cat (Dr03099094_m1), sod1 (Dr03074068_g1), gpx1a (Dr03071768_m1), tp53 (Dr03112086_m1), and casp3a (Dr03131690_m1). The reaction mixture had a total volume of 20 µL containing 10 µL TaqMan Gene Expression Master Mix (2X), 1 µL TaqMan gene expression assay (20X), 7 µL RNAse-free distilled water (Invitrogen), and 2 µL cDNA sample dilution. Samples and process controls (cDNA synthesis control and no-template control) were performed using a 7500 Real-Time PCR system (Applied Biosystems) at the manufacturer’s standard thermal cycling conditions: initial incubation at 50 °C (2 min), activation at 95 °C (10 min), 40 cycles of denaturation at 95 °C (15 s) and annealing and extension at 60 °C (1 min). Transcript levels were normalized by the cDNA concentrations previously quantified according to Valencia et al. [120]:

$$RQ = \left[ {\frac{{\left( {1 + Efficiency} \right)^{ - \Delta CT} }}{cDNA (ng)}} \right]$$

where ΔCT = CT sample − CT plate internal control.

Biochemical marker analysis

Livers, gills and brains of 12 female zebrafish per treatment were dissected, individually transferred to cryovials and shock frozen in liquid nitrogen. Samples were stored at − 80 °C until further usage. A pool of 3 livers, gills or brains were homogenized in 300 µL cold homogenization buffer (50 mM potassium phosphate buffer, pH 7.5, 1 mM ethylenediaminetetraacetic acid, 0.5 mM dithiothreitol, 0.4 mM phenylmethylsulfonyl fluoride) for 15 s on ice using an electric disperser (Kimble Kontes) resulting in 4 samples per treatment. Subsequently, homogenates were centrifuged at 10,000 g and 4 °C for 15 min (Eppendorf 5415R refrigerated centrifuge, Sigma-Aldrich). Afterwards, the supernatant was carefully transferred to new tubes and aliquoted on ice for different enzymatic and protein measurement in order to avoid repeated thawing and freezing. Supernatants were stored at − 80 °C until further usage.

7-Ethoxyresorufin-O-deethylase (EROD) activity in livers and gills

Measurement of EROD activity in liver and gill supernatants was performed according to Brinkmann et al. [25] with minor modifications. Briefly, 20 µL of sample (in triplicate) and a resorufin (Sigma-Aldrich) calibration series (1:2 dilution series from 0.004 to 1 µM) in duplicates were transferred to a 96-well plate. Resorufin standard and stock solutions were prepared in Tris-KCl buffer (pH 7,4; Trizma base 0.1 M, KCl 0.15 M). 200 µL of 7-ethoxyresorufin (0.5 µM, Sigma-Aldrich) were added to each well followed by 10 min incubation at room temperature in darkness. Shortly, before kinetic measurement of fluorescence for 25 min (kinetic intervals: 30 s) in a microplate reader (FLx800, BioTek Instruments), 20 µL NADPH (1 mM, Sigma-Aldrich) were added. Substrate deethylation was determined by measuring the formed resorufin at 540 nm excitation and 590 nm emission wavelengths. Quantification of EROD activity was performed based on the resorufin calibration and expressed in pmol resorufin mg−1 min−1.

Glutathione-S-transferase (GST) activity in liver and gills

Measurement of GST activity in zebrafish liver and gills was performed according to Habig et al. [49] with modifications regarding the adaption to 96-well plates described in Velki et al. [123]. 12 µL of supernatant as well as 180 µL of 1-chloro-2,4-dinitrobenzene (1 mM, dissolved in phosphate buffer pH 7.2, Sigma-Aldrich) and 50 µL of reduced glutathione (25 mM, Sigma-Aldrich, dissolved in MilliQ water) were added to a 96-well plate. Immediately thereafter the increase in absorbance as a result of S-(2,4-dinitrophenyl) glutathione formation was measured in triplicates at 340 nm for 15 min in 10 s intervals using a microplate reader (Multiskan Spectrum, Thermo Fisher Scientific). Resulting data were controlled for linearity in absorbance increase (R2 ≥ 0.98) and minimum increase of absorbance over time (Δt3min ≥ 0.1). Only data fulfilling these criteria were used for further calculations. Enzymatic activity was calculated as nmol conjugated glutathione min−1 mg−1. The molar extinction coefficient of 9600 M−1 cm−1 was used.

Catalase (CAT) activity in liver and gills

Measurement of catalase (CAT) activity in zebrafish liver and gill supernatants was performed according to the initial protocol developed by Claiborne [28] adapted to UV 96-well plates (Thermo Fisher Scientific). 5 µL supernatant were added to 295 µL of H2O2 solution (20.28 mM, Sigma-Aldrich). Immediately thereafter the decrease in absorbance was measured kinetically at 240 nm for 5 min in 10 s intervals using a microplate reader (Multiskan Spectrum). In addition, the absorption of an H2O2 dilution series (0.4–20.28 mM) was measured for quantification of H2O2 consumption. Calibration series as well as samples were measured in quadruplicates. Based on the increase of measurement-disturbing O2 bubbles, the linear part of the reaction (until 1 min) was used for calculations. Enzyme activity was expressed as µmol H2O2 consumption min−1 mg−1 using the calibration series.

Acetylcholinesterase (AChE) activity in brain

Measurement of AChE in brain tissues was conducted according to the initial protocol established by Ellman et al. [43] with modifications according to Velki et al. [123] for 96-well plates. Briefly, 7.5 µL sample supernatant as well as 180 µL potassium phosphate buffer (0.1 M, pH 7.2), 10 µL 5,5′-dithiobis(2-nitrobenzoic acid (1.6 mM, Sigma-Aldrich), and 10 µL acetylcholine iodide (156 mM, Sigma-Aldrich) were added to a 96-well plate. The increase in absorbance was immediately measured in triplicates at 412 nm for 25 min in 10 s intervals using a microplate reader (Multiskan Spectrum). Resulting data were controlled for linearity in absorbance increase (R2 ≥ 0.98) and minimum increase of absorbance over time (Δt3min ≥ 0.1). Only data fulfilling these criteria were used for further calculations. Enzymatic activity was calculated as nm acetylcholine hydrolyzed min−1 mg−1. For the calculations, the molar extinction coefficient of 13,600 M−1 cm−1 was used.

Protein measurement

Whole protein of supernatants (brain, liver, gill) was measured in triplicates using a DC protein assay kit (BioRad, Hercules, USA) according to the manufacturer’s instructions and quantified with a dilution series of bovine serum albumin as an external standard (1.5–0.15 mg mL−1) measured in quadruplicates.

Micronucleus frequency in peripheral erythrocytes

Peripheral blood samples of 7–11 individual males per treatment were taken from the caudal blood vessel of euthanized individuals after cutting the fish tail. The assessment of micronuclei in males and not in females was related to the sampling logistics and not to sex. One smear per individual was immediately prepared on microscopy slides. After drying, samples were fixed using a Hemacolor rapid staining kit according to manufacturer instructions (Merck KGaA) and mounted using Kaiser’s glycerine gelatine (Merck KGaA).

A microscope (Eclipse 50i Nikon Instruments, Düsseldorf, Germany) with 100 × magnification and the associated software NIS element (v.5.11, Nikon, GmbH) was used to generate images of erythrocytes. The proportion of micronuclei in peripheral erythrocytes was determined for 1000 cells per individual sample with the following scoring criteria according to Carrasco et al. [26] and Bolognesi and Hayashi [21] for the identification of micronucleated cells: (a) cells with oval appearance and intact cytoplasm, (b) micronuclei less than or equal to one third of the nucleus, (c) micronuclei clearly separated from the main nuclei and d) the micronuclei had the same staining intensity as normal nuclei.

Data analysis

Statistical analyses were performed using SPSS Statistics 26th version (IBM, New York, USA; p < 0.05). Metric data including gene transcription level, enzyme activity and micronucleus frequency analysis were tested for normal distribution (Shapiro–Wilk) and equal variance (Levene). For data fulfilling those criteria, a one-way ANOVA with HSD Tukey’s post-hoc test was performed to identify significant differences between treatment groups. Data that were not normally distributed and/or of unequal variances were logarithmically transformed. If the transformation failed, data were analyzed using non-parametric Kruskal–Wallis on ranks with Dunn–Bonferroni test. For statistical comparison of exposure regimes (3, 21 d) Student’s T test was performed for normally distributed and homoscedastic data. For data not fulfilling the criteria, the Welch test or non-parametric Mann–Whitney’s U test were applied depending on the normality of data.


Chemical analysis of target PAHs

Target PAHs in exposure media

PAH levels in the tanks were measured over two consecutive exposure cycles. In control tanks, concentrations of all target PAHs were below the limits of detection (LOD). The concentration of total PAHs was reduced for all treatments along the exposure cycle in the open test systems (Fig. 1). After exposure solutions were exchanged, a new peak in dissolved PAH concentration was measured. The concentration of total PAHs at the beginning of the cycle increased from 224.4 ± 17.7 ng L−1 (5% WAFOIL, 6 h) over 1,724.3 ± 497.5 ng L−1 (25% WAFOIL, 12 h) up to 5,744.1 ± 119.8 ng L−1 (5% WAFOIL+D, 6 h). With medium exchange, the chemically dispersed WAF (WAFOIL+D) contained target PAH in concentrations > 20-fold higher than the corresponding dilution without the dispersant application (5% WAFOIL). Details of the concentration of individual target PAHs in corresponding exposure media can be found in Additional file 1: Tables S2–S4.

Fig. 1

Sum of target PAHs in exposure media over two exposure cycles. Exposure media were exchanged every 72 h. PAHs were extracted from media using stir bar sorptive extraction and analyzed using GC–MS

The most dominant PAH across all WAF exposure media was phenanthrene followed by pyrene and fluoranthene (Additional file 1: Tables S2–S4). Phenanthrene was detected in concentrations up to 2674.4 ± 229.8 ng L−1 in chemically dispersed WAF within the first 6 h sampling interval after medium exchange. The highest concentrations of phenanthrene shortly after medium exchange (start of exposure cycles) in the case of the WAFOIL dilutions were 147.8 ± 18.0 ng L−1 (6 h) and 1251.7 ± 324.1 ng L−1 (12 h) in 5% and 25% WAFOIL, respectively. Overall, the concentration of dissolved target PAHs increased with decreasing log Kow of individual compounds. High molecular weight PAHs were not quantifiable or even below the detection limits in both WAFOIL groups, while the application of the chemical dispersant resulted in the detection of benzo[a]anthracene, chrysene and benzo[b]fluoranthene in ng L−1 ranges.

