water
Article
Interacting Effects of Polystyrene Microplastics and
the Antidepressant Amitriptyline on Early Life Stages
of Brown Trout (Salmo trutta f. fario)
Hannah Schmieg 1,* , Janne K.Y. Burmester 1, Stefanie Krais 1, Aki S. Ruhl 2,3, Selina Tisler 4,
Christian Zwiener 4, Heinz-R. Köhler 1and Rita Triebskorn 1,5
1Animal Physiological Ecology, Institute of Evolution and Ecology, University of Tübingen, Auf der
[email protected] (R.T.)
2Chair of Water Quality Control, Technische Universität Berlin, Sekr. KF 4, Str. des 17. Juni 135,
3German Environment Agency (UBA), Section II 3.1 (National and International Development of Drinking
Water Quality and Resource Protection), Schichauweg 58, D-12307 Berlin, Germany
4Environmental Analytical Chemistry, Center for Applied Geoscience, University of Tübingen,
[email protected] (C.Z.)
5Steinbeis Transfer Center for Ecotoxicology and Ecophysiology, Blumenstr. 13,
D-72108 Rottenburg, Germany
*Correspondence: [email protected]
Received: 21 July 2020; Accepted: 20 August 2020; Published: 22 August 2020
Abstract:
Whether microplastics themselves or their interactions with chemicals influence the health
and development of aquatic organisms has become a matter of scientific discussion. In aquatic
environments, several groups of chemicals are abundant in parallel to microplastics. The tricyclic
antidepressant amitriptyline is frequently prescribed, and residues of it are regularly found in surface
waters. In the present study, the influence of irregularly shaped polystyrene microplastics (<50
µ
m),
amitriptyline, and their mixture on early life-stages of brown trout were investigated. In a first
experiment, the impacts of 100, 10
4
, and 10
5
particles/L were studied from the fertilization of eggs until
one month after yolk-sac consumption. In a second experiment, eggs were exposed in eyed ova stages
to 10
5
, 10
6
particles/L, to amitriptyline (pulse-spiked, average 48
±
33
µ
g/L) or to two mixtures for two
months. Microplastics alone did neither influence the development of fish nor the oxidative stress
level or the acetylcholinesterase activity. Solely, a slight effect on the resting behavior of fry exposed
to 10
6
particles/L was observed. Amitriptyline exposure exerted a significant effect on development,
caused elevated acetylcholinesterase activity and inhibition of two carboxylesterases. Most obvious
was the severely altered swimming and resting behavior. However, effects of amitriptyline were not
modulated by microplastics.
Keywords:
microplastics; amitriptyline; brown trout; development; behavior; oxidative
stress; acetylcholinesterase
1. Introduction
Microplastic particles (MP) are detected worldwide from densely populated and rural areas to
remote regions [
1
–
4
]. The presence of MP has globally been reported for sediment, surface water
and even for air samples [
5
–
8
]. Representing the recent state of knowledge in freshwater systems,
Water 2020,12, 2361; doi:10.3390/w12092361 www.mdpi.com/journal/water
Water 2020,12, 2361 2 of 25
MP concentrations range from 0.00012 particles/L up to 2867 particles/L (according to [
9
]). However,
the potential risk for organisms and ecosystems caused by MP is still a matter of discussion. MP were
shown to be ingested and egested by fish [
10
–
12
] and small particles (mostly nanoplastics) can even
transfer into tissues [
13
–
16
]. Microplastic particles can injure organisms mechanically resulting in
inflammation and other histopathological effects in contact epithelia [
17
–
19
], disturb the energy
metabolism [
19
,
20
], and induce oxidative stress [
13
,
18
,
19
]. Early life stages of fish are considered as
very sensitive for pollutants [
21
]. In this context, high concentrations of MP have been shown to
reduce and delay hatching as well as negatively influence growth and heart rate of marine medaka
(Oryzias melastigma) [
11
]. Moreover, Malafaia, et al. [
22
] reported that exposure of zebrafish (Danio
rerio) to polyethylene (PE) MP reduced the hatching time and survival rates and led to morphological
changes. In contrast, LeMoine, et al. [
23
] found no effects of PE MP on hatching, mortality, and growth
rates of zebrafish. However, zebrafish exposed to MP exhibit transcriptomic changes as, for example,
downregulation of genes involved in the neural development. Mazurais, et al. [
12
] observed that a diet
that incorporated about 200 PE microbeads per day caused a slightly higher mortality rate in sea bass
larvae (Dicentrarchus labrax). Apart from that MP had only limited effects on the development of sea
bass larvae in the experiment [12].
The evaluation of the risk of MP for aquatic organisms in general is complex since different
polymer types with manifold additives in various sizes and shapes are present in the environment [
24
].
With our study we therefore solely address a selected aspect of MP aquatic ecotoxicology.
The topic is even more complex since not only MP themselves, but also their interaction with
chemicals have to be regarded. For example, polymerization solvents, residual monomers, plasticizers,
or other additives can leak from the particles and affect MP-exposed organisms [
25
,
26
]. In addition,
MP has the potential to ad- or absorb organic pollutants (reviewed by [
27
] and [
28
]). The sorption
can modulate the toxicity of the pollutants in different ways: If the particles are ingested and excreted
together with an adherent pollutant this would be without consequences for the organism. However,
the bioavailability of otherwise free pollutants may be reduced due to sorption what can led to less
negative effects in organisms [
11
,
29
,
30
]. On the other hand, pollutants ingested together with MP can
desorb in the digestive track, for example due to different pH conditions. In such cases, MP act as a
vector and adverse effects can be enhanced by the presence of MP [
31
–
33
]. Batel, et al. [
10
] showed that
MP and associated benzo[a]pyrene can also be transported along an artificial food web. Nevertheless,
the relevance of MP as vectors for organic pollutants in comparison to other exposure pathways in the
environment remains a matter of discussion [
34
–
36
]. Since the concentrations of persistent organic
pollutants in continental environments are expected to be higher than in marine ecosystems, sorption
of hydrophobic organic pollutants to MP might be especially important for freshwater ecosystems [
37
].
One group of chemicals commonly found in aquatic environments are
pharmaceuticals [38,39].
Residues of these or their metabolites enter surface waters mainly via wastewater treatment
plants [39,40].
Non-selective monoamine reuptake inhibitors more known as tricyclic antidepressants
are one of the oldest groups of pharmaceuticals to treat depression. Amitriptyline is the most prescribed
drug of this group [
41
–
43
]. Beside depression, amitriptyline is also used for migraine prophylaxis and
to treat chronic pain [
43
]. In comparison to other antidepressants, its mode of action is rather unspecific:
In addition to the inhibition of the reuptake of the neurotransmitters serotonin and noradrenaline,
amitriptyline acts as muscarinic acetylcholine receptor antagonist [
44
]. Furthermore, it has been shown
to bind to histamine receptors [
45
] as well as to neurotrophic tyrosine kinase A/B receptors resulting
in an upregulation of acetyl transferase and an influence on cell differentiation [
46
]. Amitriptyline is
mainly metabolized in the liver by cytochrome P450 [
47
]. An important metabolite is nortriptyline
which is also in use as an antidepressant itself [
41
,
47
,
48
]. Amitriptyline has been found in surface
waters around the world [
38
,
48
,
49
]. The highest concentration of 71.0 ng/L was reported by Baker and
Kasprzyk-Hordern [
38
] for a large river in the UK. Mean surface water concentrations are normally in
the low nanogram per liter range up to 22 ng/L [
49
–
51
]. Togola and Budzinski [
52
] reported that in
France residues of amitriptyline (1.4 ng/L) were even found in drinking water. Pharmaceuticals are
Water 2020,12, 2361 3 of 25
designed to be bioactive at low concentrations, and it can be, therefore, not excluded that they also
may affect non-target organisms at low environmental concentrations [39,40].
Demin, et al. [
53
] showed that amitriptyline increases the serotonin-triggered neurotransmission
in the brain of adult zebrafish in a dose-dependent way, with significantly higher 5-hydroxyindoleacetic
(5-HIAA)/serotonin ratios in fish exposed to 5 and 10 mg/L amitriptyline for a few hours. No effects
on the noradrenaline level were shown in this study. In contrast, Meshalkina, et al. [
54
] found a
significant reduction of the 5-HIAA/serotonin ratio and an increased dopamine and noradrenaline level
in the brain of adult zebrafish exposed for two weeks to 10 and 50
µ
g/L amitriptyline. Moreover, the
antidepressant was shown to alter the immune response in
in vitro
studies with primary macrophages
of common carp (Cyprinus carpio) as well as
in vivo
studies with zebrafish [
55
,
56
] and to affect the
oxidative stress level of both fish species [
56
–
58
]. Furthermore, amitriptyline was found to affect
the swimming behavior of zebrafish, by reducing for example the swimming activity as well as the
covered distance, and high concentrations of the antidepressant even led to side or vertical swimming
behavior [
53
,
54
,
59
,
60
]. In addition, exposure of common carp and zebrafish to the antidepressant
stimulated hatching and caused retarded development and malformations in common carp as well as
reduced body length of zebrafish larvae [
56
,
57
]. In contrast, high amitriptyline concentrations were
found to extend the time until hatch and to decrease the heart rate in zebrafish [
60
]. Most studies
about the ecotoxicity of amitriptyline were performed with model species, but ecotoxicological studies
with feral aquatic species are lacking. Brown trout are known to be a sensitive test organism and as
important predators are of ecological relevance [61,62].
In the present study, effects of polystyrene (PS) MP at an environmental relevant concentration of
100 particles/L and higher concentrations of 10
4
and 10
5
particles/L on the development of brown trout
were investigated. Fish were exposed for 182 days from freshly fertilized eggs until about one month
after the fry completed their yolk sac consumption. In a second experiment, eggs were exposed from
eyed ova stage until one week after yolk sac consumption. In addition to PS MP (10
5
and 10
6
particles/L)
fish were also exposed to amitriptyline and co-exposed to the mixtures of PS MP and the antidepressant.
The co-exposure allows to investigate a potential modulation of the effects of amitriptyline by PS MP.
In both experiments, the impact of exposure on the development and biomarkers for oxidative stress
(activity of superoxide dismutase (SOD) and the level of lipid peroxidation (LPO)) were analyzed.
Since the chorion of zebrafish has been reported to act as an effective protective barrier against carbon
nanotubes [
63
], we examined the structure of the chorion of brown trout in the first experiment by
means of scanning electron microscopy. In the second experiment, also the behavior of larvae and
endpoints for neurotoxicity (activity of acetylcholinesterase (AChE) and two carboxylesterases (CbE))
were investigated.
2. Materials and Methods
2.1. Test Organism
Eggs of brown trout (Salmo trutta f. fario) were obtained from a commercial fish breeder
(Forellenzucht Lohmühle, D-72275 Alpirsbach-Ehlenbogen, Germany). According to the EC Council
Directive, the breeding facility is listed as category 1, disease-free [
64
]. All experiments started directly
after purchase of the eggs: In experiment 1, eggs (in total 360 eggs) were exposed on the same day of
their fertilization, in experiment 2 (in total 540 eggs), exposure started 47 days post fertilization (dpf) in
the eyed ova stage.
2.2. Test Substances
In both experiments, transparent PS pellets (Polystyrol 158 K, BASF, Ludwigshafen, Germany,
density 1.05 g/mL) were cryo-milled (CryoMill, Retsch, Haan, Germany) according to the method
of Eitzen, et al. [
65
]. The resulting irregularly shaped particles were suspended in ultra-pure water
(without any surfactant), fractionated using a micro-sieve (polyamide monofilament) with nominal
Water 2020,12, 2361 4 of 25
mesh-size of 50
µ
m and the permeate was used as stock suspension. The particle concentration in the
stock suspensions were analyzed with a particle counter (SVSS, PAMAS, Rutesheim, Germany) by light
extinction in a laser-diode sensor (type HCB-LD-50/50). Exemplary particle numbers with the analyzed
size ranges are provided in Figure 1and in Table S1 in the supplement. The stock suspensions were
diluted with respective ratios to obtain the target particle concentrations for the exposure experiments.
Water 2020, 12, 2361 4 of 26
light extinction in a laser-diode sensor (type HCB-LD-50/50). Exemplary particle numbers with the
analyzed size ranges are provided in Figure 1 and in Table S1 in the supplement. The stock
suspensions were diluted with respective ratios to obtain the target particle concentrations for the
exposure experiments.
Figure 1. Size distribution of the used polystyrene microplastic particles (PS MP).
