fish cell lines and their potential uses in ecotoxicology: from … · 2015-08-18 · vivo testing...

Fish cell lines and their potential uses in ecotoxicology: from cytotoxicity studies and mixture assessment to a co-culture

model and mechanistic analyses

Fabienne Roux

Master in Ecotoxicology

2014-2015

Department of Biological and Environmental Sciences

University of Gothenburg

Examiner: Thomas Backhaus

Supervisor: Joachim Sturve

Fabienne Roux Master thesis in Ecotoxicology 2014-2015

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I. CONTENT

I. CONTENT 1

II. ABSTRACT 2

III. LIST OF ABBREVIATIONS 3

1 INTRODUCTION 4 1.1 TESTING AND ALTERNATIVE METHODS 4 1.2 PRIMARY CULTURES VS. CELL LINES 5 1.3 FISH CELL LINES 6 1.4 BIOMARKERS 7 1.5 CHALLENGES AND POSSIBLE SOLUTIONS 10 1.5 THESIS AIMS 12 2 MATERIALS AND METHODS 13 2.1 CHEMICALS AND CONSUMABLES 13 2.2 TEST SYSTEM 13 2.4 METHODS 16 3 RESULTS 21 3.1 MIXTURE TOXICITY ASSESSMENT (NICE) – CYTOTOXICITY 21 3.2 MECHANISTIC ANALYSES – SINGLE CELL AND CO-CULTURE SYSTEMS 25 4 DISCUSSION 31 4.1 MIXTURE TOXICITY ASSESSMENT (NICE) – CYTOTOXICITY 31 4.2 MECHANISTIC ANALYSES – SINGLE CELL AND CO-CULTURE SYSTEMS 33 5 CONCLUSION 38 6 ACKNOWLEDGEMENTS 39 7 REFERENCES 40


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II. ABSTRACT

The environmental risk assessment of chemicals is mostly based on in vivo single

compound experiments. However, due to the increasingly high amount of chemicals,

metabolites and mixtures to be tested, optimization of this assessment is required. Interest

in in vitro methods has been growing greatly in the recent years for economical, practical

and ethical reasons, and the use of cell lines as alternatives to in vivo testing is being

seriously considered. This thesis focused on two different aspects of methods using fish

cell lines. The first was the use of quick and simple cytotoxicity assays as screening

method for assessment of mixture toxicity in the specific case of Swedish coastal waters.

Three compounds, copper (II) ions, metoprolol and tributylphosphate, were tested on

RTgill-W1, singularly and in combination based on their EC50 and environmental ratios.

Both the Concentration Addition and the Independent Action models were used to predict

mixture toxicity. The predictions were in a factor of 2 compared with the observed

EC50s. The assay itself showed variability but could be optimized for further mixture

testing. The second, more complex assay consisted of a co-culture model as an attempt to

improve basic cell line assays. Two different cell lines (RTgill-W1 and RTL-W1)

separated by a permeable insert mimic the direct exposure of the gill epithelium to

environmental toxicants and indirect exposure of the liver through uptake and

metabolism. The effects on gene expression level, EROD activity and glutathione content

of benzylbutylphthalate and bisphenol A were analyzed. The obtained results indicate

indirect exposure of RTL-W1. The properties of this co-culture model are promising but

need to be analyzed more in depth before this model can be used in toxicity tests. The

characterization of relevant biomarkers and pathways in these cell lines is needed, in

particular for endocrine disruption.

Keywords: cell lines, in vitro alternatives, RTgill-W1, RTL-W1, NICE, cytotoxicity,

mixture, concentration addition, independent action, co-culture, gene expression,

biomarkers, endocrine disruption


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III. LIST OF ABBREVIATIONS

AB

BBP

BNF

BPA

CFDA-AM

Cu

CYP1A

CYP3A

Dio2

DMSO

EC50

EDC

EDTA

ER

ERα

ERRγ

EROD

FBS

GPx

GSH

GSSG

GST π

L-15

L-15/ex

Met

mRNA

Nrf-2

PBS

PEPCK

PPARγ

PS

REACH

RT-qPCR

SE

SULT

TBP

TER

tGSH

TTR

TRα/β

UGT

3R

AlamarBlue

benzylbutylphthalate

β-naphthoflavone

bisphenol A

5-carboxyfluorescein diacetate acetoxymethyl ester

copper (II)

cytochrome P450 1A

cytochrome P450 3A

deiodinase 2

dimethyl sulfoxide

effective concentration 50

endocrine disrupting chemical/compound

ethylenediaminetetraacetic acid

7-ethoxyresorufin

estrogen receptor α

estrogen related receptor γ

7-ethoxyresorufin-O-deethylase

fetal bovine serum

glutathione peroxidase

glutathione

oxidized glutathione

glutathione-S-transferase π

Leibovitz's L-15 medium

Leibovitz's L-15 medium/exposure

metoprolol

messenger ribonucleic acid

NF-E2-related factor 2

phosphate buffered saline

phosphoenolpyruvate carboxykinase

peroxisome proliferator-activated receptor γ

penicillin streptomycin

Registration, Evaluation and Authorisation of Chemicals

reverse transcriptase - quantitative polymerase chain reaction

standard error

sulfonyl transferase

tributylphosphate

transepithelial electric resistance

total glutathione

transthyretin

thyroid receptor α/β

UDP-glucuronosyltransferase

reduce, refine, replace


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1 Introduction

Our society relies more and more on a

wide variety of chemicals present in

every action of our daily life. For

years, these chemicals have been put

on the market with an only limited

knowledge of their hazard to human

health and to the environment.

However, due to the very high amount

of chemicals produced and to their

ever increasing presence in the

environment, it has become of major

importance to determine the effects of

these compounds. In Europe, the

implementation of REACH in 2006,

the European legislation for the

Registration, Evaluation and

Authorization of Chemicals, is

supposed to increase the knowledge

about these chemicals by improving their risk assessment. From June to December 2008,

about 150 000 compounds were pre-registered [1] to be tested and evaluated. Even

though all of these chemicals are not expected to be registered and put on the market, the

amounts are still substantial.

Most of these chemicals, once entered into the environment, end up in the aquatic

compartments, through waste water treatment plant (WWTP) and industrial effluents,

agricultural run-offs, accidental spills etc. (see Figure 1). This forms unpredictable

cocktails of contaminants with effects that are impossible to grasp today. Development of

methods to assess mixture toxicity is greatly required but challenging due to the

complexity of the situation, meaning that regulatory testing does not allow proper

consideration of this issue. Projects such as NICE (Novel Instruments for effect-base

assessment of chemical pollution in Coastal Ecosystems, www.nice.gu.se) aim to develop

new methods to specifically assess mixture toxicity1.

1.1 Testing and alternative methods

Traditional toxicity testing is mostly based on in vivo testing of single compounds on

three aquatic species representing different trophic levels: algae, daphnia and fish.

REACH would lead to an infinite amount of work to test all chemicals as is required, not

even taking into account biotransformation or degradation products, as well as the

endless amount of possible mixtures. In vivo testing is extremely time-consuming and

costly, requiring much maintenance and a high number of animals, which is ethically

1 NICE is composed of exposure assessment of the Swedish west coast and toxicity testing in different

systems (fish, invertebrates, microbes), using traditional approaches with dose-response curves but also

omics and bioinformatics. A legal contribution to assess actual legislation and to implement environmental

quality standards and monitoring approaches taking into account mixture exposure is included.

Figure 1 Multiple stressor exposure of the aquatic

environment. Toxicants enter the aquatic

environment directly or indirectly, for example

through industrial and WWTP effluents, leachates

from landfills, agricultural run-offs, atmospheric

deposition, boating activities, accidental spills etc…

http://www.nice.gu.se/


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debated. Therefore REACH supports development of alternative methods. The EURL-

ECVAM, the European Union Reference Laboratory for alternatives to animal testing,

former European Center for Validation of Alternative Methods, is actively working on

their development, according to the 3R strategy, Reduce, Refine, Replace, concept which

was coined by Russel and Burch in 1959 [2].

An important alternative to in vivo methods are in vitro methods, literally meaning “in

glass”. Instead of living whole organisms, subsets of these are used for the experiments,

such as organs and tissues, primary cell cultures and cell lines. The justification behind

this comes from the fact that the first interaction of a toxicant with an organism happens

at the cellular level. The changes provoked in cells due to this interaction can then

possibly translate to higher levels of organization, finally impacting the whole organism

[3]. Many in vitro models have already been validated by the OECD (Organisation for

Economic Co-Operation and Development) for regulatory purposes as alternatives to in

vivo testing in the case of human health studies2. The use of in silico methods such as

QSARs and toxicokinetic models is also recognized for its provision of important

information, adding insights into the mechanistic dimension of chemical interaction and

allowing optimization of toxicity testing.

1.2 Primary cultures vs. cell lines

Primary cultures and cell lines are major

components of in vitro methods. Primary

cultures are cells, tissues or organs directly

obtained from the organism and maintained in

laboratory conditions for a certain number of

days (Figure 2). If these cells can be kept alive

through passaging, they become cell lines,

which in some cases can be maintained

indefinitely [3]. The use of these cell lines has

many advantages. It avoids the testing of

contaminants on living animals or even the

regular sampling of cells for primary cultures

since, if immortalized, they constitute an infinite

supply. Their maintenance is less demanding since the only requirements are cell medium

and an incubator at the right temperature and CO2 concentration (which is even

unnecessary in the case of piscine cell lines). The costs in money and time are thus

greatly diminished, and the testing in itself uses very limited amounts of the test

chemicals and creates little toxic waste (compared to, for example, a flow-through

aquarium system). Results present little variability since the cell lines are relatively

homogeneous and used in a very controlled environment, the complex interactions

happening in a whole organism being avoided [3-5]. They are also a very interesting

alternative if the species considered cannot be used for in vivo experiments, eg if it cannot

be kept in laboratory conditions; or to test and compare the sensitivities of many different

species on a wide scale when exposed in the exact same way [5].