Bioaccumulation of target PAHs in fish tissue

At the end of the experiment (21 d) target PAHs were detected in fish carcass across all treatments including the unexposed control (Fig. 2). Overall, with the exception of phenanthrene and pyrene, PAH levels in fish exposed to both undispersed oil treatments were in a comparable range of the unexposed control. A markedly increased bioaccumulation of target PAHs was found only for the dispersed oil treatment (5% WAFOIL+D). Low molecular weight PAHs such as phenanthrene and fluoranthene were detected up to 782.1 ng g−1 and 238.4 ng g−1, respectively. Furthermore, higher molecular weight PAHs including benzo[a]anthracene, chrysene, indeno[1,2,3-cd]pyrene or benzo[ghi]pyrelene, which were scarcely detected or even below detection limits in the water phases of all treatments (Additional file 1Tables S2–S4), were detected in the fish tissues, with the highest concentrations up to 604.0 ng g−1 (benzo[a]anthracene + chrysene) (Additional file 1: Table S5).

Fig. 2

Target PAHs detected in fish carcass after 21 d of exposure to WAF dilutions of crude oil (5% and 25% WAFOIL) and chemically dispersed crude oil (5% WAFOIL+D) relative to unexposed control. Target PAHs were analysed by GC–MS. Bars represent PAH concentration of a pool of collected carcasses. Red line indicates PAH levels in the control group set to 100%

Survival and peculiar behaviour

The chemically dispersed oil treatment (5% WAFOIL+D) led to a mortality rate of 23% at 21 d, whereas 100% survival was observed for the 5% and 25% WAFOIL as well as the control. Up to 6 deceased fish were found from day 3 of exposure until the end of the experiment in 5% WAFOIL+D group. Overall, chemically dispersed oil exposed fish presented a cyclic stunned behaviour, with fish swimming abnormally close to the surface in a tilted position and apparently ignoring the provided food shortly after each medium exchange. This demeanour was progressively attenuated within the first 24 h of the exposure cycle. Fish exposed to 5% and 25% WAFOIL did not show abnormal swimming behaviour and were comparable to the control individuals.

Gene transcription levels

In an overview, only the exposure to 5% WAFOIL+D caused statistically significant alterations on mRNA levels after short- and long-term exposure (Fig. 3).

Fig. 3

Transcription levels of biotransformation (cyp1a, gstp1), oxidative stress response (cat, sod1, gpx1a), cell cycle (tp53) and apoptosis (casp3a) genes in male adult zebrafish livers after 3 and 21 d of exposure to WAF dilutions of crude oil (5% and 25% WAFOIL) and chemically dispersed crude oil (5% WAFOIL+D). Data are represented as median (line) with boxes showing the 25–75 percentile and whiskers representing min to max values of 4–5 replicate groups per treatment. Letters indicate statistically significance between treatments within the same sampling point and asterisks indicate statistically significance between sampling timepoints (p < 0.05)

Both investigated biotransformation metabolism-related genes (cyp1a, gstp1) were significantly up-regulated for the 5% WAFOIL+D treatment after 3 d of exposure (Fig. 3A, B). After 21 d the cyp1a and gstp1 mRNA levels of fish exposed to chemically dispersed WAF were significantly decreased compared to the first sampling timepoint but still remained significantly higher compared to the control (Fig. 3A, B).

The oxidative stress response genes sod1 and gpx1a were statistically significant up-regulated by the 5% WAFOIL+D treatment after 3 d of exposure (Fig. 3D, E). Transcription levels of cat did not show statistically significant differences between treatment groups. The highest induction of cat was recorded in fish exposed to 25% WAFOIL at 3 d (Fig. 3C). Overall, the up-regulation of oxidative stress genes (sod1, gpx1a and cat) significantly declined from short- (3 d) until the end of long-term (21 d) exposure, albeit not differing from the control. Fish exposed to 5% WAFOIL+D still showed the highest values (Fig. 3C–E).

In contrast, no differences on the transcription levels of cell cycle- and apoptosis-related genes (tp53, casp3a) were found across treatments or compared to unexposed fish at any sampling timepoint (Fig. 3F, G).

Biochemical marker responses

Overall, the chemically dispersed crude oil (5% WAFOIL+D) showed the strongest effects on enzymatic biomarker levels. Enzyme activities corresponding to the xenobiotic metabolism phase I (EROD) and II (GST) as well as the antioxidant defence mechanism (CAT) were investigated in zebrafish liver and gill.

No notable changes in EROD activity were recorded for both undispersed crude oil treatments (5 and 25% WAFOIL) independent of the sampled tissues or the exposure period (Fig. 4A, B). In contrast, the chemically dispersed crude oil exposure (5% WAFOIL+D) resulted in significantly increased EROD activity in liver after both exposure periods up to tenfold compared to the unexposed control. Gill GST was found to be significantly increased after the exposure to chemically dispersed crude oil at both exposure times (Fig. 4D). For liver, no alterations were recorded after 3 d of exposure in any treatment. However, after 21 d all treated groups showed higher activity, which was significantly higher compared to the previous sampling in the case of 5% WAFOIL and 5% WAFOIL+D.

Fig. 4

Biomarker activity in adult female zebrafish liver and gill tissue after 3 and 21 d of exposure to WAF dilutions of crude oil (5% and 25% WAFOIL) and chemically dispersed crude oil (5% WAFOIL+D). 7-ethoxyresorufin-O-deethylase (EROD) activity (A, B), glutathione S-transferase (GST) activity (C, D) and catalase (CAT) activity (E, F). Data representation as in Fig. 3. Letters indicate statistically significance between treatments within the same sampling point and asterisks indicate statistically significance between sampling timepoint (p < 0.05)

While for liver tissue the 5% WAFOIL treatment did not lead to significant differences in CAT activity after 21 d when comparing treatments to the control group, in gills the chemically dispersed WAF presented significant higher CAT activity than the control group after 21 d (Fig. 4E, F). However, long-term exposure induced an overall trend of markedly increased CAT activity in fish gills exposed to the remaining treatments compared to control.

As a biomarker of neurotoxicity, the acetylcholinesterase (AChE) activity in brain was significantly decreased in fish exposed to the chemically dispersed oil (5% WAFOIL+D) compared to the corresponding non-dispersed WAF dilution (5% WAFOIL) at both short and long-term exposure (Fig. 5). Nevertheless, AChE activity was not significantly different for any treatment compared to the control group.

Fig. 5

Acetylcholinesterase (AChE) activity in adult female zebrafish brain after 3 and 21 d of exposure to WAF dilutions of crude oil (5% and 25% WAFOIL) and chemically dispersed crude oil (5% WAFOIL+D). Data representation as in Fig. 3. Letters indicate statistically significance between treatments within the same sampling point and asterisk indicates statistically significance between sampling timepoints (p < 0.05)

Relative induction of enzyme activity individually normalized to control can be found in the Additional file 1: Table S6.

Micronuclei frequency in peripheral erythrocytes

Both WAFOIL exposure dilutions resulted in micronuclei formation in a comparable range to the unexposed control (Fig. 6). The long-term exposure to the chemically dispersed oil WAF (21 d) was the only treatment inducing significantly more micronuclei in erythrocytes than the 5% WAFOIL exposure with a resulting induction factor of 2.2 (all induction factors can be found in Additional file 1: Table S7).

Fig. 6

Micronuclei in peripheral erythrocytes of adult male zebrafish after 3 and 21 d of exposure to WAF dilutions of crude oil (5% and 25% WAFOIL) and chemically dispersed crude oil (5% WAFOIL+D). Results are presented as mean ± standard deviation for n = 7–11. Letters indicate statistically significance between treatments within the same sampling point and asterisks indicate statistically significance between sampling timepoints (p < 0.05)


PAH exposure and bioaccumulation

With the time-resolved chemical analysis of the water phase across 2 exposure cycles we were able to show the dynamic partitioning of target PAHs in 2 ways. First, a fast concentration decrease of analyzed PAHs in the water phase of up to 92% was observed within 3 d, which is most likely associated with physical (evaporation, absorption) and biological (uptake by fish) responses of the system [115]. These findings demonstrate the dynamic peak exposure in semi-static open systems and hence the overall importance of continuous exposure solution analysis as well as careful selection of the exposure setup according to the research question.

Second, the present results showed the strong impact of the dispersant on dissolved PAHs in the water column. The observed oil compound mobilization caused by the addition of the chemical dispersant has already been reported in several studies [5, 33, 75, 89]. This phenomenon seems to be the result of the ability of dispersants to break the oil into a high number of small droplets, incrementing the contact surface of oil and water, while at the same time, the reduced size and the high surface-to-volume ratio of these oil droplets contribute to the partitioning of PAHs from the oil into the water column [88, 105]. Furthermore, the chemical dispersant altered the composition of the PAH mixture present in WAF, which might further affect the complex toxicity. Within this, high molecular weight PAHs such as benzo[a]pyrene and benzo[g,h,i]perylene were only quantifiable in the chemically dispersed oil WAF, showing that the dispersants do not interact equally with all hydrocarbons and are prone to ease the dissolution of the less hydrophilic compounds [86].

Lower molecular weight PAHs such as fluorene and phenanthrene are known to be more easily dissolved in water and hence more bioavailable for fish than PAHs of higher molecular weight [16] leading to the assumption that those cause the observed adverse effects. In fact, tricyclic PAHs detected in the WAFs, such as phenanthrene as the most dominant target PAH, are well known toxicity drivers in fish resulting in cardiotoxicity, malformations and behavioral changes (reviewed, e.g., in [52, 54]). However, though detected in relatively low concentrations in the present study, high molecular weight PAHs are potent agonists of the aryl hydrocarbon receptor, the molecular initiating event of several mechanisms related to toxic effects in fish, such as cardiotoxicity [54]. Hence, the observed adverse effects in zebrafish are likely a response to the complex mixture toxicity of high and low molecular weight PAHs and even other crude oil constituents, such as alkylated and heterocyclic derivatives, which have been shown to cause even stronger effects in different fish models compared to parent PAHs [4, 24, 82, 110]. Each crude oil is a unique complex sample that can induce various degrees of toxicity via diverse pathways depending on the mutual interference of constituents [76]. It has to be considered that only a limited fraction of oil constituents was investigated. However, even a very comprehensive chemical profile would not guarantee the explanation of the observed mixture toxicity by individual compounds, since some previous studies successfully established such a correlation between biological effects and chemical profiles [73], while others did not establish a very clear correlation [56].

