Does the presence of microplastics influence the acute toxicity of chromium(VI) to early juveniles of the common goby (Pomatoschistus microps)? A study with juveniles from two wild estuarine populations

Aquatic Toxicology 164 (2015) 163–174

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Does the presence of microplastics influence the acute toxicity of chromium(VI) to early juveniles of the common goby (Pomatoschistus microps)? A study with juveniles from two wild estuarine populations Luís G. Luís a,b , Pedro Ferreira a,b , Elsa Fonte a,b , Miguel Oliveira a,b , Lúcia Guilhermino a,b,∗ a ICBAS – Institute of Biomedical Sciences of Abel Salazar, University of Porto, Department of Population Studies, Laboratory of Ecotoxicology, Rua de Jorge Viterbo Ferreira, 228, 4050-313 Porto, Portugal b CIIMAR/CIMAR-LA – Interdisciplinary Centre of Marine and Environmental Research, Research Group of Ecotoxicology, Stress Ecology and Environmental Health, University of Porto, Rua dos Bragas, 289, 4050-123 Porto, Portugal

a r t i c l e

i n f o

Article history: Received 2 September 2014 Received in revised form 9 April 2015 Accepted 15 April 2015 Available online 5 May 2015 Keywords: Microplastics Cr(VI) Combined effects Pomatoschistus microps Predatory performance Biomarkers

a b s t r a c t Toxicological interactions between microplastics (MP) and other environmental contaminants are of grave concern. Here, the potential influence of MP in the short-term toxicity of chromium to early juveniles of Pomatoschistus microps was investigated. Three null hypotheses were tested: (1) exposure to Cr(VI) concentrations in the low ppm range does not induce toxic effects on juveniles; (2) the presence of microplastics in the water does not influence the acute toxicity of Cr(VI) to juveniles; (3) the environmental conditions of the natural habitat where fish developed do not influence their sensitivity to Cr(VI)-induced acute stress. Fish were collected in the estuaries of Minho (M-est) and Lima (L-est) Rivers (NW Iberian Peninsula) that have several abiotic differences, including in the water and sediment concentrations of various environmental contaminants. After acclimatization to laboratory conditions, two 96 h acute bioassays were carried out with juveniles from both estuaries to: (i) investigate the effects of Cr(VI) alone; (ii) investigate the effects of Cr(VI) in the presence of MP (polyethylene spheres 1–5 ␮m ∅). Cr(VI) alone induced mortality (96 h-LC50 s: 14.4–30.5 mg/l) and significantly decreased fish predatory performance (≤74%). Thus, in the range of concentrations tested (5.6–28.4 mg/l) Cr(VI) was found to be toxic to P. microps early juveniles, therefore, we rejected hypothesis 1. Under simultaneous exposure to Cr(VI) and MP, a significant decrease of the predatory performance (≤67%) and a significant inhibition of AChE activity (≤31%) were found. AChE inhibition was not observed in the test with Cr(VI) alone and MP alone caused an AChE inhibition ≤21%. Mixture treatments containing Cr(VI) concentration ≥3.9 mg/l significantly increased LPO levels in L-est fish, an effect that was not observed under Cr(VI) or MP single exposures. Thus, toxicological interactions between Cr(VI) and MP occurred, therefore, we rejected hypothesis 2. In the presence of MP, the negative effect caused by high concentrations of Cr(VI) on the predatory performance was significantly reduced in L-est fish but not in M-est fish, and Cr(VI) concentrations higher than 3.9 mg/l caused oxidative damage in L-est fish but not in M-est fish. The acclimatization and test conditions were similar for fish from the two estuaries and these ecosystems have environmental differences. Thus, long-term exposure to distinct environmental conditions in the natural habitat during previous developmental phases influenced the sensitivity and responses of juveniles to Cr(VI), therefore, we rejected hypothesis 3. Overall, the results of this study indicate toxicological interactions between MP and Cr(VI) highlighting the importance of further investigating the combined effects of MP and other common contaminants. © 2015 Elsevier B.V. All rights reserved.

1. Introduction ∗ Corresponding author at: ICBAS – Institute of Biomedical Sciences of Abel Salazar, University of Porto, Department of Population Studies, Laboratory of Ecotoxicology, Rua de Jorge Viterbo Ferreira, 228, 4050-313 Porto, Portugal. Tel.: +351 220428000. E-mail addresses: [email protected] (L.G. Luís), [email protected] (P. Ferreira), [email protected] (E. Fonte), [email protected] (M. Oliveira), [email protected] (L. Guilhermino). http://dx.doi.org/10.1016/j.aquatox.2015.04.018 0166-445X/© 2015 Elsevier B.V. All rights reserved.

Over decades, plastics of different types and sizes have been intensively and widely used by the human society for several purposes. Today, plastics have most important applications in modern medicine and technology, and are components of goods and equipment used daily all over the world (e.g. food and drink packages,

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plastic bags, clothes, shoes, computers). In 2013, the estimated global plastic production was 288 million tonnes (PlasticsEurope, 2013), with previsions for further global market growth and expansion. In general, plastics have a long environmental life and have been accumulating in the environment (Andrady, 2011; Cole et al., 2011; Galgani et al., 2013; Wright et al., 2013), despite the efforts to reduce their environmental inputs. In the marine environment, plastics are in general slowly and gradually broken into smaller fragments as the result of several processes, eventually becoming microscopic sized particles known as microplastics (Andrady, 2011; Cole et al., 2011; Wright et al., 2013). Because they result from the fragmentation of larger plastics, such particles are often called secondary microplastics to distinguish them from those specifically manufactured to have a microscopic (or lower) size which are called primary microplastics (Andrady, 2011; Cole et al., 2011; Wright et al., 2013). Primary microplastics are components of a wide range of products used for several purposes, such as scrubbers in personal care products and air-blasting technology, vectors for drug delivery in medicine, abrasives in maritime industry, among others (Andrady, 2011; Derraik, 2002; Desforges et al., 2014; Fendall and Sewell, 2009; Gregory, 1996; von Moos et al., 2012). As the result of their manufacture and use, they enter the marine environment by direct input, through rivers, soil runoff, and components of industrial, urban and domestic effluents (Andrady, 2011; Browne et al., 2011; Derraik, 2002; Fendall and Sewell, 2009; Galgani et al., 2013). Either primary or secondary microplastics have been found everywhere in the marine environment, including in remote regions, reaching the highest concentrations in ocean gyres, and in highly anthropogenically impacted estuaries and other coastal areas (Antunes et al., 2013; Ballent et al., 2013; Barnes et al., 2009; Carpenter and Smith, 1972; Carpenter et al., 1972; Collignon et al., 2012; Dantas et al., 2012; Desforges et al., 2014; Doyle et al., 2011; Ivar do Sul et al., 2009; Martins and Sobral., 2011; Morét-Ferguson et al., 2010). Microplastics may be ingested by several marine organisms, including invertebrates (Besseling et al., 2013; Browne et al., 2008; Cole et al., 2013; Lee et al., 2013; von Moos et al., 2012; Setälä et al., 2014), fish (Boerger et al., 2010; Lusher et al., 2013; Possatto et al., 2011), and birds (van Franeker et al., 2011). Evidences also suggest the uptake of microplastics by the Mediterranean fin whale (Baleanoptera physalus) (Fossi et al., 2012). Moreover, recent studies demonstrated that microplastics can be transferred from lower to higher food web levels (Setälä et al., 2014). In marine organisms, microplastics have been found to cause physical adverse effects, such as false food satiation, damage of gills and other organs, predatory performance and efficiency reduction, and other adverse effects potentially leading to death (Besseling et al., 2013; Lee et al., 2013; von Moos et al., 2012; Oliveira et al., 2013; Sá et al., 2015; Wright et al., 2013). Microplastics may contain toxic chemicals incorporated during their manufacture (e.g. bisphenol A, phthalates, nonylphenol, polybrominated diphenyl esters), use (e.g. metals incorporated during microplastic use as air-blasting media) and/or permanence in the environment (e.g. metals, polycyclic aromatic hydrocarbons, polychlorinated biphenyls) (Rios et al., 2007; Rochman et al., 2014). These chemicals may be transferred to organisms after ingestion of microplastics, to higher trophic levels along with microplastics (Setälä et al., 2014), and/or be accumulated and biomagnified into trophic chains (Fossi and Depledge, 2014) increasing the risk of toxic effects on top predators and humans consuming contaminated species. Therefore, microplastics are now considered ubiquitous pollutants of high concern, their concentrations in the marine environment should be monitored in the scope of the European Marine Strategy Framework Directive, and their effects on the marine biota should be further investigated (Fossi and Depledge, 2014; Galgani et al., 2013; Wright et al., 2013).

