Endocrine Disrupting Effects of Bisphenol A, Bisphenol S and Benzyl Butyl-Phthalate on the Thyroid System of Juvenile Brown Trout (Salmo Trutta)

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Endocrine disrupting effects of bisphenol A, bisphenol S and benzyl butyl-phthalate on the thyroid system of juvenile brown trout (Salmo trutta)

Joan Martorell Ribera

Degree project for Master of Science (Two Years) in Biology

Zoophysiology 60 hec Spring 2015

Department of Biological and Environmental Sciences University of Gothenburg

Examiner: Michael Axelsson Department of Biological and Environmental Sciences University of Gothenburg

Supervisor: Elisabeth Jönsson Bergman Biological and Environmental Sciences/Zoophysiology University of Gothenburg

Table of Contents Table of Contents ...... 2 Abstract ...... 1 Introduction ...... 1 Thyroid hormones ...... 1 The thyroid axis ...... 2 Hypothalamus and pituitary ...... 3 Processes in the thyroid follicles ...... 3 Plasma thyroid hormones ...... 3 Inside the cell ...... 4 Endocrine disruptors ...... 4 Bisphenol A and S ...... 5 Benzyl butyl phthalate (BBP) ...... 6 Biomarkers ...... 6 Cyp1a induction ...... 7 Plasma vitellogenin ...... 7 Aim ...... 7 Materials & Methods ...... 7 Exposure ...... 7 Experimental fish ...... 8 Cholesterol pellets and implantation ...... 8 Sampling ...... 8 Vtg-ELISA ...... 9 EROD activity and protein content ...... 10 Preparation of the S9 cytosolic fraction ...... 10 EROD activity measurement ...... 10 Protein content measurement ...... 11 T3 and T4 RIA ...... 11 Equipment ...... 12 Chemicals...... 12 Statistical analysis ...... 12 Ethical permit ...... 12 Results ...... 12 Biometric data ...... 13

Vitellogenin in blood plasma ...... 13 EROD activity in liver S9 fraction ...... 14 Thyroid hormones in plasma ...... 15 Plasma triiodothyronine (T3) ...... 16 Plasma thyroxine (T4) ...... 17 T3/T4 ratio ...... 18 Discussion ...... 20 Plasma Vitellogenin (Vtg) ...... 20 Cyp1a induction/EROD activity ...... 20 Plasma thyroid hormones ...... 21 Conclusions ...... 22 References ...... 22 Acknowledgments ...... 30 General public abstract ...... 30 Appendix ...... 31

Abstract Through the action of the thyroid hormones, the active form triiodothyronine (T3) and the prohormone thyroxine (T4), the thyroid system is involved in multiple metabolic processes that are essential for development, differentiation and metamorphosis. Endocrine disrupting compounds (EDCs) can interfere at many levels of the thyroid system, acting as antagonists for the thyroid receptors in the cell membrane, affecting the development of the thyroid gland, etc. Bisphenol A (BPA) and benzyl butyl phthalate (BBP) are known EDCs used in the manufacture of plastics and present in the natural environment. Bisphenol S (BPS) has recently appeared as a substitute for BPA in some plastic products. The aim of this study was to investigate the potential of these compounds to interfere on the thyroid hormone system and toxicology markers of juvenile brown trout. Duplicate groups of juvenile brown trout (Salmo trutta) were exposed to BPA, BPS and BBP by the implantation of cholesterol pellets containing a compound dose of 2mg/kg fish or 20mg/kg fish for either 2 weeks or 8 weeks.

Plasma T3 and T4 levels were measured as an indicator of the thyroid disruption, together with toxicological markers like plasma vitellogenin (Vtg) and Cyp1A induction in the liver (EROD activity). Plasma T3 remained similar in all treatment groups and exposure time. Plasma T4 showed non- significant tendency to decrease in all the exposed groups after 8 weeks compared to control fish. There was an increase in plasma Vtg showed in the BPA and BPS groups exposed to 20mg/kg dose for 2 and 8 weeks. Cyp1A induction showed lower activity in the BPA group (2 mg/kg) for the 2 weeks exposure and in BPA and BPS groups (20mg/kg) after 8 weeks exposure. There was no effect of BBP exposure on plasma Vtg or in Cyp1A induction.

Despite the fact that the results on T4 were not significant, there was still a trend supporting the idea of an endocrine disrupting effect on the plasma T4 as seen in other studies with bisphenols and phthalates. However, more research at different levels of the thyroid system is needed for a solid explanation.

Introduction

Thyroid hormones The thyroid system is present in all the subphylum vertebrates, and the roles and structures of the thyroid hormones, the active form triiodothyronine (T3) and the pro-hormone thyroxin (T4) (Bentley., 1998) are highly conserved among the organisms forming this group, although we can see some functional differences within vertebrates (Power et al., 2001). Among non-vertebrates, tissues with equivalent activity as the thyroid gland, synthesis of iodothyronines, have been found in protochordates and ascidians (Bentley., 1998). As pointed by P.J. Bentley in his book Comparative vertebrate endocrinology (1998), “the thyroid appears to have the longest phylogenetic history of any endocrine gland”.

Thyroid hormones are involved in many vertebrate metabolic processes affecting numerous tissues. In teleost fish, these hormones have a key role in early development and differentiation, including somatic growth, osmoregulation and larval to juvenile transition or metamorphosis (Blanton et al.,

2007; Power et al., 2001; Youson et al., 1988). However, since the thyroid gland is not present in fish embryos, these development processes in early stages have to be driven by the female thyroid hormones (Power et al., 2001). A clear example for performance of the thyroid hormones in fish development are the metamorphosis processes that lead the flatfish larvae to juvenile, and how disruptions of the thyroid axis result in jaw abnormalities and failed eye migration (Gomes et al., 2014; Power et al., 2001).

In mammals, the T3 and T4 secretion and its concentration in blood is regulated by the Brain- pituitary-thyroid axis. Contrary in fish, the Brain-pituitary-thyroid axis mainly controls the release of T4 by the thyroid follicles and T4 homeostasis, while T3 activation and homeostasis is regulated by enzymatic deiodination activity in the peripheral tissues, together with its release into the bloodstream, see figure 1 and further description below (Eales et al., 1999; 1993; Morin et al., 1993). In teleost, the thyroid gland is generally situated in the basibranchial region and surrounding the ventral aorta. In brown trout, it is located between the first and the third branchial arches (Eales et al., 1999; Gomes et al., 2014). In most of the fishes the thyroid is not encapsulated like in other vertebrates. It is rather formed by a diffused net of vascularized follicles. Thyroid follicles are structured in a single cell layer around an extracellular lumen, these cells are known as thyrocytes. In the lumen, thyrocytes excrete the glycoprotein thyroglobulin (TG) forming a colloid (Eales et al., 1999; Power et al., 2001; Blanton et al., 2007).

Thyrocytes are in charge of the iodide (I-) uptake from blood, the limiting factor for the synthesis of Thyroxin (T4), which is the major hormone secreted by the thyroid in Teleost (Bentley., 1998). Since there is a great storage of iodine in sea water, fish obtain most of the iodide they need via the gills and in less proportion through the diet (Eales et al., 1999). In the other hand, fresh water fish obtain iodine through the diet, being absorbed in the gastrointestinal tract (Watanabe et al., 1997). This abundance of iodine in salty water makes goiter (enlarged thyroid gland) events in sea fish rare and mainly observed in fresh water species (Eales et al., 1999, 1993; Gomes et al., 2014).

The thyroid axis 2. 1. Deiodination Brain Tissues Hypothalamus T4T3 s THR T4/T3 T3-TR

T4/T3 Gene Pituitary feedback Blood 3. stream expression TSH

Thyroid Growth Development T4/T3 follicle

Figure 1: The thyroid axis is divided in three parts: 1. The brain-pituitary-thyroid axis which regulates the synthesis, storage and secretion of thyroxin (T4) and preservation of T4 levels in the bloodstream. In other

vertebrates than fish it controls as well the excretion of T3 in the thyroid gland. 2. Peripheral activation of triiodothyronine (T3) via deiodination of T4 (e.g. liver). T3 can either remain in the tissues or be released into the bloodstream. 3. Interaction between T3 and its receptors (TR) in cells and tissues to modulate development, growth and reproduction (Eales et al. 1999, 1993; Blanton et al. 2007).