Due to a high PAH biotransformation capacity of fish [76, 121] an efficient metabolization and elimination was expected [116]. This was experimentally indicated in the present study by the activity of corresponding molecular and enzymatic biomarkers (e.g., CYP1A activity, detailed discussion in Sect. 3 below). However, bioaccumulation of low and high molecular weight PAHs was notable in fish exposed to the chemically dispersed WAF. The accumulation of individual PAHs has been observed in several previous studies addressing dispersed oil toxicity [16] or toxicokinetic of PAH accumulation in fish [126, 127]. The accumulation of PAHs might be related to waterborne uptake via the gills and the skin as demonstrated for exposure of adult zebrafish to phenanthrene (100 ng L−1) reaching an equilibrium concentration in all tissues measured within 4 d of exposure [127]. Though detected in low concentrations in the water phase, the relatively strong enrichment of high molecular weight compounds (benzo[a]anthracene + chrysene, benzo[ghi] pyrelene, indene[1,2,3-cd]pyrene in carcass tissues after 21 d of exposure was potentially related to lower detoxification rates compared to lighter PAHs with lower logKow. Furthermore, it has to be kept in mind that bioaccumulation of PAHs could have been caused also by oil droplet ingestion followed by desorption through the gut [15, 139], since stable micro-droplets were expected in the water column of the present set up due to constant water circulation. This contribution of oil droplets for bioaccumulation could have been more relevant for PAHs of higher than for those of lower molecular weight. This has been indicated in previous studies [51] and shown in the present study in particular for benzo[g,h,i]pyrelene, which was scarcely detected in the water phase of WAFOIL+D (54.2 ng L−1 at 60 h, see Table S4) but in concentrations up to 215 ng g−1 in the WAF OIL+D fish tissues.

Mortality and peculiar behaviour

The mortality observed in the chemically dispersed oil WAF exposure tanks might have been caused by the increased concentration of PAHs compared to the undispersed WAFs. Despite the diversity of oils or chemical dispersants and the variety of tested dilutions, which limits a direct comparison between studies, similar results of elevated toxicity after the application of a chemical dispersant have been reported for zebrafish [75, 101]. In addition, the results of no mortality in 5% and 25% WAF approaches were in compliance to a previous study with zebrafish exposed to the same oil type even in higher exposure concentrations compared to the present WAFs [10].

The administration of the exposure media containing dispersant triggered a certain cyclical stunned behaviour in exposed fish of occupying the upper part of the tank for 1 d every time the solution was renewed. This behaviour correlated with the measured PAH concentration in the tanks, which sharply decreased after 24 h. Acute and chronic exposure to PAHs found in high concentrations in our study (fluorene, phenanthrene and pyrene) has been shown to reduce locomotion, induce lethargy in fish and increase boldness behaviour in exploration tests [47, 50]. In addition, the exposure to water-soluble fractions of native and chemically dispersed crude oil significantly reduced the critical swimming speed of fish in a concentration-dependent manner [135]. Moreover, the mixture of oil and the chemical dispersant FINASOL OSR 52 induced several alterations on fish exploration behaviour [7]. The fish posture description indicated by Aimon et al. [7] of still and anaesthetic-like sedative state agrees with our observations.

Different studies link altered swimming behaviour caused by oil constituents to cardiac alterations as an initiating adverse effect leading to altered metabolic rates and the observed behaviour [77, 124]. Another explanation might be an avoidance behaviour of fish due to odor and medium turbidity in combination with hypoxia. Though oxygen was not limited in the water phase (data not shown), chemical dispersants have been found to affect gill ion regulation and overall increase gill lesions in oil WAF exposed fish indicating the disturbance of oxygen transport [6, 39]. Gill membrane functionality could also simply be affected by the physical accumulation or disruption of oil droplets and hydrophobic oil constituents such as the high molecular weight PAHs. However, this previously suggested baseline toxicity of membranes (narcosis) [37, 84] is controversially discussed [54]. It is important to consider behavioural disruptions, since they may imply severe consequences to fish survival and reproductive fitness.

Biotransformation response

Biotransformation biomarkers have been applied in several studies focusing on the toxicity pathways behind PAH exposure in zebrafish [53, 64]. Representative for biotransformation phase I oxygenase activity, mRNA levels of cyp1a1 and protein levels of EROD activity have been found to be up-regulated in fish after the exposure to individual PAHs, such as phenanthrene [80, 97]. Besides, Arukwe et al. [10] observed an increment in cyp1a1 mRNA levels and EROD activity in a comparable experiment to the present study applying the identical oil type and zebrafish. In our study, the biotransformation activities were not affected by exposure to the two dilutions of undispersed WAF, emphasizing again the role of the dispersant for the toxicity of crude oil by enhancing the bioavailability of PAHs. A concurrent induction of both mRNA and enzymatic levels of CYP1A suggests a de novo synthesis of the protein most likely initiated by AhR activation by typical receptor agonists [54, 78]. In addition, in accordance with the present findings, up-regulation of phase I biotransformation strongly initiated by the application of a chemical dispersant has been observed in several previous studies [5, 29, 101, 105] including in particular experiments with the presently used FINASOL OSR 52 [35, 41], which has also mainly been related to the increment of PAHs in the exposure solution [105].

After the transformation of PAHs to more hydrophilic compounds by phase I enzymes, the resultant molecules experience a conjugation by phase II enzymes, such as GST, in order to promote its excretion [57]. In this context, levels of gstp1 were also significantly increased at both sampling timepoints in fish exposed to the chemically dispersed crude oil. However, the enzymatic activity of GST in liver was not correspondent with the increased transcription, but its lack of significance could be related to instable measurement on protein level with not all replicates meeting the established quality criteria. The discrepancy could also be related to the endpoints itself, since previous studies concluded that the induction responses of phase II systems, such as GST activity, are less uniform and, therefore, less reliable compared to phase I enzymes [68, 121]. In this context, while different PAHs have been shown to increase GST activity in fish [33, 53, 58, 92], other studies found that PAHs decreased GST activity [80, 98, 129], some showed contrary results of both inhibition and induction [116], or even no alterations of GST activity by PAH exposure [68] in liver, gills or kidney of fish.

As the most immediate organ in proximity with the exposure media and, therefore, the compounds present in it, gills can rapidly metabolize PAHs before reaching the liver and thus playing an important role in the biotransformation process [64]. Hence, an elevated biotransformation activity might be expected in gills. However, when comparing both liver and gill tissues, no consistent trend of biotransformation activity was observed. Even though its induction has been described as a biomarker for crude oil contaminants in fish [2, 3, 64], in our study, the EROD activity in gills of fish exposed to dispersed oil was nearly three times lower than in livers and not significantly different from control. In contrast, for GST activity we observed a significantly higher induction only in gills. These findings of variable biotransformation activities in different tissues highlight the complexity of biological responses and in particular the sensitivity of liver samples in chronic exposure studies with petroleum constituents.

Oxidative stress response

Transcriptional and biochemical results of the present study indicate that the chemically dispersed oil induced oxidative stress in fish, which is in compliance with the overall knowledge on WAF-dominating PAHs inducing ROS [78, 92, 111, 116, 118]. For catalase, no direct correlation between transcriptomic and enzymatic responses could be observed. The significant increase in CAT activity in gills but not in liver in our study indicates that gills represent an important organ for detoxification acting as a primary defence line for fish. The lack of CAT induction in liver might be explained by its suppression or impairment of the system caused by a severe oxidative damage, as indicated in a previous study [116]. Overall, several studies indicate that CAT could play a secondary role on enzymatic defence against the oxidative stress [33, 98]. The varying results after oil WAFs or PAH exposure have led to the assumption that CAT is a less stable biomarker and rather enzymes such as GPX are principal actors responding to ROS, such as H2O2 [33, 91, 98]. In this context, zebrafish embryos exposed to a WAF of Arabian Light crude oil for 96 h showed the same pattern as our results, with an upregulation of gpx and sod genes but no regulation of cat [100]. However, while other studies on different fish species observed comparable trends [11, 97], Milinkovitch [87, 89] did not report significant differences of CAT, SOD and GPx activities in fish exposed to chemically dispersed crude oil. All these different results reflect how the antioxidant response is conditioned by the organ or species studied and, certainly, by the conditions of the exposure to crude oil compounds. Our findings support the higher relevance of SOD and GPx as oxidative stress biomarker over CAT due to the concomitant and strong transcriptomic induction of sod1a and gpx1a. In addition, the high EROD activity observed in the present study is in line with oxidative stress, as superoxide anion radicals (O2•−) are a type of ROS that can be produced by a non-well-functioning process of the phase I mixed function oxygenase system [121]. Moreover, Luch [78] indicates that some PAHs can generate superoxide or hydroxyl radicals through autoxidation or PAH-mediated inflammatory processes. Thus, the increase in cellular superoxide (O2•−) production would be appeased by the SOD activity generating in return an excess of H2O2 that GPX enzymes would transform into water using the reductive power of a glutathione (GSH) molecule [116, 121]. Further studies on GSH content, SOD and GPX activity or the assessment of lipid peroxidation could aid to prove these assumptions.

Cellular stress including oxidative stress and DNA damage can lead to the activation of cell cycle arrest, programmed cell death (apoptosis), and repair pathways [38]. Associated marker genes were not significantly regulated in the present study independent of the exposure solution or sampling timepoint. As previously reviewed, PAHs are well known for their potential to induce apoptosis in fish [9]. In this context, p53 acting as a key transcription factor to initiate the different cascades including extrinsic apoptosis pathways involving caspases [9] has been found to be strongly upregulated by PAH [136] or petroleum oil exposure [30, 109]. In contrast, also a decrease of p53 activity compared to control has been reported for adult fish after dietary exposure to BaP [131]. It has to be considered that only two marker genes were investigated in the present study, which do obviously not cover the whole complexity of p53-regulated pathways including apoptosis.