2 http://www.oecd-ilibrary.org/environment/oecd-guidelines-for-the-testing-of-chemicals-section-4-health-

effects_20745788 (May 2015).

Figure 2 Primary culture. Isolation of

hepatocytes from a rainbow trout. Picture

taken by Fabienne Roux.


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However, these strengths of cell lines are also their major limitation. Indeed, the system

being extremely simplified, the relevance of results obtained in vitro can be questioned.

Cells are grown in an artificial environment, often in monolayers and thus lacking their

normal 3D disposition. It is mostly only one cell type that is considered, thus all

interactions between different cell types, as would normally take place in any organism,

are lost. Immortalization often requires them to be isolated from cancerous tissue, or to

treat them with UV light and viruses [6], knowing that the behavior of these cells could

therefore be modified compared to cells in the normal tissue. Generally, cell lines are also

much less sensitive than whole organisms, which is a major drawback to their use as

alternatives to in vivo testing. Importantly, their differential status is low and they tend to

lose capacities (eg expression of certain receptors, metabolic enzymes, etc.) over time.

An in depth characterization is therefore necessary if they are to be used for mechanistic

studies.

One way to circumvent these problems is to use freshly isolated primary cultures. These

are more differentiated, contain different types of cells and are thus thought to respond

more similarly to a living animal, while at the same time reducing the amount of work

and animals required for in vivo testing. However, these preparations are quite

complicated and due to their increased complexity level compared to cell lines, the results

tend to show higher variability. They also require the organism of interest to be available,

maintainable in laboratory conditions and of sufficient size to allow primary culture

isolation [5].

1.3 Fish cell lines

In the context of aquatic ecotoxicology, fish are the most diverse group of vertebrates, as

they account for roughly 33 000 species3 distributed in all aquatic niches [5]. It is the

dominant vertebrate species for the regulatory evaluation of ecotoxicity [3], in particular

for the Fish Acute Lethality Test (OECD), and is granted the same legal protection as

mammals when it comes to laboratory testing. Interestingly, only two alternative methods

to in vivo testing have been validated up to date in this field: the Threshold Approach for

Acute Fish Toxicity Testing, using a tiered approach to reduce the number of fish used,

and the Fish Embryo Acute Toxicity Test. The development of methods using cell lines

are ongoing but have not been validated yet. This is in particular due to the lesser

sensitivity of these test systems [7].

The first interest in fish cell lines was mostly driven by fisheries and their economic

interest in fish viruses, but by the 1980s these cell lines started to be used for

toxicological studies as well [4, 8]. Today, more than 280 cell lines have been

established, 43 being commercialized [7].

Fish cell lines are considered as being better representatives for fish than the more

commonly used mammalian cell lines, since they express specific enzymes and proteins

and can be exposed at more relevant temperatures. They do not require specific

incubators supplemented with carbon dioxide and can be kept for extended periods at

4°C, avoiding unnecessary freezing and thawing steps. Interestingly, they also tend to

3 According to FishBase, http://www.fishbase.se/search.php, June 2015.

http://www.fishbase.se/search.php


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immortalize spontaneously, which means that they do not require to be taken from

cancerous tissue or manipulated after isolation [4, 5]. This seems to come from a

relatively high telomerase expression in fish tissue [9].

Fish cell lines are therefore extremely relevant and economically important when it

comes to fish research, in ecotoxicology or more general fish health studies. They are

already used to assess mixture toxicity of chemicals such as PAHs, pharmaceuticals and

personal care products [10-13], but also WWTP effluent samples [14]. Moreover, they

might even be interesting to use instead of mammalian cell lines for general toxicity

studies as fish cell lines are easier to maintain and handle than mammalian cell lines [4]

and as cytotoxicity results prove to be similar. The basal cytotoxicity concept coined by

Ekwall in 1983 [15], which states that acute death of a cell occurs through interference of

toxicants with its fundamental functions in the majority of cases, irrespective of the cell

type, has proven very efficient [16].

1.4 Biomarkers

Interest in fish cell lines is not limited to general toxicity. More in-depth mechanistic

analyses using biochemical assays are common if the pathway considered has been

shown to be functioning in the cell line of interest. The development of biomarkers is for

example an important field. In ecotoxicology, biomarkers are biochemical, physiological

or histological markers such as metabolites, enzymes etc. whose presence can be related

to exposure and often effect of xenobiotics [17].

General pathways which are typically considered are oxidative stress and metabolism.

More specific pathways are also analyzed depending on the type of chemicals that are

considered. Endocrine disrupting compounds (which will be used as model compounds in

one part of this thesis), by definition exogenous compounds with the potential to disturb

any hormonal regulation and the normal endocrine system, are of importance today due

to their widespread presence and potential effects at very low concentrations. They have

been shown to not only interact with the estrogen and androgen receptors (ER and/or AR)

but also with different hormone and nuclear receptors (thyroid, glucocorticoid and

estrogen-related receptors TR, GR and ERR), xenosensors (pregnane X, constitutive

androstane and aryl hydrocarbon receptors PXR, CAR and AhR) and peroxisome

proliferator-activated receptors (PPARs) [18, 19]. Markers related to these systems could

be interesting biomarkers. The pathways considered in this thesis are introduced below.


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1.4.1 Oxidative stress

Oxidative stress (Figure 3) is an

imbalance between the production

of reactive oxygen species (ROS)

and antioxidant defenses. One major

antioxidant molecule is glutathione

(GSH), as it serves as reducing

agent by being oxidized to GSSG.

For example, glutathione peroxidase

(GPx) can catalyze its oxidation to

reduce hydrogen peroxide, a

byproduct of oxidative stress. GSSG

can then be recycled and returned to

the antioxidant pool through its

reduction by glutathione reductase

(GR). Glutathione synthetase (GS)

and glutamate-cysteine ligase

(GCL) are both involved in de novo

synthesis of GSH. Many of the genes

for oxidative stress, but also for phase II metabolism, are controlled by the transcription

factor Nrf-2. Total glutathione, GSSG:GSH ratio and the analysis of the gene expression,

amount and activity of the fore-mentioned proteins are all biomarkers for oxidative stress.

1.4.2 Xenobiotic metabolism

The metabolism of xenobiotics

is divided in different phases

(Figure 4). In phase I, the

xenobiotic is oxidized, reduced

or hydrolyzed to increase its

reactivity and hydrophilicity.

The enzymes from the CYP

superfamily are mainly

responsible for phase I

oxidative metabolism. CYP1A

is primarily responsible for the

metabolism of PAHs, dioxins

and PCBs and is induced by the

AhR. Its activity is a major

biomarker in the field of

ecotoxicology. CYP3A is the

most versatile CYP enzyme and

is responsible for the

metabolism of several types of

chemicals, such as

pharmaceuticals but also steroid hormones. Phase II consists of the conjugation of

Figure 3 Oxidative stress. A major constituent of the

antioxidant defense is GSH. Used as reduction unit by

GPx, oxidized glutathione, GSSG, is recycled by GR to

return to cellular pool of GSH. GS and GCL are

involved in de novo synthesis of GSH. Diagram by

Fabienne Roux.

Figure 4 The 3 phases of metabolism. Phase I is an activation

phase, allowing conjugation of hydrophilic molecules or groups to

the activated xenobiotic during phase II. The compound is

excreted during phase III. The expression of CYP1A, a major

phase I metabolic enzyme, is under transcriptional control of the

AhR receptor. Diagram by Fabienne Roux.


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hydrophilic molecules to the xenobiotic to facilitate its excretion, such as glutathione

(GSH), sulfonic acid or glucuronic acid. The enzymes responsible for this are

glutathione-S-transferases (GST), sulfotransferases (SULT) and UDP-

glucuronosyltransferases (UGT). Finally, in phase III, the metabolized compound is

pumped out of the cells. Metabolism occurs mainly in the liver but has also been

observed in other organs such as the gills, intestine or kidneys [20].

1.4.3 Sugar metabolism

Glucocorticoids interact with the

glucocorticoid receptor (GR) in

response to physical and

emotional stresses, inducing

gluconeogenesis and oxidation of

fatty acids [18] (see Figure 5). A

key enzyme in hepatic

gluconeogenesis is the

phosphoenolpyruvate

carboxykinase (PEPCK) as it

controls its rate-limiting step

[21]. The estrogen-related

receptors (ERR), targets of

bisphenol A, have been shown to

regulate PEPCK [21, 22].

1.4.4 Thyroid system

The thyroid hormone (TH) plays an

important role in vertebrate

development, growth and reproduction

[23]. In fish, T4, the inactive form, is

secreted by the thyroid gland and then

transported to peripheral tissues, mostly

bound to transport proteins such as

transthyretin (TTR) [24], where it can be

activated to T3, its active form, which

can bind to the thyroid receptor (TR)

(Figure 6). This activation is performed

by iodothyronine deiodinases. In fish,

only two deiodinases can perform this,

type I and type II. T3 can be deactivated

by type I and type III deiodinases and by

UGT, an important phase II metabolic

enzyme. Thyroid disruption has been

observed at all levels in fish and fish

embryos when exposed to phthalates and

BPA [25-28].

Figure 6 The thyroid hormone pathway. The TH is

synthesized by the thyroid gland as T4, the inactive

form, through TSH stimulation. T4 is then activated

to T3, the hormone form which can bind to the TR, in

peripheral tissues such as the liver. Dio2 is enzyme

responsible for this activation in the liver. Diagram

by Fabienne Roux.

Figure 5 The glucocorticoid receptor pathway.

Glucocorticoids interact with the GR to induce transcription of

genes related to gluconeogenesis or oxidation of fatty acids.

PEPCK is under transcriptional control of the GR, but has also

been shown to be regulated by ERRγ. Diagram by Fabienne

Roux.