Amitriptyline hydrochloride was purchased from Sigma Aldrich (CAS Number: 549–18–8; Lot:
BCBV1175; molecular formula: C20H23N · HCl; purity ≥ 98%; molecular weight 313.86). Amitriptyline
hydrochloride in the used concentration is water soluble without adding organic solvents. For the
stock solutions, 8.5 mg/L amitriptyline hydrochloride were solved in bidestilled water. Bottles with
stock solutions were covered in aluminum foil to protect them from light. All further given
amitriptyline concentrations refer to pure amitriptyline not amitriptyline hydrochloride. The
predicted logP octanol-water coefficient (pH 7.4) of amitriptyline is 4.92 [66].
2.3. Exposure and Sampling of Brown Trout
In both experiments, each treatment was tested in triplicates in a semi-static three-block design.
Exposures took place in a thermostat-controlled chamber with a light/dark cycle of 10/14 h. Petri
dishes and aquaria were shaded from direct light. Aquaria were aerated with glass pipettes
connected via silicone tubes to compressed air. Test suspensions were prepared from defined PS MP
stock suspensions (56,240 particles/mL). Vessels containing the respective stock suspension were
rinsed four times to avoid loss of particles. After consumption of the yolk sacs, fish were fed daily
approximately 3% of their body weight with commercial fish feed (0.5 mm, Biomar, Brande,
Denmark). At the end of the experiments, brown trout were anesthetized and killed by an overdose
of tricaine methanesulfonate ((MS-222), 1 g/L, buffered with NaHCO3). Death was ensured by
severance of the spine. Length and weight of each fish were recorded. The level of LPO, the activity
of SOD and the activity of AChE and CbE had to be analyzed in different tissues, due to the small
size of the fish.
2.3.1. Experiment 1a: Exposure of Embryos and Sac-Fry Stages
The first part of experiment 1 was conducted according to the OECD guideline 212 for exposing
fish embryos and sac-fry stages to dissolved chemicals [67]. Freshly fertilized eggs (fertilization and
start of experiment 07 December 2016) were exposed to 0 particles/L (C1), 100 particles/L (MP1h), 104
particles/L (MP1tt) and 105 particles/L (MP1ht). This first part of the experiment 1 was performed in
glass Petri dishes containing 200 mL of the respective test suspension. To achieve the final
concentration, the stock suspension was diluted with aerated artificial water (294 mg/L CaCl2 × 2 H2O,
Figure 1. Size distribution of the used polystyrene microplastic particles (PS MP).
Amitriptyline hydrochloride was purchased from Sigma Aldrich (CAS Number: 549–18–8; Lot:
BCBV1175; molecular formula: C
20
H
23
N
·
HCl; purity
≥
98%; molecular weight 313.86). Amitriptyline
hydrochloride in the used concentration is water soluble without adding organic solvents. For the stock
solutions, 8.5 mg/L amitriptyline hydrochloride were solved in bidestilled water. Bottles with stock
solutions were covered in aluminum foil to protect them from light. All further given amitriptyline
concentrations refer to pure amitriptyline not amitriptyline hydrochloride. The predicted logP
octanol-water coefficient (pH 7.4) of amitriptyline is 4.92 [66].
2.3. Exposure and Sampling of Brown Trout
In both experiments, each treatment was tested in triplicates in a semi-static three-block design.
Exposures took place in a thermostat-controlled chamber with a light/dark cycle of 10/14 h. Petri dishes
and aquaria were shaded from direct light. Aquaria were aerated with glass pipettes connected via
silicone tubes to compressed air. Test suspensions were prepared from defined PS MP stock suspensions
(56,240 particles/mL). Vessels containing the respective stock suspension were rinsed four times to
avoid loss of particles. After consumption of the yolk sacs, fish were fed daily approximately 3% of
their body weight with commercial fish feed (0.5 mm, Biomar, Brande, Denmark). At the end of the
experiments, brown trout were anesthetized and killed by an overdose of tricaine methanesulfonate
((MS-222), 1 g/L, buffered with NaHCO
3
). Death was ensured by severance of the spine. Length and
weight of each fish were recorded. The level of LPO, the activity of SOD and the activity of AChE and
CbE had to be analyzed in different tissues, due to the small size of the fish.
2.3.1. Experiment 1a: Exposure of Embryos and Sac-Fry Stages
The first part of experiment 1 was conducted according to the OECD guideline 212 for exposing
fish embryos and sac-fry stages to dissolved chemicals [
67
]. Freshly fertilized eggs (fertilization and
start of experiment 07 December 2016) were exposed to 0 particles/L (C1), 100 particles/L (MP1
h
),
10
4
particles/L (MP1
tt
) and 10
5
particles/L (MP1
ht
). This first part of the experiment 1 was performed in
Water 2020,12, 2361 5 of 25
glass Petri dishes containing 200 mL of the respective test suspension. To achieve the final concentration,
the stock suspension was diluted with aerated artificial water (294 mg/L CaCl
2×
2 H
2
O, 123.25 mg/L
MgSO
4×
7 H
2
O, 64.75 mg/L NaHCO
3
, 5.75 mg/L KCl in pure water). In each Petri dish, 30 brown trout
eggs were exposed to the test suspensions (90 eggs per treatment). Until the eyed ova stage, the eggs
were kept in complete darkness. The temperature in the Petri dishes was 6.7
±
0.2
◦
C. To maintain good
water quality, 25 to 50% of the test suspensions were renewed every second day (detailed information
in the supplement Table S2). From day 138 dpf, filtered aerated tap water (iron filter, particle filter,
activated charcoal filter) was used to prepare the test suspensions to habituate the growing larvae
to the water used in the second part of the experiment. The first part of the experiment ended after
150 days when the fry had completely consumed their yolk sacs (07 December 2016–05 May 2017).
Investigated parameters were time of development until eyed ova stage, time until hatch, heart rate
93 dpf, and mortality (excluding unfertilized eggs). After the first part of the experiment, ten larvae
were sampled from each Petri dish. For determination of the LPO level, heads of the larvae were
immediately frozen in liquid nitrogen and stored at −80 ◦C until further usage.
2.3.2. Experiment 1b: Exposure of Fry
The second part of experiment 1 lasted 33 days until 182 dpf (05 May 2017–06 June 2017).
The remaining fry of experiment 1a were transferred into 12 L aquaria with 5 L of the corresponding
test suspensions (in total C1: n =57, MP1
h
: n =57, MP1
tt
: n =53 and MP1
ht
: n =58). PS MP
stock suspensions were diluted with filtered tap water. To ensure good water quality, half of the
test suspension was renewed twice a week. Water parameters were checked 177 dpf and at the end
of the experiment (average values: Temperature 6.38
±
0.45
◦
C, pH 8.5
±
0.1, oxygen concentration
12.71 ±0.22 mg/L,
oxygen saturation 107.33
±
1.60%; conductivity 493.33
±
7.30
µ
S/cm; see supplement
Table S3). Samples for analysis of SOD activity (muscle/kidney) as well as for determination of the
LPO level (head) were frozen in liquid nitrogen and stored at −80 ◦C.
2.3.3. Experiment 2
In experiment 2, embryos/larvae were exposed in total for 60 days from eyed ova stage (47 dpf)
until one week after yolk sac consumption (29 December 2017–26/27. February 2018). Exposure groups
included a control group (C2) and groups exposed to 10
5
particles/L (MP2
ht
), 10
6
particles/L (MP2
mio
),
pulse-spiked amitriptyline (AMI2, nominal concentration 300
µ
g/L, average concentration calculated
and given as follows), 10
5
particles/L+pulse-spiked amitriptyline (MIX2
ht
, nominal concentration
300
µ
g/L amitriptyline, average concentration calculated and given as follows), and 10
6
particles/L
+pulse-spiked amitriptyline (MIX2
mio
, nominal concentration 300
µ
g/L amitriptyline, average
concentration calculated and given as follows). Test media were prepared with filtered tap water.
Exposure took place in 12 L aquaria filled with 5 L of the corresponding test media. Per aquarium,
30 individuals were exposed (3
×
30 per treatment group). 2.5 L of the test media were exchanged on
average once a week (see Table S2 in the supplement). Water parameters were determined to control
water quality at the start, after 55 days of exposure and at the end of the experiment. The average
values were pH 8.3
±
0.2, temperature 7.06
±
0.20
◦
C, conductivity 430.83
±
17.24
µ
S/cm, oxygen
content 10.92
±
0.10 mg/L, oxygen saturation 95.06
±
0.79% (see supplement Table S4). Nitrite (NO
2−
)
values did not exceed 0.05 mg/L. The heart rate was counted 21 days after the start of the experiment.
Samples for LPO (head), SOD (muscle/kidney), AChE and CbE (muscle) were frozen in liquid nitrogen
and stored at −80 ◦C.
2.4. Chemical Analyses
At the start of the second experiment as well as prior and past a water exchange (17 January
2018) mixed samples of all three blocks of each treatment group (4 mL per aquarium 12 mL in total)
were taken and frozen at
−
20
◦
C until further analysis. The water concentrations of amitriptyline
were determined using LC-MS with a 1290 Infinity HPLC system (Agilent Technologies, Waldbronn,
Water 2020,12, 2361 6 of 25
Germany) and a triple quadrupole mass spectrometer (6490 iFunnel Triple Quadrupole LC/MS, Agilent
Technologies, Waldbronn, Germany) in ESI (+) mode. An Agilent Poroshell-120-EC-C18 (2.7
µ
m, 2.1
×
100 mm) column at a flow rate of 0.4 mL/min was used for separation, and column temperature was
maintained at 40
◦
C. Eluent A and B were water (+0.1% formic acid) and acetonitrile (+0.1% formic
acid), respectively. Gradient elution was used: 0–1 min 5% B, linear increase to 100% B within 7 min,
hold for 7 min at 100% B. After switching back to the starting conditions, reconditioning time of 3 min
was employed. Samples were kept in the autosampler at 10
◦
C, the injection volume was 10
µ
L. Samples
of the control experiments were measured undiluted and samples of the experiments pulse-spiked
with amitriptyline were measured after 50-times dilution. The detection limit of amitriptyline (mass
transitionm/z278.2
→
117.1)forundilutedsamples was 10ng/L(10
µ
Linjectionvolume). Furtherdetails
on operating parameters of the triple quadrupole are provided in Tables S5 and S6 in the supplement.
2.5. Development Parameters
Mortality, malformations, eye pigmentation (only experiment 1), and hatch were checked daily.
Coagulated eggs, dead fish and remains of chorions were removed. To determine the heart rate,
five animals of each Petri dish/aquaria were transferred into a Petri dish with fresh test medium.
The heart rate was counted under a stereo microscope for 20 s and water temperature was measured.
Subsequently, fish were placed back into the corresponding Petri dish/aquaria.
2.6. Scanning Electron Microscopy
After hatching in experiment 1, chorions were immediately fixed in 2% glutardialdehyde in 0.1 M
cacodylate buffer (pH 7.6) for several days. Specimens were rinsed three times with 0.1 M cacodylate
buffer and subsequently incubated in 1% osmium tetroxide overnight. The next day, chorions were
transferred in a graded series of ethanol for dehydration. Subsequently, samples were fixed to specimen
holder stubs and sputter-coated with gold. Analyses were conducted using a scanning electron
microscope EVO LS 10 (Zeiss, Jena, Germany).
2.7. Behavior
In the second experiment, resting behavior of fry was determined after 42 days of exposure.
In each tank, the positions of all fish were recorded (resting on the side/resting in ventral position).
Additionally, the swimming behavior under stressful conditions (bright illumination, no aeration)
was quantified at the end of the experiment (27 February 2018). For this, five fish per replicate were
transferred into small tanks (17 cm
×
17 cm
×
8.5 cm) filled with 0.5 L of the corresponding test medium.
Four of these tanks were measured simultaneously. Tanks were surrounded with white polystyrene
plates and indirectly illuminated with lamps (one lamp per tank, 2700 K, 1521 Lm per lamp) facing the
top plate. After a habitation period of 2 min, swimming behavior of fish was recorded for 18 min with
four cameras (Basler acA 1300–60 gm, 1.3 megapixels resolution, Basler AG, Ahrensburg, Germany, lens:
4.5–12.5 mm; 1:1.2; IR 1/2”) positioned 32 cm above the water surface of each tank. Fry were center-point
tracked individually, and total distance moved, mean velocity over time, time of no movement and
body contact were assessed using the EthoVision 12 XT software (Noldus Information Technology
bv, Wageningen, The Netherlands). Whenever the system exhibited difficulties in automatic tracking,
data were manually corrected for swaps between tracked individuals. After the video tracking fish
were sampled as described above.