AChE activity

While previously the brain AChE activity has not been found to be altered in wild fish populations of heavily petroleum contaminated areas [67], the present study observed a significant AChE inhibition after both short- and elongated exposures to chemically dispersed crude oil compared to the same dilution of undispersed WAFOIL but not when compared to the control. In addition to complex petroleum mixtures, also individual oil constituents had led to ambiguous results regarding AChE inhibitory potential in fish [96]. Selected PAHs (e.g., phenanthrene) were found to inhibit either isolated AChE enzymes from fish [60] or brain homogenates [59] in exposure concentrations comparable to the present study, whereas another study reported no influence of PAHs on AChE activity [117]. In addition, our previous experiments with early life stages of zebrafish exposed to WAFs prepared from the identical NNS crude oil batch did not indicate AChE inhibition [62], even though embryonic stages have been described to be reliably responsive to neurotoxic exposure [123, 134]. However, since a high variability of AChE activity across the different treatments was observed in the current study, the observed AChE inhibition might be related to a generally reduced fitness of fish. Nonetheless, neurotoxicity of crude oil cannot be excluded simply by the measurement of one enzyme involved in neurotransmission. Previous studies focusing on whole transcriptome analysis, physiological changes in brain tissue, monoamine neurotransmitter regulation, and locomotor behaviour observed that petroleum exposure can affect the neuronal system in fish early life stages [45, 125, 133], indicating a neurotoxic potential of petroleum constituents.


Genotoxicity is a commonly addressed endpoint in crude oil toxicity assessment and can be investigated on several levels of biological organization [1, 20]. The micronucleus assay showing clastogenic or aneugenic chromosomal aberrations is a very sensitive indicator for DNA damage [21, 95]. This endpoint has proven the ability to reveal long-term genotoxic effects in fish several years after large oil spills during past decades in various laboratory and monitoring studies [23, 102, 103]. In the present study, only the chemically dispersed crude oil exposure resulted in significantly elevated micronuclei formation. This might be related to the increased bioavailability of higher molecular weight PAHs caused by the dispersion. Low molecular weight PAHs such as the most dominant phenanthrene have inconsistently been reported as inducers for chromosomal aberration. While Peng et al. [99] reported no significant micronuclei formation even for higher exposure concentrations compared to the present study [99], other studies detected significantly increased micronuclei rates for concentration ranges comparable to our work [97, 113]. In contrast, high molecular weight PAHs such as benzo[a]pyrene are well known to initiate a genotoxicity cascade via biotransformation activity (e.g., CYP1A, resulting in the formation of DNA adducts and oxidative radicals that can lead to DNA strand breaks and thus micronuclei formation [79, 108, 119, 138]. Molecular and biochemical results of the present study indicate a high biotransformation activity in oil exposed fish supporting the genotoxicity pathway. Overall, the micronucleus rate observed for 21 d exposed zebrafish exposed to dispersed oil (IF = 2.2) was in the lower range of reported micronucleus rates in erythrocytes found in petroleum product WAF-exposed fish in comparable laboratory studies (IF = 2–5) [14, 22, 85, 122]. Though a comparison using induction factors accounts for inter-species variability [21], a direct comparison of different oil exposure studies is limited due to a variety of selected exposure conditions and oil types. Nonetheless, in a previous study using the identical NNS crude oil batch we also found stronger micronucleus formation in acutely exposed (48 h) permanent zebrafish liver cells [61]. The higher sensitivity of the cell line compared to erythrocytes demonstrates the lack of toxicokinetics in in vitro bioassays and highlights the in situ micronucleus assay as a relevant endpoint for a more reliable ecotoxicological risk assessment of oil WAFs.

Critical considerations for dispersant toxicity interpretation in chronic exposure studies

Overall, our results highlight the importance of considering the effect of chemical dispersants in the toxicity assessment of oil spill response measures, very noticeable when comparing the biological endpoints of identical oil loadings with and without the addition of FINASOL OSR 52. As concluded in several previous studies (e.g., [107]), further supported by the current time-resolved chemical analysis, the dispersant-induced increased toxicity has been attributed to an enhanced bioavailability of toxic compounds present in oil. However, the overall toxicity interpretation should be considered cautiously as oil and dispersant toxicity seems to be species-specific and is further influenced by the exposure methods and oil types. In a worst-case scenario, this might lead to an underestimation of the risk towards more sensitive species compared to zebrafish.

While the present chemical analysis supported the overall theory of a dispersant-induced increase in dissolved PAH fractions, a toxicity of the dispersant itself cannot be excluded with the experimental setup. According to previous studies FINASOL OSR 52 has been considered as moderately toxic [17]. In particular, compared to Corexit EC9500A, another chemical dispersant frequently used as oil spill response measure, a higher toxicity has been found for FINASOL OSR 52 in different fish species (Menidia beryllina, Cyprinodon variegatus) after short and long-term exposure [13, 35]. It is important to mention that an additional control group with the dispersant alone was not included within the current study to experimentally address the role of the dispersant. From our previous studies with the dispersant Finasol OSR 51, which shows high comparability to Finasol OSR 52 used in the present study, we have to emphasize that the toxicity of the dispersant alone or in combination with an inert oil cannot be excluded per se [61,62,63]). Nonetheless, the dispersant-induced general toxicity (oxidative stress induction) or acute toxicity to zebrafish embryos and larvae were found in exposure concentrations highly elevated compared to the exposure concentrations used in the present study. Hence, it could be assumed that in the present scenario no effects from the dispersant itself might be expected.

Another aspect that needs to be considered for toxicity interpretation is the applied dispersant to oil ratio (DOR). While the DOR used in the current study (1:10) is rather recommended for heavy oils due to relatively high viscosity and hence a higher dosage need for dispersion success [74], also other studies working with much lower DORs (e.g., 1:800) observed adverse effects in fish [39]. Hence, the current DOR can be interpreted as a worst-case scenario with environmental relevance. Besides species sensitivity variations and DOR the exposure concentrations need to be critically addressed. The exposure concentrations of the present study can be considered as environmentally relevant. PAH concentrations detected in the water phase (200–5700 ng L−1) were in the lower range of concentrations detected in the water column after oil spills (up to 600,000 ng L−1) as summarized by Perrichon et al. [100]. After the Deep Water Horizon blowout, concentrations up to 189,000 ng L−1 were reported [19]. However, it has to be considered that even though environmentally relevant exposure concentrations were used, the chronic exposure remains a worst-case scenario. A previous study acutely exposing fish to chemically dispersed crude oil in a concentration range of tenfold increase compared to the current setup did conclude no ecological impact after a long-term recovery phase of fish population in semi-natural mesocosms [81]. This might indicate that the observed sublethal effects in the current study are maximum temporary without severe consequences. However, long-term studies are still scarce, and moreover, recovery phases need to be addressed in future research. Especially additional stressors under real field conditions such as ecological or anthropogenic pollution might impact the overall stress tolerance and survival and alter the overall biomarker responses complicating reliable forecasts. In this context, it has been shown that, e.g., global warming-induced temperature increase has a huge impact on acute and chronic stress physiology in fish [8].

Conclusion—combining multilevel endpoints for risk assessment

Due to the high complexity of oil exposure and the high variability of biological responses, it can be concluded that for a reliable ecotoxicological risk assessment, it is important to combine different biomarkers across different levels of biological organization. With significant molecular and enzymatic alterations of biotransformation and antioxidant defences, the present study supports previously reported endpoints for oil toxicity testing as sensitive biomarkers in adult zebrafish. In addition, while the micronucleus induction in peripheral blood erythrocytes is only one apical endpoint in genotoxicity and several other mechanisms can lead to DNA damage, the observed oxidative stress and biotransformation system activation supports the hypothesis of genotoxicity indicated by this biomarker. Comparing biomarker analyses between different studies can be a difficult task due the complexity discussed above. However, it is undeniable that a comprehensive investigation possesses an inherent value. Especially the low exposure concentrations as well as the elongated exposure period up to 21 d are valuable information for the scientific community in order to characterize oil toxicity and can assist decision making for oil spill response strategies. Nonetheless, future experiments should include an elongated recovery phase to take into account an environmentally relevant short-term peak exposure scenario. The present study further emphasizes the importance of a temporally resolved chemical characterization of the exposure water phase for effect interpretation. In addition, further research should different toxicity drivers among oil constituents and study results must be cautiously interpreted leading always to protective actions based on the essential precautionary principle that ensures the environment preservation.

Availability of data and materials

Data of the current study are available from the corresponding author on request.







Dispersant to oil ratio




Gas chromatography–mass spectrometry


Glutathione peroxidase


Glutathione S-transferase


Limit of detection


Naphthenic North Sea (crude oil)


Polycyclic aromatic hydrocarbons


Reactive oxygen species


Stir bar sorptive extraction


Water accommodated fraction


Water accommodated fraction of oil


Water accommodated fraction of chemically dispersed oil


  1. 1.

    Abdel-Massih RM, Melki PN, Afif C et al (2013) Detection of genotoxicity in hospital wastewater of a developing country using SOS chromotest and Ames fluctuation test. J Environ Eng Ecol Sci 2(4):1–8

    Google Scholar 

  2. 2.

    Abrahamson A, Andersson C, Jönsson ME et al (2007) Gill EROD in monitoring of CYP1A inducers in fish—a study in rainbow trout (Oncorhynchus mykiss) caged in Stockholm and Uppsala waters. Aquat Toxicol 85(1):1–8

    CAS  Article  Google Scholar 

  3. 3.

    Abrahamson A, Brandt I, Brunström B et al (2008) Monitoring contaminants from oil production at sea by measuring gill EROD activity in Atlantic cod (Gadus morhua). Environ Pollut 153(1):169–175

    CAS  Article  Google Scholar 

  4. 4.

    Adams J, Sweezey M, Hodson PV (2014) Oil and oil dispersant do not cause synergistic toxicity to fish embryos. Environ Toxicol Chem 33(1):107–114

    CAS  Article  Google Scholar 

  5. 5.

