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Accepted Manuscript
Title: Enzymatic activity and gene expression changes in
zebrafish embryos and larvae exposed to pesticides diazinon
and diuron
Authors: Mirna Velki, Henriette Meyer-Alert,
Thomas-Benjamin Seiler, Henner Hollert
PII:
DOI:
Reference:
S0166-445X(17)30307-7
https://doi.org/10.1016/j.aquatox.2017.10.019
AQTOX 4781
To appear in:
Aquatic Toxicology
Received date:
Revised date:
Accepted date:
9-9-2017
23-10-2017
24-10-2017
Please cite this article as: Velki, Mirna, Meyer-Alert, Henriette, Seiler, ThomasBenjamin, Hollert, Henner, Enzymatic activity and gene expression changes in zebrafish
embryos and larvae exposed to pesticides diazinon and diuron.Aquatic Toxicology
https://doi.org/10.1016/j.aquatox.2017.10.019
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Aquatic Toxicology
Enzymatic activity and gene expression changes in zebrafish embryos and larvae exposed to
pesticides diazinon and diuron
Mirna Velki1,2*, Henriette Meyer-Alert1, Thomas-Benjamin Seiler1, Henner Hollert1
1
Department of Ecosystem Analysis, Institute for Environmental Research, RWTH Aachen University,
Worringerweg 1, 52074 Aachen, Germany
2
Department of Biology, Josip Juraj Strossmayer University of Osijek, Cara Hadrijana 8/A, 31000
Osijek, Croatia
Dipl.-Biol. Henriette Meyer-Alert, [email protected]
Dr. rer. nat. Thomas-Benjamin Seiler, [email protected]
Prof. Dr. rer. nat. Henner Hollert, [email protected]
*Corresponding author
Mirna Velki, PhD, assistant professor
Josip Juraj Strossmayer University in Osijek
Department of Biology
Cara Hadrijana 8/A, 31000 Osijek
Croatia
E-mail: [email protected] ; [email protected]
Highlights
•
Suborganismic responses of zebrafish early life stages to diazinon and diuron.
•
Significant effects on enzymatic responses and gene expression changes.
1
•
Different sensitivity of different developmental stages of zebrafish.
•
Partial recovery of enzyme activity 48 h after the end of exposure.
Abstract
The zebrafish as a test organism enables the investigation of effects on a wide range of
biological levels from molecular level to the whole-organism level. The use of fish embryos
represents an attractive model for studies aimed at understanding toxic mechanisms and the
environmental risk assessment of chemicals. In the present study, a zebrafish (Danio rerio) in vivo
model was employed in order to assess the effects of two commonly used pesticides, the insecticide
diazinon and the herbicide diuron, on zebrafish early life stages. Since it was previously established
that diazinon and diuron cause effects at the whole-organism level, this study assessed the
suborganismic responses to exposure to these pesticides and the enzymatic responses (biochemical
level) and the gene expression changes (molecular level) were analyzed. Different exposure scenarios
were employed and the following endpoints measured: acetylcholinesterase (AChE),
carboxylesterase (CES), ethoxyresorufin-O-deethylase (EROD), glutathione-S-transferase (GST),
catalase (CAT) and glutathione peroxidase (GPx) activities; and gene expressions of the
corresponding genes: acetylcholinesterase (ache), carboxylesterase (ces2), cytochrome P450 (cyp1a),
glutathione-S-transferase (gstp1), catalase (cat), glutathione peroxidase (gpx1a) and additionally
glutathione reductase (gsr). Significant changes at both the biochemical and the molecular level were
detected. In addition, different sensitivities of different developmental stages of zebrafish were
determined and partial recovery of the enzyme activity 48 h after the end of the exposure was
observed. The observed disparity between gene expression changes and alterations in enzyme
activities points to the necessity of monitoring changes at different levels of biological organization.
Different exposure scenarios, together with a comparison of the responses at the biochemical and
molecular level, provide valuable data on the effects of diazinon and diuron on low organizational
levels in zebrafish embryos and larvae.
Keywords: Danio rerio; diazinon; diuron; enzymatic activity; gene expression
1. Introduction
Intensive use of pesticides worldwide leads to extensive production and application of
compounds of diverse modes of action. Around 131,000 tons of herbicides and 21,000 tons of
insecticides were sold in the EU in 2014 (Eurostat, 2016). But despite the benefits that result from
2
pesticide usage (e.g. protection of crop plants and increased production of food, prevention of
spoilage of stored foods, control of the populations of disease vector organisms), their usage can also
lead to the contamination of the environment to a considerable extent (Aktar et al., 2009). Pesticides
enter the environment via periodic inputs from accidental or controlled sources (urban and industrial
discharges) and from diffuse sources originating from domestic and agricultural activities (Akcha et
al., 2012). Depending on their chemical properties, nature of the soil and method of application, the
compounds can contaminate the soil or become subject to leaching due to runoff from rain water,
which consequently leads to contamination of surface and ground water (Domagalski and Dubrovsky,
1992). Consequently, pesticides pose a potential hazard to a number of non-target organisms not
only in terrestrial ecosystems, but also in the aquatic ecosystems.
In recent years, the usage of the zebrafish (Danio rerio) embryo test has gained increasing
popularity as a valuable model in aquatic ecotoxicology for the assessment of the effects of various
pollutants. Among various practical advantages, such as relatively easy maintenance and breeding,
the embryos are transparent and fully develop most organ systems within 96 h post-fertilization
(hpf), making them an ideal vertebrate model system (Jaja-Chimedza et al., 2012). The usage of
zebrafish embryos is in accordance with the 3Rs principle of animal research (Replacement,
Reduction and Refinement) (Russell and Burch, 1959) and protocol using fish eggs under short-term
exposure has replaced the former mandatory acute fish test for the toxicity evaluation of wastewater
in Germany (DIN, 2001). In addition to being accepted as a tool in waste water monitoring, zebrafish
embryos are being used for the assessment of adverse effects of various groups of substances,
including pesticides. The utilization of zebrafish as a test organism enables the investigation of effects
on a wide range of biological levels – from the molecular to the whole-organism level – and the use
of fish embryos represents an attractive model for studies aimed at understanding toxic mechanisms
and environmental risk assessment of chemicals (Scholz et al., 2008).
In the present study, a zebrafish in vivo model was employed in order to investigate the toxic
mechanisms of two commonly used pesticides – the insecticide diazinon and the herbicide diuron –
3
on zebrafish early life stages. With respect to the current study, the zebrafish embryos and larvae
have been used previously to investigate the effects of these pesticides and it was determined that
both diazinon and diuron significantly affect behavioral endpoints (Velki et al., 2017). Since the
effects on the whole-organism level were established, the current study assessed the suborganismic
responses in order to gain more knowledge on the mechanisms of action and to enable a better
characterization of the impacts of these pesticides on zebrafish. To obtain a detailed insight into the
effects of the investigated pesticides, the responses at the biochemical (enzymatic responses) and
molecular (gene expression changes) level were analyzed. The quantification of enzyme responses
and analyses of gene expression in homogenates of embryos/larvae enabled a relatively fast
screening since in both types of samples (i.e. for enzyme activities and for gene expression) multiple
endpoints could be assessed.