2.8. Level of Lipid Peroxides
The degree of LPO was quantified with the ferrous oxidation xylenol orange (FOX) assay. The assay
was performed according to Hermes-Lima, et al. [
68
] and Monserrat, et al. [
69
], slightly modified for
96-well plates. In pre-tests, the dilution factor with methanol as well as sample volume and incubation
time were adjusted for optimal output (see Table S7). Frozen heads of fry were homogenized with
HPLC grade methanol. Samples were centrifuged (15,000 rcf, 5 min, 4
◦
C) and the supernatants were
Water 2020,12, 2361 7 of 25
stored at
−
80
◦
C. For the final assay, the following compounds were added to each well: 50
µ
L of
0.75 mM FeSO
4
-solution, 50
µ
L of 75 mM sulfuric acid, and 50
µ
L of 0.3 mM xylenol orange solution.
Subsequently, the corresponding sample volume was added. To be able to correct for potential Fe in
the samples, additionally, a sample blank in which the FeSO
4
-solution was replaced by bidistilled
water was performed. Bidistilled water was used to achieve a total volume of 200
µ
L in each well.
Well plates were incubated at room temperature and the absorbance at 570 nm (ABS570) was measured
in a photometer (Bio-Tek Instruments, Winooski, VT, USA). In a next step, 1
µ
L of 1 mM cumene
hydroperoxide solution (CHP) was added into each well and the plates were incubated for another
30 min (at room temperature). Afterwards, the absorbance of the samples with CHP was measured
at 570 nm. Data were related to the corresponding sample blanks. Each sample was analyzed in
triplicates. CHP equivalents were calculated according to the following equation:
CHPequiv.=ABS570
ABS570 CHP ×volume CHP (1µL)×total volume (200 µL)
sample volume ×dilution factor (1)
2.9. Activity of Superoxide Dismutase
Samples containing muscle and kidney tissue were rinsed in phosphate buffered saline (PBS;
pH 7.4) before they were frozen. Superoxide dismutase (Cu/Zn SOD, Mn SOD and Fe SOD) activity
was determined with a superoxide dismutase assay kit (item no. 706002, Cayman Chemical Company,
Ann Arbor, MI, USA). Samples were homogenized with 1:5 20 mM HEPES buffer (pH 7.2) and stored at
−
80
◦
C. Prior to the assay, samples were diluted 5:150 with TRIS buffer (50 mM TRIS-HCl, pH 8.0). In the
assay formazan dye is formed as a product of the reduction of tetrazolium salt by superoxide radicals
(generated by xanthine oxidase and hypoxanthine). SOD catalyzes the dismutation of the superoxide
anion to hydrogen peroxide and molecular oxygen. After incubation for 30 min, the absorbance at
450 nm (Bio-Tek Instruments, Winooski, VT, USA) was measured and the SOD activity was calculated.
All samples were analyzed in duplicates.
2.10. Neurotoxicity
Muscle tissue was homogenized in TRIS buffer (20 mM TRIS
base
, 20 mM NaCl, inhibitor mix,
pH 7.3) in a ratio of 1:5 and centrifuged (5000 rcf, 10 min, 4
◦
C). Subsequently, 50% glycerol (1/4 of the
volume of the supernatant) were added to the supernatant. Until final analysis, samples were stored at
−
20
◦
C. The Lowry method [
70
] modified by Markwell, et al. [
71
] was used to determine the total protein
content in the samples. The activity of acetylcholinesterase (AChE) was photometrically analyzed at
405 nm (Bio-Tek Instruments, Winooski, VT, USA) according to the method of
Ellman, et al. [72]
and
modified by Rault, et al. [
73
]. Additionally, the activity of two carboxylesterases (CbE) with 5 mM
4-nitrophenyl acetate (pnpa) and 5 mM 4-nitrophenyl valerate (pnpv) were determined according
to Sanchez-Hernandez, et al. [
74
]. All samples were analyzed in triplicates. Reported are specific
activities of the enzyme per mg of total protein content. One unit corresponds to one
µ
mol substrate
hydrolyzed per min.
2.11. Statistical Analysis
All analyses were performed with the software R 3.6.2. The
α
-level was set to 0.05. Developmental
time until eyed ova stage, time until hatch and mortality were analyzed with mixed effects Cox models
(package coxme) including the treatment as fixed effect and the Petri dish/tank as random effect to
consider potential position effects and influences among the fish in the Petri dish/tank. Post-hoc
comparisons with the control were performed with Dunnett’s test (experiment 1), and among all
groups with Tukey-HSD test (experiment 2). If necessary, data were transformed to achieve normal
distribution (see supplement Table S8). Length (experiment 1a) and CbE-pnpa (experiment 2) could
not be transformed to normal distributed data. In these cases, a Kruskal-Wallis test was performed.
Data for weight, AChE activity, CbE activity, SOD activity, LPO level, body contact, total distance,
Water 2020,12, 2361 8 of 25
and length (experiment 1b and experiment 2) were analyzed with a linear mixed model (package
lme4) including treatment as fixed effect and Petri dish/aquarium as random effect. Heart rate was
similarly analyzed but with temperature during measurement as additional random effect. For mean
velocity and no movement beside treatment, recording time was included in the model as fixed effect,
and again the aquarium as random effect. Resting behavior was analyzed with a likelihood-ratio test
followed by Fisher’s exact tests. The method of Benjamini and Hochberg [
75
] was used to correct for
multiple testing. All p-values not mentioned in the manuscript are given in Table S9 in the supplement.
2.12. Animal Welfare
The animal welfare committee of the regional council of Tübingen, Germany has approved the
experiments (authorization number ZO 2/16).
2.13. Credibility of Data
Details on the fulfillment of the criteria for reporting and evaluation of ecotoxicity of data (CRED)
proposed by Moermond, et al. [76] are provided in the supplement.
3. Results
Both experiments were considered valid as the survival of fish in the control groups as well as the
oxygen saturation were above 95%, and the difference in temperature between the aquaria in each test
was smaller than 1.5 ◦C. Prevalence of malformations was negligible in both tests.
3.1. Experiment 1a: Exposure of Embryos and Sac-Fry Stages
The fertilization rate of the eggs was 98%. Table 1summarizes the results of the first part of
experiment 1. Eye pigmentation started 37 dpf and was completed after 55 dpf. Over 99% of the larvae
hatched between 68 and 84 dpf. No significant differences in time until eyed ova stage and hatch
were found between the groups exposed to MP (MP1
h
, MP1
tt
, MP1
ht
) and the control group (eyed
ova:
d.f. =3,
n=352, X
2
=16.401, p<0.001, C1/MP1
h
p=0.977, C1/MP1
tt
and C1/MP1
ht
p=1; hatch
d.f. =3, n=349, X
2
=171.66, p<0.001, C1/MP1
h
p=0.230, C1/MP1
tt
p=0.999, C1/MP1
ht
p=0.814).
It became evident that the chorion of the eggs consists of several layers and exhibits no pores in the
micrometer range (Figure 2). The heart rate of brown trout larvae was not affected by the exposure to
PS MP (d.f. =3/6.649, F=3.971, p=0.066).
Table 1.
Summary of data for the investigated endpoints in experiment 1a. All data are given as
arithmetic means ±standard deviation.
Control
(C1)
100 Particles/L
(MP1h)
104Particles/L
(MP1tt)
105Particles/L
(MP1ht)
Mortality
(%)
1
±21
±21
±21
±2
Time until eyed ova stage
(dpf)
39
±139
±140
±340
±2
Time to hatch
(dpf)
75
±373
±373
±273
±3
Heart rate
(beats/min)
51
±256
±352
±253
±3
Length
(cm)
2.7
±0.2 2.7
±0.2 2.7
±0.2 2.8
±0.2
Body mass
(g)
0.15
±0.03 0.14
±0.02 0.14
±0.03 0.16
±0.02
Lipid peroxidation
(CHP-equiv.)
57.89
±10.20 60.12
±13.70 62.42
±12.17 61.23
±13.01
Water 2020,12, 2361 9 of 25
Water 2020, 12, 2361 8 of 26
random effect. For mean velocity and no movement beside treatment, recording time was included
in the model as fixed effect, and again the aquarium as random effect. Resting behavior was analyzed
with a likelihood-ratio test followed by Fisher`s exact tests. The method of Benjamini and Hochberg
[75] was used to correct for multiple testing. All p-values not mentioned in the manuscript are given
in Table S9 in the supplement.
2.12. Animal Welfare
The animal welfare committee of the regional council of Tübingen, Germany has approved the
experiments (authorization number ZO 2/16).
2.13. Credibility of Data
Details on the fulfillment of the criteria for reporting and evaluation of ecotoxicity of data
(CRED) proposed by Moermond, et al. [76] are provided in the supplement.
3. Results
Both experiments were considered valid as the survival of fish in the control groups as well as
the oxygen saturation were above 95%, and the difference in temperature between the aquaria in each
test was smaller than 1.5 °C. Prevalence of malformations was negligible in both tests.
3.1. Experiment 1a: Exposure of Embryos and Sac-Fry Stages
The fertilization rate of the eggs was 98%. Table 1 summarizes the results of the first part of
experiment 1. Eye pigmentation started 37 dpf and was completed after 55 dpf. Over 99% of the larvae
hatched between 68 and 84 dpf. No significant differences in time until eyed ova stage and hatch
were found between the groups exposed to MP (MP1h, MP1tt, MP1ht) and the control group (eyed
ova: d.f. = 3, n = 352, Χ2 = 16.401, p < 0.001, C1/MP1h p = 0.977, C1/MP1tt and C1/MP1ht p = 1; hatch d.f.
= 3, n = 349, Χ2 = 171.66, p < 0.001, C1/MP1h p = 0.230, C1/MP1tt p = 0.999, C1/MP1ht p = 0.814). It became
evident that the chorion of the eggs consists of several layers and exhibits no pores in the micrometer
range (Figure 2). The heart rate of brown trout larvae was not affected by the exposure to PS MP (d.f.
= 3/6.649, F = 3.971, p = 0.066).
Figure 2. SEM images of the chorion of a recently hatched brown trout. (a): Overview with
distinguishable layers at the opening. (b): Detailed view of the chorion’s surface. No pores in µm
range are present.
Table 1. Summary of data for the investigated endpoints in experiment 1a. All data are given as
arithmetic means ± standard deviation.
Control(chenchen
)
(C1)
100
Particles/L(chenchen)
(MP1h)
104
Particles/L(chenchen)
(MP1tt)
105
Particles/L(chenchen)
(MP1ht)
B
A
Figure 2.
SEM images of the chorion of a recently hatched brown trout. (
A
): Overview with
distinguishable layers at the opening. (
B)
: Detailed view of the chorion’s surface. No pores in
µ
m range
are present.
The mortality did not differ between the treatment groups and was below 2% (d.f. =3, n=354,
X2=0.0017,
p=1) in all groups. At 150 dpf, fish were 2.72
±
0.18 cm long and weighted
0.15 ±0.03 g
on average. No significant differences were measured in length and body mass compared to the control
group (length: n=120, d.f. =3, X
2
=4.857, p=0.183; body mass: d.f. =3/8, F=0.928, p=0.471).
The level of LPO was alike in all treatment groups (d.f. =3/8.072, F=0.928, p=0.471).
3.2. Experiment 1b: Exposure of Fry
No fish died during the second part of the experiment. The fish were 2.92
±
0.21 cm long and
weighted 0.19
±
0.04 g on average. MP had neither a significant effect on biometric values nor on the
oxidative stress level compared to the control (Table 2; length: d.f. =3/7.819, F=2.511, p=0.134; body
mass: d.f. =3/220, F=3.576, p=0.015, C1/MP1
h
:p=0.064, C1/MP1
tt
:p=0.729, C1/MP1
ht
:p=0.997;
SOD: d.f. =3/115, F=0.341, p=0.795; LPO: d.f. =3/116, F=2.904, p=0.038, C1/MP1
h
:p=0.355,
C1/MP1tt:p=0.286, C1/MP1ht:p=0.286).
Table 2.
Summary of data for the investigated endpoints in experiment 1b. All data are given as
arithmetic means ±standard deviation.
Control
(C1)
100 Particles/L
(MP1h)
104Particles/L
(MP1tt)
105Particles/L
(MP1ht)
Mortality
(%)
0
±00
±00
±00
±0
Length
(cm)
3.0
±0.1 2.9
±0.2 3.0
±0.2 2.9
±0.2
Body mass
(g)
0.20
±0.02 0.17
±0.04 0.20
±0.04 0.19
±0.04
Lipid peroxidation
(CHP-equiv.)