    Adeyemo OK, Kroll KJ, Denslow ND (2015) Developmental abnormalities and differential expression of genes induced in oil and dispersant exposed Menidia beryllina embryos. Aquat Toxicol 168:60–71

    CAS  Article  Google Scholar 

  6. 6.

    Agamy E (2013) Impact of laboratory exposure to light Arabian crude oil, dispersed oil and dispersant on the gills of the juvenile brown spotted grouper (Epinephelus chlorostigma): a histopathological study. Mar Environ Res 86:46–55

    CAS  Article  Google Scholar 

  7. 7.

    Aimon C, Lebigre C, Le Bayon N et al (2021) Effects of dispersant treated oil upon exploratory behaviour in juvenile European sea bass (Dicentrarchus labrax). Ecotoxicol Environ Saf 208:111592

    CAS  Article  Google Scholar 

  8. 8.

    Alfonso S, Gesto M, Sadoul B (2020) Temperature increase and its effects on fish stress physiology in the context of global warming. J Fish Biol 2020:1–13

    Google Scholar 

  9. 9.

    AnvariFar H, Amirkolaie AK, Jalali AM et al (2018) Environmental pollution and toxic substances: cellular apoptosis as a key parameter in a sensible model like fish. Aquat Toxicol 204:144–159

    CAS  Article  Google Scholar 

  10. 10.

    Arukwe A, Nordtug T, Kortner TM et al (2008) Modulation of steroidogenesis and xenobiotic biotransformation responses in zebrafish (Danio rerio) exposed to water-soluble fraction of crude oil. Environ Res 107(3):362–370

    CAS  Article  Google Scholar 

  11. 11.

    Avci A, Kaçmaz M, Durak İ (2005) Peroxidation in muscle and liver tissues from fish in a contaminated river due to a petroleum refinery industry. Ecotoxicol Environ Saf 60(1):101–105

    CAS  Article  Google Scholar 

  12. 12.

    Baird WM, Hooven LA, Mahadevan B (2005) Carcinogenic polycyclic aromatic hydrocarbon-DNA adducts and mechanism of action. Environ Mol Mutagen 45(2–3):106–114

    CAS  Article  Google Scholar 

  13. 13.

    Barron MG, Bejarano AC, Conmy RN et al (2020) Toxicity of oil spill response agents and crude oils to five aquatic test species. Mar Pollut Bull 153:110954

    CAS  Article  Google Scholar 

  14. 14.

    Baršienė J, Dedonytė V, Rybakovas A et al (2006) Investigation of micronuclei and other nuclear abnormalities in peripheral blood and kidney of marine fish treated with crude oil. Aquat Toxicol 78:S99–S104.

    CAS  Article  Google Scholar 

  15. 15

    Baussant T, Sanni S, Jonsson G et al (2001) Bioaccumulation of polycyclic aromatic compounds: 1. Bioconcentration in two marine species and in semipermeable membrane devices during chronic exposure to dispersed crude oil. Environ Toxicol Chem 20(6):1175–1184

    CAS  Article  Google Scholar 

  16. 16

    Baussant T, Sanni S, Skadsheim A et al (2001) Bioaccumulation of polycyclic aromatic compounds: 2. Modeling bioaccumulation in marine organisms chronically exposed to dispersed oil. Environ Toxicol Chem 20(6):1185–1195

    CAS  Article  Google Scholar 

  17. 17.

    Bejarano AC (2018) Critical review and analysis of aquatic toxicity data on oil spill dispersants. Environ Toxicol Chem 37(12):2989–3001

    CAS  Article  Google Scholar 

  18. 18.

    Bender ML, Frantzen M, Camus L et al (2018) Effects of acute exposure to dispersed oil and burned oil residue on long-term survival, growth, and reproductive development in polar cod (Boreogadus saida). Mar Environ Res 140:468–477

    CAS  Article  Google Scholar 

  19. 19.

    Beyer J, Trannum HC, Bakke T et al (2016) Environmental effects of the Deepwater Horizon oil spill: a review. Mar Pollut Bull 110(1):28–51

    CAS  Article  Google Scholar 

  20. 20.

    Bickham JW, Sandhu S, Hebert PD et al (2000) Effects of chemical contaminants on genetic diversity in natural populations: implications for biomonitoring and ecotoxicology. Mutat Res Rev Mutat Res 463(1):33–51

    CAS  Article  Google Scholar 

  21. 21.

    Bolognesi C, Hayashi M (2011) Micronucleus assay in aquatic animals. Mutagenesis 26(1):205–213

    CAS  Article  Google Scholar 

  22. 22.

    Bolognesi C, Perrone E, Roggieri P et al (2006) Assessment of micronuclei induction in peripheral erythrocytes of fish exposed to xenobiotics under controlled conditions. Aquat Toxicol 78:S93–S98

    CAS  Article  Google Scholar 

  23. 23.

    Bolognesi C, Perrone E, Roggieri P et al (2006) Bioindicators in monitoring long term genotoxic impact of oil spill: Haven case study. Mar Environ Res 62:S287–S291

    CAS  Article  Google Scholar 

  24. 24.

    Bornstein JM, Adams J, Hollebone B et al (2014) Effects-driven chemical fractionation of heavy fuel oil to isolate compounds toxic to trout embryos. Environ Toxicol Chem 33(4):814–824

    CAS  Article  Google Scholar 

  25. 25.

    Brinkmann M, Hudjetz S, Cofalla C et al (2010) A combined hydraulic and toxicological approach to assess re-suspended sediments during simulated flood events. Part I–multiple biomarkers in rainbow trout. J Soils Sediment 10(7):1347–1361

    CAS  Article  Google Scholar 

  26. 26.

    Carrasco KR, Tilbury KL, Myers MS (1990) Assessment of the piscine micronucleus test as an in situ biological indicator of chemical contaminant effects. Can J Fish Aquat Sci 47(11):2123–2136

    CAS  Article  Google Scholar 

  27. 27.

    Chen H, Sheng L, Gong Z et al (2018) Investigation of the molecular mechanisms of hepatic injury upon naphthalene exposure in zebrafish (Danio rerio). Ecotoxicology 27(6):650–660

    CAS  Article  Google Scholar 

  28. 28.

    Claiborne A (1985) Catalase activity. In: Greenwald RA (ed) CRC Handbook of methods for oxygen radical research. CRC Press, Boca Raton, pp 283–284

    Google Scholar 

  29. 29.

    Cohen C, Gagnon MM, Nugegoda D (2006) Oil spill remediation techniques can have different impacts on mixed function oxygenase enzyme activities in fish. Bull Environ Contam Toxicol 76(5):855

    CAS  Article  Google Scholar 

  30. 30.

    Costa PM, Miguel C, Caeiro S et al (2011) Transcriptomic analyses in a benthic fish exposed to contaminated estuarine sediments through laboratory and in situ bioassays. Ecotoxicology 20(8):1749–1764

    CAS  Article  Google Scholar 

  31. 31.

    Dai Y, Jia Y, Chen N et al (2014) Zebrafish as a model system to study toxicology. Environ Toxicol Chem 33(1):11–17

    CAS  Article  Google Scholar 

  32. 32.

    Dasgupta S, Choyke S, Ferguson PL et al (2018) Antioxidant responses and oxidative stress in sheepshead minnow larvae exposed to Corexit 9500® or its component surfactant, DOSS. Aquat Toxicol 194:10–17

    CAS  Article  Google Scholar 

  33. 33.

    Dasgupta S, DiGiulio RT, Drollette BD et al (2016) Hypoxia depresses CYP1A induction and enhances DNA damage, but has minimal effects on antioxidant responses in sheepshead minnow (Cyprinodon variegatus) larvae exposed to dispersed crude oil. Aquat Toxicol 177:250–260

    CAS  Article  Google Scholar 

  34. 34.

    De Soysa TY, Ulrich A, Friedrich T et al (2012) Macondo crude oil from the Deepwater Horizon oil spill disrupts specific developmental processes during zebrafish embryogenesis. BMC Biol 10(1):40

    Article  Google Scholar 

  35. 35.

    DeLorenzo ME, Key PB, Chung KW et al (2018) Comparative toxicity of two chemical dispersants and dispersed oil in estuarine organisms. Arch Environ Contam Toxicol 74(3):414–430

    CAS  Article  Google Scholar 

  36. 36

    Di Giulio RT, Hinton DE (2008) The toxicology of fishes. Crc Press Taylor and Francis group, Boca Raton

    Book  Google Scholar 

  37. 37.

    Di Toro DM, McGrath JA, Stubblefield WA (2007) Predicting the toxicity of neat and weathered crude oil: toxic potential and the toxicity of saturated mixtures. Environ Toxicol Chem 26:24–36

    Article  Google Scholar 

  38. 38.

    Dos Santos N, Vale Ad, Reis M et al (2008) Fish and apoptosis: molecules and pathways. Curr Pharm Des 14(2):148–169

    Article  Google Scholar 

  39. 39.

    Duarte RM, Honda RT, Val AL (2010) Acute effects of chemically dispersed crude oil on gill ion regulation, plasma ion levels and haematological parameters in tambaqui (Colossoma macropomum). Aquat Toxicol 97(2):134–141

    CAS  Article  Google Scholar 

  40. 40.

    Dupuis A, Ucán-Marín F (2015) A literature review on the aquatic toxicology of petroleum oil: an overview of oil properties and effects to aquatic biota. DFO Can Sci Advis Sec Res Doc 6:52

    Google Scholar 

  41. 41.

    Dussauze M, Camus L, Le Floch S et al (2014) Impact of dispersed fuel oil on cardiac mitochondrial function in polar cod Boreogadus saida. Environ Sci Pollut Res 21(24):13779–13788

    CAS  Article  Google Scholar 

  42. 42.

    Dussauze M, Danion M, Le Floch S et al (2015) Innate immunity and antioxidant systems in different tissues of sea bass (Dicentrarchus labrax) exposed to crude oil dispersed mechanically or chemically with Corexit 9500. Ecotoxicol Environ Saf 120:270–278

    CAS  Article  Google Scholar 

  43. 43.

    Ellman GL, Courtney KD, Andres V Jr et al (1961) A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem Pharmacol 7(2):88–95

    CAS  Article  Google Scholar 

  44. 44.

    Fingas M (2011) Chapter 15-oil spill dispersants: a technical summary. In: Fingas M (ed) Oil spill science and technology. Gulf Professional Publishing, Boston, pp 435–582

    Chapter  Google Scholar 

  45. 45.