Hypothalamus and pituitary The thyroid axis starts with the stimulation of the brain and the hypothalamus via external (photoperiod, temperature, etc.) and internal cues (hormones, metabolism, etc.). These stimuli lead the hypothalamus to secrete thyrotropin-releasing hormone (TRH) which, in turn, will stimulate the thyrotrophic cells in the pituitary gland to secrete thyroid stimulating hormone (TSH). In fish, the process that involves the secretion of TRH by the hypothalamus to stimulate the secretion of TSH in the Pituitary is not clear (Gomes et al., 2014; de Groef et al. 2006). TSH targets the thyrocytes in the thyroid follicles to control the uptake of Iodine (I-) from the blood stream and regulate the release of thyroxine (T4). T4 will travel in the bloodstream to the peripheral tissues where it will be deiodinated to the active form triiodothyronine (T3) (Blanton et al. 2007). There is a negative feedback of T3 and T4 modulating the transcription of the β unit of TSH in the pituitary gland. This regulates the release of TSH and, this way, the maintenance of stable T4 concentrations in the blood stream (Eales et al., 1999; Pradet-Balade et al., 1999, 1997).

Processes in the thyroid follicles Iodide (I-) is transported from the blood vessels close to the thyroid follicles to the cytoplasm of the thyrocytes using the Na/I- symporter (NIS). Iodine is then released to the follicle lumen via a pendrin transporter where will join the thyroid peroxidase (TPO) to iodinate specific thyrosyn residues of the thyroglobuline (TG) colloid, excreted in the lumen by the thyrocytes (Eales et al, 1999; Gomes et al., 2014). Thyroid peroxidases can iodinate thyroxines with one or two Iodide molecules, forming monoiodotyrosines (MIT) or diiodotyrosines (DIT) respectively. The union, driven by TPO, of DIT with MIT will produce a triiodothyronine (T3). Instead, the union of two DIT will form a thyroxine (T4) (Degroot et al, 1977).

These newly synthesized T3 and T4 remain attached to the TG colloid structure, from which they have to be freed to become actual thyroid hormones. The iodinated colloid is endocytosed by the thyrocytes, where lysosomes with cathepsins will digest it and release T3 and T4 to the blood stream(Eales et al. 1999). In teleost T4 is the main hormone secreted by the thyroid gland (Bentley, 1998). In the other hand, the liver is in charge of the major part of T3 release in blood plasma (Morin et al., 1993)

Plasma thyroid hormones Once released in blood serum, thyroid hormones quickly conjugate with carrier proteins (Eales et al, 1999), these are: thyroxine-binding globulin (TBG), transthyretin (TTR) and serum albumin (ALB).In humans the main thyroid hormone transporter is TBG, while some studies show that in fish and amphibians, T4 is transported by albumin and T3 is carried by transthyretin (Power et al., 2000). From the blood vessels, thyroid hormones enter into the body tissues using several transporter proteins in the cell membrane. In mammals, the main transporters are Na+/taurocholate cotransporting polypeptide (SLC), heterodimeric L-type amino acid transporters (LAT), monocarboxylate transporters (MCT) and organic anion-transporting polypeptide (OATP) (Visser et

al., 2011). In Teleost, experiments have shown that an increase in T3 levels down regulates the expression of some MCT and OATP genes (Muzzio et al., 2014).

Inside the cell Already in the cytosol, T4 needs to be deiodinated to T3 to become biologically active. At the same time T3 and T4 can be inactivated by the same deiodination principle. This process is driven by the deiodinases (Dio), a family of integral membrane enzymes. Three groups of Dio are present in mammals: type 1, type 2 and type 3 (Bianco et al., 2006). Fish deiodinases show some variations to those in mammals, even though all share the characteristic of having a selenocystein in their active site (Eales et al.,1999). Although the expression of Dio enzymes depends on the tissue and the species (Gomes et al., 2014), several works have shown that in teleost, T4 deiodination occurs mainly in the liver among other organs like brain and kidney and T3 inactivation occurs mainly in the brain (Mol et al., 1998; Eales et al.,1999). In the liver takes place the major part of T4 degradation and excretion via UDP-glucuronosyltransferase (UDP-GT). This enzyme adds a glucoronic group to the T4 that increases the water solubility of the molecule and facilitates the excretion to the bile (Barter et al., 1992; Clarke et al., 1992)

Deiodination can take place in the inner or outer ring of the thyroid hormone structure by the deletion of one iodine (Eales et al., 1999). In mammals, Dio type 1 (Dio1) together with Dio type 2 (Dio2) are in charge of the outer ring deiodinations (ORD). In this pathway, ORD, is where T4 is deiodinated to the active T3, and T3 to the inactive form T2. Dio type 3 (Dio3) and Dio1 perform inner ring deiodinations (IRD). This pathway transform T4 to reverse T3 (rT3) an inactive form, as well, it can deiodinate rT3 to T2 (Bianco et al., 2006). In rainbow trout, the enzymes that perform T4 ORD shares analogy to the mammal Dio2, at the same time the enzymes leading T4 IRD and T3 IRD are similar to the mammal Dio3 (Eales et al., 1999). Enzymatic activity similar to that of mammal Dio1, has been found in Tilapias kidney (Mol et al., 1998).

Once inside the cell, Thyroid hormones conjugate with nuclear receptors (TR) promoting gene transcription. When there is no presence of thyroid hormones, TR acts as a repressor, blocking the transcription of target genes (Oetting et al., 2007). In mammals, two genes encode for the two forms of TR: TRα and TRβ. In comparison, fish have one gene encoding for TRβ and two encoding for TRα (TRα1 and TRα2) (Yamano et al., 1994). TR bind at the same time to T3 and to thyroid hormone- response elements (TREs), found in the promoter sequence of target genes. Normally, the complex T3-TR forms a heterodimer with retinoic acid receptors (RXR) prior to bind TREs (Oetting et al., 2007).

The union of T3 to the different TR and the interaction of the T3-TR-RXR complex with TRE, together with other modulating molecules in the target gene, regulate the effect of thyroid hormones in the tissue and its response (Brent et al., 2012).

Endocrine disruptors Chemicals, toxicants and pharmaceutical products, many of them present in the environment, have shown their capacity to act as endocrine disruptors in one or several points of the thyroid system, that is mimicking, inhibiting, enhancing or affecting in any way the function of the thyroid hormones. (Brown et al.,2004). A number of studies in different kind of fish and many different toxicants have shown various effects on the fish thyroid physiology. For instance, bisphenol A (BPA), bisphenol S

(BPS) and benzyl butyl phthalate (BBP) which are known to have endocrine disruptor effects in fish (Naderi et al., 2014; Knudsen et al., 1998 and Lindholst et al., 2001).

BPA structure BBP structure BPS structure PCBs structure

17,β-estradiol T3 structure T4 structure structure

Figure 2. Molecular structures for BPA, BPS, BBP, PCBs, T3, T4 and estradiol.

Bisphenol A and S BPA have been largely used in the plastic industry for the production of mainly polycarbonate hard clear plastic, but as well epoxy resins, flame retardants, PVCs (polyvinyl chloride), etc. (Staples et al., 1998; Naderi et al., 2014). Around 6 billion of BPA pounds are manufactured every year and more than 100 tones expelled to the atmosphere (Vandenberg et al., 2008). The effects of BPA as an endocrine disruptor in humans and animals have been studied for a long time (Vandenberg et al., 2008; Flint et al., 2012), and nowadays several countries including the European Union, have banned BPA from baby bottles (European commission, 2011). Since more regulation on BPA is appearing worldwide, new compounds are being developed to substitute BPA, BPS is one of them (Liao et al., 2012).

Several studies have shown the effects of BPA at different levels of the thyroid system: In humans, BPA have shown inhibitory activity on the TPO enzyme of the thyroid follicles, a key enzyme involved in the synthesis of T3 and T4 (Schmutzler et al., 2007). BPA have presented several in vitro effects on the thyroid receptor β physiology (TRβ), repressing the transcription of TRβ in monkey cells (Cercopithecus aethiops), after a low dose exposure (Sheng et al., 2012), or having an antagonistic role on the TRβ in rat pups, prior exposure of pregnant females to BPA, suggesting an interference on the negative feed-back that the thyroid hormones carry out on TSH release in the pituitary (Zoeller et al., 2005). Once more, a low exposure to BPA accelerated embryonic development and advanced the hatching in medaka fish (Oryzias latipes) through its effect on the thyroid receptor (Ramakrishnan et al., 2007). Moreover, BPA interfered with T3 action during tadpole (Xenopus laevis) metamorphosis processes (Iwamuro et al., 2003).

BPS is a new compound named to be the substitute of BPA in some industrial products (Liao et al., 2012) like the production of baby bottles (Grignard et al., 2012), in which formula BPA has been banned in several countries (European commission, 2011). However, chemical similarity between these two compounds makes expect a similar role as an endocrine disruptor (Fig. 2; Naderi et al.,

2014; Ji et al., 2013; Grignard et al., 2012). BPS has better resistance against high temperature, sunlight and organic solvents (Molina-Molina et al., 2013; Liao et al., 2012). It has shown much less sensitivity to degradation in sea water than BPA, being able to accumulate in water environments in the future (Danzl et al., 2009). Contrary to BPA, in BPS there is not a lot of research done regarding its effects as an endocrine disruptor in animals. However, since the use of BPS in industry is increasing some studies have been done: After zebrafish (Danio rerio) embryos were long term exposed to different concentrations of BPS, males showed a significant decrease in body length, weight, sperm count, plasma thyroid hormones and plasma testosterone. Females showed decreased egg ratio and plasma thyroid hormones (Naderi et al., 2014; Ji et al., 2013). In another study testing the interactions between BPS and human nuclear receptors, BPS showed antagonist activity against human androgen receptor (hAR) and an agonist behavior interacting with human estrogen receptors (hER) and pregnant X receptor (hPXR) (Molina-Molina et al., 2013).