In several regions around the world, estuaries and coastal areas are under strong anthropogenic pressure being contaminated by complex mixtures of persistent organic pollutants (POPs), metals, several other known and unknown compounds, and microplastics. Thus, organisms inhabiting these ecosystems, as well as humans consuming contaminated species, are likely to be simultaneously exposed to microplastics and other environmental contaminants. Recently, interactions between microplastics and other common environmental contaminants were found (Holmes et al., 2014; Oliveira et al., 2013) and more knowledge on such interactions are needed to improve environmental and human risk assessments. Chromium is an ubiquitous environmental contaminant that is toxic to the biota (Ahmed et al., 2013; Domingues et al., 2010; Mishra and Mohanty, 2012; Sadeghi et al., 2014) and humans (Wu and Liu, 2014) at ecologically relevant concentrations. It may be accumulated in different tissues (Fatima et al., 2014), and some of its species (e.g. Cr(VI)) are carcinogenic (IARC, 1990). It is a common contaminant of anthropogenically impacted estuaries and other coastal areas (e.g. Guimarães et al., 2012; Torres et al., 2015) where microplastics are also important contaminants (Wright et al., 2013). Therefore, simultaneous exposure of the biota to chromium and microplastics in real scenarios is likely to occur. Microplastics are able to accumulate chromium over time (Holmes et al., 2012, 2014; Rochman et al., 2014). Therefore, organisms inhabiting chromium-contaminated areas may be exposed to chromium through ingestion of microplastics or uptake by other ways, such as from water, sediments and/or contaminated preys. Recent studies showed that chromium adsorption to polyethylene pellets in seawater follows a pseudo-first-order reaction, with a very rapid initial phase followed by a second one leading to equilibrium, and that aged pellets accumulate more chromium than virgin ones (Holmes et al., 2012). Likely, chromium adsorption occurs through interactions between chromium oxyanions with polar or charged sites of the pellet surface (Holmes et al., 2012). Several factors may influence the process, including abiotic conditions, aging of microplastics, plastic additives, biofilms at plastic surface, and other contaminants present in the environment (Holmes et al., 2012, 2014; Rochman et al., 2014). Chromium was found to be accumulated by several types of plastics (Rochman et al., 2014), increasing the concerns on the impacts of the phenomena in real scenarios. After ingestion of chromium-contaminated microplastics by marine organisms, the metal may be released from microplastics (Holmes et al., 2012) becoming more available to induce toxic effects. Considering the gaps of the knowledge still existing (Holmes et al., 2012), their ecological importance and their potential implications to human health, especially through the consumption of microplastics and chromium-contaminated marine edible species, more studies are clearly needed. The main goals of this study were to investigate if the presence of microplastics in the water can interact with the acute toxicity of Cr(VI) to juveniles (early 0+ age group) of the common goby (Pomatoschistus microps) and if the environmental conditions of the original habitat where fish developed may influence their response to chromium induced stress. The following null hypotheses were tested: (H01 ) exposure to Cr(VI) concentrations in the water (low ppm range) does not induce toxic effects on P. microps early juveniles; (H02 ) the presence of microplastics in the water (ppb range) does not influence the acute toxicity of chromium to P. microps early juveniles; and (H03 ) long-term exposure to different environmental conditions during pre-developmental phases in the natural habitat does not influence the sensitivity and responses of juveniles to Cr(VI)-induced acute stress. P. microps was selected as model species for this study mainly because it is an important intermediary predator in several estuaries and other coastal areas of Europe and North Africa (Mehner, 1992; Leitão et al., 2006), including some located in urban and

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industrial regions where microplastics and chromium are important environmental contaminants, and is a suitable species for both field biomonitoring and laboratory toxicity testing (Monteiro et al., 2005, 2007; Guimarães et al., 2012; Vieira et al., 2009; Oliveira et al., 2013). To investigate the third hypothesis, the effects induced by the environmental contaminants on fish from the estuaries of Minho and Lima Rivers (NW Iberian Peninsula) were compared. These populations were selected because they have been studied for several years (e.g. Monteiro et al., 2007; Guimarães et al., 2012), and the Minho and Lima estuaries are neighbouring ecosystems with several comparable hydromorphic characteristics but with important differences in the levels of some environmental contaminants and other abiotic factors (Guimarães et al., 2012). Polyethylene spheres were selected as microplastics model for this study because this polymer is one of the most produced and used globally (PlasticsEurope, 2013), and is one of the predominant types of microplastics found in the marine environment (Andrady, 2011). Moreover, they were found to be a convenient microplastic model in a previous study with juveniles of the common goby (Oliveira et al., 2013), and the interactions of chromium with polyethylene pellets were previously investigated (Holmes et al., 2012; Rochman et al., 2014).

part), the L-est is more contaminated than the corresponding area of the M-est, mainly due to the presence of an harbour and a paper mill, among other sources of pollution (Guimarães et al., 2012). In a previous study with P. microps juveniles, older than those used in the present study, a poorer health condition of the L-est population in relation to the M-est population was found (Guimarães et al., 2012). After capture, fish were immediately transported to the laboratory in thermally isolated boxes containing water from the corresponding sampling site with aeration. They were progressively acclimatized to artificial salt water (ASW) for three weeks. ASW was prepared by dissolving ocean fish salt in distilled water, with salinity adjusted to 18 ± 1 g/l, water temperature of 20 ± 1 ◦ C, and a 16 h light (L): 8 h dark (D) photoperiod. Fish were maintained in 65 l glass aquaria containing aerated and filtered (Eheim filters) ASW with about 80–100 fish per aquarium. Water was renewed 3 times a week. Water temperature and other parameters to control the adequacy of the conditions for fish were measured at the time of water renewal. Fish were fed ad libitum twice a day with commercial fish food (Tropic Mix Aquapex Produtos, Portugal). The mortality recorded during the acclimation period was lower than 10%. Dead fish were removed as soon as they were detected.

2. Material and methods

2.4. Determination of actual concentrations of Cr(VI) and its potential decay in test media

2.1. Chemicals MP (polyethylene fluorescent microspheres) were purchased from Cospheric – Innovations in Microtechnology (USA). Fluorescent MPs were selected because this property allows the determination of their concentrations in test media in a costeffective way. According to the manufacturer, MP are spherical and opaque, have 1–5 ␮m diameter and a density of 1.2 g/cc; excitation wavelength of 365, 460 and 470 nm and emission wavelength of 588 nm. Potassium dichromate was used as Cr(VI) source. It was ≥99.9% pure and purchased from Merck (Germany). The salt (ocean fish salt, Prodac) used to prepare the artificial sea water (ASW) was from Tropical Marine Centre (Italy). The chemicals used in biomarker determinations were all of the highest purity available and purchased from Sigma–Aldrich (USA) or Merck (Germany). The Bradford reagent used for protein determinations was from Biorad (Germany). 2.2. Ethical issues Fish capture, transport and maintenance in the laboratory, and bioassays were carried out according to the legal and ethical requirements of Portuguese and European regulations, except regarding the use of chemical anaesthetics that were not used to avoid potential interactions with biomarker determinations (fish were sacrificed by decapitation under ice-cold induced anaesthesia). L. Guilhermino is accredited as investigator/coordinator (equivalent to FELASA category C) to carry animal experimentation by the Portuguese National Animal Health Authority. 2.3. Fish collection, transport and acclimation to laboratory conditions Early 0+ age group P. microps juveniles were collected in the spring 2013 in specific areas of the estuary of the Minho River (Mest) and the estuary of the Lima River (L-est) (NW Iberian coast), at low tide with a hand-operated net. These neighbouring estuaries with some comparable hydromorphological characteristics were selected because they present several differences in environmental conditions, including in the levels of diverse environmental contaminants. In the areas where fish were collected (lower estuarine