58.28
±15.46 65.46
±22.14 52.30
±18.77 58.57
±15.51
SOD
(U/mL)
95.43
±28.47 105.60
±42.84 96.15
±31.87 98.60
±30.56
3.3. Experiment 2
At every time point sampled, the concentration of amitriptyline in C2, MP2
ht
, MP2
mio
was below
the limit of detection (10 ng/L). The nominal concentration of amitriptyline in the groups AMI2, MIX2
ht
and MIX2
mio
was 300
µ
g/L. However, at the beginning of the experiment the real concentration was
only between 76% and 56% of the nominal concentration (Table 3).
Water 2020,12, 2361 10 of 25
Table 3.
Nominal and measured amitriptyline concentrations at the start of the experiment as well as
prior and after a water exchange.
Nominal
Concentration Measured Concentration
Start of
Experiment
Prior Water
Exchange
After Water
Exchange
C2 0µg/L<0.01 µg/L<0.01 µg/L<0.01 µg/L
MP2ht 0µg/L<0.01 µg/L<0.01 µg/L<0.01 µg/L
MP2mio 0µg/L<0.01 µg/L<0.01 µg/L<0.01 µg/L
AMI2 300 µg/L 214 µg/L 40 µg/L 82 µg/L
MIX2ht 300 µg/L 169 µg/L 35 µg/L 67 µg/L
MIX2mio 300 µg/L 229 µg/L 19 µg/L 68 µg/L
Prior to the water exchange, the measured concentration of amitriptyline was only between 6%
and 13% of the nominal concentration in all three groups. After the exchange of half of the media with
freshly prepared amitriptyline solutions, it was between 22% and 27% of the nominal concentration.
Due to the strong depletion of amitriptyline during the experiment, we modeled the concentration
fish were exposed to over time (Figure 3). Based on the measured concentrations, the removal of
amitriptyline from the aqueous phase was assumed as pseudo-first order:
y=c0×e−k×texp . (2)
where c
0
is the initial concentration and t
exp
are the days of exposure. To adjust the model to the
measured amitriptylineconcentrations, for kvalues between0.19 and0.29 were presumed. Accordingto
this model, the average amitriptyline concentration during the experiment was 48 ±33 µg/L.
Water 2020, 12, 2361 10 of 26
3.3. Experiment 2
At every time point sampled, the concentration of amitriptyline in C2, MP2ht, MP2mio was below
the limit of detection (10 ng/L). The nominal concentration of amitriptyline in the groups AMI2,
MIX2ht and MIX2mio was 300 µg/L. However, at the beginning of the experiment the real concentration
was only between 76% and 56% of the nominal concentration (Table 3).
Table 3. Nominal and measured amitriptyline concentrations at the start of the experiment as well as
prior and after a water exchange.
Nominal
Concentration
Measured Concentration
Start of
Experiment
Prior Water
Exchange
After Water
Exchange
C2
0 µg/L
< 0.01 µg/L
< 0.01 µg/L
< 0.01 µg/L
MP2ht
0 µg/L
< 0.01 µg/L
< 0.01 µg/L
< 0.01 µg/L
MP2mio
0 µg/L
< 0.01 µg/L
< 0.01 µg/L
< 0.01 µg/L
AMI2
300 µg/L
214 µg/L
40 µg/L
82 µg/L
MIX2ht
300 µg/L
169 µg/L
35 µg/L
67 µg/L
MIX2mio
300 µg/L
229 µg/L
19 µg/L
68 µg/L
Prior to the water exchange, the measured concentration of amitriptyline was only between 6%
and 13% of the nominal concentration in all three groups. After the exchange of half of the media
with freshly prepared amitriptyline solutions, it was between 22% and 27% of the nominal
concentration. Due to the strong depletion of amitriptyline during the experiment, we modeled the
concentration fish were exposed to over time (Figure 3). Based on the measured concentrations, the
removal of amitriptyline from the aqueous phase was assumed as pseudo-first order:
𝑦 = 𝑐0× 𝑒−𝑘×𝑡𝑒𝑥𝑝.
(2)
where c0 is the initial concentration and texp are the days of exposure. To adjust the model to the
measured amitriptyline concentrations, for k values between 0.19 and 0.29 were presumed.
According to this model, the average amitriptyline concentration during the experiment was 48 ± 33
µg/L.
Figure 3.
Modelled amitriptyline concentration over the duration of the experiment.
Measured concentrations of the three exposure groups (AMI2, MIX2
ht
, and MIX2
mio
) are depicted
as points.
Results of the second experiment are summarized in Table 4. No mortality occurred in the control
group (C2) and it was below 5% in all treatment groups (X2=0.3102, d.f. =5, p=0.9974).
Fish hatched on average 11 days after the start of the experiment (Figure 4). While no differences
were found between C2 and MP2
ht
as well as between C2 and MP2
mio
, all exposure groups with
amitriptyline (AMI2, MIX2
ht
and MIX2
mio
) hatched significantly earlier than the control group and the
Water 2020,12, 2361 11 of 25
two MP treatment groups (X
2
=164.6, d.f. =5, p<0.0001). Neither trout treated with MP nor those
exposed to amitriptyline or the mixture of both showed an influence on their heart rate (d.f. =5/7.5129,
F=0.2909, p=0.9044).
Table 4.
Summary of data for the investigated endpoints in experiment 2. All data are given as
arithmetic means
±
standard deviation. p-values in comparison to the control group are given if
significant differences occurred.
Control
(C2)
105Particles/L
(MP2ht)
106Particles/L
(MP2mio)
Amitriptyline
(AMI2)
Amitriptyline +
105Particles/L
(MIX2ht)
Amitriptyline
+106Particles/L
(MIX2mio)
Mortality
(%) 0±0 1 ±1 0 ±0 3 ±42
±14±3
Time to hatch
(days of exposure)
12 ±2 12 ±1
p=0.786 11 ±2
p=0.580
10 ±2
p<0.001
10 ±2
p<0.001
10 ±2
p<0.001
Heart rate
(beat /min) 56 ±4 55 ±2 56 ±2 56 ±1 54 ±4 55 ±2
Larvae resting on their
side (%) 10.0 ±11.9 5.6 ±4.2
p=0.405
1.1 ±1.6
p=0.018
97.8 ±3.1
p<0.001
98.9 ±1.6
p<0.001
100.0 ±0.0
p<0.001
Total distance moved (cm) 2135 ±862 1617 ±1008
p=0.385 1772 ±1103
p=0.756
224 ±95
p<0.001
253 ±136
p<0.001
345 ±165
p<0.001
Body contact (s) 75 ±31 77 ±49
p=156 ±17
p=0.999 217 ±137
p=0.082
241 ±137
p=0.023
222 ±63
p=0.066
Mean velocity (cm/s) 2.0 ±0.8 1.5 ±0.9
p=0.129 1.6 ±1.0
p=0.505
0.2 ±0.1
p<0.001
0.2 ±0.1
p<0.001
0.3 ±0.2
p<0.001
No movement (s) 579 ±207 706 ±243
p=0.628 668 ±282
p=0.902
1050 ±20
p<0.001
1049 ±28
p<0.001
1025 ±36
p<0.001
Length
(cm) 2.7 ±0.1 2.6 ±0.1
p=0.603 2.7 ±0.1
p=1
2.4 ±0.1
p<0.001
2.4 ±0.1
p<0.001
2.4 ±0.1
p<0.001
Body mass
(g) 0.14 ±0.03 0.14 ±0.02
p=10.15 ±0.02
p=0.963
0.11 ±0.02
p<0.001
0.11 ±0.02
p<0.001
0.12 ±0.02
p<0.001
Lipid peroxidation
(CHP-equiv.) 18.59 ±2.62 19.61 ±4.01 18.04 ±2.46 19.05 ±2.40 19.24 ±3.71 19.29 ±2.53
SOD
(U/mL) 117.54 ±28.89 122.62 ±28.11 121.55 ±24.58 137.88 ±33.58 133.66 ±29.16 129.09 ±31.48
AChE activity
(mu/mg protein) 52.46 ±12.66 54.48 ±13.04
p=0.997 53.69 ±12.39
p=0.999
68.29 ±13.59
p=0.008
68.53 ±14.70
p=0.008
68.87 ±14.78
p=0.008
CbE-pnpa activity
(mu/mg protein) 69.21 ±22.85 70.25 ±17.96
p=0.915 70.15 ±17.54
p=0.915
59.23 ±14.10
p=0.001
54.11 ±23.21
p=0.001
57.22 ±15.89
p=0.008
CbE-pnpv activity
(mu/mg protein) 70.28 ±24.04 66.15 ±21.79
p=0.859 70.43 ±21.51
p=1
39.12 ±26.11
p<0.001
38.27 ±27.86
p<0.001
31.33 ±23.59
p<0.001
Significant differences from the control group are highlighted in bold.
Water 2020, 12, 2361 12 of 26
CbE-pnpa
activity(chenche
n)
(mu/mg
protein)
69.21 ± 22.85
70.25 ±
17.96(chenche
n)
p = 0.915
70.15 ±
17.54(chen
chen)
p = 0.915
59.23 ±
14.10(chenchen)
p = 0.001
54.11 ±
23.21(chenche
n)
p = 0.001
57.22 ±
15.89(chenche
n)
p = 0.008
CbE-pnpv
activity(chenche
n)
(mu/mg
protein)
70.28 ± 24.04
66.15 ±
21.79(chenche
n)
p = 0.859
70.43 ±
21.51(chen
chen)
p = 1
39.12 ±
26.11(chenchen)
p < 0.001
38.27 ±
27.86(chenche
n)
p < 0.001
31.33 ±
23.59(chenche
n)
p < 0.001
Significant differences from the control group are highlighted in bold.
Fish hatched on average 11 days after the start of the experiment (Figure 4). While no differences
were found between C2 and MP2ht as well as between C2 and MP2mio, all exposure groups with
amitriptyline (AMI2, MIX2ht and MIX2mio) hatched significantly earlier than the control group and the
two MP treatment groups (Χ2 = 164.6, d.f. = 5, p < 0.0001). Neither trout treated with MP nor those
exposed to amitriptyline or the mixture of both showed an influence on their heart rate (d.f. = 5/7.5129,
F = 0.2909, p = 0.9044).
Figure 4. Percent of brown trout hatched in the different treatment groups at the different days after
exposure. All fish exposed to amitriptyline (AMI2, MIX2ht, MIX2mio) hatched significantly earlier than
the control group and the groups containing solely MP (MP2ht and MP2mio).
MP had no effect on body mass and length of the fry while fish exposed to AMI2, MIX2ht and
MIX2mio weighted significantly less and were significantly smaller than C2, MP2ht and MP2mio (Table
4; body mass: d.f. = 5/8.6811, F = 36.602, p < 0.0001; length: d.f. = 5/10.75, F = 65.262, p < 0.0001).
The resting behavior was influenced by the different treatments: Compared to C2, significantly
less fish exposed to MP2mio were resting on their side, while nearly all fish exposed to AMI2, MIX2ht
and MIX2mio showed this behavior (Figure 5; Χ2 = 604.081, d.f. = 5, p < 0.001). Furthermore, abnormal
swimming behavior was observed as larvae exposed to AMI2, MIX2ht and MIX2mio showed looping
behavior as well as side swimming.
Figure 4.
Percent of brown trout hatched in the different treatment groups at the different days after
exposure. All fish exposed to amitriptyline (AMI2, MIX2
ht
, MIX2
mio
) hatched significantly earlier than
the control group and the groups containing solely MP (MP2ht and MP2mio).
Water 2020,12, 2361 12 of 25
MP had no effect on body mass and length of the fry while fish exposed to AMI2, MIX2
ht
and
MIX2
mio
weighted significantly less and were significantly smaller than C2, MP2
ht
and MP2
mio
(Table 4;
body mass: d.f. =5/8.6811, F=36.602, p<0.0001; length: d.f. =5/10.75, F=65.262, p<0.0001).
The resting behavior was influenced by the different treatments: Compared to C2, significantly
less fish exposed to MP2mio were resting on their side, while nearly all fish exposed to AMI2, MIX2ht
and MIX2
mio
showed this behavior (Figure 5;X
2
=604.081, d.f. =5, p<0.001). Furthermore, abnormal
swimming behavior was observed as larvae exposed to AMI2, MIX2
ht
and MIX2
mio
showed looping
behavior as well as side swimming.
Water 2020, 12, 2361 13 of 26
Figure 5. Resting behavior of brown trout fry after 42 days after exposure. Compared to the control
significantly less fish exposed to MP2mio are resting on their side while significantly more fish exposed
to amitriptyline (AMI2) or the mixture of amitriptyline and MP (MIX2ht and MIX2mio) are resting on
their side (Χ2 = 604.0806, d.f. = 5, p < 0.0001, C2/MP2ht: p = 0.4048, C2/MP2mio: p = 0.01816, MP2ht /MP2mio:
p = 0.2108, C2/AMI2, C2/MIX2ht, C2/MIX2mio, MP2ht /MIX2ht and MP2mio/MIX2mio p < 0.001, AMI2/
MIX2ht and MIX2ht/MIX2mio: p = 1, AMI2/ MIX2mio: p = 0.4972). Different letters indicate significant
differences.