The endpoints for both enzymatic responses and gene expression were chosen based on the
main mode of action of the investigated pesticides. The organophosphate diazinon inhibits the
enzyme acetylcholinesterase (AChE) leading to the inactivation of the neurotransmitter acetylcholine
which is essential for the onward transmission of stimuli (Ecobichon and Joy, 1994). Diuron is a
phenylurea herbicide that blocks the electron transfer to the photosystem II which inhibits
photosynthetic oxygen and energy production in plants and algae (Wessels and Van der Veen, 1956),
and it was determined that in aquatic organisms diuron can cause an increased production of
reactive oxidative species (ROS) and oxidative stress (Behrens et al., 2016). Therefore, the following
enzyme responses were measured: AChE activity (biomarker of organophosphate exposure),
carboxylesterase (CES) activity (biomarker of organophosphate exposure and phase I detoxification
enzyme), ethoxyresorufin-O-deethylase (EROD) activity (phase I detoxification enzyme), glutathioneS-transferase (GST) activity (phase II detoxification enzyme and oxidative stress biomarker), catalase
(CAT) activity (oxidative stress biomarker) and glutathione peroxidase (GPx) activity (oxidative stress
biomarker). Regarding the gene expression, the following corresponding genes were selected:
acetylcholinesterase (ache), carboxylesterase (ces2), cytochrome P450 (cyp1a), glutathione-S4
transferase (gstp1), catalase (cat), glutathione peroxidase (gpx1a) and additionally glutathione
reductase (gsr).
An additional goal in this study was to assess the sensitivity of different developmental
stages and the possible recovery after exposure. In order to achieve this goal different exposure
scenarios were employed. Considering the previously determined effects of investigated pesticides
on whole-organism level and taking into account the main mode of actions, it was expected that
diazinon and diuron would cause changes in measured parameters in zebrafish on both enzymatic
and gene expression level. The implemented experimental setup, together with a comparison of the
responses at biochemical and molecular level, provide valuable data on the effects of diazion and
diuron on low organizational levels in zebrafish embryos and larvae.
2. Materials and methods
2.1. Chemicals
The following chemicals were used in this study and purchased from Sigma-Aldrich (Sigma
Aldrich Chemie GmbH, Steinheim, Germany): O,O-diethyl-O-[2-isopropyl-4-methyl-6-pyrimidyl]
phosphorothioate (diazinon 100 %, CAS Number: 333-41-5); N-(3,4-dichlorophenyl)-N,N-dimethylurea (diuron ≥98 %, CAS Number: 330-54-1); dimethyl sulfoxide (DMSO) (CAS Number: 67-68-5);
calcium chloride dihydrate (CAS Number: 10035-04-8); magnesium sulfate heptahydrate (CAS
Number: 10034-99-8); sodium hydrogen carbonate (CAS Number: 144-55-8); potassium chloride (CAS
Number: 7447-40-7); 5,5′-dithiobis(2-nitrobenzoic acid) (DTNB) (CAS Number: 69-78-3);
acetylthiocholine iodide (CAS Number: 1866-15-5); hydrogen peroxide (CAS Number: 7722-84-1); 4nitrophenyl acetate (CAS Number: 830-03-5); acetonitrile (CAS Number: 75-05-8); resorufin (CAS
Number: 635-78-9); 7-ethoxyresorufin (CAS Number: 5725-91-7); methanol (CAS Number: 67-56-1);
β-Nicotinamide adenine dinucleotide 2′-phosphate reduced tetrasodium salt hydrate (β-NADPH)
(CAS Number: 2646-71-1); ethylenediaminetetraacetic acid (EDTA) (CAS Number: 60-00-4); sodium
azide (CAS Number: 26628-22-8); glutathione reductase from baker's yeast (S. cerevisiae) (CAS
5
Number: 9001-48-3); L-glutathione reduced (GSH) (CAS Number: 70-18-8); 1-chloro-2,4dinitrobenzene (CDNB) (CAS Number: 97-00-7); bicinchoninic acid (BCA) kit for protein determination
(BCA1).
For the gene expression analysis the following kits/chemicals were purchased: NucleoSpin
RNA Isolation Kit (Macherey and Nagel, Düren, Germany); M-MuLV reverse transcriptase, random
primers mix, dNTPs (New England Biolabs, Frankfurt a.M., Germany); Power SYBR Green PCR Master
Mix (Life Technologies/Thermo Fisher Scientific, Darmstadt, Germany).
Diazinon and diuron stock solutions were prepared by dilution in DMSO of up to 2 mg mL-1
and stored at 4 °C.
2.2. Test organism
Zebrafish maintenance and egg production was performed according to standard procedures
(description available in Lammer et al., 2009) and is described in detail in a previous study (Velki et
al., 2017). In short, zebrafish adults were kept at 26 ± 1 °C and light-dark cycle of 14:10 hours at the
Fraunhofer IME Institute (Aachen, Germany). The eggs were obtained from community mating and
only fertilized eggs undergoing cleavage and showing no obvious irregularities during cleavage (e.g.,
asymmetry, vesicle formation) or injuries of the chorion were selected for the exposure tests (OECD,
2013). The selected fertilized eggs were placed at 27 ᵒC in pre-aerated artificial water (294.0 mg/L
CaCl2 · 2H2O; 5.5 mg/L KCl; 123.3 mg/L, magnesium sulfate MgSO4 · 7 H2O; 63.0 mg/L NaHCO3; after
aeration the pH was adjusted to 7.8) which was prepared according to ISO standards (ISO, 2007) and
exposures were conducted when the embryos were in a 32-cell-stage (ca. 2 hpf). The usage of
zebrafish was in accordance with the German law for Animal Protection (“Tierschutzgesetz”) and EU
Directive on the protection of animals used for scientific purposes (EU D2010/63/EU). Namely, in the
experiments zebrafish embryos and larvae of up to 98 hpf were used and since zebrafish younger
than 120 hpf are not considered animals (Strähle et al., 2012) no animal test authorization was
6
required according to German legislation. The used term “larva” refers to 98 hpf hatched embryos
that are using up the yolk-sac reserves and still do not feed externally.