The actual concentrations of Cr(VI) in test solutions were determined by spectrophotometry (Sena et al., 2000), using Spectra Max M2E spectrophotometer. For this, three independent initial solutions of potassium dichromate with a nominal Cr(VI) concentration of 40 mg/l were prepared in ASW (pH 8.4, salinity 18). The UV–vis spectra of these solutions were made to determine the predominant species of Cr(VI) present in ASW. Each initial solution was serial diluted in ASW 1:2 (v/v) to obtain solutions with concentrations of 20.0, 10.0, 5.0, 2.5, 1.3 and 0.6 mg/l and their absorbance was read at 370 nm (the highest absorbance value obtained in our experimental conditions). The Pearson correlation coefficient was used to measure the correlation between Cr(VI) concentration and absorbance. The absorbance values were then plotted against the corresponding Cr(VI) concentrations and a linear regression model was fitted to the data (independent variable: absorbance values; dependent variable: Cr(VI) concentration) to obtain the absorbance versus concentration calibration curve. Using the equation of the calibration curve, the actual concentrations of Cr(VI) in test solutions were determined at the beginning of the bioassay and the potential decay of test substance along the bioassays was calculated at 0, 24, 48, 72 and 96 h from the absorbance readings at 370 nm. 2.5. Determination of actual concentrations of MP and its potential decay in test media The actual concentrations of MPs and their decay were determined by spectrofluorometry (Jasco FP-6200 spectrofluorometer). Six independent colloidal solutions of MP in ASW with a concentration of 12 mg/l were prepared. Each solution was serially diluted 1:2 (v/v) in ASW to obtain additional solutions with concentrations between 12 mg/l and 0.012 mg/l. The fluorescence of blank (ASW only) and of MP solutions were read at 470/588 nm (excitation and emission wavelength, respectively). After discounting the blank values, the fluorescence values were plotted against the corresponding nominal concentrations of MP. The Pearson correlation coefficient was used to measure the correlation between the two variables, and a linear regression model was fitted to the data (independent variable: fluorescence values; dependent variable: MP concentrations). The actual MP concentrations in test solutions were determined at the beginning of the MP/chromium bioassay

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from the equation of the regression model. The potential decay of MP solutions in test media along the bioassays was assessed directly from the fluorescence after discounting the fluorescence of blanks. 2.6. Bioassays and exposure conditions Bioassays were carried out in temperature- and photoperiodcontrolled chambers (Bronson climate) under a 16 h L: 8 h D photoperiod and a water temperature of 20 ± 1 ◦ C. Feeding was stopped 48 h before the beginning of the bioassays. Fish measuring about 1.5–2 cm long (total length) were visually selected. Due to the high sensitivity of early 0+ juveniles to manipulation, fish were only measured (digital calliper) and weighed (Kern ABS-N, Germany) at the end of each bioassay (after the post-exposure predatory performance assay). The exposure period was 96 h, no food was provided, and test medium was not renewed. A first bioassay was carried out to investigate the effects of Cr(VI) alone, including the following treatments: 0 (control, ASW only), 5.6, 8.4, 12.6, 18.9 and 28.4 mg/l of Cr(VI) (nominal concentrations). A second bioassay was performed with the following treatments (MP and Cr(VI) nominal concentrations): control (ASW water only); MP alone (0.184 mg/l); 5.6 mg/l of Cr(VI) + 0.184 mg/l of MP; 8.4 mg/l of Cr(VI) + 0.184 mg/l of MP; 12.6 mg/l of Cr(VI) + 0.184 mg/l of MP; 18.9 mg/l of Cr(VI) + 0.184 mg/l of MP; 28.4 mg/l of Cr(VI) + 0.184 mg/l of MP. These concentrations were selected to investigate potential interactions between microplastics and Cr(VI) and do not intend to be ecologically relevant. The MP concentration was selected based on the results of a previous study with early juveniles of P. microps (Oliveira et al., 2013). In each bioassay, 18 fish (9 from each estuary) were individually exposed in 500 ml of ASW, prepared as previously described (Section 2.3), in 1000 ml glass beakers with continuous aeration. Fish were considered dead when neither opercular movement nor response to mechanical stimuli could be detected. Dead fish were removed as soon as they were noticed. At the end of the exposure period, the predatory performance and biomarkers were assessed as indicated in the Sections 2.7 and 2.8, respectively. Water temperature, pH, conductivity, salinity and dissolved oxygen were measured at time 0, 24, 48, 72 and 96 h (multi-parameter meter HACH, HQ40d; Seawater Refractometer, Hanna, HI96822), and test media samples collected to determine the concentrations of Cr(VI) and/or MP (as indicated in Sections 2.4 and 2.5). 2.7. Post-exposure predatory performance assay At the end of each exposure period, a post-exposure predatory performance assay was carried out using Artemia franciscana nauplii (24 h old) as prey. Each fish was transferred to a prey-exposure chamber (14 cm diameter, 11.5 cm high) containing 300 ml of clean ASW. After 5 min, 12 preys were offered to the fish. The number of preys ingested by the fish in a 3 min time interval was recorded. The fish was then removed and the number of preys remaining in the chamber was counted by visual observation to confirm the number of prey ingested. The fish post-exposure predatory performance, hereafter indicated as predatory performance, was expressed as the percentage of preys ingested relative to the total number of preys offered (12). After the post-exposure predatory performance assay, fish were put back into their original test beakers where they remained for another 2 h. 2.8. Determination of sub-individual biomarkers The number of sub-individual biomarkers used as effect criteria was limited by the small size of the fish (availability of tissues), the availability of fish in the wild, and ethical issues related with the total number of fish sacrificed. Therefore, four widely used biomarkers, involved in mechanisms affecting the survival and

performance of organisms in the wild were used, namely: the activity of acetylcholinesterase (AChE), gluthatione S-transferase (GST) and ethoxyresorufin-O-deethylase (EROD) enzymes; and the levels of lipid peroxidation (LPO). AChE, the enzyme that is responsible for the degradation of the neurotransmissor acetylcholine in cholinergic synapses, was used as indicative of neurotoxicity. EROD was used as indicative of phase I biotransformation changes. GST was used because it is involved in the phase II of biotransformation and also in anti-oxidant defences. LPO levels were selected among other oxidative stress-related parameters because they provide indication of oxidative damage rather than response to oxidative stress only. The concentrations of chromium in the body of fish were not determined because they would have required a considerably higher number of fish to be used and sacrificed. Therefore, considering the ethical and conservational reasons previously indicated, these determinations were not performed. After the 2 h resting period, fish were sacrificed by decapitation under ice-cold induced anaesthesia. Each head was placed in 1 ml of ice cold phosphate buffer (0.1 M, pH 7.2) for posterior AChE determination, and each remaining body was put in 1 ml of icecold phosphate buffer (0.1 M, pH 7.4) for the determination of the other biomarkers. All the samples were homogenised on ice with a Ystral GmbH d-7801 Dottingen homogeniser. Body homogenates were divided in samples for LPO, and EROD and GST determination. All the samples were frozen at −80 ◦ C until further analysis. In the day of biomarker analysis, samples were unfrozen on ice, and the homogenates for enzymatic analysis were centrifuged at 4 ◦ C (Eppendorf 5810R centrifuge): head homogenates were centrifuged at 3300 × g for 3 min; and body homogenates at 10,000 × g for 20 min. The supernatants were carefully collected. AChE was determined in individual head supernatant samples, at 25 ◦ C, assessing the rate of acetylcholine degradation at 412 nm (Bio Tek Power Wave spectrophotometer microplate reader) by the Ellman technique (Ellman et al., 1961) adapted to microplate (Guilhermino et al., 1996), after protein standardization to 0.5 mg/ml (Bradford, 1976; with modifications as indicated in Frasco and Guilhermino, 2002). P. microps head samples prepared as previously indicated contain mainly AChE (Monteiro et al., 2005). EROD activity was determined in pooled samples (3 fish each) of body supernatant at 25 ◦ C in the presence of NADPH using 7-ethoxyresorufin as substrate generally as described by Burke and Mayer (1974), in a Jasco FP-6200 spectrofluorometer at 530/585 excitation/emission wavelength, respectively. GST activity was determined in individual body supernatant samples at 25 ◦ C using 1-chloro-2,4-dinitribenzene as substrate (Habig et al., 1974; with some adaptations as described in Frasco and Guilhermino, 2002) in a Bio Tek Power Wave spectrophotometer microplate reader. LPO levels were determined in body homogenates through the measurement of the thiobarbituric acid reactive substances (TBARS), following Ohkawa (1979) with some modifications (Bird and Draper, 1984; Torres et al., 2002). At the end of LPO level and enzyme activity measurements, the protein content of the samples was determined again (Bradford, 1976; adapted to microplate as indicated in Frasco and Guilhermino, 2002). The enzyme activities and LPO levels were expressed as a function of protein content, i.e. nanomoles per mg of protein (nmol/min/mg protein) for AChE and GST activities; picomoles per mg of protein (pmol/min/mg protein) for EROD activity, and nanomoles of TBARS per mg of protein (nmol TBARS/mg protein) for LPO. 2.9. Data analysis Probit transformed mortality percentages were plotted against the Cr(VI) concentrations to obtain the toxicity curves, and the median lethal concentrations (LC50 ) were determined (Finney,