Benzyl butyl phthalate (BBP) Phthalates are a family of chemicals used in industry as emollients and supplements to enhance plastics flexibility (Boas et al., 2012; Jugan et al., 2010). Around 800 000 tones are produced every year in the European countries (Heudorf et al., 2007). Some of the products where BBP is used are: Vinyl tiles, artificial leather and traffic cones. Phthalates are metabolized and eliminated via urine in hours and do not accumulate in the organism (Boas et al., 2012).

In humans, a negative correlation between mono butyl phthalate and T4 in urine was found in pregnant woman (Huang et al., 2007). In men and children, phthalate metabolites negatively altered T4 and T3 levels in blood serum (Meeker et al., 2007; Boas et al., 2010). Rats treated with various phthalate esters, alone or combined, showed a reduction in the thyroid follicle size and cells changed to columnar shape, linked with an increase of thyroglobulin turnover (Howarth et al., 2001). Zebrafish embryos under a seven days exposure to phthalates showed a decrease in the whole body T4 levels (Zhai et al. 2014).

Focusing on BBP, this chemical has shown inhibitory effects on the L-type amino acid transporters (LAT) affecting the uptake of thyroid hormones by the erythrocyte’s membrane in tadpoles (Rana catesbeiana) (Shimada et al., 2004). Regarding to the iodide uptake by the thyrocytes to synthesize T3 and T4, BBP has induced an increasing effect of the sodium/iodide symporter (NIS) activity and NIS mRNA expression (Breous et al., 2005). BBP appeared to compete with T3 for the binding site of the thyroid blood transporter Transthyretin (TTR), but showed no interactions with the thyroid receptor β (TRβ) binding site (Ishihara et al., 2003). However, it inhibited the gene expression of TRβ in Xenopus laevis cells (Sugiyama et al., 2005). In the other hand, BBP has shown no effect on the estradiol receptor (ER) in rainbow trout but a decrease in the zona radiata proteins (Zrp) in plasma (Knudsen et al. 1998).

Biomarkers Biomarkers in ecotoxicology are defined as measureable biological parameters covering from molecular level to behaviour, in which an alteration of their natural performance indicates the presence of toxic compounds in the organism (Depledge et al., 1993). In this study, induction of cytochrome P450 1A (Cyp1a) in liver and vitellogenin levels in plasma have been used as biomarkers to analyse the toxicological capacities of BPA, BPS and BBP on the brown trout organism.

Cyp1a induction The induction of the Cyp1a enzymes in fish liver has been largely used as biomarker for the exposure to environmental contaminants such as polycyclic aromatic hydrocarbons (PAH), planar halogenated hydrocarbons (PHH) among others (Payne et al., 1987; Payne & Penrose, 1975; Whyte et al., 2000). Cyp1a enzymes participate in the metabolization and clearance of toxicant compounds by increasing their solubility in water (Hahn & Karchner, 1995; Andersson & Förlin, 1992). However, the increase of Cyp1a enzymes to deal with some toxicant compounds can produce reactive oxygen species (Schlezinger et al., 2000). When the organism is exposed to toxicant compounds, Cyp1a shows an increase in numbers inside the cells (Stegeman and Lech, 1991), this induction of the Cyp1a levels is driven by the interaction of toxicant compounds with an aryl hydrocarbon receptor (AhR) present in the cytosol, which enhance the gene expression of Cyp1a and other enzymes (Nebert et al.,1993).

Plasma vitellogenin Vitellogenin (Vtg) levels in fish plasma are used as an indicator of endocrine disruption (Hansen et al., 1998; Sumpter & Jobling, 1995). Vtg is the main hormone involved in the yolk synthesis in developing oocytes and necessary for embryogenesis (Sumpter & Jobling., 1995). Vitellogenesis in the liver is controlled by several hormones, although the most important is 17,β-estradiol from the estrogens group (Carnevali et al., 1992; Wahli et al., 1981). Vtg is present in female plasma and increases when the fish is sexually mature (Scott & Sumpter., 1983). Instead, Vtg levels in male plasma are scarce and in many cases not detectable due to silencing of the Vtg gene (Flouriot et al., 1995). However, the detection of high levels of plasma Vtg in males and juvenile fish can occur when exposed to natural estrogens or EDCs present in the environment, because estrogenic compounds interact with the estrogen receptors in the liver mimicking the endogenous hormones action and enhancing Vtg synthesis (Sumpter & Jobling, 1995).

Aim In this project, juvenile individuals of Brown trout (Salmo trutta) have been exposed to three synthetic compounds, bisphenol A (BPA), bisphenol S (BPS) and benzyl butyl phthalate (BBP) to investigate the effects on the thyroid hormone levels, an indicative of the potential thyroid hormone system disruption, and toxicological markers such as plasma vitellogenin and Cyp1A induction in the liver (EROD).

Materials & Methods

Exposure During the experimental design two term exposures were planned involving the three compounds of interest, BPA, BPS and BBP. A short term exposure with two weeks duration and a long term exposure with eight weeks duration. Moreover, within the two time terms the fish were exposed to the each compound in two different concentrations, a high dose with 20 mg/kg fish and a low dose with 2mg/kg fish.

Fish were placed in 14 aquaria in groups of 10 fish per tank. After an acclimation time of 2 weeks, each fish was weighed, measured and injected with a cholesterol pellet of 20 mg weight containing

either the chemical or a sham implant (see below) and a PIT tag for individual identification of the fish. For each toxicant there were 4 groups of fish exposed, corresponding to high and low dose exposures and its replicates. A pair of groups was settled as controls.

After two weeks of the implant injections, according to the short term exposure, five fish of each aquarium were sacrificed, measured and sampled for tissues. Subsequently, after eight weeks of the implant injections, the five fish remaining in the aquaria were sacrificed, measured and sampled for tissues.

In several aquaria some fish died before the samplings due to unknown or natural reasons. In these cases the fish were not included in the experiment.

Experimental fish The fish used for this experiment were juvenile individuals of Salmo trutta with an average weight of 100g per fish. Around 200 fish from Vänneåns fiskodling, Laholm, Sweden, were purchased and housed in two large freshwater flowthrough tanks. Fish were kept at 10ºC temperature and 12/12 hours of light/dark. The individuals were fed 1% of the body weight with commercial dry food (Biomar, Norway) three times a week during the acclimation and the experiment.

Cholesterol pellets and implantation An implant made following the method from Jönsson et al. 1998 was used to continuously release the toxicants through the duration of the experiment, two and eight weeks respectively.

Using a final weight of 20 mg per pellet, it contained a 95% of Cholesterol and a 5% of molten cocoa butter for the control pellets. For the BPA and BPS low dose pellets it contained a 94% of cholesterol, 5% of cocoa and 1 % of BPA or BPS. For the BBP low dose it contained 95% cholesterol, 4 % of cocoa and 1% of BBP. For the high dose pellets with BPA, BPS and BBP it contained 85 % of cholesterol, 5 % cocoa and 10 % of BPA, BPS or BBP.

After mixing the components of the pellet, it was compressed using a mold to obtain the final pellet. Before implantation, fish were sedated with 2-phenoxyethanol 4ml/l water and fork length (cm) and weight (g) were noted. Implant injections were then performed by a small cut on the ventral zone of the fish using an scalpel and subsequently introducing the pellet into the intraperitoneal region.

With the data from weight and fork length, the condition factor (CF) for each individual was calculated using CF = body weight (g)/length3 (cm)*100.

Sampling After the exposure period fish were sacrificed with a sharp blow to the head and blood was immediately extracted from the caudal vein with a heparinized syringe. Id tag, weight (g) and fork length (cm) were noted for each fish. Using the freshly extracted blood, haemoglobin and glucose levels were measured using a cuvette system from Hemocue with assayed haemoglobin (HemoTrol) and glucose (GlucoTrol-AQ) as quality controls. Part of the blood was centrifuged in a hematocrit capillary centrifuge for 2 min. using capillary tubes to obtain the haematocrit level. The main part of

the extracted blood was centrifuged at 6000 rpm for 5 min to separate the plasma fraction which was transferred to several eppendorf tubes (5) and stored on dry ice temporally until final storage at -80˚C.

For each fish, liver, kidney, head kidney, pituitary gland, hypothalamus and thyroid gland were extracted and stored separately in liquid nitrogen. The bile content and the cholesterol pellet were recovered and stored in separate glass vials on dry ice temporally until transferred to -80˚C.