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1.5 Challenges and possible solutions

The major challenge when it comes to in vitro methods is extrapolation. Indeed, results

obtained at the cellular level might not translate to any effect at the organism level, since

the level of complexity is much higher and compensation processes can occur [29]. In the

case of ecotoxicology, where extrapolation to the whole environment is required, this

challenge becomes very relevant.

The fact that cell lines prove to be much less sensitive to toxicants than whole organisms

has been shown to be due to a number of factors. The bioavailability of compounds in a

test system can be significantly reduced in the case of lipophilic (binding to the plastic,

serum proteins, etc.) and volatile compounds (evaporation), leading to overestimation of

exposure and thus underestimation of toxicity [3, 30]. Using serum free medium and

accounting for the bioavailable fraction are efficient improvements [3]. The dosing

procedure is also very important. Moreover, when toxicity is not only triggered by basal

cytotoxicity but by a specific pathway or at a specific target site, the test system must be

chosen accordingly [3, 8]. By keeping these limitations in mind, the sensitivity of the

assay can be greatly improved. Tanneberger et al [8] showed that by taking these factors

into account, it is possible to predict in vivo fish acute toxicity for up to 73% of the 35

organic chemicals tested on RTgill-W1 with less than a 5-fold difference, which is very

promising and supports the idea that in vitro cytotoxicity assays could eventually replace

the Fish Acute Lethality Test [7]. Unfortunately, when it comes to specific pathways,

better characterization of cell lines is required, especially since their differentiation status

is low. It is often not known if a cell line possesses the pathway of interest at all, or if it is

functioning. The use of recombinant cell lines is of true interest when it comes to the

study of mechanisms, but has the limitation that it is an artificial system.

Other attempts to render in vitro systems more realistic have been for example using 3D

cell structures such as hepatocyte spheroids to take into account 3D cell interactions. [31]

suggest indeed that rainbow trout hepatocyte spheroids retain a high biochemical,

morphological and functional status and that they could prove to be more realistic models

for toxicity testing than the traditional 2D monolayers.

Additional promising approaches are the co-culture

methods. In these methods, different cell types are grown

together [32], to account for their direct/indirect interactions.

This can be performed by growing the cell lines together,

but also on permeable inserts to keep them separated (see

Figure 7), which can be used to mimic for example uptake

of a toxicant and subsequent exposure of internal organs.

The in vitro epithelium can be exposed to two different

media on the basolateral and apical side, representing the

natural conditions it is usually exposed to. Full culture

medium supplemented with serum can be used to mimic the

physiological conditions on the basolateral side, whereas

more simple and serum free media, sometimes even

environmental samples, can be added on the apical side. Figure 7 Transwell inserts

from Greiner Bio-one.

Image from http://www.

greinerbioone.com/nl/belgiu

m/articles/news/42/ (May

2015)


11

This system should allow in vitro analysis of physiological barrier properties, such as ion

flow but also uptake, metabolism and excretion of toxicants, with their effects on the

second cell type.

The assembly of such in vitro epithelia has been successfully performed with primary

cultures of gill epithelium, which can be exposed to water on the apical side (reviewed in

[33]). The use of cell lines has also been considered. The capacity of RTgill-W1 to form

tight monolayers has been assessed by [34], who reported the presence of functional tight

junctions, suggesting that this cell line could be used to test physiological properties of

gills. However, they noted a few important drawbacks needing more investigation, such

as the low transepithelial resistance (TER), the lack of detection of the tight junction

proteins on Western Blot and the low expression of certain key epithelial proteins.

Drieschner [35] evaluated an intestinal barrier model composed of a rainbow trout

intestinal cell line, RTgutGC. Indications of the presence of tight junction proteins were

found and even though no differentiation of the cells was observed, the results are

promising and further research could lead to the development of the first piscine cell line

intestinal barrier model. The established culture protocol was also used by Catinot [36] to

analyze a co-culture model using RTL-W1 underneath the inserts. These studies indicate

a promising potential of these insert methods and encourage their further evaluation.


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1.5 Thesis aims

The general aim of this thesis was to assess different uses of in vitro fish cell systems in

ecotoxicology, and was divided in two parts. During the first part, the use of RTgill-W1

for the screening of single compounds and mixture cytotoxicity was applied in the

context of the NICE project. The 3 chosen compounds were copper ions, metoprolol and

tributylphosphate (Figure 8), based on the results of an environmental exposure

assessment of the Swedish West Coast.

The aim of this first part was to assess whether this strategy was suitable for the analysis

of the toxicity of each compound and of mixtures, and thus whether it could be an

interesting tool to use for mixture assessment.

During the second part, a co-culture system of a gill cell line, RTgill-W1, and a liver cell

line, RTL-W1, was tested to analyze its use as tool for in depth mechanistic studies4

based on specific biomarker responses. The model compounds chosen were 2 endocrine

disrupting chemicals (EDCs), benzylbutylphthalate (BBP) and bisphenol A (BPA), due to

their high ecological relevance. The aim of this second part was to analyze the properties

of this system and its suitability for toxicity testing, in particular in the case of EDCs.

4 V3R application, Joachim Sturve.

Cu2+

Figure 8 The three compounds chosen for the mixture toxicity assessment on RTgill-W1 according to the

chemical monitoring on the Swedish west coast (NICE): from left to right, copper ions, metoprolol and

Tributylphosphate.


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2 Materials and methods

2.1 Chemicals and consumables

Acetonitrile, AlamarBlue (AB), benzylbutylphthalate (BBP), betanaphthoflavone (BNF),

bisphenol A (BPA), 5-carboxyfluorescein diacetate acetoxymethyl ester (CFDA-AM),

copper sulfate (CuSO4), dimethyl sulfoxide (DMSO), 5,5'-dithio-bis-(2-nitrobenzoic

acid) (DTNB), 7-ethoxyresorufin (ER), ethylenediaminetetraacetic acid (EDTA),

fluorescamine, glutathione (GSH), glutathione reductase (GR), methanol, metoprolol

(MET), nicotinamide adenine dinucleotide phosphate (NADPH), 5-sulfosalicylic acid (5-

SSA), tributyl phosphate (TBP) were all purchased from Sigma-Aldrich (Stockholm,

Sweden). Fetal bovine serum (FBS), Leibovitz's L-15 medium (L-15), penicillin

streptomycin (PS) and trypsin were purchased from Gibco (Stockholm, Sweden). 6-well

plates Cellstar® and 6-well inserts ThinCerts TM were purchased from Greiner Bio-One

(Stockholm, Sweden). 48-well plates were from VWR (Stockholm, Sweden). 6-well

plates, 96-well plates and the culture flasks for RTgill-W1 were from Sarstedt

(Helsinborg, Sweden). Culture flasks for RTL-W1 were from Techno Plastic Products

AG (Trasadingen, Switzerland). RNeasy® Plus Minikit and QIAshredder were purchased

from Qiagen (Sollentuna, Sweden). Experion, iScript cDNA synthesis and SSOAdvanced

TM Universal SYBR® Green Supermix were from BioRad (Sundbyberg, Sweden).

Stock solutions of the used chemicals were prepared and stored as follows. In sterile

MilliQ water, 5 mg/mL Cu2+ (4°C) and 10 mM metoprolol (4°C). In DMSO, 400 mg/mL

TBP (4°C), 10 mM BNF (-20°C), 10 mM BBP (freshly prepared), 1 M BPA (freshly

prepared).

2.2 Test system

2.2.1 Rainbow trout as a model

Rainbow trout (Oncorhynchus mykiss) is the most commonly studied coldwater species

[37]. Easily farmed and maintained in laboratory conditions, it is a useful model and a

standard species for toxicity testing. It is considered as one of the most sensitive species

when it comes to acute toxicity (OECD5). One major advantage of the rainbow trout is

the great diversity of cell lines that have been established from this organism [5, 35]

(Figure 9). This gives the possibility to study a variety of endpoints using a combination

of different cell lines. The improvement of this concept could lead to the creation and use

of a “virtual fish” in ecotoxicology [35]. During this thesis, two different cell lines from

rainbow trout were used: RTgill-W1 for the gill, RTL-W1 for the liver.

5 https://books.google.se/books?id=I-PvBQAAQBAJ&lpg=PA72&ots=5WbCeZKn1W&dq=OECD%20

rainbow% 20trout&pg=PA119#v=onepage&q=OECD%20rainbow%20trout&f=true

https://en.wikipedia.org/wiki/Nicotinamide_adenine_dinucleotide_phosphate


14

The gill is an organ of vital importance for fish. Responsible for gas exchange,

osmoregulation, ionic diffusion, pH regulation, nitrogen balance [38], it is the primary

site of uptake of waterborne contaminants. This organ is also metabolically active. Due to

the pivotal role of gills, any damage can be expected to have severe consequences on the

health status of the organism and it is thus of great interest to ecotoxicology. The study of

gill has mainly been performed in vivo but primary cultures of gill epithelium have been

maintained on permeable inserts, allowing formation of confluent monolayers and tight

epithelium with physiological and morphological similarities to in vivo gill epithelium

[33, 39]. The use of RTgill-W1 instead of primary cultures is being considered [34]. A

second organ of major importance in the ecotoxicology of fish is the liver due to its high

metabolic capacities and crucial detoxification role. It is also very important for

vitellogenesis. A wide array of biomarkers related to this organ has been developed [20].

Freshly isolated hepatocytes are a common in vitro model and hepatic cell lines can be

found for a number of fish.

2.2.2 Cell lines

RTgill-W1 RTgill-W1 is an epithelial cell line

derived from gill explants of a normal

adult rainbow trout (Figure 10) [39,

40]. It has been shown to contain

pavement cells, mitochondria-rich cells

and goblet cells depending on the

culture conditions [39]. This cell line

can tolerate simple buffers without

serum, such as L-15/ex, and can even be

exposed directly to environmental

samples, without extraction or

concentration steps [14, 39].

Tanneberger et al [8] showed that

cytotoxicity results on RTgill-W1, by

using simple buffers and taking into

Figure 9 Main cell lines prepared from rainbow trout tissue. Diagram by Fabienne Roux.