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Table 1 Mean and standard deviation (within brackets) of the predatory performance, activity of the enzymes acetylcholinesterase (AChE), ethoxyresorufin-O-deethylase (EROD) and glutathione S-transferases (GST), and lipid peroxidation levels (LPO) determined in the control groups of all the bioassays. PRED – predatory performance expressed in percentage (%) of the total number of prey offered. N – number of replicates (pooled samples for EROD – 6 per estuary; individual fish for the other parameters – 18 per estuary). Parameter

N

Fish from the Minho estuary

Fish from the Lima estuary

Students’ t-test

Total length (cm) Weight (mg) PRED (%) AChE (nmol/min/mg protein) EROD (pmol/min/mg protein) GST (nmol/min/mg protein) LPO (nmol TBARS/mg protein)

36

1.6 (±0.29) 55.5 (±34.1) 70.8 (±14.65) 128.7 (±16.56) 0.05 (±0.016) 40.3 (±16.97) 0.44 (±0.16)

1.5 (±0.35) 47.6 (±38.3) 64.8 (±14.45) 100.8 (±16.82) 0.07 (±0.02) 36.7 (±9.51) 0.33 (±0.13)

t34 = 0.938, p > 0.05

36 36 36 12 36 36

1971). The Cr(VI)-induced mortality curves obtained with fish from distinct estuaries were compared using an analysis of covariance (ANCOVA), with the estuary from where fish came from as fixed factor and Cr(VI) concentrations as covariate. Predatory performance data, expressed in percentages, were arcsine transformed (Zar, 1999). Data from enzyme activity and LPO level analysis and were checked for normality of distribution and homogeneity of variances; they were transformed using the log (x + 1) transformation when necessary. Data were then analysed using one-way analysis of variance (ANOVA). The Dunnett’s test or the Tukey test were used to discriminate statistically significant treatments. The Student’s t-test was used to make other comparisons. The SPSS statistical package (version 22) was used for all the statistical analysis. The significance level was set to 0.05. 3. Results and discussion 3.1. General conditions and parameters of fish in the control groups The abiotic factor variation during the bioassays and the mortality recorded in the control groups fulfil the validity criteria of OECD guidelines for acute testing with juvenile fish (OECD,

t34 = 0.516, p > 0.05 t34 = 1.241, p > 0.05 t34 = 5.095, p < 0.05 t10 = −2.503, p < 0.05 t34 = 0.331, p > 0.05 t34 = 2.143, p < 0.05

1992), as indicated and discussed in the Supplementary material (Section S1). The mean and SD of the morphometric parameters, post-exposure predatory performance, LPO levels, and enzymatic activities determined in fish of all the control groups are shown in Table 1. Fish from the L-est showed significantly lower levels of AChE activity and LPO, and higher EROD activity than those of the M-est. No significant differences were found in any of the other parameters. The lower AChE activity of L-est fish relative to M-est fish collected in the spring is in good agreement with the findings of a previous field monitoring study carried out with older P. microps juveniles collected in the same estuaries (Guimarães et al., 2012). It suggests previous long-term developmental exposure to inhibitors of this enzyme with the enzyme activity levels not being completely recovered during the acclimatization period. L-est fish showed significant lower EROD activity than M-est fish. Although these results should be considered carefully because the low enzyme activity levels found in fish from both estuaries, they suggest previous long-term exposure to CYP1A inducers in L-est fish. The L-est has a harbour and a paper mill, among other sources of contamination, and therefore the long-term exposure to environmental contaminants resulting from these industrial activities (e.g. PAHs, PCBs, resins acids and other persistent pollutants) is likely to occur because several of these compounds were found in

Table 2 Mortality (%) induced by Cr(VI) (Cr) and microplastics (MP) alone and in mixture on fish from the estuaries of Minho (M-est) and Lima (L-est) Rivers after 96 h of exposure. The 96 h median lethal concentrations of chromium (LC50 ) with the corresponding 95% confidence intervals within brackets are shown. Ctr – control; mort – mortality (%). Bioassay testing the effects of Cr(VI) alone Ctr

Cr(VI) concentration (mg/l)

Cr(VI) 96 h-LC50

0

5.6

8.4

12.6

18.9

28.4

(mg/l)

M-est fish mort

0

11.1

0

0

22.2

44.4

L-est fish mort

0

33.3

44.4

33.3

44.4

100

30.5 (23.31–62.16) 14.4 (4.30–51.01)

Bioassay testing the effects of Cr(VI) in the presence of MP Ctr

Cr(VI) concentration

+0.216 mg/l of MP

MP alone

Cr(VI) 96 h-LC50

0

1.8

3.9

8.0

18.9

28.4

0.216

(mg/l)

M-est fish mort

0

0

0

11.1

11.1

66.7

44.4

L-est fish mort

0

0

11.1

11.1

22.2

88.9

33.3

25.8 (19.58–43.12) 20.9 (14.50–34.78)

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Table 3 Effects of Cr(VI) alone on the post-exposure predatory performance (PEPP), activity of the enzymes acetylcholinesterase (AChE), ethoxyresorufin-O-deethylase (EROD) and glutathione S-transferase (GST), and lipid peroxidation levels of early Pomatoschitus microps juveniles from the Minho and Lima Rivers estuaries. The analysis were conducted with fish that survived until de end of the bioassay. All the fish from the Lima River estuary exposed to the highest concentration of Cr(VI) tested died before the end of the bioassay and thus no parameters were determined in these fish. For predatory performance, AChE, GST and LPO, the values are the mean of 5–9 individual fish. For EROD, the values are the mean of 3 pooled samples (2–3 fish each). Cr(VI) conc. (mg/l)

0

5.6

8.4

12.6

18.9

28.4

Minho River estuary fish N PEPP (%) AChE activity (nmol/min/mg prot.) EROD activity (pmol/min/mg prot.) GST activity (nmol/min/mg prot.) LPO levels (nmol TBARS/mg prot.)