2.3. Exposure experiments for the assessment of enzymatic activity and gene expression
For the assessment of pesticide effects on enzymatic activities and the determination of
sensitivity and recovery of different developmental stages, fertilized zebrafish eggs were exposed to
diazinon and diuron applying 4 exposure scenarios: 1) fertilized eggs were exposed for 48 h – “early
exposure” (2 hpf – 50 hpf exposure), 2) fertilized eggs were exposed for 96 h – “prolonged exposure”
(2 hpf – 98 hpf exposure), 3) fertilized eggs were first placed in artificial water for 48 h and then
exposed for 48 h – “late exposure” (50 hpf – 98 hpf exposure), 4) fertilized eggs were exposed for 48
h and then placed in artificial water for 48 h recovery – “recovery exposure” (2 hpf – 50 hpf
exposure, 50 hpf – 98 hpf recovery).
The pesticides were tested in five different concentrations prepared as dilutions of respective
DMSO stock solution in artificial water, with all conditions containing a final DMSO concentration of
0.1 %. The exposure concentrations were sublethal and were selected based on the no-observableeffect-concentration (NOEC) determined in a previous study (Velki et al., 2017). Namely, NOEC for
both pesticides was 2 mg L-1, equaling 6.6 µM diazinon and 8.6 µM diuron, and this was chosen as
the highest exposure concentration. Besides NOEC, four additional concentrations were assessed for
both pesticides, i.e. 6.6, 3.3, 1.65, 0.33 and 0.066 µM diazinon, and 8.6, 4.3, 2.15, 0.43 and 0.086 µM
diuron were tested. Exposure scenarios were conducted in borosilicate glass vessels with 50 embryos
exposed in 50 mL of solution at 26 ± 1°C and 14 h photoperiod.
In order to determine the effects of the pesticides on gene expression and to compare it to
the enzyme responses, additional sets of exposures were conducted. Specifically, for the assessment
of gene expression changes, fertilized zebrafish eggs were exposed to diazinon and diuron employing
the early (48 h exposure, 2 hpf – 50 hpf) and the prolonged (96 h exposure, 2 hpf – 98 hpf) exposure
scenarios. Zebrafish were exposed to three concentrations of diazinon and diuron, i.e. to 6.6, 1.65
7
and 0.066 µM diazinon and to 8.6, 2.15 and and 0.086 µM diuron (the lowest, middle and the highest
concentration used in assessment of enzyme responses). Again, pesticides were prepared as dilutions
of respective DMSO stock solutions in artificial water (the final DMSO concentration was 0.1 %).
Exposures were conducted in borosilicate glass vessels with 30 embryos exposed in 30 mL of solution
at 26 ± 1°C and 14 h photoperiod.
In all experiments the semi-static renewal technique was applied, i.e. the pesticide test
solutions were regularly renewed every 24 h to ensure the maintenance of the pesticide
concentrations. Since DMSO was used as a solvent, the analyses of pesticide effects on both
enzymatic activity and gene expression were made in comparison to the DMSO solvent control
condition. For that it was ensured that the same DMSO concentration of 0.1% was kept in all
pesticide exposures and solvent control conditions (hereinafter abbreviated as “control”).
Experiments were repeated five times and repetitions were conducted as independent experiments
(for every repetition zebrafish embryos from different batches were used).
2.4. Sample preparation
At the end of each exposure scenario, embryos/larvae (50 for the enzyme activity and 30 for
the gene expression) were transferred in a glass vessel containing benzocaine (saturated solution in
water). After ca. 30 s in benzocaine, embryos/larvae were washed twice in phosphate buffered
saline. Embryos/larvae were then transferred to 2 mL tubes and the excess of phosphate buffered
saline was removed.
For enzyme activity measurements, 600 μL of buffer (1.8 L 0.1 M Na2HPO4 adjusted with 0.5 L
0.1 M NaH2PO4 to pH 7.8) in tubes with 50 embryos/larvae was added. Samples were snap frozen in
liquid nitrogen and stored at –80 °C until homogenization. When all sample replicates were collected,
they were thawed on ice, homogenized for ca. 20 s (homogenizer VDI 12, S12N-5S, VWR
International GmbH, Darmstadt, Germany) and centrifuged at 10000 g, 15 min, 4 °C. Supernatants
were aliquoted and used for enzyme activity measurements.
8
In the case of samples for gene expression measurements, tubes with 30 embryos/larvae
were snap frozen in liquid nitrogen and stored at -80 °C. When all sample replicates were collected,
they were thawed on ice and used for isolation of RNA.
2.5. Enzyme activity measurements
2.5.1. Measurement of acetylcholinesterase (AChE) activity
The measurement of the AChE activity was conducted according to the method of Ellman et
al. (1961) and was adapted for measurement in 96-well plate. The reaction medium consisted of 180
μL of sodium phosphate buffer (0.1 M, pH 7.2), 10 μL of DTNB (1.6 mM), 10 μL of acetylthiocholine
iodide (156 mM) and 7.5 μL of sample. The increase in absorbance was measured as triplicates at 412
nm for 5 min in 10 s steps at 25°C by using an Infinite® 200 micro plate reader (Tecan, Männedorf,
Switzerland). The enzymatic activity was expressed as nmol of acetylthiocholine hydrolysed per min
per mg of protein and for calculations the molar extinction coefficient of 13600 M-1 cm-1 was used.
2.5.2. Measurement of carboxylesterase (CES) activity
The activity of CES was measured according to Hosokawa and Satoh (2002) with some
modifications and was adapted for measurement in 96-well plate. The reaction medium was
comprised of 15 µL of sample and 150 μL of 4-nitrophenyl acetate (1 mM). The increase in
absorbance was measured as triplicates at 405 nm for 3 min in 10 s steps at 25°C by using an Infinite®
200 micro plate reader (Tecan, Männedorf, Switzerland). The enzymatic activity was expressed as
nmol of 4-nitrophenol formed per min per mg of protein and for calculations the molar extinction
coefficient of 16400 M−1 cm−1 was used.
2.5.3. Measurement of ethoxyresorufin-O-deethylase (EROD) activity
The measurement of the EROD activity was conducted based on the methodology described
by Schiwy et al. (2015) with some modifications. First, 35 μL of sample (or resorufin in the case of
9
standard curve) and 60 μL of 7-ethoxyresorufin (0.6 μM in methanol) were added to 96-well plate
and incubated for 10 min at 28 °C in darkness to allow the enzyme to attach to its substrate. Then 45
μL of β-NADPH (3.35 mM) was added and incubated for 2 min at room temperature in darkness. The
increase in fluorescence was measured as triplicates at 530 nm excitation and 590 nm emission for
30 min in 30 s steps at 25°C by using an Infinite® 200 micro plate reader (Tecan, Männedorf,
Switzerland). The enzymatic activity was calculated based on the standard curve with resorufin and
expressed as fmol of resorufin formed per min per mg of protein.