Figure 10 RTgill-W1 cell line. Image taken by

Britt Wassmur.


15

account the true exposure, correlate with in vivo LC50s with less than a five-fold

difference in 73% of cases, including basal cytotoxicity but also some more specific

mechanisms. This cell line is presently being investigated for its use as an approved

alternative method by EURL-ECVAM in the context of fish acute toxicity testing

(CellSens project, [7]).

RTgill-W1 has already been used in many toxicological studies, such as for metals, waste

water and industrial effluents [3, 41], nanoparticles, viruses. It has been shown to form

tight monolayers when grown in transwell membrane chambers, expressing tight junction

proteins to some degree [34, 41] which allows it to be exposed to different media in each

compartment. Even seawater exposure can be performed as long as L-15 is supplemented

in the basolateral compartment [42, 43].

The metabolic capacities of RTgill-W1 have not been assessed thoroughly. Schirmer et al

recommend the use of RTgill-W1 for cytotoxicity testing as it lacks an inducible CYP

system, avoiding metabolism to influence exposure and cytotoxicity [44, 45]. However,

CYP1A gene expression has been shown to be induced and low EROD activity induction

is seen when exposed to BNF (see Results). The presence of other metabolic enzymes

(phase I, phase II) has not been reported.

RTL-W1

RTL-W1 is a liver epithelial cell line

derived from the normal liver of a 4-

year old adult male rainbow trout

(Figure 11). It underwent spontaneous

immortalization. This cell line was

specifically developed for its inducible

EROD activity and use in toxicological

studies [46].

RTL-W1 has been much used to test

AhR agonists and CYP1A induction, but

also to analyze cytotoxicity of

pharmaceuticals and personal care

products [12] and genotoxicity of

environmental samples [47, 48].

The metabolic capacities of RTL-W1 have been assessed by Thibaut et al [49]. The

activities of phase I enzymes such as CYP3A, CYP2K and CYP2M were very low

compared to freshly isolated hepatocytes. The same was observed for the phase II

enzymes as only low levels of UGT and of phenolic SULT activities were detected.

CYP1A, xenobiotic reductase and GST activities are better conserved [46, 49]. It has also

been reported that expression of Vtg or ER was not induced by estradiol and no increase

in vitellogenin was observed [29] which could suggest a non-functional ER pathway.

Figure 11 RTL-W1 cell line. Image taken by Britt

Wassmur.


16

2.2.3 Co-culture system

The co-culture system used in this thesis consists of two monolayers of cells separated by

a permeable insert (see Figure 7, Figure 12). The cell line grown on the insert (here

RTgill-W1) represents the epithelial biological barrier exposed to the toxicant, whereas

the cell line grown underneath (here RTL-W1), in the multiwell plate, represents the

target inner organs. The purpose of this system is to mimic uptake of toxicants through

the gill, possible metabolism and subsequent exposure of the liver.

2.4 Methods

2.4.1 Cell line maintenance

Both cell lines were obtained from Kristin Schirmer (EAWAG). They are routinely

maintained in 75 cm2 flasks containing L-15 medium with 10% (v/v) fetal bovine serum

(FBS) and without antibiotics, at 19°C.

2.4.2 Mixture toxicity assessment (NICE) – cytotoxicity

Cytotoxicity was determined using cell viability assays following the method detailed by

Dayeh et al [50], with both AlamarBlue (AB) and 5-carboxyfluorescein diacetate

acetoxymethyl ester (CFDA-AM). These probes are non-fluorescent and are taken up and

transformed to fluorescent metabolites by living cells, which can then be detected by

fluorimetry. AB is composed of resazurin and gives a measure of cellular metabolic

activity, since it is taken up by the cells and metabolized to the fluorescent resorufin in

living cells by cytoplasmic and mitochondrial oxidoreductases. CFDA-AM is

metabolized by non-specific esterases to carboxyfluorescein and gives a measure of cell

plasma membrane integrity since the latter maintains the necessary cytosolic conditions

for esterase activity. Also specific impairment of esterase activity or uptake of the dye

could influence the results [5, 50, 51]. These assays allow insight into the general health

status of the cells and possibly as well into the mode of action leading to cytotoxicity [5].

Figure 12 Co-culture system used in this thesis. A monolayer of RTgill-W1 is grown on the insert and

directly exposed to the toxicants. The RTL-W1 grown underneath can be indirectly exposed if the

toxicants make it through the first monolayer. Diagram by Fabienne Roux.


17

RTgill-W1 were seeded in 48 or 96-well plates at a density of 150 000 cells in 500 uL (L-

15 5% FBS) or 50 000 cells in 200 µL. The outer wells were filled with medium only to

avoid plate effects, as it has been noted that cells in outer wells grow much slower. The

plates were sealed with Parafilm® to avoid evaporation. Cells were allowed to grow for

48h to attain confluence. The 48h exposure was performed in L-15/ex. Single compounds

were first tested in 2 independent experiments to obtain toxicity curves and determine

EC50s. The mixtures were based on the EC50 ratio and on the environmental ratio.

DMSO was added to all compounds at a concentration of 0.1% (v/v). Cell toxicity was

determined by measuring fluorescence (excitation/emission wavelength for AB 530/590

and CFDA-AM 485/530 nm) using VICTORTM 1420 multilabel counter.

Data analysis was performed on the AB results as it showed to be the most sensitive

endpoint. Dose-response curves were modeled using the best fit approach (Generalized

Logit 1, Generalized Logit 2, Weibull, Gompertz, Morgan-Mercier Flodin, Box-Cox

Weibull, Brain-Cousens). The Concentration Addition (1) and Independent action (2)

models were used to analyze mixture toxicity.

CA: (1)

IA: (2)

Both models assume that the toxicants do not interfere with each other and that their

effects can be measured on a common endpoint. The additional assumptions of CA are

that the compounds have the same mode of action, and thus similar shapes and slopes of

concentration response curves, simply shifted depending on the sensitivity of the test

system to the compound. IA assumes on the contrary that the toxicants target totally

independent pathways through different modes of action [52].

2.4.3 Mechanistic analyses – single cell and co-culture systems

2.4.3.1 Gene expression

Seeding and exposure of RTL-W1

RTL-W1 were seeded in 6-well plates Cellstar® at a density of 500 000 in 3 mL L-15 5%

FBS 1% PS. They were exposed after 24h of growth to BBP and BPA. Both compounds

were first diluted in DMSO and final DMSO content was 0.1% (v/v). Concentrations

were based on preliminary cytotoxicity studies and literature search for BBP, and on

literature search for BPA. Exposure lasted for 24h, after which cells were trypsinized and

collected in eppendorf tubes.

𝐸(𝑐𝑚𝑖𝑥) = 𝐸(𝑐1 + ⋯ + 𝑐𝑛) = 1 − ∏[1 − 𝐸(𝑐𝑖)]

𝑛

𝑖=1

𝐸𝐶𝑥(𝑚𝑖𝑥) = (∑𝑝𝑖

𝐸𝐶𝑥(𝑖)

𝑛

𝑖=1

)

−1


18

RNA isolation

RNA was directly isolated using the QIAGEN RNeasy® Plus Mini kit, lysate

homogenization being performed using QIAshredderTM spin columns. The RNA

concentration and 260/280 ratios were determined using a Thermo Scientific Nanodrop

2000c spectrophotometer. The purified RNA was stored at -80C.

cDNA synthesis

cDNA was synthesized from 500ng RNA using the iScriptTM cDNA synthesis kit on the

MyCyclerTM instrument from BioRad and stored at -20C.

qPCR

New primers were designed using PRIMER-BLAST6 from NCBI and Primer3Plus7,

following the recommendations issued by [53]. The secondary structure was checked

using Beacon DesignerTM Free Edition8. As a rule of thumb, any primer with any ΔG < -

3,5 kcal/mol or a ΔG < -1 kcal/mol in the case of 3’ hairpins were discarded. Rating

according to NetPrimer9 was also taken into account for the final choice.

Amplification reactions were performed in duplicates using the SsoAdvancedTM

Universal SYBR® Green Supermix. Each reaction of qPCR contained the amount of

cDNA equivalent to 10 ng of total RNA, with 300 nM to 500 nM of each primer (

Table ), in a total volume of 10 μL in 96-well plates. Reactions were performed in the

CFX ConnectTM Real-Time system from BioRad, with 10 min denaturation at 95°C,

followed by 40 cycles of 95°C for 15 s and 60°C for 1 min (55°C in the case of CYP1A

and CYP3A, 58°C for Dio2 and PEPCK).

The obtained Cq values were normalized with the geometric mean of the 2 reference

genes, β-actin and ubiquitin. Fold change was determined using the ΔΔCq equation:

𝐹𝑜𝑙𝑑 𝑐ℎ𝑎𝑛𝑔𝑒 = 2−∆∆𝐶𝑞

Where ΔΔCq = ΔCqsample – ΔCqcontrol

ΔCqsample = difference between sample Cq and the reference Cq for each replicate,

ΔCqcontrol = mean of the difference between control Cq and the reference Cq.

Seeding and exposure of inserts

RTgill-W1 were seeded in inserts placed in 6-well plates Cellstar® at a density of

280 000 cells in 3 mL L-15 containing 5% FBS and 1% PS. The lower compartment was

filled with 3 mL of the same medium. Medium was changed once a week in both

compartments and cells were allowed to grow for at least 3 weeks. About one week

before exposure, the medium in the upper compartment was changed to L15-ex 1% PS.

Transepithelial resistance was measured after a few days. RTL-W1 were seeded in a

6 http://www.ncbi.nlm.nih.gov/tools/primer-blast/ 7 http://primer3plus.com/cgi-bin/dev/primer3plus.cgi 8 http://www.premierbiosoft.com/qpcr/ 9 http://www.premierbiosoft.com/netprimer/

http://www.ncbi.nlm.nih.gov/tools/primer-blast/

http://primer3plus.com/cgi-bin/dev/primer3plus.cgi

http://www.premierbiosoft.com/qpcr/

http://www.premierbiosoft.com/netprimer/


19

separate 6-well plate Cellstar® at least 48 h prior to exposure at a density of 500 000 cell

in 3 mL, in L-15 5% FBS 1% PS. The inserts were moved to the plates containing RTL

and RTgill-W1 were exposed to BPA in L-15/ex 1% PS for 24 h. Gene expression

analysis was performed as explained above.