Specific growth rate for length (SGRL) and weight (SGRW) were calculated for the two exposure times 14 days and 60 days respectively. The implanting day was regarded as starting point of the experiment (length=L1, and weight=W1) and the sampling days were regarded as end point (length=L2, and weight=W2). Specific growth rate was calculated using the formula ln(W2/W1)*100/days of exposure for SGRW and ln(L2/L1)*100/days of exposure for SGRL.

Vtg-ELISA To elucidate the levels of vitellogenin (Vtg) in blood plasma, a quantitative enzyme linked immune sorbent assay (ELISA) with a competitive technique was used (Specker et al. 1994). Several buffers were prepared including Coating buffer: 0,1M Na2CO3, pH 9,6; Washing buffer: 0,05 M phosphate buffered saline (PBS), pH 7,4 containing 0,1% Tween 20; Sample buffer: Washing buffer containing 1% (v/v) BSA or other equivalent protein; Substrate solution: Horseradish peroxidase substrate kit (Bio Rad); Quenching solution: 4% oxalic acid.

For each sample, plasma was thawed and diluted 1:40 in a glass vial with sample buffer obtaining a final volume of 400 µl. Max binding sample containing only sample buffer was prepared, together with an 11 steps standard curve serial dilution, following a 1:3 dilution factor and starting with Vtg standard diluted 1:500. The final volume for the max binding sample and the standard curve serial dilution was 400 µl. Glass vials and buffers used were kept in ice during the whole procedure.

Once all dilutions were done, primary antibody (PO-1), rabbit anti-arctic char, was diluted 1:3000 in sample buffer and 400 µl was added to the previous prepared samples, max binding and standard curve serial dilution. With a final volume of 800 µl, vials were vortex and incubate overnight at 4 ºC. Microtiter plate wells (Greiner) were coated with 200 µl Vtg standard diluted 1:3000 in coating buffer and kept overnight at 4ºC.

Once completed the overnight incubation, plate wells were washed three times with washing buffer and unspecific binding was blocked with sample buffer, followed by 30 minutes incubation at 37 ºC. To initiate the competitive stage of the protocol, 200 µl of the PO-1 treated samples, serial dilution and max binding were distributed in the Vtg-coated plate wells and incubated 90 minutes at 37ºC. During this incubation, the remaining PO-1 antibodies not bounded to sample Vtg molecules, bound to the Vtg molecules coating the plate well.

Following the competition incubation, plates were washed to eliminate antibodies not bound to the well Vtg-surface. Subsequently, 1:2000 dilution of enzyme-labelled secondary antibody (2ary Ab) goat anti-rabbit was added to wells and incubated at 37ºC for 45 minutes.

Plates were washed to eliminate unbound 2ary Ab and horseradish peroxidase substrate was added to plate wells in order to start the colouring enzyme reaction, which was quenched after 5 minutes with oxalic acid 4% distributed to wells. At this point, the plate was situated in the absorbance reader and colour levels were measured at 405 nm.

EROD activity and protein content Ethoxyresorufin-O-deethylase (EROD) and protein content are some of the techniques used to measure the enzymatic activity of Cyp1a (Whyte et al., 2000). Cyp1a drives the catalytic reaction of 7-ethoxyresorufin to resorufin, the concentration of resorufin produced in a determined period of time is measured with fluorometry and divided by the total protein content (Kennedy and Jones, 1994). In order to quantify the (EROD) activity in the liver of the experimental fish, several procedures were performed following the method of Förlin, 1980.

Preparation of the S9 cytosolic fraction Each liver sample, which had been stored in liquid nitrogen, was thawed in buffer (0,1 M Na/K- phosphate with 0,15 M KCl in miliQ water) and homogenized in glass beakers with a Potter-Elvehjem glass-teflon tissue homogenizer. The whole procedure was carried out on ice to avoid loss of enzymatic activity. Once homogenized, samples were centrifuged for 20 minutes at 4˚C and 10000 rcf. The supernatant obtained (S9) was distributed in several eppendorf tubes and stored at -80˚C. In this study we measured the EROD activity in S9 fraction without further separations of the microsomal fraction and cytosol.

EROD activity measurement Prior to quantifying EROD activity in the spectrofluorometer, S9 fractions were thawed and mixed with EROD buffer (0,1 M Na-phosphate pH 8), 10 µM 7-ethoxyresorufin (7-ER) and 10 mM NADPH in quartz vials. The final solution was vortex and measured every x seconds for 1 minute at 530 nm, obtaining a slope representing the enzymatic activity in the sample.

The activity analysed is the transformation of 7-ethoxyresorufin to resorufin (Fig. 3) catalysed by the CYP1A enzymes found in the S9 fraction. Resorufin is excited at 530 nm and emits fluorescence at 585 nm that is measured by the spectrofluorometer. A vial containing a dilution of Rhodamine B in EROD buffer (Rho Std) was measured for 5 minutes as a fluorometry standard for Resorufin.

7-ethoxyresorufin resorufin + light (585nm)

Figure 3. Enzymatic reaction measure in the EROD activity assay.

Protein content measurement To obtain the EROD activity in nmol/mg protein*minute, the protein content in the S9 fraction of the liver samples was measures following the method of Lowry et al. 1951.

Starting the procedure, an standard curve for protein was prepared diluting a standardized sample of bovine serum albumin (BSA) in miliQ water, obtaining a 5 points standard curve (0,1; 0,3; 0,5; 0,7 and 0,9 mg/ml). S9 fraction samples were thawed and diluted 1:40 in miliQ water.

Standard curve, miliQ blanks and diluted samples were distributed in a 96 wells plate and mixed with a 1:25 solution CuSO4, K-Na-Tartrat, NaOH with Na2CO3 and NaOH. The plate was then incubated for 10 minutes at room temperature.

Following the incubation, a dilution of Folin’s reagent was added to wells and incubated at room temperature for 30 minutes. The levels of protein were measured at 750 nm absorbance in the spectrophotometer. Once gathered all the data corresponding to EROD activity and protein content, the following formula was applied, (0,8/ (Rho std/3))*60* EROD activity slope)/(ml of sample*protein (mg/ml)).

T3 and T4 RIA The levels of thyroid hormones triiodothyronine (T3) and thyroxine (T4) in blood plasma were quantified using a radioimmuno assay (RIA) technique as described in Rotllant et al. 2003.

Four different buffers were used in this assay, barbital buffer: Na-barbital 0,11 M, pH 8,6, 0,1 % Na- azid; RIA buffer: 0.25 % bovine gammaglobulin in 0.11 M barbital buffer; 20% PEG: 20% PEG in 0.11M barbital buffer, pH 9.0; ANS buffer: 1 mg/l ANS in RIA buffer.

Several stocks were made prior to start the assay, standard T3 and T4: 1000 ng/ml in barbital buffer; reference T3 and T4: R1, 0,4 ng/ml in barbital buffer and R2, 3,1 ng/ml in barbital buffer; 125I-T3 and 125I-T4 label stock: 6000 cpm/µl in ANS buffer. All stocks were stored in the freezer until use. Each plasma sample was diluted 1:10 in RIA buffer and kept in the refrigerator until the start of the assay. The standard curve was prepared in RIA buffer with an 11 steps serial dilution and a 1:2 dilution factor, starting with T3, T4 standard diluted 1:20. Antibody (Ab) working solutions were made in RIA buffer with a 1:144 dilution for T3Ab and a 1:400 dilution for T4Ab. For each RIA assay label solution was prepared from the stock 125I-T3 or 125I-T4, with a final count of 100 cpm/µl.

To start the assay, diluted samples, serial dilution and reference R1 and R2 were mixed in different glass vials with RIA buffer, antibody AbT3 or AbT4 and label for T3 or T4, with a final volume of 270 µl per tube. Moreover, non-specific binding (NSB) with buffer and label was made, and max binding (B0) with buffer, antibody and label.

All tubes were vortex and incubate for 2 hours at 27ºC. During the incubation, labelled thyroid hormones compete to bind the specific antibodies with the unknown thyroid hormones present in the plasma. Total count control (T) was prepared with only label and kept in the fridge until the gamma-counter measurements.

After the incubation, tubes were cooled down in ice water for 30 minutes and 20% PEG solution was added in each tube and vortex, followed by a 25 minutes centrifugation at 4ºC and 3000 rpm. Subsequently, the liquid fraction in the vials was aspirated and counts per minute (cpm) were measured in the gamma counter.

Equipment Gamma-counter, Wallac Wizard 1470 automatic gamma counter. Spectrophotometer, Wallac Victor 1420 multilabel counter. Spectrofluororometer, Photon technology international.