ANOVA results

9 75.0 (±5.20) 139.9 (±4.23) 0.05 (±0.00) 38.9 (±3.41) 0.4 (±0.06)

8 45.8* (±10.32) 137.0 (±8.36) 0.09 (±0.03) 34.02 (±2.92) 0.6 (±0.16)

9 45.4* (±9.01) 140.3 (±4.54) 0.10 (±0.01) 34.0 (±3.36) 0.6 (±0.18)

9 39.8* (±4.11) 145.1 (±7.23) 0.10 (±0.03) 36.1 (±4.45) 0.4 (±0.05)

7 39.3* (±6.48) 129.3 (±9.94) 0.06 (±0.02) 32.0 (±3.78) 0.5 (±0.05)

5 46.9* (±7.91) 123.2 (±9.07) 0.14 (−) 29.2 (±2.48) 0.5 (±0.01)

Lima River estuary fish N PEPP (%)

9 63.9 (±5.20)

6 40.3* (±8.45)

5 31.7* (±7.17)

6 31.9* (±5.86)

4 16.7* (±4.81)

0 –

F4,25 = 7.491, p < 0.05

AChE activity (nmol/min/mg prot.) EROD activity (pmol/min/mg prot.) GST activity (nmol/min/mg prot.) LPO levels (nmol TBARS/mg prot.)

109.2 (±6.33) 0.07 (±0.01) 33.9 (±1.07) 0.4 (±0.05)

114.4 (±7.99) 0.16 (−) 37.7 (±4.57) 0.6 (±0.06)

111.4 (±7.64) 0.16 (−) 28.5 (±3.79) 0.6 (±0.15)

85.4 (±10.65) 0.10 (−) 36.8 (±5.90) 0.5 (±0.10)

93.7 (±7.78) 0.02 (−) 40.3 (±4.70) 0.5 (±0.12)

–

F4,25 = 2.277, p > 0.05

–

–

–

F4,25 = 1.072, p > 0.05

–

F4,25 = 1.124, p > 0.05

F5,41 = 4.539, p < 0.05 F5,41 = 1.078, p > 0.05 F4,10 = 1.104, p > 0.05 F5,41 = 0.753, p > 0.05 F5,41 = 1.075, p > 0.05

Conc. – concentration; N – number of fish used for the analysis; prot. – protein. (−) less than 3 samples, not included in the statistical analysis. * Indicates statistically significant differences from the control group (ANOVA and the Dunnett’s test, p ≤ 0.05). Standard errors are given within brackets.

sediments, water and fish from this estuary (Guimarães et al., 2012). The lower LPO levels found in L-est fish relative to the mean of Mest juveniles indicates lower lipid damage in L-est fish, possibly as the result of activated anti-oxidative stress defences in response to long-term exposure to oxidants. In summary, the results of Table 1 and the findings of previous studies (Guimarães et al., 2012) indicate that P. microps juveniles from the two estuaries were exposed to different environmental conditions during their previous development in the original habitats. 3.2. H01 : exposure to chromium concentrations in the water (low ppm range) does not induce toxic effects on P. microps early juveniles A representative spectrum of Cr(VI) solutions (40 mg/l) in ASW (Fig. S-1), and the calibration curve for chromium (Fig. S-2) are shown and discussed in the Supplementary material (Section S2). The actual concentrations of Cr(VI) in test media determined at the beginning of the bioassay testing the effects this metal alone on P. microps early juveniles using the linear regression model fitted to calibration curve data (chromium concentration in mg/l = −0.348 + 12.822 × absorbance in D.O. units) are shown in Table S-1 (Supplementary material, Section S2). Because the deviation of the actual concentrations from the corresponding nominal ones was always lower than 20%, the biological results were expressed relatively to nominal concentrations (OECD, 1992). The decay of Cr(VI) in test media was always lower than 20% (Supplementary material, Section S2, Table S-1), thus one can assume that the concentrations of the test substance were satisfactorily maintained during the bioassay (OECD, 1992). In the range of concentrations tested, Cr(VI) alone induced mortality, reaching 44% in juveniles from the M-est, and 100% in fish from the L-est at the highest concentration tested. The calculated 96 h-LC50 s were 30.5 mg/l and 14.4 mg/l for M-est and L-est, respectively (Table 2), with overlapping 95% confidence intervals

(95% CI). Despite this overlap, significant differences between the toxicity curves of fish from the different estuaries were found (ANCOVA, F2,9 = 6.827, p < 0.05) indicating differences of sensitivity to Cr(VI) between M-est and L-est fish. The LC50 s determined in the present study (Table 2) compare with corresponding values found in the literature for other fish species. For example, potassium dichromate 96 h-LC50 s of 35.7 mg/l and 375.8 mg/l were calculated for the freshwater stinging catfish Heteropneustes fossilis (Ahmed et al., 2013), and the common carp Cyprinus carpio (Kumar et al., 2013), respectively, whereas the 96 h-LC50 of Cr(VI) to the zebrafish (Danio rerio) was 39.4 mg/l (Domingues et al., 2010). The effects of Cr(VI) on the predatory performance and selected subindividual biomarkers of early P. microps juveniles are shown in Table 3. Relative to the corresponding control groups, fish from both estuaries exposed to Cr(VI) showed a significantly decreased predatory performance, reaching 47% at the highest concentration tested (28.4 mg/l) in M-est fish, and 74% at 18.9 mg/l in L-est fish (all the L-est fish died when exposed to 28.4 mg/l of Cr(VI)). The decrease of predatory performance is likely to result in difficulties in getting food and thus to decrease individual fitness. Less ingested food may also induce growth and reproduction delay and increase susceptibility to diseases, parasites and predation, with potential negative effects on population fitness. In both M-est fish and L-est fish, Cr(VI) did not induce any significant changes in the other parameters analysed relative to the control groups (Table 3). Therefore, in the range of concentrations tested, Cr(VI) did not cause significant anticholinesterase activity or lipid oxidative damage effects, and GST and EROD enzymes seem not to be significantly involved in the defences against Cr(VI)-induced stress. In fish, Cr(VI) has the ability to enter into the cells and react with intracellular components, liberating unstable radicals which can induce oxidative stress, among other toxic effects (Ahmad et al., 2006). In previous studies with other fish species, Cr(VI) was found to induce oxidative stress after acute exposure (Dogan et al., 2014; Kumar et al., 2013; Pacheco et al., 2013). The role of GST enzymes on the anti-oxidant

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3.3. H02 : the presence of microplastics in the water (ppb range) does not influence the acute toxicity of chromium to P. microps early juveniles The nominal and the actual concentrations of Cr(VI) in test media of the bioassay carried out to investigate the influence of MP on Cr(VI) toxicity to early juveniles of P. microps are shown in Table S-2 (Supplementary material, Section S3). Contrary to what was found in the bioassay testing the effects of Cr(VI) alone where the deviations of actual concentrations from the nominal ones was all ways lower than 20% (Table S-1), higher deviations of actual Cr(VI) from nominal concentrations of 5.6, 8.4 and 12.6 mg/l were found in the presence of MP (Supplementary material, Section S3, Table S-2). Thus, in these cases, the biological results (Table 4, Figs. 1 and 2) were expressed relative to actual Cr(VI) concentrations (1.8, 3.9 and 8 mg/l), keeping the nominal concentrations for the two highest concentrations (18.8 and 28.4 mg/l) because their deviations did not exceed 20% (OECD, 1992). The higher deviations of Cr(VI) actual concentrations from nominal ones found in the mixture bioassay relative to the bioassay testing the effects of the metal alone suggest that at least part of the Cr(VI) bound to MP in test media. Moreover, in the presence of MP, these deviations decreased with the increase of Cr(VI) concentration (Supplementary material, Section S3, Table S-2). The concentration of MP was the same in all treatments and relatively low in comparison to Cr(VI) concentrations. Thus, the decreasing trend of deviations (68.5–8.4%) with the increase of Cr(VI) concentration in test media suggests that the amount of Cr(VI) bound to MP was constant, possibly because MP have a limited capability of binding the metal. In aquatic ecosystems and aqueous solutions, polyethylene plastic debris are able to accumulate chromium, with the accumulation increasing over time and approaching an equilibrium (Holmes et al., 2012, 2014; Rochman et al., 2014). These findings provide support to our hypothesis that some Cr(VI) bound to MP present in test media. Chromium adsorption to MP probably occurs through interactions between chromium oxyanions with polar or charged sites of the MP particle surface (Holmes et al., 2012), with several factors potentially influencing the process, including abiotic conditions,

90

A

Predatory performance (%)