Video tracking revealed considerable differences in the behavior of fry (Figure 6, Table 4). For
all investigated parameters, fish exposed to MP2ht and MP2mio did not differ from the control group.
In contrast, fish exposed to AMI2, MIX2ht, and MIX2mio covered only 11–19% of the distance compared
to the control (Figure 6A). In addition, in the three groups containing amitriptyline, fry swam
significantly (84–90%) slower than the control fish (Figure 6B). In the control group and in MP2ht and
MP2mio, the mean velocity increased over the recording time. This effect did not occur in the groups
exposed to AMI2 or the mixtures. It is especially noticeable that the fry in the three groups containing
amitriptyline had more body contact. Compared to the control group, the time fry had body contact
to another fish was tripled in AMI2, MIX2ht, and MIX2mio (Figure 6C). Nevertheless, this behavior was
highly variable among the different aquaria. A significant difference was only found between C2 and
MIX2ht, MP2ht, and MIX2ht as well as MP2mio and MIX2mio. AMI2 and MIX2mio only showed a trend
towards more body contact than the fry in the control group. Furthermore, fry exposed to AMI2,
MIX2ht and MIX2mio spent significantly more time in inactivity than the fish in the control group and
the two exposure groups containing solely MP (Figure 6D; distance: d.f. = 5/84, F = 21.25, p < 0.001;
velocity: d.f. = 5/12, F = 35.290, p < 0.001, time: d.f. = 1/1601, F = 33.115 p < 0.001; no movement: d.f. =
5/12, F = 30.028, p < 0.001, time: d.f. = 1/2.0794 x 1021, F = 40.300, p < 0.001; body contact: d.f. = 5/12, F =
5.302, p = 0.008).
Figure 5.
Resting behavior of brown trout fry after 42 days after exposure. Compared to the control
significantly less fish exposed to MP2
mio
are resting on their side while significantly more fish exposed
to amitriptyline (AMI2) or the mixture of amitriptyline and MP (MIX2
ht
and MIX2
mio
) are resting on
their side (X
2
=604.0806, d.f. =5, p<0.0001, C2/MP2
ht
:p=0.4048, C2/MP2
mio
:p=0.01816, MP2
ht
/MP2
mio
:p=0.2108, C2/AMI2, C2/MIX2
ht
, C2/MIX2
mio
, MP2
ht
/MIX2
ht
and MP2
mio
/MIX2
mio
p<0.001,
AMI2/MIX2
ht
and MIX2
ht
/MIX2
mio
:p=1, AMI2/MIX2
mio
:p=0.4972). Different letters indicate
significant differences.
Video tracking revealed considerable differences in the behavior of fry (Figure 6, Table 4). For all
investigated parameters, fish exposed to MP2
ht
and MP2
mio
did not differ from the control group.
In contrast, fish exposed to AMI2, MIX2
ht
, and MIX2
mio
covered only 11–19% of the distance compared
to the control (Figure 6A). In addition, in the three groups containing amitriptyline, fry swam
significantly (84–90%) slower than the control fish (Figure 6B). In the control group and in MP2
ht
and MP2
mio
, the mean velocity increased over the recording time. This effect did not occur in the
groups exposed to AMI2 or the mixtures. It is especially noticeable that the fry in the three groups
containing amitriptyline had more body contact. Compared to the control group, the time fry had
body contact to another fish was tripled in AMI2, MIX2
ht
, and MIX2
mio
(Figure 6C). Nevertheless, this
behavior was highly variable among the different aquaria. A significant difference was only found
between C2 and MIX2
ht
, MP2
ht
, and MIX2
ht
as well as MP2
mio
and MIX2
mio
. AMI2 and MIX2
mio
only showed a trend towards more body contact than the fry in the control group. Furthermore, fry
exposed to AMI2, MIX2
ht
and MIX2
mio
spent significantly more time in inactivity than the fish in the
control group and the two exposure groups containing solely MP (Figure 6D; distance: d.f. =5/84,
F=21.25, p<0.001; velocity: d.f. =5/12, F=35.290, p<0.001, time: d.f. =1/1601, F=33.115 p<0.001;
Water 2020,12, 2361 13 of 25
no movement:
d.f. =5/12,
F=30.028, p<0.001, time: d.f. =1/2.0794 x 10
21
,F=40.300, p<0.001; body
contact: d.f. =5/12, F=5.302, p=0.008).
Water 2020, 12, 2361 14 of 26
Figure 6. Behavior of brown trout during video tracking (n = 15 in each group). (a): Total distance
moved in 18 min. (b): Mean velocity. (c): Total time individuals had body contact. (d): Time of no
movement over recording time. (a) and (c): The box plots display the median, the 25th and 75th
percentiles as well as minimum and maximum values (whiskers); the dots indicate outliers. Different
letters indicate significant differences.
No differences occurred in the activity of SOD and the level of LPO between all exposure groups
(SOD: d.f. = 5/173, F = 2.081, p = 0.070; LPO: d.f. = 5/12.287, F = 1.010, p = 0.452). The AChE activity was
increased about 30% in all treatment groups with amitriptyline (Figure 7a). Contrarily, CbE-pnpa
was reduced between 14% and 22% in the three exposure groups with the antidepressant. Outliers
with very low activity of CbE-pnpa were found in all groups but occurred cumulatively in AMI2,
MIX2ht, MIX2mio (Figure 7b). The activity of CbE-pnpv was even more reduced in the AMI2 and the
mixture exposure groups (Figure 7c). Compared to the control, the activity of CbE-pnpv was 44 - 55%
lower in AMI2, MIX2ht, MIX2mio (AChE: d.f. = 5/11.822, F = 6.081, p = 0.005; CbE-pnpa: d.f. = 5, Χ2 =
39.211, p < 0.001; CbE-pnpv: d.f. = 5/177, F = 22.135, p < 0.001).
Figure 6.
Behavior of brown trout during video tracking (n =15 in each group). (
A
): Total distance
moved in 18 min. (
B
): Mean velocity. (
C
): Total time individuals had body contact. (
D
): Time of
no movement over recording time. (
A
) and (
C
): The box plots display the median, the 25th and
75th percentiles as well as minimum and maximum values (whiskers); the dots indicate outliers.
Different letters indicate significant differences.
No differences occurred in the activity of SOD and the level of LPO between all exposure groups
(SOD: d.f. =5/173, F=2.081, p=0.070; LPO: d.f. =5/12.287, F=1.010, p=0.452). The AChE activity was
increased about 30% in all treatment groups with amitriptyline (Figure 7a). Contrarily, CbE-pnpa was
reduced between 14% and 22% in the three exposure groups with the antidepressant. Outliers with
very low activity of CbE-pnpa were found in all groups but occurred cumulatively in AMI2, MIX2
ht
,
MIX2
mio
(Figure 7b). The activity of CbE-pnpv was even more reduced in the AMI2 and the mixture
exposure groups (Figure 7c). Compared to the control, the activity of CbE-pnpv was 44 - 55% lower in
AMI2, MIX2
ht
, MIX2
mio
(AChE: d.f. =5/11.822, F=6.081, p=0.005; CbE-pnpa: d.f. =5, X
2
=39.211,
p<0.001; CbE-pnpv: d.f. =5/177, F=22.135, p<0.001).
Water 2020,12, 2361 14 of 25
Water 2020, 12, 2361 15 of 26
Figure 7. Enzyme activity in brown trout fry exposed to microplastic particles (MP), amitriptyline, or
the mixture of both. (a): Specific activity of acetylcholinesterase (AChE). (b): Specific activity of
carboxylesterases (CbE) with the substrate pnpa. (c): Specific activity of CbE with the substrate pnpv.
The box plots display the median, the 25th and 75th percentiles as well as minimum and maximum
values (whiskers); the dots indicate outliers. Different letters indicate significant differences.
4. Discussion
4.1. Amitriptyline Concentration
The measured amitriptyline concentrations were considerably lower than the nominal
concentration of 300 µg/L. Due to the strong depletion of the amitriptyline concentration, fish were
not constantly exposed to the same concentration. The concentration was considerably lower and
fluctuated due to the water exchange design. Amitriptyline was found to be one of the most stable
pharmaceuticals at pH 7 [77]. However, degradation of the antidepressant via photolysis occurs at
low or high pH [77]. In our experiment, the pH was 8.3 ± 0.2 and thereby slightly alkaline. At pH 9,
Baena-Nogueras, et al. [77] found a half-life of amitriptyline of 9.63 h (wavelength 300-800 nm,
irradiance 500 W/m2). In our experiment stock solutions were covered in aluminum foil and aquaria
were shaded from direct light which reduces the impact of photodegradation. Another factor that
likely influenced the concentration of amitriptyline is a possible sorption of the chemical to the glass
of the aquaria. Amitriptyline has a high adsorption capacity to kaolinite and Ca-montmorillonite
[78,79]. One day prior to the start of the experiment, the tanks were filled with the corresponding
amitriptyline solutions to saturate sorptive surfaces, and the test media were renewed before the start
of the experiment. Nonetheless, it cannot be excluded that, still, sorption of amitriptyline to the glass
occurred. A third process that has likely reduced the measured amitriptyline concentration is the
uptake and metabolization of the antidepressant by the fish. Amitriptyline was found to
bioconcentrate in brain tissue, gills, liver, blood plasma as well as in bile and muscle tissue of fish
with bioconcentration factors between 4 up to 198 [80–83]. Furthermore, in gilt-head bream (Sparus
aurata), amitriptyline was shown to be degraded to a broad range of metabolites including
nortriptyline which is known to be also a bioactive antidepressant [83]. In our experiment,
amitriptyline exposure led to severe effects in early life stages of brown trout which supports the
assumption that fish had taken up the antidepressant. To counteract the decline of amitriptyline
Figure 7.
Enzyme activity in brown trout fry exposed to microplastic particles (MP), amitriptyline,
or the mixture of both. (
A
): Specific activity of acetylcholinesterase (AChE). (
B
): Specific activity of
carboxylesterases (CbE) with the substrate pnpa. (
C
): Specific activity of CbE with the substrate pnpv.
The box plots display the median, the 25th and 75th percentiles as well as minimum and maximum
values (whiskers); the dots indicate outliers. Different letters indicate significant differences.
4. Discussion
4.1. Amitriptyline Concentration
The measured amitriptyline concentrations were considerably lower than the nominal
concentration of 300
µ
g/L. Due to the strong depletion of the amitriptyline concentration, fish were
not constantly exposed to the same concentration. The concentration was considerably lower and
fluctuated due to the water exchange design. Amitriptyline was found to be one of the most
stable pharmaceuticals at pH 7 [
77
]. However, degradation of the antidepressant via photolysis
occurs at low or high pH [
77
]. In our experiment, the pH was 8.3
±
0.2 and thereby slightly alkaline.
At pH 9,
Baena-Nogueras, et al. [77]
found a half-life of amitriptyline of 9.63 h (wavelength 300–800 nm,
irradiance 500 W/m
2
). In our experiment stock solutions were covered in aluminum foil and aquaria
were shaded from direct light which reduces the impact of photodegradation. Another factor that
likely influenced the concentration of amitriptyline is a possible sorption of the chemical to the glass of
the aquaria. Amitriptyline has a high adsorption capacity to kaolinite and Ca-montmorillonite [
78
,
79
].
One day prior to the start of the experiment, the tanks were filled with the corresponding amitriptyline
solutions to saturate sorptive surfaces, and the test media were renewed before the start of the
experiment. Nonetheless, it cannot be excluded that, still, sorption of amitriptyline to the glass
occurred. A third process that has likely reduced the measured amitriptyline concentration is the
uptake and metabolization of the antidepressant by the fish. Amitriptyline was found to bioconcentrate
in brain tissue, gills, liver, blood plasma as well as in bile and muscle tissue of fish with bioconcentration
factors between 4 up to 198 [
80
–
83
]. Furthermore, in gilt-head bream (Sparus aurata), amitriptyline
was shown to be degraded to a broad range of metabolites including nortriptyline which is known
to be also a bioactive antidepressant [
83
]. In our experiment, amitriptyline exposure led to severe
effects in early life stages of brown trout which supports the assumption that fish had taken up the
Water 2020,12, 2361 15 of 25
antidepressant. To counteract the decline of amitriptyline concentration, we performed as many water
exchanges with freshly prepared amitriptyline solution and/or cryo-milled PS MP (whose amount was
limited) as possible. However, the goal of our study, which was to show a possible modulation of
amitriptyline-induced effects in early life stages of brown trout, was not influenced by the unexpectedly
low amitriptyline concentrations, since even those induced strong reactions in the exposed fish.