Chemicals For the implants, Cholesterol, Bisphenol A, Bisphenol S and benzyl –butyl-phthalate were purchased from Sigma-Aldrich (Dorset, UK). For the RIA procedure primary antibodies were purchased from Invitro (closed company) and radioactive labels from Perkin-Elmer (Waltham, USA). Regarding the Vtg ELISA, primary antibody was obtained from Biosense laboratories (Bergen, Norway) and secondary antibody from Biorad (Solna, Sweden). The horseradish peroxidase kit was purchased from Biorad (Solna, Sweden). For the EROD activity assay, NADPH, 7-ethoxyresorufin and rhodamine were obtained from Sigma-Aldrich (Dorset, UK). Other chemicals used in this study were of the analytical grade.

Statistical analysis Data was obtained by direct measurements during the sampling times and after laboratory analysis of the extracted tissues (plasma and liver). For the statistical analysis a one-way ANOVA test (analysis of variances) was performed with SPSS. Homogeneity of variances was checked by a Levene’s test and the variables which not fulfil this assumption were transformed to log10 prior the ANOVA test. Variables were considered significant with a 95% confidence interval (P-value < 0.05). Tukey’s post- hoc test was implemented for a multiple group comparison.

Ethical permit The ethical permit was obtained through the Laboratory of Animal Science for Researchers in the University of Gothenburg. Joachim Sturve 220-2013

Results

Biometric data The measurements for weight, fork length and liver weight noted for each fish during the implanting day or the two and eight weeks sampling days had no statistical significance (ANOVA, p>0,05) between groups or with the control (ANOVA, p>0,05; Appendices 1 and 2). The calculated SGRW, SGLW and CF showed no differences between the groups or with the control (ANOVA, p>0,05; Appendices 1 and 2). Blood measurements for haematocrit, haemoglobin, and blood glucose showed no statistical significance between groups or with the control in the ANOVA tests (ANOVA, p>0,05; Appendices 1 and 2).

Vitellogenin in blood plasma In the two weeks exposure (Fig. 4), fish exposed to a high concentration of BPA showed a significant higher amount of Vtg in plasma compared with control fish (Fig. 4, ANOVA, p<0,05; Tukey’s, P<0,05). Vtg measurements after BPA exposure in low concentration showed no difference compared with the control (Fig. 4). Phthalate exposure did not significantly affect the levels of Vtg in blood plasma neither in high nor low exposure (Fig. 4). For the two weeks exposure to Bisphenol S, Vtg levels showed a slightly higher concentration for the BPS high dose exposure compared with controls but not statistically significant (Fig. 4; ANOVA, p>0,05; Tukey’s, p=0,182) . The plasma Vtg amount for BPS in low dose exposure remained close to the control levels (Fig. 4). Vtg, 2 weeks exposure

3,5 * 3 2,5

µg/ml 1,5 1 0,5 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H

Figure 4. Group average with standard error bars after two weeks exposure (N=67), showed as µg/ml of Vitellogenin in plasma. Control group (C, n=10) and toxicant exposed groups (BPA, n=9-10; BBP, n=9-10; BPS, n=10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish). * significant difference with the control group (n=10)(ANOVA, p-value<0,05; Tukey's post-hoc test, p-value<0,05). Data was logarithmically transformed to fulfill variance homogeneity.

For the eight weeks exposure BPA exposed fish indicated a higher amount of Vtg (Fig. 5; ANOVA, p>0,05; Tukey’s, p=0,664) compared with the control, when treated with the higher dose test. BPA

exposure in low concentration remained at control Vtg levels in plasma (Fig. 5). In the BBP treatment both exposure doses, high and low, had no significant difference with the control group (Fig. 5). With the BPS exposure, the higher concentration of toxicant induced a modest increase in plasma Vtg (Fig. 5).

Vtg, 8 weeks exposure 3,5 3 2,5 2

µg/ml 1,5 1 0,5 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 5. Group average with standard error of mean bars after eight weeks exposure (N=58), showed as µg/ml of Vitellogenin in plasma. Control group (C, n=5) and toxicant exposed groups (BPA,n=8-10; BBP,n=9-10; BPS,n=6-10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

EROD activity in liver S9 fraction For the two weeks exposure (Fig. 6), BPA exposed in low dose, indicated a lower activity trend than the control group, however not significant in the ANOVA test (Fig. 6; ANOVA, p>0,05). The group exposed to BPA in high dose presented lower EROD activity trend than the control but higher than BPA low dose group (Fig. 6). The BBP high dose exposed group had a slightly higher EROD activity trend than the control group (Fig. 6), while BBP exposed in low dose showed an almost equal activity to that of the control (Fig.6). The group exposed to BPS in low dose had a similar EROD activity compared to the control group (Fig. 6). In the BPS high dose treated group the activity remained slightly lower compared with the control (Fig. 6).

EROD activity, 2 weeks exposure 0,025

minute 0,02 · · 0,015

0,01

0,005

nmol/mgprotein 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 6. Group average with standard error of the means bars after 2 weeks exposure (n=67), showing EROD activity as nmol/mg protein·minute. Control group (C, n=10) and toxicant exposed groups (BPA, n=9-10; BBP, n=9-10; BPS, n=10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

Focusing on the eight week time exposure none of the groups showed significance in the ANOVA test (Fig. 7; ANOVA, p>0,05). Fish exposed to BPA indicated lower EROD activity trend in low and high dose treatments when compared with the control group. Fish treated with the low dose BBP had an EROD activity that was similar to control levels (Fig. 7). On the other hand, the fish exposed to BBP in high dose showed a decreased activity trend compared with the control (Fig.7). Within the BPS treatment, the low dose exposed group presented the largest decrease in EROD activity compared with the control (Fig.7; ANOVA, p>0,05; Tukey’s, p=0,286). However, fish exposed to the high dose of BPS, showed a similar EROD activity as the control group (Fig. 7).

EROD activity, 8 weeks exposure

0,025 0,02 0,015 0,01 0,005

nmol/mgprotein·minute 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H

Figure 7. Group average with standra of the mean error bars after 2 weeks exposure (N=57), showing EROD activity as nmol/mg protein·minute. Control group (C, n=5) and toxicant exposed groups (BPA,n=8-10; BBP,n=9-10; BPS,n=6- 10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

Thyroid hormones in plasma

Plasma triiodothyronine (T3) Within the two weeks exposure (Fig. 8) none of the groups showed significance in the ANOVA test (p>0,05), plasma T3 after two weeks of the BPA implant injection appeared to be similar to those in the control group for high and low concentrations of exposure (Fig. 8). For the BBP treatment, the lower dose of exposure kept the T3 levels near the control levels (Fig. 8). The higher BBP concentration induced a slightly increased trend of plasma T3 compared with the control (Fig. 8). After two weeks exposure with BPS, the higher exposure dose showed a slightly higher trend compared with the control T3 levels (Fig. 8). BPS exposure in low concentration showed T3 levels near the control levels (Fig. 8).

T3, 2 weeks exposure 16 14 12 10

6 ng/ml 4 2 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 8. Group average with standard error of mean bars after 2 weeks exposure (N=67), showed as ng/ml of triiodothyronine (T3) in plasma. Control group (C, n=10) and toxicant exposed groups (BPA, n=9-10; BBP, n=9-10; BPS, n=10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

T3 after eight weeks of exposure (Fig. 9) showed that BPA treatment did not affect T3 levels either in low or high exposure dose when compared with the control group (Fig. 9). None of the groups showed significance in the ANOVA test (p>0,05), BBP exposure in high dose had a small increasing trend on the plasma T3 concentration of the group compared with the control (Fig. 9). Plasma T3 levels after the BBP exposure in low dose remained similar to the control levels (Fig. 9). For the eight weeks exposure, BPS exposed group showed a slight increased trend in plasma T3 when treated with the low dose and compared with the control group (Fig. 9). However, no statistical significance was noted. Exposure with BPS in a high dose resulted in a T3 plasma concentration almost equal as the control group (Fig. 9).

T3, 8 weeks exposure 16 14 12

10 8 ng/ml 6 4 2 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 9. Group average with standard error of the mean bars after 8 weeks exposure (N=58), showed as ng/ml of triiodothyronine (T3) in plasma. Control group (C, n=5) and toxicant exposed groups (BPA,n=8-10; BBP,n=9-10; BPS,n=6-10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

Plasma thyroxine (T4) In the plasma T4 measurements for 2 weeks exposure (Fig. 10) none of the groups showed significance in the ANOVA test (p>0,05). The low dose BPA treated group maintained its T4 levels similar to the control group T4 concentration (Fig. 10). BPA exposed in high dose, produced a weak increased trend of T4 concentration compared with the control group (Fig. 10). Fish that were exposed to BBP in high dose indicated a slight trend to increase the T3 plasma concentration compared with the control or the BBP low dose exposed group. BPS treatment in high and low dose had no effect on the thyroxine levels compared with the control group (Fig. 10). T4, 2 weeks exposure 16 14 12

10 8

ng/ml 6 4 2 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 10. Group average with standard error of the mean bars after 2 weeks exposure (N=67), showed as ng/ml of thyroxine (T4) in plasma. Control group (C, n=10) and toxicant exposed groups (BPA, n=9-10; BBP, n=9-10; BPS, n=10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish). For the statistical analysis (ANOVA, Tukey) data was logarithmically transformed to fulfill variance homogeneity. dffgzdh

Within the eight weeks exposure (fig. 11) none of the groups showed significance in the ANOVA test (p>0,05). The quantification of T4 levels in plasma indicated that BPA low dose exposed group may

had a lower T4 concentration compared with the control (Fig. 11; Tukey’s, p=0,097).The BPA treated group with high dose indicated a T4 concentration between BPA exposed in low dose and control group (Fig. 11). BBP groups with low and high dose of exposure had almost equal concentrations of T4 in plasma, despite remaining in a lower trend than the control levels (fig. 11). As in the BPA low exposure, BPS low dose treatment produced a decreased trend in the T4 plasma levels in the exposed group compared with the control group (Fig. 11; Tukey’s, p=0,040). Compared with control, BPS exposed group in low dose indicated a lower trend for T4 levels in plasma (Fig. 11; Tukey’s, p=0,276).