2.4.3.2 Total glutathione assay

Total glutathione (tGSH) content can give an insight into metabolic disruption and

oxidative stress as glutathione is involved in both processes.

To measure tGSH, an indirect assay is used. GSH reacts with DTNB to form GSH.TNB

and free TNB. The latter can be measured spectrophotometrically by measuring

absorbance at 415 nm. Addition of glutathione reductase (GR) allows the reduction of

GSSG to GSH, so that all glutathione present in the cell is in its reduced form and can

react with the DTNB (

Figure 13).

Figure 13 Total glutathione content measurement. Oxidized glutathione (GSSG) is converted to GSH by

glutathione reductase in presence of NADPH. The total reduced glutathione (GSH) can react with DTNB to

form GS-TNB, releasing free TNB that can be detected spectrophotometrically by measuring absorbance at

415 nm.

RTL-W1 were seeded in 6-well plates at a density of 500 000 cells in 3 mL L-15 5% FBS

1% PS. Once confluent, they were exposed to BPA for 48h. Both compounds were

prepared as stocks in DMSO with a total final concentration of 0.1% DMSO in the test

medium. After 48h, the test solution was discarded, the cells were rinsed with PBS first

and then with EDTA for 3 minutes, sampled by trypsination and pelleted by

centrifugation at 300g for 5 minutes. The supernatant was discarded, and the pellets were

resuspended in 100 µL of 5% 5-sulfosalicylic acid (SSA). They were homogenized by

sonication (3s), allowed to rest on ice for 15 min and then centrifuged for 20 min at

10 000g. The supernatant was stored at -80°C.

For tGSH measurements, the samples were thawed and diluted in 5% SSA. 20 µL was

added in duplicates to a 96-well plate, along with the GSH standard curve. 200 µL of

DTNB, NADPH and buffer were added to the plate and incubated for 5 minutes. Last, 20

µL of GR was added to the plate and readings were performed at 415 nm for 7 minutes in

the SpectraMax 190 spectrophotometer from Molecular Devices.

2.4.3.3 EROD

7-ethoxyresorufin-O-deethylase (EROD) activity gives a measurement of CYP1A-

mediated phase I metabolism. This activity is measured by the deethylation of 7-

ethoxyresorufin in the presence of endogenous NADPH to resorufin, a strongly

fluorescent molecule emitting at 590 nm [54].


20

RTgill-W1 and RTL-W1 were compared for their potential EROD activity induction

through BNF exposure. The cells were seeded in 48-well plates at a density of 150 000

and 50 000 cells in 0.5 mL respectively and allowed to grow to confluence. They were

then exposed to 1 and 0.1 µM BNF (final DMSO content 0.1% v/v) in L-15 5% FBS 1%

PS for 24 hours. Cells were washed with 500 µL PBS and then exposed to 200 µL of 7-

ethoxyresorufin solution. The plate was then read every minute for 10 minutes to measure

the formation of resorufin (excitation/emission wavelength 530/590 nm) using

VICTORTM 1420 multilabel counter. After the measurements, 100 µL fluorescamine in

acetonitrile (0,3 mg/mL) was added to each well to determine the protein content, and the

plate was measured again after 10 minutes of incubation (excitation/emission wavelength

400/460 nm).

This assay was performed as well on RTL-W1 when exposed to BPA for 24h.

2.4.4 Statistical analysis

All statistical analysis was performed using IBM SPSS Statistics 22 for Windows. Data

which showed normality (Shapiro-Wilk (SW) > 0.05) and homogeneity of variance

(Levene (L) > 0.05) was analyzed by a one-way ANOVA followed by Tukey post hoc

(p<0.05). If the data was not normal (SW<0.05), analysis was performed using a non-

parametric Levene test (>0.05) followed by a Kruskal-Wallis test (p<0.05). If the data

was normal (SW>0.05) but not with homogeneous variance (L<0.05), and if taking the

log of the data was not sufficient to improve homogeneity of variance, a Welch ANOVA

followed by a Dunnet T3 post hoc (p<0.05) was used instead. Significant results are

marked with an asterisk (*). Data is shown as mean ± standard error (SE).

Table 1 Primer sequences. The basic qPCR set up used a primer concentration of 300 nM and an annealing

temperature of 60°C, with the following exceptions: a primer concentration 500 nM; b annealing

temperature 55°C; c annealing temperature 58°C.

Forward primer 5’→3’ Reverse primer 5’→3’ Ubi ACAACATCCAGAAAGAGTCCAC GCAGCCTGAGGCACACTTG β-actin TGGCATCACACCTTCTAC AATCTGGGTCATCTTCTCC CYP1A a,b TCCTGCCGTTCACCATCCCACACTGCAC AGGATGGCCAAGAAGAGGTAGACCTC CYP3A a,b GCCAGCCAGCAGAAGAGT GGATTCGTAGCCAGATTGTAAGC GSTπ ACCTGGTGCTCTGCTCCAGTT AGAGCTCAGGAAGCCCTTGAT NRF-2 a TTTGTCCCTTCCTGAGCTGC GGGCAATGGGTAGAAGCTGT GPx CCTGGGAAATGGCATCAAGT GGGATCATCCATTGGTCCATAT GCLcat TGAGGGAGTTTGTGGACAAGC AATAGTTCTGGCATCGCTCCTC ERRγ CAGCAGATGAACCTGAGCCA TCATGGAGTCCGTCCTGGAA GR ATGGGGTCAGTCAGCTTTGG AGGGAGGAAAGGAAAGCAGC PEPCKc AGACCAACCCTCATGCCATG TGGAGTTGGGATGAGCACAC PPARγ ACAGACACTTTCCCCTGACCAA CGTCAGAGACTTCATGTCATGGA ERα ACTCTGGTGCCTTCTCCTTC GCGTCGGTGATGTTGTCC Dio2 c TTGAGGCACCCAACTCCAAA AGCCGAAGTTCACCACAAGA TTR ACTGCCCGTTGATGGTTAAG TTCCCCTGTCAAGTCTGTCA TRβ TACAAAAATTATCACCCCCGCCA CTCACAGAACATAGGCAGCTTTT


21

3 Results

3.1 Mixture toxicity assessment (NICE) – cytotoxicity

3.1.1 Single compound toxicity

The environmental concentrations of the different toxicants were the following:

Table 2 Environmental concentrations in ng/L of Cu2+, MET and TBP detected on the Swedish west coast

during the chemical monitoring by NICE, june 2012 and September 2013.

June 2012 September 2013

Instöränna Fiskebäckskil Skalkorgarna Lerkil Stenungsund Björlanda Fiskebäck

Cu2+ 1100 2300 1100 840 4900 8600 5200

MET 0,23 0,2 3,5 <0.1 0,44 1,4 0,89

TBP 27 <23 47 29 45 34 37

Two independent experiments (E01 and E02) were first performed to determine the

EC50s of each chemical. Depending on the quality of the results obtained, either one or

both experiments were used to calculate the EC50s and the EC50 ratio.

Table 3 Dose-response curve parameters obtained for Cu2+, MET and TBP, EC50s and molar ratio used for

the mixture. The experiment was performed in 48-well plates and for 48h. Only AB data was used for

EC50 calculation as it was the most sensitive endpoint. The numbers of controls was 11 for Cu2+, 8 for

MET, 7 for TBP, with 4 replicates of each concentration. Results from the first (E01) or second (E02)

experiment were chosen according to their quality, except for TBP where both experiments were pooled.

Cu2+ (E01) MET (E02) TPB (Pooled)

tet1 -4,02E+00 -4,10E+00 1,40E-03

tet2 7,53E+00 1,58E-02 1,48E+01

tet3 7,72E-01 5,73E-01 3,72E+02

Model Gl1 BCW Brain-Cousens

EC50 3 µM 5360 µM 420 µM

p (EC50 ratio) 0,05% 92,69% 7,26%


22

3.1.2 Mixture toxicity

For mixture toxicity testing, 3 different independent experiments (A, B and C) were

performed. Each compound was tested singularly (to account for potential fluctuations

between cell passages) and as a mixture, based on the EC50 and environmental ratios.

The results obtained for the single compounds are the following.

Table 4 EC50s [µM] obtained for the single compounds in 3 independent experiments, A, B and C,

using the cytotoxicity assay on RTgill-W1. Exposure lasted 48h.

Figure 14 Modeled single compound dose-response curves of Cu2+, TBP, MET in the 3 mixture

experiments (AlamarBlue) on RTgill-W1. Black line = Cu2+; dashed dark grey = TBP; dashed and dotted

light grey = MET. Each experiment was performed using 4 96-well plates, with 10 controls; dilution series

of 10 concentrations with 1 replicate on each plate. Exposure lasted 48h.

The first mixture to be assessed was based on the EC50 ratios and gave the following

results:

Table 5 EC50s [µM] obtained with the mixture based on the EC50 ratio of Cu2+, TBP and MET in the 3

experiments A, B and C, using the cytotoxicity assay on RTgill-W1. Exposure lasted 48h.

-1,2

-1

-0,8

-0,6

-0,4

-0,2

0

0,2

0,4

0,6

0,8

1

1 10 100 1000 10000

Me

tab

olic

inh

ibit

ion

single compound concentration [µM]

EC50 [µM] Cu2+ MET TBP

A 35 7480 218 B 19 7140 300 C 11 6590 324

EC50 [µM] Mixture (EC50 ratio) CA IA

A 3570 2120 2980 B 2635 2510 4530 C 3608 2360 4402


23

The environmental mixture, based on the concentrations of each chemical detected in the

sampling site of Stenungsund, was also assessed in experiment C. The ratio of each compound

can be found in Table 6:

Table 6 Molar ratio of Cu2+, MET and TBP for mixture testing according to the environmental concentrations

detected in Stenungsund during the chemical monitoring in June 2012.