80

Cr Cr+MP

70 60 50 40 30 20 10 0

0

5

10

15

20

25

30

Cr(VI) concentration (mg/l) 80

B

Cr

70

Predatory performance (%)

response of fish to chromium-induced oxidative stress is not yet completely understood. For example, as found in the present study with P. microps (Table 3), no significant changes on GST activity were found in the liver of H. fossilis after 48 h of exposure to Cr(VI) via water (Dogan et al., 2014); however, a significant inhibition of gill GST activity was found in D. rerio exposed for 96 h to Cr(VI) via water (Domingues et al., 2010), while a significant increase of GST activity was found in the liver of the European eel (Anguilla anguilla) after 180 min of Cr(VI) intraperitoneal (i.p.) administration (Pacheco et al., 2013). Although the differences of experimental conditions and tissues analysed make direct comparisons difficult, the role of GST in the response to Cr(VI)-induced oxidative stress in fish may be dependent on the species studied, the exposure time, and the tissue investigated, among other factors. In P. microps, Cr(VI) did not increase LPO levels after 96 h of exposure via water (Table 3), contrary to the increase of LPO levels in A. Anguilla after i.p. administration of Cr(VI). Cr(VI) is able to disrupt cholinergic signalling (Ciacci et al., 2012), e.g. through cholinesterase activity inhibition (Domingues et al., 2013 Guilhermino et al., 1998; Elumalai et al., 2002). In the present study, and despite the decreasing trend of AChE activity with the increase of Cr(VI) concentrations, no significant differences between exposed and non-exposed fish were found (Table 3). Overall, the results of Tables 2 and 3 indicate that short-term exposure of early juveniles of P. microps to water Cr(VI) concentrations in the low ppm range induced toxic effects, namely mortality and predatory performance decrease, leading to the rejection of our first hypothesis (H01 ).

169

Cr+MP

60 50

*

40 30 20 10 0

0

5

10

15

20

25

30

Cr(VI) concentration (mg/l) Fig. 1. Predatory performance (% of Artemia nauplii ingested) of Pomatoschistus microps juveniles from the estuaries of Minho (A) and Lima (B) rivers after 96 h of exposure to microplastics (MP) alone or to the mixture MP + Cr(VI). For comparative purposes the results obtained in the bioassay with Cr(VI) alone are also shown. The values are the mean of 3–9 fish with corresponding standard errors. Cr – effects obtained in the bioassay testing Cr(VI) alone; Cr + MP – results obtained in the bioassay carried out to investigate the effects of MP on the toxicity of Cr(VI); MP – treatment containing microplastics alone. Different letters indicate statistically significant differences (ANOVA and the Tukey test, p < 0.05) among the treatments of the bioassay carried out to investigate the combined effects of MP and Cr(VI). * – indicates statistically significant differences (Student’s t test, p < 0.05) between the treatment of the Cr + MP bioassay and the corresponding treatment of the bioassay testing Cr(VI) alone (Cr).

aging of microplastics, plastic additives, biofilms at plastic surface, and other contaminants present in the environment (Holmes et al., 2012; Rochman et al., 2014). As found for the bioassay with the metal alone, no significant decay of Cr(VI) in test media occurred during the mixture bioassay (Supplementary material, Section 3, Table S-2). These results suggest an initial rapid binding of Cr(VI) to MP with no further significant accumulation of the metal by the plastic particles, thus in good agreement with previous findings by other authors (Holmes et al., 2012). The actual concentrations of MP in test media at the beginning of the mixture bioassay (Supplementary material, Section 3, Table S-3), determined from the linear model fitted to calibration curve data (MP concentrations in mg/l = −0.234 + 0.017 × fluorescence in F units) (Supplementary material, Section 3, Figure S-3) show mean deviations from the nominal concentration (0.184 mg/l) between 18.3 and 44.1%, with no significant correlation with the Cr(VI) concentration (N = 6, r = 0.452, p > 0.05). The presence of MP as a suspension in test media may contribute at least in part for these deviations. The mean concentration calculated using all the MP determinations made was 0.216 mg/l and therefore the biological results are expressed relative to this concentration. The MP decay

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Table 4 Effects of microplastics (MP) alone and Cr(VI) in the presence of MP (0.216 mg/l) on the post-exposure predatory performance PEPP, activity of the enzymes acetylcholinesterase (AChE), ethoxyresorufin-O-deethylase (EROD) and glutathione S-transferase (GST), and lipid peroxidation (LPO) levels of early Pomatoschitus microps juveniles from the estuaries of Minho and Lima rivers. The analysis were conducted with fish that survived until de end of the bioassay (N). For GST and LPO, the values are the mean of 5–9 individual fish. For EROD, the values are the mean of 3 pooled samples (2–3 fish each). Cr(VI) conc. MP conc. (mg/l)

0 0

Minho River estuary fish N 9 PEPP (%) 66.7a (±4.39) 117.4a AChE activity (nmol/min/mg (±3.92) prot.) EROD activity 0.05 (pmol/min/mg (±0.01) prot.) GST activity 41.6 (nmol/min/mg (±7.48) prot.) LPO levels 0.5 (nmol (±0.05) TBARS/mg prot.) Lima River estuary fish N 9 PEPP (%) 65.7a (±4.23) 92.4a AChE activity (nmol/min/mg (±2.99) prot-) EROD activity 0.05 (pmol/min/mg (±0.01) prot.) GST activity 39.4 (pmol/min/mg (±4.28) prot.) LPO levels 0.3a (±0.03) (nmol TBARS/mg prot.)

0 0.216

1.8 0.216

3.9 0.216

8.0 0.216

18.9 0.216

28.4 0.216

5 50.0a,b,c (±5.89) 96.4a,b (±6.84)

9 55.6a,b (±7.48) 97.7a,b (±7.21)

9 45.4a,b,c (±7.49) 87.8b (±9.35)

8 39.6a,b,c (±7.51) 83.1b (±2.95)

8 32.3 b,c (±4.29) 84.8b (±2.61)

3 22.0c (±5.56) 80.7b (±6.41)

F6,44 = 4.167, p < 0.05 F6,44 = 6.735, p < 0.05

0.04 (−)

0.06 (±0.03)

0.06 (±0.02)

0.11 (±0.03)

0.11 (±0.01)

0.08 (±0.01)

F4,10 = 1.738, p > 0.05

27.2 (±4.06)

36.4 (±6.72)

32.0 (±5.95)

32.1 (±4.06)

29.7 (±3.99)

22.9 (±5.50)

F6,44 = 0.811, p > 0.05

0.4 (±0.02)

0.7 (±0.11)

0.8 (±0.15)

0.6 (±0.08)

0.7 (±0.13)

1.0 (±0.13)

F6,44 = 1.891, p > 0.05

6 50.0a,b (±4.23) 73.5b,c (±3.87)

9 55.6a,b (±4.23) 77.8a,b,c (±4.90)

8 49.0a,b (±4.23) 85.7a,b (±3.62)

8 49.0a,b (±4.23) 63.6c (±4.31)

7 34.5b (±4.23) 72.5b,c (±4.45)

1 –

0.03 (−)

0.07 (±0.03)

0.07 (±0.02)

0.09 (±0.03)

0.10 (±0.01)

–

F4,10 = 1.072, p > 0.05

36.3 (±9.53)

41.0 (±6.54)

44.7 (±8.71)

55.7 (±4.74)

45.7 (±7.27)

–

F5,41 = 0.967, p > 0.05

0.4a (±0.05)

0.3a (±0.06)

0.6b (±0.12)

0.6b (±0.13)

0.6b (±0.09)

–

F5,41 = 4.317, p < 0.05

–

ANOVA results

F5,41 = 6.536, p < 0.05 F5,41 = 6.536, p < 0.05

Conc. – concentration; prot. – protein; N – number of fish used for the analysis. (−) less than 3 samples, not included in the statistical analysis; different letters (a,b,. . .) indicate statistically significant differences among treatments, whereas the same letter indicates no statistically significant differences (ANOVA and the Tukey’s test). Standard error bars are given within brackets.