T4, 8 weeks exposure 16 14 12

10 8

ng/ml 6 4 2 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 11. Group average with error bars after 8 weeks exposure (n=58), showed as ng/ml of thyroxine (T4) in plasma. Control group (C, n=5) and toxicant exposed groups (BPA,n=8-10; BBP,n=9-10; BPS,n=6-10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish). For the statistical analysis (ANOVA, Tukey) data was logarithmically transformed to fulfill variance homogeneity.

T3/T4 ratio The T3/T4 ratio in plasma was calculated dividing the concentration (ng/ml) of T3 by the concentration of T4 for each individual. T3/T4 ratio in plasma is one indicator for the T4 deiodination to the active T3 form (Eales et al., 1993). In the two weeks exposure (fig. 12) none of the groups showed significance in the ANOVA test (p>0,05). The T3/T4 ratio for the BPA low dose exposed group was similar to the control T3/T4 ratio (Fig. 12). The group treated with a high dose of BPA indicated a slightly lower ratio than the control (Fig.12). Following the BBP exposure, the groups treated with high and low dose of the toxicant had an almost equal T3/T4 ratio situated slightly below the control group ratio (Fig.12). The group treated with a low dose of bisphenol S indicated a similar ratio than BBP exposures (Fig. 12). In the other hand, the group exposed to BPS in high dose indicated a slightly higher ratio trend than the control but with no significance in the ANOVA test (Fig. 12).

t3/t4 ratio, 2 weeks exposure 3 2,5 2 1,5 1 0,5 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H Figure 12. Group average with standard error of the mean bars after 2 weeks exposure (N=67), showing the calculated T3/T4 ratio in plasma. Control group (C, n=10) and toxicant exposed groups (BPA, n=9-10; BBP, n=9-10; BPS, n=10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

Within the calculated T3/T4 ratio for the eight weeks exposure (Fig. 13) none of the groups showed significance in the ANOVA test (p>0,05), the control group indicated a generally lower ratio compared with the exposed, however, did not reach significance (Fig. 13). The T3/T4 ratio for the group exposed to BPA in low dose was almost two points higher than the control group (Fig. 13). Instead, the groups exposed to BPA in high dose, indicated a ratio lower than low dose BPA group but higher than the control (Fig. 13). BBP exposed groups had a higher ratio trend than the control but similar among the two dose exposures (Fig. 13). BPS exposed in low dose indicated the highest T3/T4 ratio trend in the exposed groups compared with the control, although not statistically significant in the ANOVA test (Fig. 13; Tukey’s, p=0,046). BPS exposed group in high concentration indicated a trend ratio lower than BPS low dose group and similar to the BBP exposure (Fig. 13).

t3/t4 ratio, 8 weeks exposure 4 3,5 3 2,5 2 1,5 1 0,5 0 C BPA-L BPA-H BBP-L BBP-H BPS-L BPS-H

Figure 13. Group average with error bars after 8 weeks exposure (N=58), showing the calculated T3/T4 ratio in plasma. Control group (C, n=5) and toxicant exposed groups (BPA,n=8-10; BBP,n=9-10; BPS,n=6-10) in low (L; 2mg/kg fish ) and high injected concetration (H; 20 mg/kg fish).

Discussion

Plasma Vitellogenin (Vtg) As seen in the results, for the two weeks exposure with BPA, there was a significant increase in Vtg when the fish were treated with a high dose of toxicant. A similar elevated concentration of Vtg was seen in fish after eight weeks of exposure with BPA in high dose, although not significant, may indicate a chronic effect of BPA. Fish treated with BPS also indicated a trend towards a slightly higher Vtg in plasma, for the high dose-exposed groups both after two and eight weeks. Fish sex was not considered for the analysis of the Vtg measurements, since only a few individuals of the overall exposure (N=125) were sexually mature at the time of the samplings. Nevertheless, BPA and BPS had observable trends on the plasma Vtg concentration in the fish injected with a high dose. Previous studies showed the estrogenic effect of BPA on juvenile fish when exposed to this compound, BPA acts as an estrogen receptor (ER) agonist enhancing vitellogenesis and increasing plasma Vtg. This supports our results and the endocrine disrupting capacity of BPA as a xenoestrogen in juvenile brown trout (Arukwe et al., 2000; Lindholst et al., 2000; Larsen et al., 2006; Flint et al., 2012).

Regarding the BPS exposure, the research on this compound as an endocrine disruptor in fish has just started and the literature available is scarce. In a recent study, Naderi et al., 2014, found that zebrafish embryos had a significant increase of plasma Vtg after a 75 days waterborne exposure with BPS at concentrations of 10 µg/L and 100 µg/L. Together with our results of an slight increasing trend of plasma Vtg after BPS exposure in high dose, this gives support to the BPS estrogenic capacity.

Cyp1a induction/EROD activity None of the EROD activity results for the two term exposures showed statistical significance. However, the tendency to lower EROD activity compared with the control shown by BPA exposed groups for either two or eight weeks exposure, sustains a continued decreasing effect on the Cyp1a enzymatic activity due to an inhibition of the Cyp1a expression in the liver. The down-regulation of Cyp1a enzymes occurs due to binding of estrogenic compounds such as natural estrogens or BPA with the estrogen receptor in liver cells. This ligand-receptor interaction enhances the crosstalk between the estrogen receptor and the AhR receptor, which modulates Cyp1a gene expression. This way, we can relate the increasing trend of plasma Vtg on the BPA high dose exposed group with the slightly decreasing tendency of EROD activity showed by the same group either in the two or eight weeks exposure (Maradonna et al., 2014; Olsvik et al., 2009; Klinge et al., 2000: Navas et al., 2001).

Although no studies were found about the effect of BPS on the Cyp1a system, a resemblance with BPA results could be expected due to the close chemical characteristics of the substance (Fig.2). Moreover BPA and BPS had similar effects regarding an increasing trend in plasma Vtg levels for the high dose exposed groups. BPS exposed in low dose had the highest Cyp1a inhibition trend after eight weeks of exposure. However the BPS high dose exposed group had an EROD activity similar to the control, making these results quite inconsistent. It could be hypothesized that BPS in high dose may have a weak estrogenic activity that does not inhibit the Cyp1a enzymes.

EROD activity and plasma Vtg levels for the BBP exposure remained similar to the control levels, only after the eight weeks exposure, the BBP high dose group indicated a trend for a small decrease in

EROD activity. These results could be supported by a weak estrogenic activity that slightly down- regulated Cyp1a expression. However, this should be supported by an increase of plasma Vtg levels that did not occur. Studies have shown no estrogenic activity of BBP in in vitro assays (Nomura et al., 2006). In addition Mankidy et al., 2013 found that fathead minnow embryos had no altered expression of Cyp1a and estrogen receptors after 96h exposure to BBP. Instead, a rapid metabolisation and excretion of BBP could be an explanation to these weak BBP estrogenic effects found in our study as pointed by Nomura et al., 2006 in in vitro studies. Nevertheless, research with small mammals showed enhanced EROD activity after BBP exposure (Singletary et al., 1997; Harju et al., 1996) and in vitro studies on liver have shown that BBP has an estrogenic effect (Jobling et al., 1995).

Plasma thyroid hormones For the T3 in plasma, the circulating levels remained almost equal to the control levels for the whole experiment, only BPS and BBP in highest doses indicated a tendency, but which was non-significant to induce an increase in T3 levels after the two weeks exposure. BPA has been shown to up-regulate zebrafish genes involved in the synthesis of thyroid hormones in the thyroid follicle (Gentilcore et al., 2013), that gives support to the hypothesis of an increase of T3 synthesis, considering the similar structure and function of BPS and BPA, although BPA results did not show any difference with the control T3 levels. Observing the T3/T4 ratio of the BPS high dose exposed group, it is higher than the control group indicating a higher T4 to T3 transformation activity that could explain the slight increase of T3 after exposure to BPS.