Cu2+ MET TPB

p (env ratio) 99,08% 9,90 10-3 % 0,91%

-0,6

-0,4

-0,2

0

0,2

0,4

0,6

0,8

1

1,2

1000 10000

Me

tab

olic

inh

ibit

ion

Mixture concentration [µM]

A

-0,2

0

0,2

0,4

0,6

0,8

1

1,2

1000 10000

Me

tab

olic

inh

ibit

ion


B

-0,6

-0,4

-0,2

0

0,2

0,4

0,6

0,8

1

1,2

1000 10000

Me

tab

olic

inh

ibit

ion


C Figure 15 Modeled and predicted mixture

toxicity curves of the mixture based on the

EC50 ratio. Grey line = model from

experimental data; black line = IA prediction;

black dotted line = CA prediction. Each

experiment was performed using 4 96-well

plates, with 10 controls; dilution series of 10

concentrations with 1 replicate on each plate.

Exposure lasted 48h.


24

The obtained mixture toxicity was

Table 7 EC50s [µM] obtained in experiment C for the environmental mixture of Cu2+, TBP and MET, the CA and

IA predictions, using the cytotoxicity assay on RTgill-W1. Exposure lasted 48h.

Figure 16 Modeled and predicted mixture toxicity curves of the mixture based on the environmental concentrations.

Grey line = model from experimental data; black line = IA prediction; black dotted line = CA prediction. This

experiment was performed using 4 96-well plates, with 10 controls; dilution series of 10 concentrations with 1

replicate on each plate. Exposure lasted 48h.

EC50 [µM] Mixture (env ratio) CA IA

C 14,5 9,6 9,6

-0,4

-0,2

0

0,2

0,4

0,6

0,8

1

1,2

1 10 100

Me

tab

olic

inh

ibit

ion


C


25

3.2 Mechanistic analyses – single cell and co-culture systems

3.2.1 EROD activity and cytotoxicity in RTgill-W1 and RTL-W1

EROD activity was analyzed in RTgill-W1 and RTL-W1. Inducible EROD activity was

observed in both cell lines after 24h, with the highest response at 0.1 µM BNF. At 1 µM

BNF, BNF competes with 5-ethoxyresorufin (ER) for CYP1A active sites, which limits

the metabolism of ER and thus shows a low EROD activity.

Figure 17 EROD activity in RTgill-W1 and RTL-W1 after 24h exposure to 0.1 and 1 µM BNF. The

experiment was performed in L-15 5% FBS 1% PS. Metabolism of 7-ethoxyresorufin was normalized with

the amount of protein detected by fluorescamine.

EC50s (in µM) based on AlamarBlue were the following.

Table 8 EC50s (in µM) of BBP and BPA on RTL-W1 and RTgill-W1 with different media, 48h exposure.

The experiments were performed in 48 well plates. *Due to lack of toxicity of BBP, precise determination

of EC50s has been impossible to perform. The percentages between parentheses are the effect observed at

the concentration indicated. (1) indicates the EC50 on which the BBP concentrations were based for the

gene expression experiment on RTL-W1.

BBP* BPA

(1) RTL-W1 L-15/ex >6.4 (38%) -

RTL-W1 L-15 5%FBS - 194,5

RTgill-W1 L-15/ex ~50 (55%) 59

RTgill-W1 L-15 5%FBS >50 (15%) 69

-2

0

2

4

6

8

10

12

14

RTgill-W1 RTL-W1

spe

cifi

c ER

OD

act

ivit

y [p

mo

l/m

in/m

g p

rot]

CT 0.1 µM 1 µM


26

3.2.2 Tool identification on RTL-W1

Primer design

The development of a primer for EERγ was unsuccessful, and PPARγ and TTR do not

seem to be expressed in RTL-W1. It must be noted that optimization of primers for

PEPCK and CYP3A was not ideal as the efficiency was over 110%. All other primers

were optimized with efficiencies between 90 and 110%.

Gene expression

The concentrations for the exposure to BBP of RTL-W1 were chosen based on literature

and on a preliminary cytotoxicity test (noted (1), (see Table 8). A concentration of 1 µM

was chosen as highest test concentration since the EC50 seemed to be close to 10 µM and

as 1/10 of EC50 as highest concentration is reasonable for sublethal experiments.

Moreover, no cytotoxicity was observed at 0.64 µM. BPA was added in the experiment

before any cytotoxicity data could be measured in our facilities, and the highest

concentration was based on literature, where a 24h EC50 of 125 µM had been reported

by [55] for MeWo cells. 10 µM was chosen as it is about 10% of the EC50. A factor of

50 was used between the high and low concentration for both compounds.

BBP exposure showed no significant effect on any of the genes tested (see Figure 18).

BBP BPA

A significant effect was observed at 10 µM BPA on the Dio2 gene expression (see

Figure 18). No other significant difference or general trend was detected for ERα, GSTπ,

GPx or Nrf-2. PPARγ could not be detected in the RTL-W1 cell line even though it the

primers functioned on the liver samples.

The response of Dio2 led to a second round of exposure with BPA (Figure 19), with

more focus on the thyroid system. Unfortunately, quantification of TTR and TRα

expression was unsuccessful, due to lack of expression of non-functional primers. The

concentration span was increased to see if any low concentration effects, specific to

Figure 18 Fold change of different potential new biomarker genes observed in RTL-W1 exposed to BBP and BPA

for 24h. Data is shown as mean ± SE (n =3). Asterisks (*) indicate significant changes (one-way ANOVA and

Tukey post hoc, p<0.05).

*


27

endocrine disruptors, could be found. The highest concentration was increased to 20 µM

as RTgill-W1 showed no metabolic inhibition with up to 16 µM of BPA in both media.

å

Figure 19 Fold change of different potential new biomarker genes observed in RTL-W1 exposed to

different concentrations of bisphenol A for 24h. Data is shown as mean ± SE (n =4). Asterisks (*) indicate

significant changes (one-way ANOVA and Tukey post hoc, p<0.05).

Thyroid system

In the second round of exposure, 20 µM BPA showed a statistically significant decrease

for Dio2 levels (after having logged the data, one-way ANOVA p = 0.000; Tukey p =

0.000). This decreasing trend clearly starts at 2 µM, even though no statistical

significance supports it, and confirms the results obtained during the first round. No

response was observed when it comes to TRβ.

The expression of TTR was too low in the RTL-W1 to be quantified (data not shown).

The primer was however functional on the liver samples.

Glucocorticoid receptor

Exposure to BPA had no significant effect on the glucocorticoid receptor. A very slight

increase at 0.2 µM could be observed and might have been significant if the controls had

shown less variability.

Metabolism: phase I

BPA exposure of 20 µM clearly increased the amounts of CYP1A mRNA. Welch

ANOVA indicated significance (p = 0.022), however this was not confirmed by Dunnett

T3 post-hoc. CYP3A gene expression was on the opposite significantly decreased (one-

way ANOVA p = 0.016; Tukey p = 0.045) at 20 µM BPA.

0

1

2

3

4

5

Dio2 TRβ GR CYP1A CYP3A

Fold

ch

ange

CT 0,02 µM 0,2 µM 2 µM 20 µM

*

* *


28

EROD activity in RTL-W1

Figure 20 EROD activity in RTL-W1 after 24h exposure to different concentrations of BPA. 2 plates were

used for the experiment (1 and 2) and data was compared to the control on the corresponding plate (the 5

lowest BPA concentrations on plate 1, BNF and 20 µM BPA on plate 2). The number of controls was 4 per

plate, with 3 replicates for each BPA concentration. Data is shown as mean ± SE (n =4). Welch ANOVA

indicated significant difference between the groups (p=0.001), however Dunnet T3 validated only the

exposure to BNF as statistically different from the controls.

The increased expression of CYP1A translated at the activity level of this enzyme as a 4

to 5-fold increase in EROD activity could be observed with an exposure of 6.3 and

20 µM BPA compared to the controls (Figure 20). Welch ANOVA indicated significant

difference between the groups (p=0.001), however Dunnet T3 validated only the

exposure to BNF as statistically different from the controls.

Total glutathione measurements in RTL-W1

Figure 21 Total glutathione content in RTL-W1 after 48h exposure to BPA. The number of controls was 4

per plate, with 3 replicates for each BPA concentration. Data is shown as mean ± SE (n =4). Asterisks (*)

indicate significant results (Kruskal-Wallis, p<0.05).

0

1

2

3

4

5

6

7

8

9

10

ERO

D a

ctiv

ity

vs c

on

tro

l[p

mo

l/m

in/m

g p

rote

in]

0

20

40

60

80

100

120

140

160

CT BPA 2 uM BPA 20 uM

tGSH

co

nte

nt

[µm

ol]

*


29

Exposure to 2 µM of BPA on RTL cells showed a statistically significant increase in total

glutathione content (Kruskal-Wallis p = 0.043) (Figure 21).

No effect could be noted on the GSTπ and GPx expression levels through qPCR (see

Figure 18).


TER measurements

Figure 22 TER measurements of empty inserts in L-15, in L-15/ex and of inserts with a confluent

monolayer of RTgill-W1, a few days after having changed the apical medium to L-15/ex.

TER values (Figure 22) were measured before exposure. No great difference was

observed. A slight increase can be noted with the monolayer of RTgill-W1.

Gene expression

RTgill-W1 RTL-W1

Figure 23 Fold change of CYP1A, CYP3A and Dio2 gene expression in RTgill-W1 and RTL-W1 when

exposed for 24h to BPA. Data is shown as mean ± SE (n =4). Asterisks (*) indicate significant results (one-

way ANOVA and Tukey post hoc, p<0.05).