4.2. Effects of MP
In the first experiment, brown trout eggs exposed to MP developed in a similar way as fish of the
control group. This might be due to an impeded passage of the tested MP through the egg chorion
which protects the first developmental stages of the fish embryo [
84
]. This explanation is highly
probable since our own REM images showed the chorion of brown trout eggs to be free of micropores >
1
µ
m. Similarly, Li, et al. [
11
] describes that 10
µ
m PS MP accumulated at the outside of the chorion of
marine medaka but failed to pass it. In a study with zebrafish, van Pomeren, et al. [
15
] found the uptake
of nanoplastics via chorion and epidermis to be marginal. Furthermore, LeMoine, et al. [
23
] observed
that MP could not pass the chorion of zebrafish and that the larvae started to ingest and accumulate
MP (10–45
µ
m) not earlier than 5 d post hatch, i.e., as soon as they start to feed on exogenous food.
Larvae of marine medaka ingested MP from two days post hatching [
11
]. In the second experiment
we therefore decided to expose the fish in a later developmental stage. In addition to the different
developmental stages at the start of the experiment, a slight difference in the temperature explains
the faster development and higher heart rate of the fish in the second experiment. Killeen, et al. [
85
]
investigated the influence of temperature on the development of brown trout in detail and described
that the development is generally delayed at lower temperatures.
In both experiments of our study, PS MP had no effect on hatching, growth, mortality, or heart rate.
LeMoine, et al. [
23
] and Chen, et al. [
29
] also observed that neither PE MP (10–45
µ
m, 480 particles/mL)
nor PS MP (45
µ
m, 1 mg/L) had any effect on mortality, hatching or growth rate of zebrafish embryos.
While PE MP (10–45
µ
m) incorporated into the diet of European sea bass likewise did not affect the
growth of the fish, high dosages (10
5
particles/g diet) of PE MP had a slight but significant effect on the
survival [
12
]. Furthermore, Malafaia, et al. [
22
] reported a trend to early hatching and lower survival
of zebrafish larvae exposed to PE MP (38.26
±
15.64
µ
m; up to 7.07
×
10
3
particles/L). In contrast,
Li, et al. [11]
observed that PE MP (6534.0
±
247.8 and 63,640.0
±
723.5 particles/L) delayed the hatching
and reduced the heart rate and growth of marine medaka. Overall, in most studies only minor effects
on mortality and developmental parameters were observed in fish exposed to MP. A plausible reason
for this might be that the chorion acts as protective barrier, and effects on this life stage are therefore
more likely caused by chemicals leaking from the MP than by the particles themselves.
The exposure of brown trout to PS MP during the development apparently did not lead to
oxidative stress as neither the activity of SOD nor the LPO level was influenced. Also, Chen, et al. [
29
]
did not observe any influence of PS MP on the activity of catalase (CAT) and glutathione peroxidase
(GPx) but they reported a significant decrease in the level of the reduced form of glutathione (GSH) in
zebrafish larvae. In juvenile brown trout, also no effect of PS MP on oxidative stress was found [
86
].
Furthermore, in common gobies (Pomatoschistus microps) the LPO level was not affected by PE MP
exposure in several studies [
33
,
87
–
89
]. Nevertheless, other studies showed that some MP may induce
oxidative stress in fish [
19
,
90
,
91
]. Thus, it should be further investigated which parameters of MP as for
example particle size, additives or the way of exposure can be responsible for MP-induced oxidative
stress in fish.
PS MP did neither affect the AChE activity nor the activity of two CbEs of brown trout. Similar
results were found for juvenile brown trout exposed to the same PS MP (10
4
particles/L) for 96 h [
86
].
Furthermore, PS MP (45
µ
m) did not cause alterations in the activity of AChE in zebrafish [
29
].
However, several studies showed that PE MP (1–5
µ
m) reduced the AChE activity in common gobies
significantly [
33
,
88
,
89
]. Moreover, 30 days exposure to 200
µ
g/L fluorescent PE (70–88
µ
m) led to a
significantly decreased AChE activity in the Amazonian discus fish (Symphysodon aequifasciatus) [
20
].
Water 2020,12, 2361 16 of 25
Considering the different polymer types and organisms investigated, the size of the administered
particles seems to be important as the most neurotoxic effects are reported for small MP or even
nanoplastics [13,29].
Exposure to 10
6
PS MP resulted in a minimal change in the resting behavior of trout larvae.
However, it seems rather unlikely that this slight change is biologically relevant especially when
considering the huge variation between the replicas. Furthermore, in the video tracking of brown
trout fry, no differences in swimming and shoaling behavior were observed in response to MP
exposure. Likewise, exposure of embryo and larvae of zebrafish to PE MP (10–45
µ
m) did not affect
the covered distance during darkness [
29
]. Exposure of Krefft’s frillgobies (Bathygobius kreffti) to PE
MP (38–45 mm) via diet did not affect their boldness or exploration behavior [
92
]. Moreover, Critchell
and Hoogenboom [
93
] observed that the feeding and aggression behavior of juvenile planktivorous
fish (Acanthochromis polyacanthus) was not affected by polyethylene terephthalate (PET) MP (<300
µ
m
and 2 mm). In addition, the foraging activity and survival facing predators of post-larvae surgeonfish
(Acanthurus triostegus) was not affected by PS MP (90
µ
m) exposure [
94
]. In contrast, juvenile black
rockfish (Sebastes schlegelii) stayed more closely together when exposed to about 10
6
particles/L PS MP
(15
µ
m) than the control group. Furthermore, black rockfish exposed to PS MP swam with reduced
speed and showed both a diminished explorative behavior during search for food and an increased
feeding time [
95
,
96
]. In European sea bass exposed to MP (0.69 mg/L; 1–5
µ
m) a significantly reduced
swimming velocity and resistance time (0.26 and 0.69 mg/L) until being dragged away by water flow
was observed [
97
]. Moreover, PE MP (1–5
µ
m and 420–500
µ
m) led, depending on the influence of the
environmental condition during development, to a significant reduction of the predatory performance
in early juvenile of common goby [
89
,
98
]. In two other studies investigating the effects of PE MP
(1–5
µ
m) in common gobies, a non-significant reduction of the post exposure predatory performance
was observed [
87
,
88
]. Likewise, the predatory performance of juvenile barramundi (Lates calcarifer)
was not influenced by PS MP (97
µ
m) but the fish showed more curved swimming paths than control
fish [99].
The factors that delimit effects of MP on the behavior of fish remain unclear. It does not seem that
chemical properties of the polymer have a major importance as, for example, after exposure to PE MP
and PS MP some studies found effects on the behavior of fish and some did not. It is possible that
different size ranges or additives of the particles are responsible for the reported differences. Moreover,
the discrepancies can be due to different sensitivity of the test species or the used study design.
In the present study, no biologically relevant effects of PS MP on the development and behavior of
brown trout were observed.
4.3. Effects of Amitriptyline
In our experiment the survival of fish was not affected by the exposure to amitriptyline.
However, the antidepressant led to a significantly reduced time until hatching. In common carp
Sehonova, et al. [57]
also observed a significant stimulation of hatching speed in fish exposed to 10,
100 and 500
µ
g/L amitriptyline. In addition, Yang, et al. [
56
] found that amitriptyline (1 ng up to
1 mg) reduced the time to hatch of zebrafish in a concentration-dependent manner. In contrast, higher
concentrations of 3 mg/L amitriptyline led to an increased hatching time and a higher mortality in
zebrafish [
60
]. This is, per se, not necessarily contradictory: In common carp, amitriptyline caused
either developmental stimulation or retardation depending on the developmental stage and the
concentration of the antidepressant [
57
]. Based on these data, amitriptyline seems to stimulate hatching
of fish in general but may also delay hatching at high concentrations. In the present study, no influence
was found on the heart rate of the larvae. In contrast, the considerably higher concentration of 3 mg/L
led to a reduced heart rate in zebrafish larvae [60].
After two months, brown trout larvae exposed to amitriptyline were significantly smaller and
weighted less than fish of the control group. Yang, et al. [
56
] reported that the body length of
zebrafish was significantly reduced after exposure to concentrations as low as 100 ng/L amitriptyline
Water 2020,12, 2361 17 of 25
for 120 h. Wu, et al. [
100
] found only a slight but insignificant reduction in body size of zebrafish and
a modulation of the expression of genes encoding early growth response factors after exposure to
0.1 µg/L amitriptyline for 120 h. Thus, amitriptyline affects the growth of different fish species which
can influence the survival of the larvae. Another reason for growth effects might be behavioral changes
that led to less food consumption.
Exposure to amitriptyline did not cause an alteration of the activity of SOD or the level of LPO in
early life stages of brown trout. Likewise, amitriptyline led to only moderate or minor effects on the
oxidative stress level of zebrafish during development: The transcription of glutathione S-transferase
(GST), GPx, and SOD genes in zebrafish embryos was not significantly altered in response to the
antidepressant [
58
]. Nonetheless, a significant upregulation of CAT mRNA formation in fish exposed to
30
µ
g/L amitriptyline was observed after 144 h [
58
]. In common carp, exposure to 10
µ
g/L amitriptyline
during development had no effect on the activity of CAT, GPx, GST and cytosolic SOD or the amount
of protein carbonylation [
57
]. However, the antidepressant led to an increase in the level of LPO and in
glutathione reductase activity [
57
]. Furthermore, Yang, et al. [
56
] have found a significantly positive
influence of 100 ng/L amitriptyline to the antioxidant capacity of zebrafish as the activities of both SOD
and CAT were enhanced and the formation of hydroxyl radicals and LPO was significantly suppressed.
Nevertheless, in fish treated with higher amitriptyline concentrations the activity of SOD, CAT and
peroxidase was found to be inhibited, and in zebrafish exposed to 1 mg/L amitriptyline the hydroxyl
radical formation and the LPO level were significantly increased [
56
]. Brown trout might be less
susceptible to oxidative stress caused by the tricyclic antidepressant than common carp and zebrafish
whereas the rather low amitriptyline concentrations applied in our study do not allow any predictions
about effects of higher concentrations that caused hydroxyl radical formation and an increase of LPO
level in zebrafish.
Amitriptylineisaneuroactivecompoundandhasbeenshowntoinduceneurotoxiceffectsinhuman
cell lines as well as in non-target organisms. In a neuroblastoma cell line 100
µ
M amitriptyline caused
a total loss of viability in neurons and a 30% loss of viability in astrocytes [
101
].
Sehonova, et al. [57]
observed neuronal dystrophy in common carp exposed to amitriptyline (lowest observed effect
concentration 10
µ
g/L). Moreover, mRNA expression of genes related to the development of eyes
and the central nervous system (pax 6) were significantly downregulated in zebrafish exposed to
30
µ
g/L amitriptyline [
60
]. In the present study, the antidepressant increased the activity of AChE and
significantly inhibited two other CbEs in early life stages of brown trout. The two CbE are involved in
the detoxification of pollutants and are assumed to act in a protective way against AChE inhibiting
pesticides [
74
]. An impact of amitriptyline on AChE activity is likely since amitriptyline was shown to
act as a muscarinic acetylcholine receptor antagonist [
44
]. In human serum and erythrocyte ghosts,
amitriptyline caused a decrease of the AChE activity [
102
]. However, to the best of our knowledge,
our study was the first to investigate the effects of amitriptyline on the activities of AChE, CbE-pnpv,
and CbE-pnpa in non-target organisms.
Considering its influence on neuronal processes and the general purpose of antidepressants to alter
behavior, it is not surprising that amitriptyline caused also behavioral changes in non-target organisms.
In our study fish showed altered resting and swimming behavior. Video tracking conditions in the
used artificial system are rather stressful for the fish [
103
], therefore reduced velocity and covered
distance can be interpreted as a result from an anxiolytic effect of amitriptyline. Even though it cannot
be excluded that freezing behavior and, therefore, anxiogenic behavior increased the time of inactivity,
it seems more probable that the effect was caused by the sedative effect of amitriptyline. Another
possible explanation is that amitriptyline interfered with neuronal processes resulting in ataxic behavior,
like looping or side swimming. Likewise, zebrafish exposed to 5 mg/L and 10 mg/L amitriptyline
exhibited a significantly reduced maximal swimming velocity. Furthermore, 10 mg/L amitriptyline
caused ataxic movement, like swimming vertically or on the side [
53
]. Moreover, exposure of adult
zebrafish to 50
µ
g/L, 1 and 5 mg/L amitriptyline significantly reduced the time until fish entered
the top in a novel tank test and increased the time they spent in the top region of the tank [
53
,
54
].