The trend for an increase of T3 after two weeks exposure with BBP high dose could be explained by an increase of T3 synthesis in the thyroid follicles or a higher T4 deiodination in the liver. However T3/T4 ratio remained at control levels for the BBP exposure, discarding in part a higher T4 transformation to T3. Moreover, studies with BBP have shown that it can compete with T3 to bind with the thyroid plasma transporter Transthyretin which could drive to a decrease in plasma T3 (Ishihara et al., 2003) and research in humans have shown that phthalate metabolites negatively altered T4 and T3 levels in blood serum (Meeker et al., 2007; Boas et al., 2010). However, in vitro studies have shown an increased iodide uptake by the TPO enzymes when exposed to BBP (Breous et al., 2005; Wenzel et al., 2005). This BBP induced iodide uptake could lead to higher thyroid hormone synthesis in the thyroid follicles supporting our results.

For the T4, no significant results were obtained after the two weeks or the eight weeks exposure. However, it is observable that on the two weeks exposure all the groups presented a slightly higher level of plasma T4 than the control group. This could be explained by an inhibition of the T4 down- regulation in the brain as pointed by Zoeller et al., 2005 after exposing pregnant rats to BPA, resulting in plasma T4 increase in pups after 15 postnatal days. Another explanation could be an increase of T4 synthesis in the thyroid gland, driven by higher TSH levels induced after EDC exposure. However, PCB exposure in humans that increased the levels of TSH, showed as well a decrease in plasma thyroid hormones (Osius et al., 1999; Schell et al., 2008).

On the other hand, after eight weeks of exposure all the exposed groups showed lower T4 levels than the control. One reason for the general T4 decrease after eight weeks could be an increased deiodination activity driven by Dio2 in the liver, as shown by Zhai et al. 2014 after an exposure with

phthalates; however the T3 levels after eight weeks exposure didn’t show an increase correlated to the T4 decrease. Another explanation to these results is the enhanced T4 glucoronization in the liver and therefore excretion of T4 into the bile, decreasing the amount of T4 in plasma. This has been shown in previous studies with PCB exposure (Barter et al., 1992; Hood & Klaassen, 2000). Decreased synthesis of T4 in the thyroid follicles due to a lack of iodine uptake could be another hypothesis to our results, although some studies have shown and increased iodide uptake after BBP exposure (Breous et al., 2005). Inhibition of the TPO enzyme after exposure with EDC, as shown by Schmutzler et al., 2007, would lead to a decrease of T3 and T4 synthesis in the thyroid gland, however the T3 levels are not affected after eight weeks exposure. Finally, PCBs and bisphenol derivatives have shown high affinity for TTR and prealbumin, competing with thyroid hormones for these plasma transporters and decreasing plasma thyroid levels (Rickenbacher et al., 1986; Darnerud et al., 1996; Hamers et al., 2006). These studies could support the indicated decrease in plasma T4 after the eight weeks exposure found in our study.

In this study, the disrupting effects on the thyroid system have been analysed using the plasma levels of thyroid hormones, finding a decreasing trend for the plasma T4 levels after long exposure with BPA, BPS and BBP either in high or low dose. However, the actions of the thyroid hormones are widespread in the organism and can be disrupted at many levels, pointing that our interpretations of the plasma thyroid disruption need the support from other thyroid status indicators such as deiodination activity, plasma TSH levels, thyroid receptors expression and T4 glucoronization for a complete thyroid disruption assessment.

The non-significance in the ANOVA test for the eight weeks exposure could be discussed because of the loss of individuals in the control group before the second sampling took place, five out of the original ten fish in the control group died. This fact resulted in a weaker data to be used in the ANOVA.

Conclusions As final conclusion, this study indicates the estrogenic potential of BPA as shown by the induction of VTG in juvenile brown trout, strengthening the view that BPA may act as an endocrine disruptor in fish as it does in mammals. In comparison, BPS does not seem to have as potent effects as BPA on the organism in terms of VTG. However, this should be confirmed in additional studies, since BPS is the substitute for BPA in some plastic products and could be harmful for the organism. Regarding to the thyroid system, although BPA, BPS and BBP did not show to have any statistically significant effects on the thyroid hormone levels, more analyses on other parts of the system should be carried out to investigate the potential thyroid disruption of these substances in juvenile brown trout.

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Acknowledgments I would like to thank Lisa Jönsson, Noomi Asker and Joachim Sturve for their help and guidance during this project. Thanks as well to Bethanie Almroth for her help during the sampling days and to my friends and colleagues in the zoophysiology department with whom I shared this year of science and learning.

General public abstract Triiodothyronine (T3) and thyroxine (T4) are known as the thyroid hormones, these hormones are organic molecules produced in the thyroid gland of vertebrate animals (mammals, fish, reptiles, etc.) and are involved in many biologic processes such as early development and metamorphosis. Endocrine disrupting compounds (EDCs) are man-made chemicals that can interfere with the biological function of many hormones like estrogens, thyroid hormones, etc. Bisphenol A (BPA) and benzyl butyl phthalate (BBP) are known EDCs used in the manufacture of plastics and present in the natural environment. Bisphenol S (BPS) has recently appeared as a substitute for BPA in some plastic products. The aim of this study was to investigate the potential of these compounds to interfere on the thyroid hormone system and toxicology markers of juvenile brown trout. Several groups of juvenile brown trout (Salmo trutta) were exposed to BPA, BPS and BBP by the implantation of cholesterol pellets containing a high (20mg/kg fish) or low (2mg/kg fish) compound dose for either 2 weeks or 8 weeks. The effects of these compounds on the thyroid system were measured comparing the levels of T3 and T4 in the blood of exposed fish with blood of non-exposed fish. Toxicological markers are biological parameters known to be altered when toxic compounds are present in the organism. In this study we used two toxicological markers, the hormone vitellogenin (Vtg) in blood, which normal concentration can change when exposed to chemicals with EDC activity and the activity of the Cyp1a enzymes in liver, these enzymes are in charge of processing toxic compounds in the organism and their activity can be altered depending on the presence of these compounds. Our results indicated that the levels of T3 in blood of exposed fish remained unaltered after the exposure to BPA, BPS and BBP. T4 levels showed a slight tendency to decrease in all the exposed groups after 8 weeks compared with the non-exposed fish. Regarding to the toxicological markers, there was an increase in Vtg levels in blood showed in the BPA and BPS exposed fish with the high dose of the compound for 2 and 8 weeks. Cyp1A activity indicated a tendency to decrease in the BPA low dose exposed fish for the 2 weeks exposure and in BPA and BPS exposures after 8 weeks. There was no effect of BBP exposure on the toxicological markers. Despite the fact that the results on T4 were not statistically relevant, there was still a trend supporting the idea of an alteration on the T4 levels in

blood after exposure to BPA, BPS and BBP as seen in other studies with similar compounds. However, more research on different aspects of the thyroid system is needed for a solid explanation.