2.5.4. Measurement of glutathione-S-transferase (GST) activity
The measurement of the GST activity was conducted following the method of Habig et al.
(1974) and adapted for the measurement in 96-well plate. The reaction medium consisted of 180 μL
of CDNB (1 mM), 12.5 µL of sample and 50 µL of GSH (25 mM). The increase in absorbance was
measured as triplicates at 340 nm for 3 min in 10 s steps at 25°C by using an Infinite® 200 micro plate
reader (Tecan, Männedorf, Switzerland). The enzymatic activity was expressed as nmol of conjugated
GSH in one minute per mg of proteins and for calculations the molar extinction coefficient of 9600 M1
cm-1 was used.
2.5.5. Measurement of catalase (CAT) activity
The measurement of CAT activity was conducted following the method of Claiborne (1985).
The reaction medium contained 400 μL of hydrogen peroxide (0.02 M, prepared in sodium
phosphate buffer (0.1 M, pH 7.2)) and 40 µL of sample. The decrease in absorbance was measured as
triplicates at 240 nm for 2 min in 10 s steps at 25°C by using an Infinite® 200 micro plate reader
(Tecan, Männedorf, Switzerland). To calculate hydrogen peroxide concentration, a molar extinction
coefficient of 42.6 M-1 cm-1 was used and the activity was expressed as μmol of degraded hydrogen
peroxide per min per mg of protein.
10
2.5.6. Measurement of glutathione peroxidase (GPx) activity
The activity of GPx was measured according to the method of Wendel (1980) and adapted for
the measurement in 96-well plate. The reaction cocktail contained sodium azide (1 mM), β-NADPH
(120 mM), glutathione reductase and GSH (200 mM). In each microplate well 160 µL of reaction
cocktail was mixed with 40 µL of sample. After stabilization of this mixture for 5 min, 20 µL of
hydrogen peroxide (50 mM) was added and the increase in absorbance was measured as triplicates
at 340 nm for 5 min in 10 s steps at 25°C by using an Infinite® 200 micro plate reader (Tecan,
Männedorf, Switzerland). The enzymatic activity was expressed as nmol of NADPH oxidized per min
per mg of protein and for calculations the molar extinction coefficient of 6220 M−1 cm−1 was used.
2.5.7. Measurement of protein concentration
For calculation of the specific enzyme activities (activity expressed as amount of used
substrate per min per mg of protein), the respective protein concentrations in all samples were
measured using the BCA kit. The standard curve was made using the bovine serum albumin (BSA)
standard provided in the kit. Protein concentrations of the samples were determined in a 96-well
plate by adding 200 µL of a working solution (prepared according to the kit instructions) to 10 µL of
the sample. After incubation for 30 min at 37 °C, the absorbance was recorded at 562 nm by using an
Infinite® 200 micro plate reader (Tecan). The protein concentrations were then determined by
comparison to a standard curve and used for the calculation of the specific enzyme activities. The
calculated specific activities were then expressed relative to the average value of the controls (i.e.
making control equal to 1).
2.6. Gene expression measurements
2.6.1. RNA isolation and cDNA synthesis
11
For the gene expression analysis total RNA from zebrafish embryos/larvae was extracted
using the NucleoSpin RNA kit. In tubes with zebrafish embryos/larvae 450 µL of lysis buffer (provided
in the kit) was added and the samples were homogenized for ca. 20 s (VDI 12, S12N-5S, VWR
International GmbH, Darmstadt, Germany). The obtained homogenate was centrifuged at 14000 g,
10 min, 4 °C. The supernatant was used in the following steps and the extraction was further
performed according to the kit instructions. The RNA concentration and purity were determined
spectrophotometrically using BioDrop-µLITE (BioDrop Ldt., Cambridge, UK). The average RNA yield
was 400 ng/µL, while the A260/A280 ratio ranged from 2.05 to 2.1. For the cDNA synthesis, 1 µg of
total RNA from each sample was reverse transcribed into cDNA in a 20 µL reaction mixture
containing the random primer mix, reverse transcriptase M-MuLV RT and dNTPs according to the
supplier’s instructions.
2.6.2. Quantitative real-time PCR analysis (qRT-PCR)
Gene specific primers for ache, ces2, cyp1a, gstp1, cat, gpx1a, gsr, eef1a1l2 and actb2 were
either designed using the Primer-BLAST or previously published sequences were used. The qRT-PCR
was performed on StepOnePlus Real-Time PCR System (Applied Biosystems) using the Power SYBR
Green PCR Master Mix according to the supplier’s instructions for quantification of gene expression.
After denaturation at 95 °C for 10 min, 40 cycles of amplification were carried out with denaturation
at 95 °C for 15 s, annealing and elongation at 61 °C for 1 min. To ensure that a single product was
amplified, melt curve analysis was performed on the PCR products and polyacrylamide gel
electrophoresis was performed for each gene. Samples were analyzed in technical triplicates and for
determination of PCR efficiencies, composite cDNA samples were serially diluted and run in a qPCR
reaction. The average PCR efficiencies were the following: 109 % for ache; 101 % for ces2; 102 % for
cyp1a; 91 % for gstp1; 109 % for cat; 102 % for gpx1a; 108 % for gsr; 97 % for eef1a1l2 and 96 % for
actb2. Expression of the target genes was quantified according to the model reported by Pfaffl
(2001). The expression level of the target gene was normalized to the two reference genes – eef1a1l2
12
and actb2 (for all target genes geometric mean of the expression of these two reference genes was
used as a normalization factor). The levels of gene expression were expressed relative to the average
value of the controls (i.e. making control equal to 1).
Table 1. Primer sequences, amplicon sizes, accession numbers and annealing temperatures used in
qPCR reaction.
Gene
Primer sequence (5’ – 3’)
F
CATACGCACAATACGCTGCC
R
TACACAGCACCATGCGAGTT
F
GTGGAGCTTGCATGTTTAAGG
R
GCTGATCTCCTGTGCTGAAGTA
cyp1a
F
AAAGACACCTGCGTGTTTGTAA
*
R
GAGGGATCCTTCCACAGTTCT
gstp1
F
TGGTGCTGTTTCAGTCCAAG
**
R
AGCCTCACTGTCGTTTTTGC
F
TCTCCTGATGTGGCCCGATA
R
GGTTTTGCACCATGCGTTTC
F
GAGGCACAACAGTCAGGGATT
R
CTTCATTCTTGCAGTTCTCCTGGT
F
CGGCCTCAACCTCAGTCAAA
R
TGCTTCATCAGGTGTCAGAAGG
eef1a1l2
F
GTGGTAATGTGGCTGGAGAC
***
R
TGTGAGCAGTGTGGCAATC
actb2
F
GAAGATCAAGATCATTGCCCCAC
****
R
TCGGCGTGAAGTGGTAACAG
ache
ces2
cat
gpx1a
gsr
Amplicon
size (bp)
Accession number
Annealing
temperature
118
NM_131846.2
61 °C
129
NM_001077252.1
61 °C
68
NM_131879.2
60 °C
80
NM_131734.3
60 °C
169
NM_130912.2
61 °C
126
NM_001007281.2
60 °C
142
NM_001020554.1
60 °C
138
NM_001039985.1
59 °C
179
NM_181601.4
59 °C
Gene abbreviations: ache – acetylcholinesterase; ces2 – carboxylesterase 2; cyp1a – cytochrome
P450, family 1, subfamily A; gstp1 – glutathione S-transferase pi 1; cat – catalase; gpx1a –
glutathione peroxidase 1a; gsr – glutathione reductase; eef1a1l2 – eukaryotic translation elongation
factor 1 alpha 1, like 2; actb2 – actin, beta 2.