    Gao D, Wu M, Wang C et al (2015) Chronic exposure to low benzo[a]pyrene level causes neurodegenerative disease-like syndromes in zebrafish (Danio rerio). Aquat Toxicol 167:200–208

    CAS  Article  Google Scholar 

  46. 46.

    George-Ares A, Clark JR (2000) Aquatic toxicity of two Corexit® dispersants. Chemosphere 40(8):897–906

    CAS  Article  Google Scholar 

  47. 47.

    Gonçalves R, Scholze M, Ferreira AM et al (2008) The joint effect of polycyclic aromatic hydrocarbons on fish behavior. Environ Res 108(2):205–213

    Article  CAS  Google Scholar 

  48. 48.

    González JF, Reimschuessel R, Shaikh B et al (2009) Kinetics of hepatic phase I and II biotransformation reactions in eight finfish species. Mar Environ Res 67(4–5):183–188

    Article  CAS  Google Scholar 

  49. 49.

    Habig WH, Pabst MJ, Jakoby WB (1974) Glutathione S-transferases: the first enzymatic step in mercapturic acid formation. J Biol Chem 249(22):7130–7139

    CAS  Article  Google Scholar 

  50. 50.

    Hamilton TJ, Krook J, Szaszkiewicz J et al (2021) Shoaling, boldness, anxiety-like behavior and locomotion in zebrafish (Danio rerio) are altered by acute benzo[a]pyrene exposure. Sci Total Environ 774:145702

    CAS  Article  Google Scholar 

  51. 51.

    Hansen BH, Salaberria I, Read KE et al (2019) Developmental effects in fish embryos exposed to oil dispersions—the impact of crude oil micro-droplets. Mar Environ Res 150:104753

    CAS  Article  Google Scholar 

  52. 52.

    Hodson PV (2017) The toxicity to fish embryos of PAH in crude and refined oils. Arch Environ Contam Toxicol 73(1):12–18

    CAS  Article  Google Scholar 

  53. 53.

    Huang L, Zuo Z, Zhang Y et al (2015) Toxicogenomic analysis in the combined effect of tributyltin and benzo[a]pyrene on the development of zebrafish embryos. Aquat Toxicol 158:157–164

    CAS  Article  Google Scholar 

  54. 54.

    Incardona JP (2017) Molecular mechanisms of crude oil developmental toxicity in fish. Arch Environ Contam Toxicol 73(1):19–32

    CAS  Article  Google Scholar 

  55. 55.

    Incardona JP, Swarts TL, Edmunds RC et al (2013) Exxon Valdez to Deepwater Horizon: comparable toxicity of both crude oils to fish early life stages. Aquat Toxicol 142:303–316

    Article  CAS  Google Scholar 

  56. 56.

    Incardona JP, Vines CA, Anulacion BF et al (2012) Unexpectedly high mortality in Pacific herring embryos exposed to the 2007 Cosco Busan oil spill in San Francisco Bay. Proc Natl Acad Sci USA 109 (2):E51–E58

    CAS  Article  Google Scholar 

  57. 57.

    Jakoby WB (1978) The glutathione S-transferases: a group of multifunctional detoxification proteins. Adv Enzymol Relat Areas Mol Biol 46:383–414

    CAS  Google Scholar 

  58. 58.

    Jee JH, Kang JC (2005) Biochemical changes of enzymatic defense system after phenanthrene exposure in olive flounder, Paralichthys olivaceus. Physiol Res 54(585):e591

    Google Scholar 

  59. 59.

    Jee J, Kang J (2003) Effects of phenanthrene exposure on the acetylcholinesterase activity of olive flounder (Paralichthys olivaceus). Fish Aquat Sci 6(4):225–227

    CAS  Google Scholar 

  60. 60.

    Jett DA, Navoa RV, Lyons MA Jr (1999) Additive inhibitory action of chlorpyrifos and polycyclic aromatic hydrocarbons on acetylcholinesterase activity in vitro. Toxicol Lett 105(3):223–229

    CAS  Article  Google Scholar 

  61. 61.

    Johann S, Goßen M, Behnisch PA et al (2020) Combining different in vitro bioassays to evaluate genotoxicity of water-accommodated fractions from petroleum products. Toxics 8(2):45

    CAS  Article  Google Scholar 

  62. 62.

    Johann S, Nüßer L, Goßen M et al (2020) Differences in biomarker and behavioral responses to native and chemically dispersed crude and refined fossil oils in zebrafish early life stages. Sci Total Environ 709:136174

    CAS  Article  Google Scholar 

  63. 63.

    Johann S, Esser M, Nüßer L et al (2020) Receptor-mediated estrogenicity of native and chemically dispersed crude oil determined using adapted microscale reporter gene assays. Environ Int 134:105320.

    CAS  Article  Google Scholar 

  64. 64.

    Jönsson ME, Brunström B, Brandt I (2009) The zebrafish gill model: induction of CYP1A, EROD and PAH adduct formation. Aquat Toxicol 91(1):62–70

    Article  CAS  Google Scholar 

  65. 65.

    Jørgensen KS, Kreutzer A, Lehtonen KK et al (2019) The EU Horizon 2020 project GRACE: integrated oil spill response actions and environmental effects. Environ Sci Eur 31(1):1–10

    Article  Google Scholar 

  66. 66.

    Jung J, Hicken CE, Boyd D et al (2013) Geologically distinct crude oils cause a common cardiotoxicity syndrome in developing zebrafish. Chemosphere 91(8):1146–1155

    CAS  Article  Google Scholar 

  67. 67.

    Jung J, Kim M, Yim UH et al (2011) Biomarker responses in pelagic and benthic fish over 1 year following the Hebei Spirit oil spill (Taean, Korea). Mar Pollut Bull 62(8):1859–1866

    CAS  Article  Google Scholar 

  68. 68.

    Kopecka-Pilarczyk J, Correia AD (2009) Biochemical response in gilthead seabream (Sparus aurata) to in vivo exposure to a mix of selected PAHs. Ecotoxicol Environ Saf 72(4):1296–1302

    CAS  Article  Google Scholar 

  69. 69.

    Kuhl AJ, Nyman JA, Kaller MD et al (2013) Dispersant and salinity effects on weathering and acute toxicity of South Louisiana crude oil. Environ Toxicol Chem 32(11):2611–2620

    CAS  Google Scholar 

  70. 70.

    Kujawinski EB, Kido Soule MC, Valentine DL et al (2011) Fate of dispersants associated with the Deepwater Horizon oil spill. Environ Sci Technol 45(4):1298–1306

    CAS  Article  Google Scholar 

  71. 71.

    Lacave JM, Bilbao E, Gilliland D et al (2020) Bioaccumulation, cellular and molecular effects in adult zebrafish after exposure to cadmium sulphide nanoparticles and to ionic cadmium. Chemosphere 238:124588

    CAS  Article  Google Scholar 

  72. 72.

    Lacave JM, Vicario-Parés U, Bilbao Eet al (2018) Waterborne exposure of adult zebrafish to silver nanoparticles and to ionic silver results in differential silver accumulation and effects at cellular and molecular levels. Sci Total Environ 642:1209–1220

    CAS  Article  Google Scholar 

  73. 73.

    Le Bihanic F, Morin B, Cousin X et al (2014) Developmental toxicity of PAH mixtures in fish early life stages. Part I: adverse effects in rainbow trout. Environ Sci Pollut Res. 21(24):13720-13731.

    Article  Google Scholar 

  74. 74.

    Le Floch S (2016) The use of dispersants to combat oil spills in Germany at sea. Federal Institute for Risk Assessment Press and Public Relations, Berlin

    Google Scholar 

  75. 75.

    Li X, Ding G, Xiong Y et al (2018) Toxicity of water-accommodated fractions (WAF), chemically enhanced WAF (CEWAF) of Oman crude oil and dispersant to early-life stages of zebrafish (Danio rerio). Bull Environ Contam Toxicol 101:314–319

    CAS  Article  Google Scholar 

  76. 76.

    Logan DT (2007) Perspective on ecotoxicology of PAHs to fish. Hum Ecol Risk Assess 13(2):302–316

    CAS  Article  Google Scholar 

  77. 77.

    Lucas J, Percelay I, Larcher T et al (2016) Effects of pyrolytic and petrogenic polycyclic aromatic hydrocarbons on swimming and metabolic performance of zebrafish contaminated by ingestion. Ecotoxicol Environ Saf 132:145–152

    CAS  Article  Google Scholar 

  78. 78.

    Luch A (2005) Nature and nurture—lessons from chemical carcinogenesis. Nat Rev Cancer 5(2):113–125

    CAS  Article  Google Scholar 

  79. 79.

    Łuczyński MK, Góra M, Brzuzan P et al (2005) Oxidative metabolism, mutagenic and carcinogenic properties of some polycyclic aromatic hydrocarbons. Environ Biotechnol 1(1):16–28

    Google Scholar 

  80. 80.

    Mai Y, Peng S, Li H et al (2019) Histological, biochemical and transcriptomic analyses reveal liver damage in zebrafish (Danio rerio) exposed to phenanthrene. Comp Biochem Physiol 225:108582

    CAS  Google Scholar 

  81. 81.

    Mauduit F, Domenici P, Farrell AP et al (2016) Assessing chronic fish health: an application to a case of an acute exposure to chemically treated crude oil. Aquat Toxicol 178:197–208

    CAS  Article  Google Scholar 

  82. 82.

    Martin JD, Adams J, Hollebone B et al (2014) Chronic toxicity of heavy fuel oils to fish embryos using multiple exposure scenarios. Environ Toxicol Chem 33(3):677–687

    CAS  Article  Google Scholar 

  83. 83.

    Martínez-Álvarez I, Le Menach K, Devier MH et al (2021) Uptake and effects of graphene oxide nanomaterials alone and in combination with polycyclic aromatic hydrocarbons in zebrafish. Sci Total Environ 775:145669

    Article  CAS  Google Scholar 

  84. 84.

    Meador JP, Nahrgang J (2019) Characterizing crude oil toxicity to early-life stage fish based on a complex mixture: are we making unsupported assumptions? Environ Sci Technol 53(19):11080–11092

    CAS  Article  Google Scholar 

  85. 85.