0

100

200

300

400

500

600T

ER

[Ω

·cm

2]

CT L-15 CT L-15/ex Rtgill-W1

* * *


30

In both cell lines, significant CYP1A expression induction can be observed (Figure 23).

In the case of RTgill-W1 this already happens at 2 µM, whereas it takes 20 µM for RTL-

W1. This is the same response as was observed for single cell line exposure of RTL-W1,

but to a lesser extent (2-fold instead of 4-fold). The response of CYP3A is less clear and

no trend can be shown. Dio2 expression is clearly if not significantly inhibited at 20 µM

BPA, which is also similar to what had been observed in RTL-W1 before, and to a lesser

extent (fold change of 0.60 instead of 0.47).


31

4 Discussion

This project aimed to assess the potential uses of two different fish cell systems in

ecotoxicology. The first system was a gill cell line (RTgill-W1) in an optimized

cytotoxicity assay that had been designed by Tanneberger et al [8]. This system was

tested for mixture toxicity screening in the context of the NICE project. The second

system was a co-culture model composed of both gill (RTgill-W1) and liver (RTL-W1)

cell lines, aiming to mimick environmental exposure through the gills and subsequent

exposure of inner organs. Its properties for mechanistic studies were assessed using

different biochemical assays.

4.1 Mixture toxicity assessment (NICE) – cytotoxicity

To assess the mixture toxicity screening potential of the RTgill-W1 assay, the combined

toxicity of 3 compounds found in Swedish coastal waters, copper ions, metoprolol and

tributylphosphate, was tested. Single compound toxicity was first determined to obtain

the respective EC50s, which were compared with existing literature. Two mixtures were

tested: one was based on the EC50 ratio of the compounds, and the second on their

environmental ratio according to monitoring in Stenungsund. The mixture toxicity was

modeled using the concentration addition and independent action models.

4.1.1 Single compound toxicity

Copper, an essential trace element, is widely used by the metal industry, but also as

fertilizer, wood preservative, and pesticide [56]. Highly toxic to the aquatic environment,

it is besides cadmium one of the most polluting metals in the environment [57]. During

the chemical monitoring, the highest Cu2+ concentration, 8.6 µg/L (about 0.14 µM) (see

Table ), was detected in Björlanda. Copper sulfate toxicity has been assessed by [57] on

RTgill-W1 and an EC50 of 29.2 µM was obtained after 2.5h. They also suggest that

oxidative stress could trigger cytotoxicity. In vivo EC50s generally range between 0.1 and

10 µM (ECOTOX database). The EC50 values observed in this study varied between 3

and 35 µM (Table , Fel! Hittar inte referenskälla.). The second compound, metoprolol,

is a selective β1-adrenergic receptor blocking pharmaceutical widely used to treat

cardiovascular diseases such as hypertension, heart angina and abnormal heart rates. β-

blockers are found in WWTP effluents worldwide [58], at concentrations sometimes as

high as 5 µg/L [59]. The highest detected concentration on the Swedish west coast was in

Skalkorgarna (3.5 ng/L) (Table ). Metoprolol blocks β-adrenergic receptors and seems

non-toxic to fish (exposure of up to 100 mg/L during 2 days of Japanese Medaka larvae

did not result in increased mortality, [60]). Previous cytotoxicity studies performed on

RTG-2, a rainbow trout gonadal cell line reported EC50s higher than 1800 µM [10, 58].

The EC50 values that were obtained in this study ranged from 3800 to 7500 µM (Table ,

Fel! Hittar inte referenskälla.). The last compound, tributyl phosphate, is an

organophosphorus compound. Its main use is as a solvent in rare earth extraction and

purification and as component of flame retardant for aircraft hydraulic fluids [61]. It can

enter the aquatic environment through waste water discharges, in particular from

extraction plants [62]. Due to its relatively high logKOW (=4 according to [61], it is

expected to adsorb mainly to sediments and particulate matters. Surface waters in Europe


32

have been found to contain between 100 and 3900 ng/L of TBP [61]. During this study,

the highest TBP concentrations were detected in Skalkorgarna and Stenungsund (47 and

45 ng/L respectively) (Table ). Its bioconcentration factor in fish is relatively low and its

toxicity to rainbow trout (LC50, up to 48 days) has been shown to be in the range of 1 to

15 mg/L, thus 4 to 60 µM [62, 63]. Its mode of action for toxicity is unknown [61]. TBP

has been tested on HeLa cells and an EC50 of 5600 µM was found [64]. The EC50s

obtained here ranged from 200 to 420 µM (Table , Fel! Hittar inte referenskälla.),

which is less sensitive than in vivo data showed but more sensitive than the former in

vitro study.

Since no in vivo mortality data is available for MET, no in vitro – in vivo comparison can

be made. For the other two chemicals, the EC50s obtained during this experiment are in

good general accordance with in vivo studies, with a maximum of one order of magnitude

difference. As the repeatability of each experiment was limited, it is difficult to be more

precise, but when compared to general in vitro – in vivo differences which have been

reported to reach up to 3 orders of magnitude, the use of this sensitive RTgill-W1 assay

seems promising, as was noted by Tanneberger et al [8].

4.1.2 Mixture toxicity

Copper ions, MET and TBP are dissimilarly acting when it comes to their modes of

action, which would suggest the use of the IA model. However, RTgill-W1 is not a test

system that can account for the specific modes of action of the chemicals considered here

and the observed cytotoxicity should mostly be due to disruption of basal cellular

functions. It has been shown that both models give fairly similar results with mixtures

that are not composed of purely similarly or dissimilarly acting compounds, CA being an

efficient worst-case scenario with, in a majority of cases, not more than a 2-fold

underestimation of the EC50 [65, 66]. Both models seem thus reasonable to consider.

One important limitation to keep in mind is that none is able to account for hormesis in

the present situation, IA due to its intrinsic properties and CA since not all compounds

show hormesis in the same span [67].

The experimental mixture EC50 based on the EC50 ratio (Fel! Hittar inte

referenskälla.) is consistently underestimated by a factor of 1.1 to 1.7 by the CA. The

IA tends to overestimate it, by a maximum factor of 1.7, but does also underestimate it

one time by a factor of 1.2 (Figure 15). No antagonism or synergism can be noted.

Similarly, in the case of the mixture based on the environmental ration, both models

overestimate the toxicity by a factor of 1.5.

Interestingly, the EC50s of each compound varied significantly between each experiment

(up to a factor of 10 for copper ions, 1.5 for MET, 2 for TBP, 1.5 for the mixture, Figure

14, Figure 15) and thus between the mixture experiments and the initial single compound

assessment on which the EC50 ratio was based. Therefore, the toxicity of each compound

in each experiment was taken into account for the prediction. Except in the case of

copper, the variability is in expected and acceptable ranges. In the case of the mixtures,

considering the wide range of toxicity of these compounds, these variations seem

relatively small and both models give reasonably good estimates of the mixture EC50.


33

The observed variability can be due to several factors. Each experiment was performed

on a different passage of RTgill-W1, and it is known that differences in response can then

be observed. The use of 96-well plates instead of 48-well plates also leads to an increase

of variability. In this experimental setup, only 4 replicates per concentration were used,

and each one was on a different plate. This set up was necessary in order to assess the

toxicity of all 3 single compounds and of the mixture at the same time, to analyze

sensitivity differences between the passages. Moreover, only 64 wells out of 96 were

used per plate, as it has been observed that the cells in the outer wells grew significantly

slower. Finally, no chemical analysis was performed during this short study and nominal

concentrations were used, thus actual exposure was not taken into account, which could

also be a factor of variation.

To improve this assay, it would be necessary to evaluate how much variability comes

from the actual splitting (passaging) of the cells and how their behavior evolves with

regard to the EC50s, especially in cases of high variation such as was observed with

copper. Chemical analysis would be necessary to determine the actual exposure and

verify that the stocks remain stable. As TBP is lipophilic, the true exposure might be

much lower than is expected according to the nominal concentration. Optimization of the

dosing procedure could be helpful.

The use of this assay seems to be applicable for the analysis of mixture toxicity in fish.

However, this assay is considered as a good alternative to acute in vivo testing, but might

overlook important chronic effects happening via the specific modes of action of the

chemicals. This is relevant in case of single compound exposures and becomes even more

critical in case of multiple stressor situations. Metoprolol is here a perfect example as it

seems non-toxic to the system, however chronic exposure and potential effects on the

cardiac system could be very significant in case of multiple stressor situations. It is thus

of true importance to keep in mind that specific sublethal modes of action are a challenge

to assess with a cell line.

4.2 Mechanistic analyses – single cell and co-culture systems

Two endocrine disrupting chemicals, bisphenol A (BPA) and benzylbutylphthalate

(BBP), were chosen as model compounds to test on the co-culture system. An initial

identification of different biochemical tools (gene expression, EROD activity, glutathione

content) was first performed on RTL-W1 only, before the actual co-culture system was

used.

4.2.1 Tool identification on RTL-W1

The gene expression analysis was limited by the lack of or low expression of certain

genes, even though they are expressed in the normal liver. ERRγ, PPARγ and TTR could

thus not be assessed. This is unfortunate but can be expected when cell lines are used.

BBP did not show any significant effect on the genes that were analyzed. For some of

them, it is consistent with what has been found in the literature [68] for developing

zebrafish embryos. Higher concentrations of BBP could probably have been used due to


34

its probably very limited bioavailability in the test system, especially with the presence of

serum proteins. Additional genes having shown response to some degree could also be

considered.

Thyroid system

The only effect that could be observed when it comes to the thyroid system is a

significant decrease in Dio2 gene expression when exposed to BPA (see Figure 18,

Figure 19). Deiodinases are considered promising biomarkers for thyroid disruption in

fish due to their central role in thyroid hormone activation and metabolism, see Figure 6

[23]. Dio2 has already been shown to be responsive to endocrine disruptors such as

MEHP, a phthalate; however in that particular case its gene expression was significantly

increased [28]. This could be due to different properties of the compounds. The phthalate

which was tested in this study, BBP, did not have any effect on Dio2 (Figure 18). It is

interesting to see that Dio2 is expressed in RTL-W1 at low level (Cq values of 25-26).