* Bräunig et al. (2015)
** designed within the DanTox project at Institute for Environmental Research, RWTH Aachen University
*** Kühnert (2015)
**** Primers were designed at Institute for Environmental Research, RWTH Aachen University by Sina Volz
using the NCBI Primer Blast BeaconPrimer and NetPrimer.
13
2.7. Data analysis
Data analyses and plotting were performed using the software R version 3.1.2. (normality
testing) and GraphPad Prism 5 (other analyses and plotting). The data were checked for normality
prior to further analysis (Shapiro-Wilk normality test). The enzymatic responses were normally
distributed (in several cases data were log-transformed to obtain normality) and in order to
determine the significance level reached between the control and treatment groups, one-way
analysis of variance (ANOVA) was applied followed by Dunnett’s multiple comparison test. For the
gene expression results the distribution of some response variables departed significantly from
normality, so the non-parametric statistics were considered adequate for the statistical analysis of
the data obtained. Gene expression data were therefore analyzed using the non-parametric Kruskal–
Wallis one-way analysis of variance by ranks followed by Dunn´s multiple comparison test.
The probability level for statistical significance was set to p<0.05 throughout the study. The
data analyzed using ANOVA (enzymatic responses) are presented as mean ± standard deviation,
whereas data analyzed using Kruskal–Wallis (gene expression responses) are presented using the
Tukey's boxplot (the bottom and the top of the box are 25th and 75th percentile, the horizontal line in
the box is the median and the ends of the whiskers represent 1.5 times of the interquartile distance).
3. Results
3.1. Enzyme activity results
3.1.1. Enzyme activities measured in embryos after early exposure (2 hpf – 50 hpf exposure)
The measurement of enzyme activities in zebrafish embryos exposed to diazinon and diuron
for 48 h showed significant differences in CES, EROD and GST activities (Fig. 1) compared to the
control. The changes of EROD and GST activity were observed in only one concentration – the
induction of EROD at 0.33 µM diazinon and the decrease of GST at 0.086 µM diuron. Regarding the
CES activity, the significant inhibition was observed in all diazinon concentrations applied and the
14
activity decreased up to 51 % (at 6.6 µM). AChE, CAT and GPx activity showed no significant changes
after the exposures to diazinon and diuron.
Figure 1. Relative enzyme activity measurements in zebrafish embryos exposed to diazinon (DIA) and
diuron (DIU) for 48 h (“early exposure”; 2 hpf – 50 hpf exposure). Bars represent means ± standard
15
deviations (N=5). Significant differences compared to control are labelled with * (p<0.05), ** (p<0.01)
and *** (p<0.001).
3.1.2. Enzyme activities measured in larvae after prolonged exposure (2 hpf – 98 hpf exposure)
Measurement of enzyme activities in zebrafish larvae exposed to diazinon and diuron for 96
h showed significant differences in AChE and CES activities when compared to the control (Fig. 2).
Specifically, diazinon caused significant inhibition of AChE activity of 31 and 49 % at the two highest
concentrations, i.e. at 3.3 and 6.6 µM, respectively. Regarding the CES activity, the significant
inhibition was observed for all diazinon concentrations, and the activity decreased up to 61 % (at 6.6
µM). For the CES activity, significant changes were detected also after diuron exposure. At all diuron
concentrations a tendency of an increase in CES activity was observed, even though only at two
concentrations (0.086 and 2.15 µM) significant increase in CES activity was recorded. Other
measured enzyme activities showed no significant changes after the exposures to diazinon and
diuron.
16
Figure 2. Relative enzyme activity measurements in zebrafish larvae exposed to diazinon (DIA) and
diuron (DIU) for 96 h (“prolonged exposure”; 2 hpf – 98 hpf exposure). Bars represent means ±
standard deviations (N=5). Significant differences compared to control are labelled with * (p<0.05),
** (p<0.01) and *** (p<0.001).
17
3.1.3. Enzyme activities measured in larvae after late exposure (50 hpf – 98 hpf exposure)
The measurement of enzyme activities in zebrafish larvae exposed from 50 hpf to 98 hpf are
shown in Fig. 3. Similar to the prolonged exposure (2 hpf – 98 hpf), significant differences in AChE
and CES activities were detected. Namely, diazinon caused significant inhibition of AChE activity of 31
% at the highest concentration applied (6.6 µM). Regarding the CES activity, the significant inhibition
was observed for all diazinon concentrations, and the activity decreased down to 42 %. In contrast to
the prolonged exposure (2 hpf – 98 hpf), where an increase in CES was observed after diuron
exposure (Fig. 2), here at all applied diuron concentrations showed lower activity and a significant
decrease was obtained at 0.086 and 4.3 µM. The significant changes in other measured enzymes
could not be detected.
18
Figure 3. Relative enzyme activity measurements in zebrafish larvae exposed to diazinon (DIA) and
diuron (DIU) for 48 h (“late exposure”; 50 hpf – 98 hpf exposure). Bars represent means ± standard
deviations (N=5). Significant differences compared to control are labelled with * (p<0.05), ** (p<0.01)
and *** (p<0.001).
19
3.1.4. Enzyme activities measured in larvae after recovery exposure (2 hpf – 50 hpf exposure, 50 hpf –
98 hpf recovery)
The measurement of enzyme activities in zebrafish larvae exposed for 48 h and then placed
in artificial water for 48 h are shown in Fig. 4. Significant changes were observed only for CES activity
after exposure to diazinon. Specifically, significant inhibition of CES activities was determined at 0.33,
1.65, 3.3 and 6.6 µM diazinon. Compared to CES inhibition in zebrafish embryos after early exposure
(2 hpf – 50 hpf, Fig. 1), here the significant inhibition was not detected at the lowest concentration
applied, i.e. at 0.066 µM, so partial recovery can be noticed. The level of CES inhibition was similar as
in embryos exposed for 48 h; it reached 57 % at 6.6 µM of diazinon. Significant changes in other
measured enzymes could not be detected.