during the assay ranged from 27 to 37% (Supplementary material, Section 3, Table S-3). At least four (not mutually exclusive) hypotheses may be raised to explain this decay: (i) removal from the water column due to aggregation and sedimentation (not visible to the naked eye because the MP amount was very low) and/or binding to the internal surface of the beakers; (ii) MP uptake by fish; (iii) loss of MP fluorescence during the bioassay enabling a proper detection through the method used; and (iv) binding to chromium with the resulting complex(es) showing differences of fluorescence properties and thus not being detected using the excitation and emission wavelengths used. Regardless of the reasons for the decay, the results indicate that if a constant concentration of MP is intended to be tested in further bioassays, test media should be periodically renewed or a flow-through system should be used. The mortality of P. microps juveniles recorded in the mixture bioassay is shown in Table 2. MP alone caused mortality in fish from both estuaries: 33.3% in L-est fish and 44.4% in M-est fish. In the presence of MP, the 96 h-LC50 s of Cr(VI) were 25.8 mg/l for M-est fish and 20.9 mg/l for L-est. Despite the differences found in the Cr(VI) 96 h-LC50 s determined for fish from distinct estuaries, and for juveniles of the same estuary in the absence and presence of MP, all the 95% CI overlap, indicating no significant differences among these parameters. The highest Cr(VI) concentrations tested in the bioassay with the metal alone induced 44.4% and 100% of mortality in M-est and L-est fish, respectively. The corresponding

treatment of the mixture bioassay caused a mortality of 66.7% in Mest fish and 88.9% of mortality in L-est fish. Because the mortality caused by the mixture treatment containing the highest Cr(VI) concentration was lower than the sum of the lethal effects caused by MP and Cr(VI) alone, the interaction evaluated through mortality seems to be antagonism. However, the type of interaction needs to be further investigated using toxicological interaction models. In the continuation of this study, it will be also of high interest to determine the threshold concentration of MP eliciting mortality in P. microps juveniles. The effects on the predatory performance of fish induced by Cr(VI) in the presence of MP are shown in Fig. 1, where the results obtained in fish exposed to the metal alone are also shown for comparative purposes. In fish from both estuaries significant differences of the predatory performance among treatments were found (Table 4). MP alone induced a decrease (about 25%) of the predatory performance of fish from both estuaries but no significant differences relative to the respective control groups were found (Table 4). The mixture treatments containing Cr(VI) concentrations equal or higher than 18.9 mg/l significantly decreased the predatory performance of M-est fish, reaching 67% of reduction at the highest concentration tested (Fig. 1A). In the fish from M-est, the comparison of the predatory performance under exposure to the highest chromium treatments (18.9 and 28.4 mg/l) in the absence and presence of MP indicated no significant differences (18.9 mg/l:

L.G. Luís et al. / Aquatic Toxicology 164 (2015) 163–174

Cr

AChE activity (nmol/min/mg protein)

A

Cr+MP

160

*

140

*

120 100 80 60 40 20 0

0

5

10

15

20

25

30

AChE activity (nmol/min/mg protein)

Cr(VI) concentration (mg/l)

Cr

B

140

Cr+MP

120

*

100 80 60 40 20 0

0

5

10

15

20

25

30

Cr(VI) concentration (mg/l) Fig. 2. Acetylcholinesterase activity of Pomatoschistus microps juveniles from the estuaries of Minho (A) and Lima (B) rivers that survived the 96 h bioassay carried out to investigate the effects of microplastics (MP) on the Cr(VI) toxicity. The values are the mean of 3–9 fish with corresponding standard error bars. One Unity of enzymatic activity = 1 nano mole hydrolized per minute. For comparative purposes the results obtained in the bioassay with Cr(VI) alone are also shown. Different letters indicate statistically significant differences (Tukey’s test, p < 0.05) among treatments of the bioassay carried out to investigate the combined effects of MP and Cr(VI) on P. microps juveniles (Cr + MP). * – indicates statistically significant differences (Student’s t test, p < 0.05) between the treatment of the Cr + MP bioassay and the corresponding treatment of the bioassay testing Cr(VI) alone (Cr).

t13 = 0.786, p > 0.05; 28.4 mg/l: t6 = 0.786, p > 0.05). In L-est fish (Fig. 1B), a significant reduction of the predatory performance was found at 18.9 mg/l of Cr(VI) (47%) relative to control values. In these fish, significant differences of predatory performance in the presence and absence of MP were found (t13 = −2.726, p < 0.05), with MP decreasing the predatory performance inhibition induced by chromium (48% in the presence of MP; 74% in the absence of MP). Thus, these findings suggest toxicological interactions between MP and chromium affecting the predatory performance in L-est fish, but not in M-est fish. The acclimatization and test conditions of fish from distinct estuaries were the same. Thus, the differences recorded between fish from the two estuaries may result from the exposure to distinct environmental conditions during previous developmental phases, which may influence the basal levels of factors able to alter the internal concentrations of chromium species resulting from the Cr(VI) biotransformation by P. microps and/or to cope with the chemical stress generated. For example, in A. anguilla, Cr(VI) reduction by GSH results in Cr(V), Cr(IV) and Cr(III) generating oxidative reactive species able to induce oxidative stress and damage (Ahmad et al., 2006; Pacheco et al., 2013). In fish from both estuaries, significant differences in AChE activity among treatments were found (Fig. 2, Table 4). In L-est fish

171

exposed to MP alone, a significant inhibition (20%) of AChE activity was found, indicating that MPs alone are able to decrease the AChE activity of early P. microps juveniles. In M-est fish a decrease of AChE activity (18%) was also found but the differences relative to the control group were not satistically significant. The inhibition of AChE by MPs alone found in the present study is in good agreement with the findings of previous study with P. microps juveniles where a mean of 22% of inhibition was determined (Oliveira et al., 2013). The combined treatments containing Cr(VI) at concentrations equal or higher than 8.0 mg/l caused a significant inhibition of AChE in fish from both estuaries, reaching a maximum of 31% relative to the respective control group. Significant differences of AChE activity between fish exposed to Cr(VI) alone (in the bioassay testing the effects of the metal as single toxicant) and those exposed to the corresponding mixture treatments were found (M-est fish: t13 = −4.603 and t13 = −3.270 for Cr(VI) concentrations of 18.9 and 28.4 mg/l, respectively; L-est fish: t9 = −2.567 for Cr(VI) concentrations of 18.9 mg/l; p < 0.05). Thus, whereas in the range of concentrations tested, Cr(VI) alone had no significant anti-cholinesterase effects (Table 3), under combined exposures with MP, the metal seems to increase the inhibitory effects of MP on AChE activity (MP alone: 18–20% inhibition; maximal inhibition caused by the mixture: 31%). This suggests that the type of toxicological interaction between Cr(VI) and MP on AChE activity may be potentiation. However, specific studies need to be carried out to deeper investigate the type of interaction on AChE activity. In fish, the inhibition of AChE activity by 30% or above may lead to disruption of the nervous system function (Almeida et al., 2010; Vieira et al., 2009), with a wide range of adverse effects at individual level resulting from intoxication (e.g. reduction of swimming performance, growth delay) potentially leading to death. As in the bioassay testing Cr(VI) alone, no significant differences in GST and EROD activity among treatments were found (Table 4), indicating that these enzymes were not significantly involved in the response to MP- and Cr(VI)/MP-induced stress. Under simultaneous exposure to MP and Cr(VI), a significant increase of LPO levels was found in fish from the L-est exposed to the treatments containing concentrations of Cr(VI) equal or higher than 3.9 mg/l but not in M-est fish (Table 4). These results indicate that the simultaneous exposure to MP and Cr(VI) is able to induce lipid oxidative damage on P. microps juveniles, and that previous developmental conditions of fish may influence the final outcome. In summary, the comparison of the bioassays carried out with Cr(VI) alone and in mixture with MP, indicate that toxicological interactions between the two agents occurred in early juveniles of P. microps. It also indicates that the health status of the fish and/or the conditions to which fish were exposed during pre-developmental phases in their natural habitat, including environmental contamination, influence the final outcome of such interactions. Therefore, the presence of MP in test media influences the toxicity of Cr(VI) to P. microps early juveniles by two processes: The first one are direct interactions between the two substances in test media likely influencing the Cr(VI) availability to fish, possibly through binding of chromium anions to MP surface as discussed in Section 3.2 and suggested by Tables S-2 and S-3 (Supplementary material) and Table 2. The second are toxicological interactions inside the organisms as suggested mainly by the significant AChE inhibition in the combined treatments of the mixture bioassay, reaching higher enzyme inhibition than the one induced by MP alone (Table 4), and contrary to findings of the bioassay testing Cr(VI) alone where no significant effects on AChE were found (Table 3), and the significant increase of LPO levels in L-est fish in mixture treatments, contrary to the lack of significant effects in this parameter in fish exposed to Cr(VI) or MP alone (Tables 2 and 4). These findings lead to the rejection of our second hypothesis (H02 ).