Water 2020,12, 2361 18 of 25
Meshalkina, et al. [
54
] also observed a reduction in the covered distance and mean velocity, and an
increased meandering and angular velocity. In addition, decreased swimming activity was observed
in zebrafish larvae treated with 100
µ
g/L amitriptyline in different exposure scenarios [
59
]. In another
study, Sehonova, et al. [
60
] observed that 300
µ
g/L amitriptyline resulted in a significantly decreased
swimming distance of zebrafish in the dark and, when exposed to 3000
µ
g/L amitriptyline, even in
both dark and light conditions. Common carp exposed to 100
µ
g/L and 500
µ
g/L amitriptyline just
floated apathetically and sank to the bottom of the aquarium [57].
Synoptically, our study revealed severe effects of amitriptyline on the behavior of brown trout
at concentrations in the
µ
g/L range (calculated average concentration of 48
µ
g/L amitriptyline).
Even though also other antidepressants like citalopram [
103
,
104
], fluoxetine [
105
,
106
] or
venlafaxine [
106
,
107
] exhibit an amitriptyline-analogous mode of action and have been shown
to cause similar effects in non-target organisms our study was not designed to draw conclusions on
the environmental risk of the antidepressant, since measured amitriptyline concentrations in surface
waters are largely in the range of 22 ng/L [
49
,
51
,
81
]. To evaluate the ecotoxicological risk on non-target
organisms, definitely further research on environmental relevant concentrations of amitriptyline as
single substance and in combination with other antidepressants would be necessary. The focus of our
study, however, was on the potential interactions of a psychoactive drug and MP.
4.4. Co-Exposure of MP and Amitriptyline
In our experiment brown trout were co-exposed to amitriptyline and PS MP. Like recently pointed
out by Heinrich, et al. [
108
] environmental pollutants are present both in liquid (water) and solid
(e.g., MP) compartments. Co-exposure allows to establish an equilibrium in the overall system
including water, MP, and organisms. Co-exposure of brown trout to MP plus amitriptyline led to the
same effects as the exposure to amitriptyline alone and the amitriptyline concentrations in the water
phase of the treatment groups showed no considerable difference. While the mortality and heart rate
as well as the oxidative stress level were not affected, larvae hatched significantly earlier, were smaller
and weighted less than control fish. Furthermore, the swimming behavior was altered, and the activity
of AChE was increased while the activities of the two other tested CbEs were inhibited. Therefore,
we did not find any indication for PS MP to modulate the effects of amitriptyline on the development
and behavior of brown trout. Several studies reported that MP can alleviate some negative effects
caused by other pollutants by reducing their bioavailability [
11
,
29
,
88
]. On the other hand, MP was
reported to increase negative effects of other contaminants [
32
,
33
,
109
]. In our experiments, measured
amitriptyline concentrations were only slightly lower in those exposure groups which also included MP.
Nevertheless, the data basis of the chemical analytic seems to be too limited to draw clear conclusions
about the sorption of amitriptyline to MP. Only few studies found, like the present study, no interaction
between MP and pollutants [
86
,
90
]. However, it cannot be excluded that this might have been resulting
from a bias against the publication of negative results [27].
5. Conclusions
In our study PS MP (<50
µ
m up to 10
6
particles per liter) did not influence the development of
brown trout. It is very likely that MP is not capable to pass the chorion so that the fish are protected until
hatch. Therefore, it might be reasonable to focus on life stages that have hatched and started to feed on
exogenous food. The only observed effect caused by MP in our experiment was a slight change in the
resting behavior that occurred at the highest tested (but environmentally irrelevant) concentration.
The biological relevance seems negligible considering the solely small change in behavior and the
high variability among the replicas. The antidepressant amitriptyline affected development and
behavior at considerably higher concentrations than reported in the environment. These effects were
not modulated by the co-exposure of the antidepressant with PS MP. Overall, no harmful effects were
caused by PS MP in the tested concentrations in brown trout and PS MP did not modulate the effects of
amitriptyline on fish. Nevertheless, these results do not allow to deduce a general statement about
Water 2020,12, 2361 19 of 25
the risk of MP considering the complexity caused by different polymers, size classes, and additives
reported in the environment as well as different sensitivities of affected organisms.
Supplementary Materials:
The following are available online at http://www.mdpi.com/2073-4441/12/9/2361/s1,
Table S1: Size ranges(in
µ
m)and counted particlenumbers(permL) of polystyreneparticlesin the stocksuspension.
Table S2: Volume of changed test medium in each Petri dish/tank during experiment 1 and experiment 2. Table S3:
Average measured water parameters of the experiment 1b. Table S4: Average measured water parameters of the
experiment 2. Table S5: Operating parameters of the triple quadrupole MS (Agilent 6490 QqQ) in positive mode.
Table S6: Specific measurement parameters for amitriptyline with LC-QqQ in water samples. Intraday variations
(RSD) is calculated with 1
µ
g/L standard (10
µ
L injection volume and 4 replicates (n)). Limit of quantification
=LOQ. Table S7: Parameters for the determination of the lipid peroxide content in the different experiments.
Table S8: Used data transformations for the statistical analysis. Table S9: Summary of all p-values of the single
comparisons of the different endpoints. CRED reporting. Raw data experiment 1. Raw data experiment 2.
Author Contributions:
Conceptualization: R.T., H.-R.K., and H.S.; methodology, H.S., S.K., S.T., A.S.R.; formal
analysis, H.S.; investigation, H.S., J.K.Y.B., and S.T.; resources, H.-R.K., R.T., C.Z., and A.S.R.; data curation, H.S.;
writing—original draft preparation, H.S. and S.T.; writing—review and editing, H.S., J.K.Y.B., S.K., A.S.R., S.T.,
C.Z., H.-R.K., and R.T.; visualization, H.S.; supervision, R.T., C.Z., and H.-R.K.; project administration, R.T.;
funding acquisition, R.T. All authors have read and agreed to the published version of the manuscript.
Funding:
This research was funded by the German Federal Ministry of Education and Research, Support Code:
02WRS1378. The experiments were conducted within the joint project MiWa (Microplastics in the water
cycle—sampling, sample preparation, analytics, occurrence, removal, and assessment).
Acknowledgments:
We wish to thank Martin Jekel for the initiation and coordination of the MiWa project. For the
scanning electron microscope analysis, we thank Monika Meinert and Oliver Betz. Furthermore, the authors
wish to thank Michael Ziegler, Stefanie Jacob, Katharina Peschke, Simon Schwarz, Carla Lorenz, Paul Thellmann,
Andreas Dieterich, Sabrina Wilhelm as well as Katharina Reitter, Lea Schuster and Aron Meral for help, discussion,
and technical assistance. We acknowledge support by Open Access Publishing Fund of University of Tübingen.
Conflicts of Interest:
The authors declare no conflict of interest. The funders had no role in the design of the
study; in the collection, analyses, or interpretation of data; in the writing of the manuscript, or in the decision to
publish the results.
References
1.
Allen, S.; Allen, D.; Phoenix, V.R.; Le Roux, G.; Dur
á
ntez Jim
é
nez, P.; Simonneau, A.; Binet, S.; Galop, D.
Atmospheric transport and deposition of microplastics in a remote mountain catchment. Nat. Geosci.
2019
,
12, 339–344. [CrossRef]
2.
Peeken, I.; Primpke, S.; Beyer, B.; Gütermann, J.; Katlein, C.; Krumpen, T.; Bergmann, M.; Hehemann, L.;
Gerdts, G. Arctic sea ice is an important temporal sink and means of transport for microplastic. Nat. Commun.
2018,9, 1–12. [CrossRef]
3.
Piehl, S.; Leibner, A.; Löder, M.G.J.; Dris, R.; Bogner, C.; Laforsch, C. Identification and quantification of
macro- and microplastics on an agricultural farmland. Sci. Rep. 2018,8, 17950. [CrossRef]
4.
Wang, W.; Ndungu, A.W.; Li, Z.; Wang, J. Microplastics pollution in inland freshwaters of China: A case
study in urban surface waters of Wuhan, China. Sci. Total Environ. 2017,575, 1369–1374. [CrossRef]
5.
Claessens, M.; Meester, S.D.; Landuyt, L.V.; Clerck, K.D.; Janssen, C.R. Occurrence and distribution of
microplastics in marine sediments along the Belgian coast. Mar. Pollut. Bull.
2011
,62, 2199–2204. [CrossRef]
6.
Dris, R.; Gasperi, J.; Mirande, C.; Mandin, C.; Guerrouache, M.; Langlois, V.; Tassin, B. A first overview
of textile fibers, including microplastics, in indoor and outdoor environments. Environ. Pollut.
2017
,
221, 453–458. [CrossRef]
7.
Leslie, H.A.; Brandsma, S.H.; Van Velzen, M.J.M.; Vethaak, A.D. Microplastics en route: Field measurements
in the Dutch river delta and Amsterdam canals, wastewater treatment plants, North Sea sediments and biota.
Environ. Int. 2017,101, 133–142. [CrossRef]
8.
Vianello, A.; Jensen, R.L.; Liu, L.; Vollertsen, J. Simulating human exposure to indoor airborne microplastics
using a Breathing Thermal Manikin. Sci. Rep. 2019,9, 8670. [CrossRef]
9.
Li, C.; Busquets, R.; Campos, L.C. Assessment of microplastics in freshwater systems: A review.
Sci. Total Environ. 2020,707, 135578. [CrossRef]
10. Batel, A.; Linti, F.; Scherer, M.; Erdinger, L.; Braunbeck, T. Transfer of benzo[a]pyrene from microplastics to
Artemia nauplii and further to zebrafish via a trophic food web experiment: CYP1A induction and visual
tracking of persistent organic pollutants. Environ. Toxicol. Chem. 2016,35, 1656–1666. [CrossRef]
Water 2020,12, 2361 20 of 25
11.
Li, Y.; Wang, J.; Yang, G.; Lu, L.; Zheng, Y.; Zhang, Q.; Zhang, X.; Tian, H.; Wang, W.; Ru, S. Low level
of polystyrene microplastics decreases early developmental toxicity of phenanthrene on marine medaka
(Oryzias melastigma). J. Hazard. Mater. 2020,385, 121586. [CrossRef] [PubMed]
12.
Mazurais, D.; Ernande, B.; Quazuguel, P.; Severe, A.; Huelvan, C.; Madec, L.; Mouchel, O.; Soudant, P.;
Robbens, J.; Huvet, A.; et al. Evaluation of the impact of polyethylene microbeads ingestion in European sea
bass (Dicentrarchus labrax) larvae. Mar. Environ. Res. 2015,112, 78–85. [CrossRef] [PubMed]
13.
Ding, J.; Zhang, S.; Razanajatovo, R.M.; Zou, H.; Zhu, W. Accumulation, tissue distribution, and biochemical
effects of polystyrene microplastics in the freshwater fish red tilapia (Oreochromis niloticus). Environ. Pollut.
2018,238, 1–9. [CrossRef] [PubMed]
14.
Mattsson, K.; Johnson, E.V.; Malmendal, A.; Linse, S.; Hansson, L.-A.; Cedervall, T. Brain damage and
behavioural disorders in fish induced by plastic nanoparticles delivered through the food chain. Sci. Rep.
2017,7, 11452. [CrossRef] [PubMed]
15.
Van Pomeren, M.; Brun, N.R.; Peijnenburg, W.J.G.M.; Vijver, M.G. Exploring uptake and biodistribution of
polystyrene (nano)particles in zebrafish embryos at different developmental stages. Aquat. Toxicol.
2017
,
190, 40–45. [CrossRef] [PubMed]
16.
Triebskorn, R.; Braunbeck, T.; Grummt, T.; Hanslik, L.; Huppertsberg, S.; Jekel, M.; Knepper, T.P.; Krais, S.;
Müller, Y.K.; Pittroff, M.; et al. Relevance of nano- and microplastics for freshwater ecosystems: A critical
review. Trac. Trends Anal. Chem. 2019,110, 375–392. [CrossRef]
17.
Karami, A.; Romano, N.; Galloway, T.; Hamzah, H. Virgin microplastics cause toxicity and modulate the
impacts of phenanthrene on biomarker responses in African catfish (Clarias gariepinus). Environ. Res.
2016
,
151, 58–70. [CrossRef] [PubMed]
18.
Lei, L.; Wu, S.; Lu, S.; Liu, M.; Song, Y.; Fu, Z.; Shi, H.; Raley-Susman, K.M.; He, D. Microplastic particles
cause intestinal damage and other adverse effects in zebrafish Danio rerio and nematode Caenorhabditis elegans.