20
Figure 4. Relative enzyme activity measurements in zebrafish larvae exposed to diazinon (DIA) and
diuron (DIU) for 48 h followed by 48 h recovery (“recovery exposure”; 2 hpf – 50 hpf exposure, 50 hpf
– 98 hpf recovery). Bars represent means ± standard deviations (N=5). Significant differences
compared to control are labelled with * (p<0.05), ** (p<0.01) and *** (p<0.001).
21
3.2. Gene expression results
3.2.1. qRT-PCR results measured in embryos after early exposure (2 hpf – 50 hpf exposure)
Results of the measurement of the gene expression in zebrafish embryos exposed to diazinon
and diuron for 48 h are shown in Fig. 5. Significant changes were observed in the case of ces2, cyp1a,
cat and gpx1a. Namely, diazinon caused an increase in the expression of ces2 (Fig. 5) while at the
enzymatic level the activity of the enzyme was inhibited (Fig. 1). Diuron also caused an increase in
the expression of ces2 (Fig. 5), even though at the enzymatic level no significant changes could be
observed (Fig. 1). Regarding the cyp1a expression, a significant increase after exposure to diuron was
detected (Fig. 5), even though at the enzymatic level no significant changes could be observed (Fig.
1). Changes in CAT and GPX activity were not recorded after exposure to the investigated pesticides
(Fig. 1), however the expression of cat was significantly decreased after exposure to diazinon
whereas the expression of gpx1a was significantly increased after exposure to diuron (Fig. 5).
22
Figure 5. qRT-PCR results of selected genes measured in zebrafish embryos exposed to diazinon (DIA)
and diuron (DIU) for 48 h (“early exposure”; 2 hpf – 50 hpf exposure). The results are expressed as
23
normalized gene expression and presented using Tukey's boxplot (N=5). Significant differences
compared to control are labelled with * (p<0.05), ** (p<0.01) and *** (p<0.001).
3.2.2. qRT-PCR results measured in larvae after prolonged exposure (2 hpf – 98 hpf exposure)
The results of the gene expression measurements in zebrafish larvae exposed to diazinon and
diuron for 96 h are shown in Fig. 6. Significant changes were observed in all measured target genes.
Specifically, diazinon caused an increase in the expression of ache and ces2 (Fig. 6), while at the
enzymatic level the activities of these enzymes were decreased (Fig. 2). Diuron also caused an
increase in the expression of ces2 (Fig. 6), which is in accordance with the increase in CES activity (Fig.
2). Regarding the cyp1a expression, same as for the 48 h, a significant increase after the exposure to
diuron was detected (Fig. 6), even though at the enzymatic level no significant changes could be
observed (Fig. 2). Changes in GST, CAT and GPX activity were not recorded after the exposure to the
pesticides (Fig. 2), however the expression of gstp1 was significantly increased after exposure to
diazinon, and the expression of cat, gpx1a and gsr were significantly increased after exposure to
diazinon and diuron (Fig. 6).
24
Figure 6. qRT-PCR results of selected genes measured in zebrafish larvae exposed to diazinon (DIA)
and diuron (DIU) for 96 h (“prolonged exposure”; 2 hpf – 98 hpf exposure). The results are expressed
25
as normalized gene expression and presented using Tukey's boxplot (N=5). Significant differences
compared to control are labelled with * (p<0.05), ** (p<0.01) and *** (p<0.001).
4. Discussion
In the present study, the effects of the insecticide diazinon and the herbicide diuron on
zebrafish embryos and larvae were investigated by assessing the responses at the biochemical
(enzymatic responses) and molecular (gene expression changes) level. The modes-of-action of these
pesticides are naturally different, and even though it would be expected that an exposure to the
insecticide diazinon targeting the nervous system would result in more pronounced adverse effects
in zebrafish compared to effects of herbicide diuron targeting photosystem II, a previous study
showed that both pesticides have similar acute toxicity on zebrafish larvae and cause changes in
behavior (Velki et al., 2017). Regarding the assessment of other effects, studies investigating effects
of diuron on zebrafish are very scarce and besides evaluation of developmental toxicity using the fish
embryo test (Padilla et al., 2012), no data is available. Effects of diazinon on zebrafish have been
investigated in several studies and it was shown that besides behavioral effects diazinon also causes
morphological and physiological effects, inhibits AChE activity and increases the abundance of stress
protein (Hsp70) (Modra et al., 2011; Yen et al., 2011; Watson et al., 2014). In the present study it was
shown that the exposure of zebrafish embryos and larvae to diazinon and diuron leads to changes in
enzyme activities and gene expression.
Measurements of enzyme activities were conducted after applying four different exposure
scenarios which enabled the comparison of sensitivity of different developmental stages and the
assessment of possible recovery. The early exposure of 48 h (2 hpf – 50 hpf) to diazinon showed the
inhibition of CES activity. Considering that diazinon is an organophosphate, the inhibition was
expected due to the mode of action. Surprisingly, there was no change in the activity of AChE. Both
AChE and CES are inhibited by organophosphates and AChE activity is commonly used as a biomarker
of exposure to organophosphate compounds (e.g. Rickwood and Galloway, 2004; Assis et al., 2011;
26
Ghazala et al., 2014). However, it has been previously established that CES is frequently more
sensitive to organophosphate inhibition than AChE (Escartin and Porte, 1997; O’Neill et al., 2004;
Wheelock et al., 2005; Collange et al., 2010, etc.) and the higher sensitivity of CES was also confirmed
in the present study. Both AChE and CES are inhibited by the same class of pesticide compounds, but
these enzymes have different roles in the organism. Namely, AChE is involved in the termination of
impulse neurotransmission, whereas CES is a phase I detoxification enzyme. So due to their different
roles, as well as different sensitivities to particular compounds, the activities of both enzymes should
be always assessed. Regarding the other measured enzyme activities, a significant change in EROD
and GST activity was recorded. However, since variances in the responses were high and significant
changes were recorded at only one concentration applied, no conclusions could be made.
The prolonged exposure (for 96 h, 2 hpf – 98 hpf) revealed a similar response of CES activity
after diazinon exposure, i.e. again the inhibition was recorded and longer exposure caused higher
level of inhibition compared to early exposure. Also, the prolonged exposure led to a significant
inhibition of AChE activity at the two highest concentrations of diazinon. The observed 49 % AChE
inhibition at 6.6 μM is in accordance with the inhibition recorded in the study of Yen et al. (2011)
where an exposure to 10 μM diazinon for 5 days caused over 50 % AChE inhibition in zebrafish
larvae. A comparison of AChE and CES inhibition again revealed the higher inhibition and sensitivity
of CES activity. Regarding diuron, the prolonged exposure lead to an increase in CES activity
compared to control (which was not observed after early exposure). It is known that CES enzymes
play a significant role in the metabolism and subsequent detoxification of many agrochemicals and
pharmaceuticals (Redinbo and Potter 2005; Potter and Wadkins 2006). Thus, prolonged exposure to
diuron likely induces detoxification mechanisms leading to an increase in CES activity.