BPA is expected to interact with TRβ although the exact nature of this interaction is

debated [27]. since the TR seems to be expressed at the transcriptional level even though

at low levels (Cq values of 30), different assays could have been used to follow receptor

interaction, and the transcription level of different genes downstream of the TR could

provide valuable information.

Interestingly, in the parallel study that was conducted on brown trout by Joan Martorell

Ribera [69], T4 plasma levels were decreased after 8 week exposure to BPA, BPS and

BBP, even though T3 plasma levels remained stable. The combination of these

observations could suggest a role of Dio2 in BPA-mediated thyroid disruption and

requires more investigations. The thyroid system is complex and a better overview of the

different actors is necessary. Even though RTL-W1 seems to express Dio2 and TRβ at

relatively low levels, this cell line is probably not the most suitable in order to study this.

The high concentrations of BPA needed to obtain a response at the transcriptional level

and the lack of response to BBP indicate that this cell line is not a very sensitive assay if

Dio2 is to be used as a biomarker for thyroid disruption. Many other proteins are reported

to be sensitive to endocrine disruption, which are either not expressed in RTL-W1 even

though they are expressed in the liver (as it seems to be the case for TTR) or simply

expressed in different tissues, such as the pituitary.

Metabolism

RTgill-W1 has been described as a cell line lacking inducible CYP1A by Schirmer et al

[44, 45], an advantage for cytotoxicity testing as phase I metabolism would not interfere

with cytotoxicity. As was shown when both cell lines were compared (Figure 17), EROD

activity was indeed less inducible in RTgill-W1, however a response could be observed.

CYP1A gene expression was also inducible in this cell line.

Phase I metabolism: CYP

Interestingly, the results obtained on RTL-W1 (Figure 19, Figure 20) are opposite to

most in vivo results. Indeed, a general down-regulation of 5 out of 6 different CYP

enzymes was observed with the exposure of juvenile atlantic cod to BPA (50 µg/L) for 3

weeks. In particular, down-regulation of CYP1A gene expression, decrease in CYP1A


35

enzyme levels and inhibition of EROD activity was noted. Similarly, CYP3A mRNA

levels decreased, as well as CYP3A activity even though enzyme levels remained stable

[70]. Brown trout exposed to BPA, BPS and BBP (2 and 20 mg/kg) for 2 to 8 weeks

show a general decrease in EROD activity for BPA [69]. Arukwe et al [71] also observed

a very clear decrease in EROD activity in the liver of juvenile atlantic salmon when

exposed to increasing concentrations of BPA (from 1 to 125 mg/kg) or E2 (5 mg/kg).

This inhibition of the hepatic CYP1A seems to be the most common in vivo response in

the case of exposure of fish to estrogens, however the opposite response has also been

observed. It is not exactly known why BPA has this effect on phase I metabolism but

involvement of the ER has been suggested [7]. Navas et al [72] studied the effects of E2

(endogenous ER agonist) on rainbow trout hepatocytes. They also noted a decreased

EROD activity and CYP1A mRNA levels due to E2 exposure but interestingly, the co-

exposure of E2 and an ER antagonist, tamoxifen, cancelled the inhibitory effect of E2 on

the CYP1A system. They therefore suggest that E2 inhibitory effect on CYP1A

expression and EROD activity is somehow mediated by the estrogen receptor. Grans et al

[73] obtained a similar inhibition of CYP1A in EE2 exposed rainbow trout hepatocytes,

but the use of fulvestrant, a specific ER antagonist, did not cancel that effect. It should be

noted however that ICI did not cancel EE2-mediated Vtg induction in this study, which is

surprising for an ER antagonist.

One hypothesis to partly explain the induction of CYP1A mRNA expression and EROD

activity observed here could be that as ER is probably non-functional in RTL-W1 [29],

this liver cell line cannot respond properly to estrogen exposure.

Unfortunately, studies involving the ER and CYP1A have not been performed with BPA

and it is therefore difficult to draw conclusions. However, the lacking ER pathway could

be related to this unexpected response. This highlights the importance of choosing

relevant test systems if one wants to analyze mechanistic effects of toxicants.

Phase II metabolism

Chemical analysis The two major metabolites that have been found for BPA in fish are the BPA sulfonate

and the BPA glucuronic acid [74, 75]. The main metabolite in both zebrafish and rainbow

trout is the BPA glucuronic acid [75]. [70] found a decreasing trend in UGT gene

expression when juvenile atlantic cod were exposed to BPA. An increase in GSTπ

expression was also noted, even though its activity was not altered.

Total glutathione

In this study, only tGSH was shown to be impacted by BPA (Figure 21), whereas GSTπ

expression levels remained stable (Figure 18). In case of increased metabolism, one

would expect increased clearance of GSH and thus lower tGSH levels. It is thus possible

that the increased tGSH is not related to phase II metabolism but more to oxidative stress

response, especially since GSTπ does not seem to be involved in BPA metabolism. It

would probably be informative to analyze the GSSG:GSH ratio since it is modified in

case of oxidative stress due to GSSG formation. Unfortunately, the amount of cells


36

required to detect GSSG is quite high and so this step has not been performed in this

study.


Tightness of the membrane

It is difficult to determine whether the gill cells form a tight monolayer on the inserts in

this study. As the shape of these cells is quite elongated it is more difficult to see if they

are truly fully confluent. In addition to this, it is hard to detect them at all with a

microscope when they are seeded on the inserts due to the background formed by the

membrane. TER measurements, the presence of tight junction proteins and permeability

to dyes are all methods that can be used to determine how tight an epithelium is [33].

Only TER was measured in this study (Figure 22). This was performed on empty inserts

with both media, and after medium change to L-15/ex, shortly before exposure. The

values obtained were in the range of 102 Ω·cm2. A small increase in TER could be noted,

however the meaning of it is difficult to determine.

The study by Trubitt et al [34] suggests the presence of functioning tight junction proteins

by the measurement of TER and analysis of the response of confluent monolayers of

Rtgill-W1 on transwell inserts when exposed to different hormones. However the TER

values were significantly smaller than the ones found in primary cell cultures (in the

range of 10 Ω·cm2 vs. 103-104 Ω·cm2) [33, 34]. It states that tight junctions have also

been observed by TEM, even though they were undetectable on Western Blots. Their

expression at the gene level was also relatively low, indicating important differences

between in vivo epithelia and this in vitro model. However, they conclude that the

monolayer formed by RTgill-W1 should be tight.

In our case, the best would have been to measure the evolution of TER over time, which

was not performed due to contamination risks and lack of time. Moreover, an increase in

TER does not mean that the monolayer is totally tight, and should be backed up with

another method of analysis, such as tight junction protein detection. The permeability of

the membrane to the dye Lucifer Yellow will be analyzed in later experiments.

BPA exposure

Exposure of RTgill-W1 in the co-culture system increased CYP1A gene expression

(Figure 23). It is known that phase I metabolism also occurs in gills, and EROD activity

in the gill tissue is considered as a potential sensitive biomarker for exposure to CYP1A

inducers [2, 7]. Interestingly, it has been observed that estrogen exposure increases

EROD activity in the gills, which is in accordance with the CYP1A mRNA increase

noted here. The mechanisms of BPA are thus different in both organs, however to which

extent and why is not known [7]. CYP3A did not seem very impacted. Interestingly,

RTL-W1 responded (Figure 23), indicating indirect exposure to BPA. CYP1A gene

expression was induced, and Dio2 gene expression reduced, similarly as in the single cell

line exposure but to a lesser extent. CYP3A did not respond clearly.

The indirect exposure of the liver cell line could have happened via different ways. First

of all, it is not clear whether the RTgill-W1 was as tight as hoped for, which means that

leakage of BPA could have occurred. BPA could also have been taken up and excreted to


37

the basolateral side, maybe even metabolized. BPA is considered to be mainly

metabolized by the UGTs and SULTs in rainbow trout. It would have been interesting to

measure the gene expression of these enzymes in the RTgill-W1 but no sequence for

these genes was annotated for rainbow trout. The chemical analysis of both media will be

performed in the summer and might give interesting insights into the metabolic capacities

of RTgill-W1.


38

5 Conclusion

To conclude, the results obtained during this thesis confirm the good potential of cell

lines to be used for general cytotoxicity studies but their better characterization is needed

for more specific studies.

RTgill-W1 proved efficient and sensitive for the toxicity assessment of copper,

metoprolol and tributylphosphate. Single compound EC50s were in the range of in vivo

data (less than one order difference) and both Concentration Addition and Independent

Action predicted the mixture EC50s with less than a 2-fold difference compared to the

observed EC50. However this assay is limited for chronic toxicity and sublethal

endpoints, which is of critical importance for environmental mixture assessment. This

cell line could be of use for compounds acting through pathways that are known to be

expressed.

The co-culture system needs to be better characterized before it can be used for

toxicological studies, but promising indications of monolayer integrity have been

reported for RTgill-W1. In the specific case of BPA, inhibition of the Dio2 gene

expression was observed, indicating that this enzyme could potentially be an interesting

biomarker for endocrine disruption, encouraging further research in this direction. With

established in vitro biomarkers, capacities and limitations, this co-culture model could

prove very useful.


39

6 Acknowledgements

I would like to thank Joachim, my supervisor, who gave me the possibility to perform my

one year thesis in his lab; who somehow always managed to find time to listen, help and

support, even on weekends, and who guided me during this whole year with good advice.

Britt, who has been there every day of this thesis and has shown me everything, with

great patience, availability and clarity. Thank you ! Thomas, sometimes supervisor,

sometimes examiner, who was very helpful for the first part of this work and the

modeling. The whole ÅL group for its support (and nice fika!). Finally, the whole lunch

clan, my Fridays will be very boring from now on.


40

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