    Medeiros LCC, Delunardo FAC, Simões LN et al (2017) Water-soluble fraction of petroleum induces genotoxicity and morphological effects in fat snook (Centropomus parallelus). Ecotoxicol Environ Saf 144:275–282

    CAS  Article  Google Scholar 

  86. 86.

    Mielbrecht EE, Wolfe MF, Tjeerdema RS et al (2005) Influence of a dispersant on the bioaccumulation of phenanthrene by topsmelt (Atherinops affinis). Ecotoxicol Environ Saf 61(1):44–52

    CAS  Article  Google Scholar 

  87. 87.

    Milinkovitch T, Godefroy J, Théron M et al (2011) Toxicity of dispersant application: biomarkers responses in gills of juvenile golden grey mullet (Liza aurata). Environ Pollut 159(10):2921–2928

    CAS  Article  Google Scholar 

  88. 88.

    Milinkovitch T, Kanan R, Thomas-Guyon H et al (2011) Effects of dispersed oil exposure on the bioaccumulation of polycyclic aromatic hydrocarbons and the mortality of juvenile Liza ramada. Sci Total Environ 409(9):1643–1650

    CAS  Article  Google Scholar 

  89. 89.

    Milinkovitch T, Ndiaye A, Sanchez W et al (2011) Liver antioxidant and plasma immune responses in juvenile golden grey mullet (Liza aurata) exposed to dispersed crude oil. Aquat Toxicol 101(1):155–164

    CAS  Article  Google Scholar 

  90. 90.

    Molinier V, Le Goué E, Rondón-Gonzáles M et al (2018) Optimization of chemical dispersants effectiveness in case of subsurface oil spill. Colloids Surf A Physicochem Eng Asp 541:43–51

    CAS  Article  Google Scholar 

  91. 91.

    Nahrgang J, Camus L, Carls MG et al (2010) Biomarker responses in polar cod (Boreogadus saida) exposed to the water soluble fraction of crude oil. Aquat Toxicol 97(3):234–242

    CAS  Article  Google Scholar 

  92. 92.

    Nahrgang J, Camus L, Gonzalez P et al (2009) PAH biomarker responses in polar cod (Boreogadus saida) exposed to benzo (a) pyrene. Aquat Toxicol 94(4):309–319

    CAS  Article  Google Scholar 

  93. 93.

    Navarro P, Etxebarria N, Arana G (2009) Development of a focused ultrasonic-assisted extraction of polycyclic aromatic hydrocarbons in marine sediment and mussel samples. Anal Chim Acta 648(2):178–182

    CAS  Article  Google Scholar 

  94. 94.

    Nigro M, Frenzilli G, Scarcelli V et al (2002) Induction of DNA strand breakage and apoptosis in the eel Anguilla anguilla. Mar Environ Res 54(3–5):517–520

    CAS  Article  Google Scholar 

  95. 95.

    Obiakor MO, Okonkwo JC, Nnabude PC et al (2012) Eco-genotoxicology: micronucleus assay in fish erythrocytes as in situ aquatic pollution biomarker: a review. J Anim Sci Adv 2(1):123–133

    Google Scholar 

  96. 96.

    Olivares-Rubio HF, Espinosa-Aguirre JJ (2020) Acetylcholinesterase activity in fish species exposed to crude oil hydrocarbons: a review and new perspectives. Chemosphere 264(1):128401

    Google Scholar 

  97. 97.

    Oliveira M, Pacheco M, Santos MA (2007) Cytochrome P4501A, genotoxic and stress responses in golden grey mullet (Liza aurata) following short-term exposure to phenanthrene. Chemosphere 66(7):1284–1291

    CAS  Article  Google Scholar 

  98. 98.

    Oliveira M, Pacheco M, Santos MA (2008) Organ specific antioxidant responses in golden grey mullet (Liza aurata) following a short-term exposure to phenanthrene. Sci Total Environ 396(1):70–78

    CAS  Article  Google Scholar 

  99. 99.

    Peng C, Muthusamy S, Xia Q et al (2015) Micronucleus formation by single and mixed heavy metals/loids and PAH compounds in HepG2 cells. Mutagenesis 30(5):593–602

    CAS  Article  Google Scholar 

  100. 100.

    Perrichon P, Le Menach K, Akcha F et al (2016) Toxicity assessment of water-accommodated fractions from two different oils using a zebrafish (Danio rerio) embryo-larval bioassay with a multilevel approach. Sci Total Environ 568:952–966

    CAS  Article  Google Scholar 

  101. 101.

    Philibert DA, Lyons D, Philibert C et al (2019) Field-collected crude oil, weathered oil and dispersants differentially affect the early life stages of freshwater and saltwater fishes. Sci Total Environ 647:1148–1157

    CAS  Article  Google Scholar 

  102. 102.

    Pilcher W, Miles S, Tang S et al (2014) Genomic and genotoxic responses to controlled weathered-oil exposures confirm and extend field studies on impacts of the Deepwater Horizon oil spill on native killifish. PLoS ONE 9(9):e106351

    Article  CAS  Google Scholar 

  103. 103.

    Pietrapiana D, Modena M, Guidetti P et al (2002) Evaluating the genotoxic damage and hepatic tissue alterations in demersal fish species: a case study in the Ligurian Sea (NW-Mediterranean). Mar Pollut Bull 44(3):238–243

    CAS  Article  Google Scholar 

  104. 104.

    Prieto A, Zuloaga O, Usobiaga A et al (2007) Development of a stir bar sorptive extraction and thermal desorption–gas chromatography–mass spectrometry method for the simultaneous determination of several persistent organic pollutants in water samples. J Chromatogr A 1174(1–2):40–49

    CAS  Article  Google Scholar 

  105. 105.

    Ramachandran SD, Hodson PV, Khan CW et al (2004) Oil dispersant increases PAH uptake by fish exposed to crude oil. Ecotoxicol Environ Saf 59(3):300–308

    CAS  Article  Google Scholar 

  106. 106.

    Ramachandran SD, Sweezey MJ, Hodson PV et al (2006) Influence of salinity and fish species on PAH uptake from dispersed crude oil. Mar Pollut Bull 52(10):1182–1189

    CAS  Article  Google Scholar 

  107. 107.

    Redman AD, Parkerton TF (2015) Guidance for improving comparability and relevance of oil toxicity tests. Mar Pollut Bull 98(1):156–170

    CAS  Article  Google Scholar 

  108. 108.

    Regoli F, Gorbi S, Frenzilli G et al (2002) Oxidative stress in ecotoxicology: from the analysis of individual antioxidants to a more integrated approach. Mar Environ Res 54(3–5):419–423

    CAS  Article  Google Scholar 

  109. 109.

    Ruiz P, Orbea A, Rotchell JM et al (2012) Transcriptional responses of cancer-related genes in turbot Scophthalmus maximus and mussels Mytilus edulis exposed to heavy fuel oil no. 6 and styrene. Ecotoxicology 21(3):820–831

    CAS  Article  Google Scholar 

  110. 110.

    Scott JA, Incardona JP, Pelkki K et al (2011) AhR2-mediated, CYP1A-independent cardiovascular toxicity in zebrafish (Danio rerio) embryos exposed to retene. Aquat Toxicol 101(1):165–174

    CAS  Article  Google Scholar 

  111. 111.

    Shi H, Sui Y, Wang X et al (2005) Hydroxyl radical production and oxidative damage induced by cadmium and naphthalene in liver of Carassius auratus. Comp Biochem Physiol 140C(1):115–121

    CAS  Google Scholar 

  112. 112.

    Shimada T (2006) Xenobiotic-metabolizing enzymes involved in activation and detoxification of carcinogenic polycyclic aromatic hydrocarbons. Drug Metab Pharmacokinet 21(4):257–276

    CAS  Article  Google Scholar 

  113. 113.

    Shirmohammadi M, Salamat N, Ronagh MT et al (2017) Using cell apoptosis, micronuclei and immune alternations as biomarkers of phenanthrene exposure in yellowfin seabream (Acanthopagrus latus). Fish Shellfish Immunol 72:37–47

    Article  CAS  Google Scholar 

  114. 114.

    Singer MM, Aurand D, Bragin GE et al (2000) Standardization of the preparation and quantitation of water-accommodated fractions of petroleum for toxicity testing. Mar Pollut Bull 40(11):1007–1016

    CAS  Article  Google Scholar 

  115. 115.

    Spaulding ML (2017) State of the art review and future directions in oil spill modeling. Mar Pollut Bull 115(1):7–19

    CAS  Article  Google Scholar 

  116. 116.

    Sun Y, Yu H, Zhang J et al (2006) Bioaccumulation, depuration and oxidative stress in fish Carassius auratus under phenanthrene exposure. Chemosphere 63(8):1319–1327

    CAS  Article  Google Scholar 

  117. 117.

    Tang Y, Donnelly KC, Tiffany-Castiglioni E et al (2003) Neurotoxicity of polycyclic aromatic hydrocarbons and simple chemical mixtures. J Toxicol Environl Health A 66(10):919–940

    CAS  Article  Google Scholar 

  118. 118.

    Timme-Laragy AR, Van Tiem LA, Linney EA et al (2009) Antioxidant responses and NRF2 in synergistic developmental toxicity of PAHs in zebrafish. Toxicol Sci 109(2):217–227

    CAS  Article  Google Scholar 

  119. 119.

    Tung EWY, Philbrook NA, Belanger CL et al (2014) Benzo[a]pyrene increases DNA double strand break repair in vitro and in vivo: A possible mechanism for benzo[a]pyrene-induced toxicity. Mutat Res Gen Toxicol Environ Mutagen 760:64–69

    CAS  Article  Google Scholar 

  120. 120.

    Valencia A, Andrieu J, Nzioka A et al (2020) Transcription pattern of reproduction relevant genes along the brain-pituitary-gonad axis of female, male and intersex thicklip grey mullets, Chelon labrosus, from a polluted harbor. Gen Comp Endocrinol 287:113339

    CAS  Article  Google Scholar 

  121. 121.

    Van der Oost R, Beyer J, Vermeulen NP (2003) Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environ Toxicol Pharmacol 13(2):57–149

    Article  Google Scholar 

  122. 122.