The late exposure (50 hpf – 98 hpf) enabled the comparison of the response with an early
exposure (2 hpf – 50 hpf) and determination of the difference in sensitivity between stages. The
obtained results showed that even though the duration of exposure was the same (48 h), different
stages of development of the zebrafish embryos gave different responses. In the case of the late
27
exposure, a significant inhibition of both AChE and CES activity was observed. Since after the early
exposure for 48 h no inhibition of AChE activity could be detected and the inhibition of CES was
lower, it can be concluded that the later stage had a higher sensitivity to diazinon exposure. So the
later zebrafish life stage was more susceptible to the pesticide exposure, and the response
resembled the one of the prolonged exposure more to. The possible explanation for the difference in
susceptibility can be ascribed to the difference in the ontogenetic stage. Namely, several
investigations support the idea that fish early stages are commonly most sensitive to chemicals
exposure (e.g. Luckenbach et al., 2001; Hallare et al., 2004) and Léonard et al. (2005) reported that
the eleutheroembryo may be more sensitive than the embryo. Our results also showed that larvae
were more vulnerable to the pesticide exposure than the embryo stage. Similar results have been
obtained in studies assessing effects of other pesticides. For example, it was determined that the
larval stage of zebrafish had a higher sensitivity compared to the embryo stage after exposure to the
organophosphate insecticides phoxim and chlorpyrifos (Wang et al., 2016), acetanilide herbicide
butachlor (Wang et al., 2016) and aryloxyphenoxypropionate herbicide cyhalofop-butyl (Cao et al.,
2016), and triazole fungicide difenoconazole (Mu et al., 2013). The sensitivity of the larval stage may
be due to the fact that the embryonic membrane and chorion prevent the transmission of chemicals
into the embryo (e.g. Xu et al. 2008; Embry et al. 2010), while larvae lack the chorion protection (e.g.
Fent and Meier, 1992). So in the early exposures, embryos were still protected by the chorion,
whereas in the post-hatch life-stage which was used for the late exposure, the lack of a chorion could
lead to higher susceptibility to pesticide exposure. Also, as reviewed in Embry et al. (2010), in
comparison to the juvenile and adult fish, early life stages may not have the fully developed
metabolic pathways to degrade xenobiotics, e.g. lower levels of some cytochrome P450 enzymes in
trout embryos compared to adult fish have been recorded (Buhler et al., 1997).
Likewise, the late exposure to diuron led to a significant decrease in CES activity, whereas at
the early exposure no changes could be detected. This result also points to the different
susceptibility of different life stages. More interestingly, the prolonged exposure (2 hpf – 98 hpf) to
28
diuron caused a significant increase in CES which is in contrary to the observed decrease in the late
exposure (50 hpf – 98 hpf). The possible explanation for the observed changes could be in the
different amounts of pesticide that could entered in zebrafish depending on different developmental
stage. In particular, due to the protection of chorion at the early stage, the transmission of pesticide
was limited and no significant effect could be observed. Still, the low amounts of diuron could enter
even in the embryo stage and the gradual increase of diuron intake in prolonged exposure possibly
triggers the stronger metabolizing processes and consequently increase in activity. On the other
hand, in the late exposure the absence of chorion led to a rapid intake of diuron which likely caused a
decrease in enzyme activity. The involved mechanisms for the observed effect are unknown, even
though the obtained results could indicate some form of an adaptive response. For example, one of
these responses is hormesis, a phenomenon by which adaptive responses to low doses of otherwise
harmful conditions improve the functional ability of organisms (Calabrese, 2008). Even though in this
study the characteristic low dose stimulation and high dose inhibition typical for hormetic effects
were not observed, differences in uptake of pesticide may have actually led to different exposure
concentrations in applied exposure scenarios. The phenomena of hormesis and toxicity thresholds
are likely related to activation of adaptive pathways responsible for cellular and physiological
homeostasis (Zhang et al., 2008) and it was shown that hormetic effect can be time dependent
(Zhang et al., 2013). This further confirms the possibility that similar adaptive mechanism caused
observed effects.
The last applied exposure scenario was the recovery exposure where 48 h exposure to the
pesticides (2 hpf – 50 hpf exposure) was followed by a removal of the pesticide and placement of
zebrafish in artificial water for an additional 48 h (50 hpf – 98 hpf recovery). Changes were observed
only after the exposure to diazinon where the CES activity was inhibited. Compared to the early
exposure (2 hpf – 50 hpf), it could be observed that the level of CES inhibition was lower in the
recovery exposure, and at the lowest concentration of diazinon the CES activity reached the control
level (whereas at the same concentration after the early exposure significant inhibition was
29
recorded). However, the CES activity was still inhibited at other diazinon concentrations, so only a
partial recovery of this enzyme activity occurred during the 48 h in artificial water. The slow recovery
of CES is in accordance with other studies which generally showed slow recovery of AChE and CES
activities (e.g. Coeurdassier et al., 2001; Collange et al., 2010).
Besides the enzymatic responses, the changes in gene expression were also assessed by
applying the early exposure (2 hpf – 50 hpf) and prolonged exposure (2 hpf – 98 hpf) scenario. In the
early exposure significant changes were observed for ces2, cyp1a, cat and gpx1a expression, whereas
after the prolonged exposure significant changes were observed for all measured target genes, i.e.