Appendix Two weeks exposure treatment lenght1 lenght2 weight1 weight2 SGRW SGLW CF liver Ht Hgb Gluc C Mean 22,7900 22,9500 126,6000 129,1000 ,1735582 ,0482677 1,056693 1,7010 31,1250 77,0000 5,9600 N 10 10 10 10 10 10 10 10 8 10 10 Std. 1,42084 1,32014 27,78969 22,29823 ,51084236 ,13671023 ,0735324 ,39979 4,18970 14,62114 1,62768 Deviation Std. Error ,44931 ,41747 8,78787 7,05132 ,16154254 ,04323157 ,0232530 ,12642 1,48128 4,62361 ,51472 of Mean BPA_low Mean 22,7200 22,8500 127,2000 128,6000 ,1563999 ,0412237 1,050017 1,7800 29,6000 73,2000 5,8000 N 10 10 10 10 10 10 10 10 10 10 10 Std. 2,10227 1,96540 44,28895 37,45279 ,51290610 ,13982206 ,1015514 ,45651 2,87518 9,47277 1,02089 Deviation Std. Error ,66480 ,62151 14,00540 11,84361 ,16219515 ,04421562 ,0321134 ,14436 ,90921 2,99555 ,32283 of Mean BPA_high Mean 22,7778 23,1556 131,2222 137,1111 ,2694979 ,1111971 1,112100 2,0722 31,2500 77,2222 5,5625 N 9 9 9 9 9 9 9 9 8 9 8 Std. ,85114 ,68028 12,97861 18,46242 ,70000869 ,20725125 ,1074939 ,39366 4,92080 16,76885 ,79631 Deviation Std. Error ,28371 ,22676 4,32620 6,15414 ,23333623 ,06908375 ,0358313 ,13122 1,73977 5,58962 ,28154 of Mean BBP_low Mean 22,5900 22,7400 122,2000 123,6000 ,1191519 ,0465671 1,051109 1,7540 28,6000 75,4000 5,2100 N 10 10 10 10 10 10 10 10 10 10 10 Std. 1,32535 1,16447 21,85203 16,30406 ,52595922 ,13475795 ,0605662 ,50461 6,88315 19,51182 ,58585 Deviation Std. Error ,41911 ,36824 6,91022 5,15580 ,16632291 ,04261421 ,0191527 ,15957 2,17664 6,17018 ,18526 of Mean BBP_high Mean 23,0889 23,3000 135,4444 138,0000 ,1233100 ,0592967 1,093886 1,8100 34,1111 84,8889 5,9667 N 9 9 9 9 9 9 9 8 9 9 9 Std. ,97011 1,11803 20,54331 21,60440 ,40192528 ,16391065 ,0809574 ,57351 3,25747 10,05540 1,20623 Deviation Std. Error ,32337 ,37268 6,84777 7,20147 ,13397509 ,05463688 ,0269858 ,20277 1,08582 3,35180 ,40208 of Mean BPS_low Mean 22,5800 22,6500 121,5000 132,2000 ,5304133 ,0199680 1,043883 1,7930 27,4000 71,4000 4,5900 N 10 10 10 10 10 10 10 10 10 10 10 Std. 1,15258 1,20023 23,07115 28,06263 ,71517389 ,09311503 ,0703925 ,55562 7,18331 19,03914 ,72793 Std.Deviation Error ,36448 ,37955 7,29574 8,87418 ,22615784 ,02944556 ,0222601 ,17570 2,27156 6,02071 ,23019 BPS_high Meanof Mean 22,9600 23,0600 132,7000 130,5556 -,1611135 ,0292752 1,092537 1,6156 34,4000 83,8000 5,7300 N 10 10 10 9 9 10 10 9 10 10 10 Std. ,56804 ,52111 15,30468 16,44013 ,66080850 ,07817919 ,0663886 ,39950 4,55095 15,46897 2,35563 Deviation Std. Error ,17963 ,16479 4,83977 5,48004 ,22026950 ,02472243 ,0209939 ,13317 1,43914 4,89172 ,74492 of Mean Total Mean 22,7824 22,9500 127,9706 131,1343 ,1773206 ,0498156 1,070533 1,7871 30,8615 77,4559 5,5388 N 68 68 68 67 67 68 68 66 65 68 67 Std. 1,24540 1,18957 25,07361 23,56030 ,58947470 ,13647535 ,0815820 ,46887 5,50249 15,51353 1,35769 Deviation Std. Error ,15103 ,14426 3,04062 2,87835 ,07201580 ,01655007 ,0098933 ,05771 ,68250 1,88129 ,16587 of Mean

Appendix 1. Data from the implanting day and the sampling day after two weeks exposure, showing group mean (Mean), number of individuals per group (n), standard deviation (Std. Deviation) and standard error (Std. Error of Mean) for the measurements, fork length at the implanting day (length 1), fork length at the sampling day (length 2), weight at the implanting day (weight 1), weight at the sampling day (weight 2), specific growth rate in weight (SGRW), specific growth rate in length (SGLW), condition factor (CF), liver weight at sampling day (liver), haematocrit at sampling day (Ht), haemoglobin at sampling day (Hgb) and blood glucose at sampling day (Gluc) of control group (C) and

treatment groups BPA, BBP and BPS in low dose (BPA_low, BBP_low, BPS_low; 2mg/kg fish ), high dose (BPA_high, BBP_high, BPS_high; 20 mg/kg fish) and totals (Total).

Eight weeks exposure treatment lenght1 lenght2 weight1 weight2 SGRW SGLW CF liver Ht Hgb Gluc C Mean 22,4400 23,8000 123,8000 134,7560 ,4626497 ,3840970 1,091845 1,5360 26,4000 58,2000 5,5400 N 5 5 5 5 5 5 5 5 5 5 5 Std. ,81731 1,56684 13,84558 29,41149 ,92627788 ,25058014 ,0121909 ,53074 1,51658 7,19027 1,27004 Deviation Std. Error ,36551 ,70071 6,19193 13,15322 ,41424406 ,11206285 ,0054519 ,23735 ,67823 3,21559 ,56798 of Mean BPA_low Mean 22,8625 23,6625 120,8750 128,6213 ,3774265 ,2981311 1,003535 1,6925 27,6250 60,1250 7,9875 N 8 8 8 8 8 8 8 8 8 8 8 Std. 1,19156 1,46184 21,22288 27,81433 ,85274667 ,15904408 ,0653803 ,42500 4,62717 8,44203 3,63257 Deviation Std. Error ,42128 ,51684 7,50342 9,83385 ,30149148 ,05623057 ,0231154 ,15026 1,63595 2,98471 1,28431 of Mean BPA_high Mean 23,4111 24,3333 140,8889 140,0044 -,1017615 ,2299546 ,977909 1,9200 24,4444 61,3000 7,5000 N 9 9 9 9 9 10 10 10 9 10 8 Std. 1,44347 1,60857 27,95731 32,82987 ,99245450 ,22511496 ,3495284 ,77942 6,40529 11,81383 2,93987 Deviation Std. Error ,48116 ,53619 9,31910 10,94329 ,33081817 ,07118760 ,1105306 ,24647 2,13510 3,73586 1,03940 of Mean BBP_low Mean 21,3900 22,2700 106,9000 114,0830 ,3869235 ,3022729 1,071109 1,5150 23,5556 52,5000 6,4200 N 10 10 10 10 10 10 10 10 9 10 5 Std. 2,11211 2,09075 27,40418 31,29534 ######## ,22558229 ,0759583 ,44044 3,84419 12,55433 1,08259 Deviation Std. Error ,66791 ,66115 8,66596 9,89646 ,42625493 ,07133538 ,0240201 ,13928 1,28140 3,97003 ,48415 of Mean BBP_high Mean 22,6778 23,3556 126,0000 123,1511 -,3554994 ,2058770 1,066488 1,7933 28,1111 67,3333 8,0889 N 9 9 9 9 9 9 9 9 9 9 9 Std. 1,65135 2,18066 27,40438 39,96960 ######## ,44723113 ,0817746 ,84156 3,29562 8,58778 4,71154 Deviation Std. Error ,55045 ,72689 9,13479 13,32320 ,71699213 ,14907704 ,0272582 ,28052 1,09854 2,86259 1,57051 of Mean BPS_low Mean 22,5286 23,4143 117,2857 125,6214 ,4215238 ,2556108 1,020935 2,2017 23,3333 56,5714 10,2333 N 7 7 7 7 7 7 7 6 6 7 6 Std. 1,28415 1,43693 18,05283 25,14297 ,72976072 ,20133377 ,0633947 ,42527 6,65332 11,57378 3,89084 Deviation Std. Error ,48536 ,54311 6,82333 9,50315 ,27582362 ,07609701 ,0239610 ,17362 2,71621 4,37448 1,58843 of Mean BPS_high Mean 22,2300 23,4900 116,1000 127,7610 ,6259885 ,3671390 1,047321 1,7000 27,4444 66,9000 8,2900 N 10 10 10 10 10 10 10 10 9 10 10 Std. ,94874 1,03543 20,26190 23,48989 ,89930692 ,18965640 ,0761957 ,66966 5,45690 16,86845 3,22402 Deviation Std. Error ,30002 ,32743 6,40737 7,42815 ,28438582 ,05997462 ,0240952 ,21177 1,81897 5,33427 1,01953 of Mean Total Mean 22,4793 23,4310 121,3621 127,0507 ,2465017 ,2871418 1,037218 1,7572 25,9091 60,6949 7,8588 N 58 58 58 58 58 59 59 58 55 59 51 Std. 1,51655 1,71118 24,68446 30,06330 ######## ,25334720 ,1549795 ,62926 5,00101 12,41836 3,42171 Deviation Std. Error ,19913 ,22469 3,24123 3,94751 ,16205135 ,03298300 ,0201766 ,08263 ,67434 1,61673 ,47913 of Mean

Appendix 2. Data from the implanting day and the sampling day after eight weeks exposure, showing group mean (Mean), number of individuals per group (N), standard deviation (Std. Deviation) and standard error (Std. Error of Mean) for the measurements, fork length at the implanting day (length 1), fork length at the sampling day (length 2), weight at the implanting day (weight1), weight at the sampling day (weight 2), specific growth rate in weight (SGRW), specific growth rate in length (SGLW), condition factor (CF), liver weight at sampling day(liver), haematocrit at sampling day (Ht), haemoglobin at sampling day (Hgb) and blood glucose at sampling day (Gluc) of control group (C) and treatment groups BPA, BBP and BPS in low dose (BPA_low, BBP_low, BPS_low; 2mg/kg fish ), high dose (BPA_high, BBP_high, BPS_high; 20 mg/kg fish) and totals (Total). (######## data not available do to negative values of SGRW)