    Vanzella TP, Martinez CB, Cólus IM (2007) Genotoxic and mutagenic effects of diesel oil water soluble fraction on a neotropical fish species. Mutat Res 631(1):36–43

    CAS  Article  Google Scholar 

  123. 123.

    Velki M, Meyer-Alert H, Seiler T et al (2017) Enzymatic activity and gene expression changes in zebrafish embryos and larvae exposed to pesticides diazinon and diuron. Aquat Toxicol 193:187–200

    CAS  Article  Google Scholar 

  124. 124.

    Vignet C, Le Menach K, Lyphout L et al (2014) Chronic dietary exposure to pyrolytic and petrogenic mixtures of PAHs causes physiological disruption in zebrafish—part II: behavior. Environ Sci Pollut Res 21(24):13818–13832

    CAS  Article  Google Scholar 

  125. 125.

    Vignet C, Trenkel VM, Vouillarmet A et al (2017) Changes in brain monoamines underlie behavioural disruptions after zebrafish diet exposure to polycyclic aromatic hydrocarbons environmental mixtures. Int J Mol Sci 18(3):560

    Article  CAS  Google Scholar 

  126. 126.

    Wang H, Li Y, Xia X et al (2018) Relationship between metabolic enzyme activities and bioaccumulation kinetics of PAHs in zebrafish (Danio rerio). J Environ Sci 65:43–52

    Article  Google Scholar 

  127. 127.

    Wang H, Xia X, Liu R et al (2019) Multicompartmental toxicokinetic modeling of discrete dietary and continuous waterborne uptake of two polycyclic aromatic hydrocarbons by zebrafish Danio rerio. Environ Sci Technol 54(2):1054–1065

    Article  CAS  Google Scholar 

  128. 128.

    Wang M, Zhu G, Milkov AV et al (2020) Comprehensive molecular compositions and origins of DB301 crude oil from Deep Strata, Tarim Basin, China. Energy Fuels 34(6):6799–6810

    CAS  Article  Google Scholar 

  129. 129.

    Wang Y, Shen C, Wang C et al (2018) Maternal and embryonic exposure to the water soluble fraction of crude oil or lead induces behavioral abnormalities in zebrafish (Danio rerio), and the mechanisms involved. Chemosphere 191:7–16

    CAS  Article  Google Scholar 

  130. 130.

    Wernersson A, Carere M, Maggi C et al (2015) The European technical report on aquatic effect-based monitoring tools under the water framework directive. Environ Sci Eur 27(1):1–11

    CAS  Article  Google Scholar 

  131. 131.

    Williams R, Hubberstey AV (2014) Benzo(a)pyrene exposure causes adaptive changes in p53 and CYP1A gene expression in Brown bullhead (Ameiurus nebulosus). Aquat Toxicol 156:201–210

    CAS  Article  Google Scholar 

  132. 132.

    Whyte JJ, Jung RE, Schmitt CJ et al (2000) Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit Rev Toxicol 30:347–570

    CAS  Article  Google Scholar 

  133. 133.

    Xu EG, Khursigara AJ, Magnuson J et al (2017) Larval red drum (Sciaenops ocellatus) sublethal exposure to weathered Deepwater Horizon crude oil: developmental and transcriptomic consequences. Environ Sci Technol 51(17):10162–10172

    CAS  Article  Google Scholar 

  134. 134.

    Yen J, Donerly S, Levin ED et al (2011) Differential acetylcholinesterase inhibition of chlorpyrifos, diazinon and parathion in larval zebrafish. Neurotoxicol Teratol 33(6):735–741

    CAS  Article  Google Scholar 

  135. 135.

    Yu X, Xu C, Liu H et al (2015) Effects of crude oil and dispersed crude oil on the critical swimming speed of puffer fish Takifugu rubripes. Bull Environ Contam Toxicol 94(5):549–553

    CAS  Article  Google Scholar 

  136. 136.

    Yuan L, Lv B, Zha J et al (2017) Benzo[a]pyrene induced p53-mediated cell cycle arrest, DNA repair, and apoptosis pathways in Chinese rare minnow (Gobiocypris rarus). Environ Toxicol 32(3):979–988

    CAS  Article  Google Scholar 

  137. 137.

    Zahed MA, Aziz HA, Isa MH et al (2010) Effect of initial oil concentration and dispersant on crude oil biodegradation in contaminated seawater. Bull Environ Contam Toxicol 84(4):438–442

    CAS  Article  Google Scholar 

  138. 138.

    Żelazna K, Rudnicka K, Tejs S (2011) In vitro micronucleus test assessment of polycyclic aromatic hydrocarbons. Environ Biotechnol 7:70–80

    Google Scholar 

  139. 139.

    Zhai Y, Xia X, Xiong X et al (2018) Role of fluoranthene and pyrene associated with suspended particles in their bioaccumulation by zebrafish (Danio rerio). Ecotoxicol Environ Saf 157:89–94

    CAS  Article  Google Scholar 

  140. 140.

    Zhang B, Matchinski EJ, Chen B et al (2019) Chapter 21-Marine oil Spills—Oil Pollution, Sources and Effects. In: Zhang B (ed) World Seas: an environmental evaluation 2nd edition, volume III: ecological issues and environmental impacts. Elsevier, Amsterdam

    Google Scholar 

Download references


Thanks to staff at Driftslaboratoriet Mongstad, Equinor (former Statoil) for supplying the oil. The authors gratefully acknowledge Dr. Xabier Lekube at Plentzia Marine Station (PIE-UPV/EHU) for handling oil samples and Dr. José M. Lacave for help with the experimental setup and support in sample preparation and processing, as well as other personnel from CBET research group for assistance during sample collection. The authors would like to kindly thank Nikon (Nikon GmbH) for their contribution to this study as partner of the Students Lab “Fascinating Environment” at RWTH Aachen Biology and Biotechnology (ABBt) providing the imaging technology for erythrocyte evaluation.


Open Access funding enabled and organized by Projekt DEAL. The present study was funded by EU H2020-BG-2005-2 project GRACE (grant agreement #679266), Spanish MINECO (NACE project CTM2016-81130-R) and MECD (FPU16/01837 grant to A.E.), Basque Government (consolidated research groups IT810-13, IT1302-19, IT1213-19).

Author information




A.E and S.J equally contributed to the execution of the experiment, sample handling, acquisition, analysis and interpretation of data and manuscript writing, D.B and A.P performed chemical analysis of exposure media and fish tissue, H.H revised the manuscript along with T.B.S and A.O who contributed to the conception and design of the work. All authors read and approved the final manuscript.

Corresponding author

Correspondence to Sarah Johann.

Ethics declarations

Ethics approval and consent to participate

The present study was conducted with the approval of the Ethics Committee in Animal Experimentation of the University of the Basque Country (expedient number NoRefCEID: M20/2017/173) and authorization of the Local Government of Biscay according to current regulations.

Consent for publication

Not applicable.

Competing interests

The authors declare that they have no competing interests.

Additional information

Publisher's Note

Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.

Supplementary Information

Additional file 1:

Table S1. List of Abbreviations and limits of detection of target PAHs analyzed in exposure solutions and carcasses. Table S2. Concentrations of target PAHs (ng/L) in 5% WAFOIL exposure media over two cycles (exposure media was changed every 72 h) and analyzed using GC–MS. Values below detection limits are expressed as BDL. Table S3. Concentrations of target PAHs (ng/L) in 25% WAFOIL exposure media over two cycles (exposure media was changed each 72 h) and analyzed using GC–MS. Values below detection limits are expressed as BDL. Note that the value at 6 h for the first cycle is missing due to technical problems. Table S4. Concentrations of target PAHs (ng/L) in 5% WAFOIL+D exposure media over two cycles (exposure media was changed each 72 h) and analyzed using GC–MS. Values below detection limits are expressed as BDL. Note that the value at 24 h for the first cycle and at 72 h for the second cycle are missing due to technical problems. Table S5. Concentrations of PAHs (ng/g) in fish carcass tissue after 21 days of exposure to WAF dilutions of NNS crude oil. Fish were exposed to WAF dilutions of crude oil only (WAFOIL) and chemically dispersed crude oil (FINASOL OSR 52, WAFOIL+D). Values below detection limits are expressed as BDL. Table S6. Relative biochemical biomarker activity in brain, liver and gills of adult zebrafish after short- (3 d) and long-term (21 d) exposure to WAF dilutions of NNS crude oil. Fish were exposed to WAF dilutions of crude oil only (WAFOIL) and chemically dispersed (FINASOL OSR 52) crude oil (WAFOIL+D). Data represent the mean relative induction (IF) of 7-ethoxyresorufin-O-deethylase (EROD), glutathione S-transferase (GST), catalase (CAT) and acetylcholinesterase (AChE) in tissue homogenates of 4 replicates per treatment (1 replicate = pool of 3 individual tissues) compared to unexposed control with standard deviation. Table S7. Relative micronucleus frequency (IF) in erythrocytes from adult zebrafish blood smears after short- (3 d) and long-term (21 d) exposure to WAF dilutions of an NNS crude oil compared to unexposed to control. Data represent the mean IF of 7–10 individual fish exposed to WAF dilutions of crude oil only (WAFOIL) and chemically dispersed crude oil (FINASOL OSR 52, WAFOIL+D) with standard deviation (SD).

Rights and permissions

Open Access This article is licensed under a Creative Commons Attribution 4.0 International License, which permits use, sharing, adaptation, distribution and reproduction in any medium or format, as long as you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons licence, and indicate if changes were made. The images or other third party material in this article are included in the article's Creative Commons licence, unless indicated otherwise in a credit line to the material. If material is not included in the article's Creative Commons licence and your intended use is not permitted by statutory regulation or exceeds the permitted use, you will need to obtain permission directly from the copyright holder. To view a copy of this licence, visit

Reprints and Permissions

About this article

Verify currency and authenticity via CrossMark

Cite this article

Esteban-Sánchez, A., Johann, S., Bilbao, D. et al. Multilevel responses of adult zebrafish to crude and chemically dispersed oil exposure. Environ Sci Eur 33, 106 (2021).

Download citation


  • Bioaccumulation
  • Biomarkers
  • Biotransformation
  • Chemical dispersants
  • Crude oil
  • Genotoxicity
  • Oil spills
  • Oxidative stress
  • Water accommodated fractions
  • Zebrafish