ache, ces2, cyp1a, gstp1, cat, gpx1a and gsr. Even though the dose responses were not observed,
recorded changes point to the involvement of this genes in the responses to exposure to investigated
pesticides. Regarding the mechanisms involved in the gene expression changes, induction of cyp1a is
dependent on the response of AhR (aryl hydrocarbon receptor) that has been consistently observed
in most species (Denison and Nagy, 2003). Dioxins and dioxin-like compounds (e.g., planar
halogenated aromatic hydrocarbons (HAH) and polycyclic aromatic hydrocarbons (PAH)) represent
ligands for the AhR. Additionally, it has been shown that there is a relatively large number of AhR
ligands whose structure and physiochemical characteristics are dramatically different than that of the
“classical” HAH and PAH ligands (Denison et al., 1998; Denison and Heath-Pagliuso, 1998). In our
study exposure to diuron caused significant induction in cyp1a expression and in the study of
Takeuchi et al. (2008) diuron showed AhR-mediated transcriptional activity in mouse cell lines. Some
of the same mechanisms of induction and inhibition of cytochrome p450 enzymes affect the
expression of carboxylesterases. The regulation appears to be influenced by the pregnane x receptor
(PXR) and constitutive androstane receptor (CAR) proteins, which upon activation, move to the
nucleus and bind to DNA response elements in promoters to induce the expression of many phase I
and phase II metabolizing enzymes (Staudinger et al., 2010). Results of ces2 expression revealed
disparity between the responses at the molecular and biochemical level. Specifically, both the early
and prolonged exposures to diazinon showed an inhibition of CES activity, whereas the gene
30
expression analysis showed an increase in ces2. On the other hand, after the prolonged exposure to
diuron both CES activity and ces2 expression were increased. Even though opposite effects were
observed, both results can be potentially explained. In particular, diazinon inhibits the CES activity
and therefore stimulates a higher expression of ces2 since there is a need for a higher amount of CES
to replace the inhibited enzyme. On the other hand, an exposure to diuron induces the detoxification
and stimulates the expression of ces2 which leads to an increase in CES activity. This is only one
possible explanation, but the obtained results point to the need for the assessment of responses at
different levels of biological organization, and it is advisable to include higher-level responses.
Further, similar results were observed in the case of cat and gpx1a expression. While CAT
and GPx activities remained unchanged after exposure to the investigated pesticides, the expression
of cat and gpx1a were significantly affected in both exposure scenarios. The changes in expression of
cat, gpx1a and gsr indicate the involvement of oxidative stress in the toxicity mechanism of these
pesticides. Even though there are no studies in zebrafish, it was shown that diazinon caused changes
in the activities of antioxidative enzymes in carp (Cyprinus carpio) (Oruc and Usta, 2007) and that
exposure of oysters (Crassostrea gigas) to diuron and exposure of goldfish (Carassius auratus) to a
mixture of herbicides including diuron led to an increased production of ROS and caused oxidative
stress (Fatima et al., 2007; Behrens et al., 2016). For the genes involved in antioxidative responses,
NF‐E2‐related factor 2 (Nrf2) regulates the expression of multiple antioxidant genes and
protects against oxidative damage triggered by contaminants exposure, injury and inflammation (e.g.
Osburn and Kensler, 2008). Thus Nrf2 plays a crucial role in the process of dealing with oxidative
stress and quenching of reactive oxygen species (ROS) derived from the metabolism of
environmental pollutants (Bao et al., 2017). It was shown that some redox-sensitive proteins, which
can undergo reversible oxidative and reductive reactions, can activate or delay downstream signaling
pathways, including antioxidant enzyme systems (Den Hertog et al., 2005; Hongjun et al., 2005). This
is a possible reason for observed changes in expression of measured antioxidant genes. Specifically,
the expression of gpx1a and gsr was increased, whereas the expression of cat was reduced after
31
early exposure scenario and increased after the prolonged exposure. The difference between the
responses of antioxidative enzymes is not uncommon and at the enzymatic level, the decrease in CAT
can be caused by high concentration of superoxide radicals which are known to inhibit CAT activity
(Kono and Fridovich, 1982). Regarding the observed down-regulation of cat and up-regulation of
gpx1a, similar trends were obtained also in other studies, e.g. after toxicity assessment of
chlorpyrifos in two different life stages of zebrafish (Jeon et al., 2016) and safflower Carthamus
tinctorius L. to zebrafish embryos/larvae (Xia et al., 2017).
The disparity between mRNA abundance and enzyme activity was observed also in the study
of Glanemann et al. (2003) who concluded that it is wrong to assume that responses at the mRNA
level reflect the response at the protein level or at the level of active enzyme, i.e. that there is a static
correlation between levels of mRNA, protein, and active enzyme. Moreover, few studies showed
evidence for the existence of disparity between transcriptome and proteome (Anderson and
Seilhamer, 1997; Gygi et al., 1999). The disparity should not be surprising since the level of gene
expression is determined by different factors and numerous factors may be responsible for the
observed discrepancy between the measured gene expression and enzyme activity levels. Inferring
enzymatic activity from transcriptome data is hindered by complexity of the relationship between
RNA levels and enzyme activities (e.g. differences in protein stability, post-translational
modifications, regulation of enzyme activity) (Glanemann et al., 2003). Similar results on the
discrepancies between the gene expression and enzyme activity were also observed in other studies
with zebrafish. For example, Craig et al. (2007) determined that in zebrafish exposed to cooper there
was no association between increased transcription and antioxidant enzyme activities. Also, in the
study of Jin et al. (2010) the mRNA induction patterns in zebrafish exposed to herbicide atrazine
were not in accordance with the changes in levels of antioxidant enzymes. This all suggests that some
enzymes are not controlled by transcriptional means, yet by other some other mechanisms at the
enzymatic level. As suggested by Jin et al. (2010) it is possible that mRNA levels represent a snapshot
of the cell activity at any given time, and the protein activity might be regulated at the post32
translational level. Therefore, the issue of disparity between enzyme activities and mRNA data should
be taken into account when assessing the toxic effects of pesticides (and other pollutants) at a
molecular level.
In conclusion, the obtained results show that both diazinon and diuron cause changes at a
biochemical and molecular level in zebrafish embryos and larvae. Moreover, the sensitivity of
different developmental stages of zebrafish differed notably and zebrafish larvae proved to be more
susceptible to pesticide exposure compared to the embryo stage which is probably due to a
protective role of the chorion. Since various factors (e.g. life stage, duration of the exposure, etc.)
affect the detoxification processes and final effects of pesticides, it is crucial to include various
aspects of exposure in the toxicity assessment. A comparison of enzyme responses revealed that CES
is more sensitive to organophosphate exposure than AChE which is valuable information for future
monitoring and risk assessment studies. Moreover, 48 h of recovery yielded only a partial recovery of
CES activity. Given the differences between the enzyme responses and gene expression changes, the
assessment of toxic effects should be conducted at different levels of biological organization in order
to obtain comprehensive data that will enable realistic prediction of final toxic effects of pesticides to
aquatic organisms.
Acknowledgments
The research stay of Mirna Velki at Department of Ecosystem Analysis, Institute for Environmental
Research, RWTH Aachen University was supported by Alexander von Humboldt Foundation
(Research Fellowship for postdoctoral researchers).
The DanioVision system was purchased thanks to the 2015 RWTH Dean’s Seed Fund grant to the
project “Behavioral consequences of neurotoxicity and potential adverse effects”. The authors thank
Tecan Group Ltd. and Nikon Instruments Germany for their contribution to this study as a partner of
the Students Lab "Fascinating Environment" at Aachen Biology and Biotechnology (ABBt).
33
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