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3.4. H03 : long-term exposure to different environmental conditions during pre-developmental phases in the natural habitat does not influence the acute toxicity and responses to chromium induced stress either in the absence or presence of MP Despite the overlap of 96 h-LC50 s 95% CI, significant differences in the Cr(VI) toxicity curves of fish from M-est and L-est were found (ANCOVA, F2,9 = 6.827, p < 0.05), indicating differences of sensitivity to the metal between fish from distinct estuaries. The results of Table 2 show that the mortality induced by the highest concentration of Cr(VI) when tested alone was higher in L-est fish (100%) than in M-est fish (44.4%). These results indicate that L-est fish are more susceptible to high concentrations of Cr(VI) than M-est fish. The presence of MP slightly decreased the mortality induced by the highest concentration of Cr(VI) tested on L-est fish (from 100% to 88.9% in the absence and presence of MP) but increased the mortality among M-est fish (from 44.4 to 66.7%) (Table 2). Moreover, the presence of MP significantly decreased the negative effects of the highest Cr(VI) on the predatory performance of L-est fish (Fig. 1B) but had no significant effects on the predatory performance of Mest fish (Fig. 1A). Furthermore, under simultaneous exposure to Cr(VI) and MP, a significant increase of LPO levels was found in L-est fish but not in M-est ones. Overall these findings indicate differences of sensitivity between M-est and L-est fish to both high concentrations of Cr(VI) alone and in mixture with MP. Because the acclimatization and test conditions of fish from distinct estuaries were the same, the differences recorded between fish from distinct estuaries likely result from the influence of distinct environmental conditions to which the fish were exposed during previous developmental phases in their original habitats, including different levels of several environmental contaminants, availability of food, among conditions that are different in M-est and L-est (e.g. Guimarães et al., 2012). Thus, these findings lead to the rejection of our third hypothesis (H03 ).

4. Conclusions Cr(VI) when tested alone induced mortality in early juveniles of P. microps, with 96 h- LC50 s of 14.4 mg/l (95% CI = 4.30–51.01) in L-est fish and 30.5 mg/l (95% CI = 23.31–62.16) in M-est fish (Table 2), decreased the predatory performance of juveniles from both estuaries (up to 74%), and had no significant effects in the other parameters analysed (AChE, GST and EROD activity; LPO levels) (Table 3). These results indicate that in the range of concentrations tested (5.6–28.4 mg/l) Cr(VI) is toxic to early juveniles of P. microps, thus rejecting the first null hypothesis tested in the present study. At the beginning of the mixture bioassay, the deviation of actual lowest Cr(VI) concentrations from nominal ones was higher than 20%, an effect that was not observed in the bioassay with Cr(VI) alone nor at the highest concentrations of the mixture bioassay (Tables S1 and S2, Supplementary material). This finding suggests binding of the metal to MP particles, and thus interaction between the two substances in aqueous media, a finding previously reported by other authors (Holmes et al., 2012, 2014; Rochman et al., 2014). Under simultaneous exposure to Cr(VI) and MP, a significant decrease of the predatory performance (up to 67 %) (Fig. 1) and a significant inhibition of AChE activity (up to 31 %) (Table 4, Fig. 2), higher than the inhibition caused by MP alone (by 18–21%) and that was not observed in the bioassay investigating the effects of Cr(VI) alone (Table 3), were found in fish from both estuaries. Moreover, mixture treatments containing actual concentrations of Cr(VI) equal or higher than 3.9 mg/l significantly increased LPO levels in L-est fish, an effect that was not observed under exposures to Cr(VI) or MP singly (Tables 3 and 4). These findings indicate toxicological interactions between Cr(VI) and MP

inside the fish, thus rejecting our second null hypothesis. The presence of MP slightly decreased the mortality induced by the highest Cr(VI) tested (28.4 mg/l) on L-est fish (from 100% to 88.9%) but increased the mortality induced on M-est fish (from 44.4% to 66.7%) (Table 2). Moreover, the presence of MP significantly decreased the negative effects of the highest concentration of Cr(VI) tested on predatory performance of L-est fish but had no significant effects on M-est fish (Fig. 1). Furthermore, the simultaneous exposure to MP and Cr(VI) induced oxidative damage in L-est fish but not in M-est fish (Table 4). The acclimatization and test conditions were similar for fish from M-est and L-est estuaries which have several environmental differences. Thus, the findings of the present study indicate that long-term exposure to environmental conditions of the natural habitat during previous developmental phases influences the response and sensitivity of early juveniles of P. microps to Cr(VI)-induced stress in both the absence and the presence of MP, thus rejecting the third null hypothesis. The potential of Cr(VI) and MP alone and/or in combination to decrease the fish predatory behaviour and AChE activity are effects of grave concern because they can lead to the reduction of individual performance and ultimately to death with negative effects on the population fitness. Overall, the results of this study indicate toxicological interactions between MP and Cr(VI) highlighting the importance of further investigating the combined effects of MP and other common contaminants on wild organisms. Acknowledgements We thank to Prof. Natividade Vieira for providing parental A. franciscana and several members of the project “SIGNAL” for helping in fish collection. This study was financially supported by National and European Regional Development funds through the Portuguese Foundation for the Science and Technology and the Operational Competitiveness Programme (COMPETE) under the project “SIGNAL” (PTDC/AAC-AMB/110331/2009; FCOMP-010124-FEDER-013876), with additional contribution of the project Pest-C/MAR/LA0015/2013, and funds of the Institute of Biomedical Sciences of Abel Salazar (ICBAS), University of Porto. L.G. Luís had a BI grant in the scope of the project “SIGNAL”. The funding institutions did not participate in the scientific work and paper preparation. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox.2015.04. 018 References Ahmad, I., Maria, V.L., Oliveira, M., Pacheco, M., Santos, M.A., 2006. Oxidative stress and genotoxic effects in gill and kidney of Anguilla Anguilla L. exposed to chromium with or without pre-exposure to ␤-naphthoflavone. Mutat. Res. Genet. Toxicol. Environ. Mutagen. 608, 16–28. Ahmed, M.K., Kundu, G.K., Al-Mamum, M.H., Sarkar, S.K., Akter, M.S., Khan, M.S., 2013. Chromium(VI) induced acute toxicity and genotoxicity in freshwater stinging catfish: heteropneustes fossilis. Ecotoxicol. Environ. Saf. 92, 64–70. Almeida, J.R., Oliveira, C., Gravato, C., Guilhermino, L., 2010. Linking behavioural alterations with biomarkers responses in the European seabass Dicentrarchus labrax L. exposed to the organophosphate pesticide fenitrothion. Ecotoxicology 19, 1369–1381. Andrady, A.L., 2011. Microplastics in the marine environment. Mar. Pollut. Bull. 62, 1596–1605. Antunes, J.C., Frias, J.G.L., Micaelo, A.C., Sobral, P., 2013. Resin pellets from beaches of the Portuguese oast and adsorbed persistent organic pollutants. Estuar. Coast. Shelf Sci. 130, 62–69. Barnes, D.K.A., Galgani, F., Thompson, R.C., Barlaz, M., 2009. Accumulation and fragmentation of plastic debris in global environments. Philos. Trans. R. Soc. B: Biol. Sci. 364, 1985–1998.

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