Sci. Total Environ. 2018,619–620, 1–8. [CrossRef] [PubMed]
19.
Lu, Y.; Zhang, Y.; Deng, Y.; Jiang, W.; Zhao, Y.; Geng, J.; Ding, L.; Ren, H. Uptake and accumulation of
polystyrene microplastics in zebrafish (Danio rerio) and toxic effects in liver. Environ. Sci. Technol.
2016
,
50, 4054–4060. [CrossRef] [PubMed]
20.
Wen, B.; Zhang, N.; Jin, S.-R.; Chen, Z.-Z.; Gao, J.-Z.; Liu, Y.; Liu, H.-P.; Xu, Z. Microplastics have a more
profoundimpact than elevated temperatures on the predatory performance, digestion and energy metabolism
of an Amazonian cichlid. Aquat. Toxicol. 2018,195, 67–76. [CrossRef] [PubMed]
21.
Mohammed, A. Why are early life stages of aquatic organisms more sensitive to toxicants than adults?
In New Insights into Toxicity and Drug Testing; Gowder, S.J.T., Ed.; IntechOpen: London, UK, 2013; pp. 49–62.
[CrossRef]
22.
Malafaia, G.; De Souza, A.M.; Pereira, A.C.; Gonçalves, S.; Da Costa Ara
ú
jo, A.P.; Ribeiro, R.X.; Rocha, T.L.
Developmental toxicity in zebrafish exposed to polyethylene microplastics under static and semi-static
aquatic systems. Sci. Total Environ. 2020,700, 134867. [CrossRef] [PubMed]
23.
LeMoine, C.M.R.; Kelleher, B.M.; Lagarde, R.; Northam, C.; Elebute, O.O.; Cassone, B.J. Transcriptional
effects of polyethylene microplastics ingestion in developing zebrafish (Danio rerio). Environ. Pollut
2018
,
243, 591–600. [CrossRef] [PubMed]
24.
Wagner, M.; Scherer, C.; Alvarez-Muñoz, D.; Brennholt, N.; Bourrain, X.; Buchinger, S.; Fries, E.; Grosbois, C.;
Klasmeier, J.; Marti, T.; et al. Microplastics in freshwater ecosystems: What we know and what we need to
know. Environ. Sci. Eur. 2014,26, 12. [CrossRef] [PubMed]
25.
Lithner, D.; Damberg, J.; Dave, G.; Larsson, Å. Leachates from plastic consumer products—Screening for
toxicity with Daphnia magna.Chemosphere 2009,74, 1195–1200. [CrossRef] [PubMed]
26.
Schiavo, S.; Oliviero, M.; Romano, V.; Dumontet, S.; Manzo, S.; Liu, G. Ecotoxicological assessment of virgin
plastic pellet leachates in freshwater matrices. J. Environ. Account. Manag. 2018,6, 345–353. [CrossRef]
27.
De S
á
, L.C.; Oliveira, M.; Ribeiro, F.; Rocha, T.L.; Futter, M.N. Studies of the effects of microplastics on aquatic
organisms: What do we know and where should we focus our efforts in the future? Sci. Total Environ.
2018
,
645, 1029–1039. [CrossRef] [PubMed]
28.
Wang, F.; Wong, C.S.; Chen, D.; Lu, X.; Wang, F.; Zeng, E.Y. Interaction of toxic chemicals with microplastics:
A critical review. Water Res. 2018,139, 208–219. [CrossRef] [PubMed]
Water 2020,12, 2361 21 of 25
29.
Chen, Q.; Gundlach, M.; Yang, S.; Jiang, J.; Velki, M.; Yin, D.; Hollert, H. Quantitative investigation of the
mechanisms of microplastics and nanoplastics toward zebrafish larvae locomotor activity. Sci. Total Environ.
2017,584–585, 1022–1031. [CrossRef] [PubMed]
30.
Rehse, S.; Kloas, W.; Zarfl, C. Microplastics reduce short-term effects of environmental contaminants. Part
I: Effects of bisphenol A on freshwater zooplankton are lower in presence of polyamide particles. Int. J.
Environ. Res. Public Health 2018,15, 280. [CrossRef] [PubMed]
31.
Guilhermino, L.; Vieira, L.R.; Ribeiro, D.; Tavares, A.S.; Cardoso, V.; Alves, A.; Almeida, J.M. Uptake and
effects of the antimicrobial florfenicol, microplastics and their mixtures on freshwater exotic invasive bivalve
Corbicula fluminea.Sci. Total Environ. 2018,622–623, 1131–1142. [CrossRef] [PubMed]
32.
Nematdoost Haghi, B.; Banaee, M. Effects of micro-plastic particles on paraquat toxicity to common carp
(Cyprinus carpio): Biochemical changes. Int. J. Environ. Sci. Technol. 2017,14, 521–530. [CrossRef]
33.
Oliveira, M.; Ribeiro, A.; Hylland, K.; Guilhermino, L. Single and combined effects of microplastics and
pyrene on juveniles (0+group) of the common goby Pomatoschistus microps (Teleostei, Gobiidae). Ecol. Indic.
2013,34, 641–647. [CrossRef]
34.
Bakir, A.; O’Connor, I.A.; Rowland, S.J.; Hendriks, A.J.; Thompson, R.C. Relative importance of microplastics
as a pathway for the transfer of hydrophobic organic chemicals to marine life. Environ. Pollut.
2016
,
219, 56–65. [CrossRef] [PubMed]
35.
Beckingham, B.; Ghosh, U. Differential bioavailability of polychlorinated biphenyls associated with
environmental particles: Microplastic in comparison to wood, coal and biochar. Environ. Pollut.
2017
,
220, 150–158. [CrossRef] [PubMed]
36.
Koelmans, A.A.; Bakir, A.; Burton, G.A.; Janssen, C.R. Microplastic as a vector for chemicals in the aquatic
environment: Critical review and model-supported reinterpretation of empirical studies. Environ. Sci. Technol.
2016,50, 3315–3326. [CrossRef] [PubMed]
37.
Dris, R.; Imhof, H.; Sanchez, W.; Gasperi, J.; Galgani, F.; Tassin, B.; Laforsch, C. Beyond the ocean:
Contamination of freshwater ecosystems with (micro-) plastic particles. Environ. Chem.
2015
,12, 539–550.
[CrossRef]
38.
Baker, D.R.; Kasprzyk-Hordern, B. Spatial and temporal occurrence of pharmaceuticals and illicit drugs in
the aqueous environment and during wastewater treatment: New developments. Sci. Total Environ.
2013
,
454–455, 442–456. [CrossRef] [PubMed]
39.
Fent, K.; Weston, A.A.; Caminada, D. Ecotoxicology of human pharmaceuticals. Aquat. Toxicol.
2006
,
76, 122–159. [CrossRef] [PubMed]
40.
Sanderson, H.; Johnson, D.J.; Wilson, C.J.; Brain, R.A.; Solomon, K.R. Probabilistic hazard assessment of
environmentally occurring pharmaceuticals toxicity to fish, daphnids and algae by ECOSAR screening.
Toxicol. Lett. 2003,144, 383–395. [CrossRef]
41.
Breyer-Pfaff, U. The Metabolic Fate of Amitriptyline, Nortriptyline and Amitriptylinoxide in Man.
Drug Metab. Rev. 2004,36, 723–746. [CrossRef] [PubMed]
42.
Chockalingam, R.; Gott, B.M.; Conway, C.R. Tricyclic antidepressants and monoamine oxidase Inhibitors:
Are They Too Old for a New Look? In Antidepressants: From Biogenic Amines to Newe Mechanisms of Action;
Macaluso, M., Preskorn, S.H., Eds.; Springer: Cham, Switzerland, 2019; Volume 250.
43.
Schwabe, U.; Paffrath, D.; Ludwig, W.-D.; Klauber, J. Arzneiverordnungs-Report 2019–Aktuelle Daten, Kosten,
Trends und Kommentare; Springer: Berlin, Germany, 2019; p. 927.
44.
Snyder, S.H.; Yamamura, H.I. Antidepressants and the muscarinic acetylcholine receptor. Arch. Gen. Psychiatry
1977,34, 236–239. [CrossRef] [PubMed]
45.
Nguyen, T.; Shapiro, D.A.; George, S.R.; Setola, V.; Lee, D.K.; Cheng, R.; Rauser, L.; Lee, S.P.; Lynch, K.R.;
Roth, B.L.; et al. Discovery of a novel member of the histamine receptor family. Mol. Pharmacol.
2001
,59, 427.
[CrossRef] [PubMed]
46.
Jang, S.-W.; Liu, X.; Chan, C.-B.; Weinshenker, D.; Hall, R.A.; Xiao, G.; Ye, K. Amitriptyline is a TrkA and
TrkB receptor agonist that promotes TrkA/TrkB heterodimerization and has potent neurotrophic activity.
Chem. Biol. 2009,16, 644–656. [CrossRef] [PubMed]
47.
Rudorfer, M.V.; Potter, W.Z. Metabolism of tricyclic antidepressants. Cell. Mol. Neurobiol.
1999
,19, 373–409.
[CrossRef] [PubMed]
Water 2020,12, 2361 22 of 25
48.
Ma, L.-D.; Li, J.; Li, J.-J.; Liu, M.; Yan, D.-Z.; Shi, W.-Y.; Xu, G. Occurrence and source analysis of selected
antidepressants and their metabolites in municipal wastewater and receiving surface water. Environ. Sci.
Process. Impacts 2018,20, 1020–1029. [CrossRef] [PubMed]
49.
Ferrey, M.L.; Heiskary, S.; Grace, R.; Hamilton, M.C.; Lueck, A. Pharmaceuticals and other anthropogenic
tracers in surface water: A randomized survey of 50 Minnesota lakes. Environ. Toxicol. Chem.
2015
,
34, 2475–2488. [CrossRef] [PubMed]
50.
Lajeunesse, A.; Gagnon, C.; Sauv
é
, S. Determination of Basic Antidepressants and Their
N-Desmethyl Metabolites in Raw Sewage and Wastewater Using Solid-Phase Extraction and Liquid
Chromatography−Tandem Mass Spectrometry. Anal. Chem. 2008,80, 5325–5333. [CrossRef] [PubMed]
51.
Thomas, K.V.; Da Silva, F.M.A.; Langford, K.H.; De Souza, A.D.L.; Nizzeto, L.; Waichman, A.V. Screening for
selected human pharmaceuticals and cocaine in the urban streams of Manaus, Amazonas, Brazil. Jawra J.
Am. Water Resour. Assoc. 2014,50, 302–308. [CrossRef]
52.
Togola, A.; Budzinski, H. Multi-residue analysis of pharmaceutical compounds in aqueous samples.
J. Chromatogr. A 2008,1177, 150–158. [CrossRef] [PubMed]
53.
Demin, K.A.; Kolesnikova, T.O.; Khatsko, S.L.; Meshalkina, D.A.; Efimova, E.V.; Morzherin, Y.Y.; Kalueff, A.V.
Acute effects of amitriptyline on adult zebrafish: Potential relevance to antidepressant drug screening and
modeling human toxidromes. Neurotoxicol. Teratol. 2017,62, 27–33. [CrossRef] [PubMed]
54.
Meshalkina, D.A.; Kysil, E.V.; Antonova, K.A.; Demin, K.A.; Kolesnikova, T.O.; Khatsko, S.L.;
Gainetdinov, R.R.; Alekseeva, P.A.; Kalueff, A.V. The effects of chronic amitriptyline on zebrafish behavior
and monoamine neurochemistry. Neurochem. Res. 2018,43, 1191–1199. [CrossRef] [PubMed]
55.
Qiu, W.; Wu, M.; Liu, S.; Chen, B.; Pan, C.; Yang, M.; Wang, K.-J. Suppressive immunoregulatory effects of
three antidepressants via inhibition of the nuclear factor-
κ
B activation assessed using primary macrophages
of carp (Cyprinus carpio). Toxicol. Appl. Pharmacol. 2017,322, 1–8. [CrossRef] [PubMed]
56.
Yang, M.; Qiu, W.; Chen, J.; Zhan, J.; Pan, C.; Lei, X.; Wu, M. Growth inhibition and coordinated physiological
regulation of zebrafish (Danio rerio) embryos upon sublethal exposure to antidepressant amitriptyline.
Aquat. Toxicol. 2014,151, 68–76. [CrossRef] [PubMed]
57.
Sehonova, P.; Plhalova, L.; Blahova, J.; Doubkova, V.; Marsalek, P.; Prokes, M.; Tichy, F.; Skladana, M.;
Fiorino, E.; Mikula, P.; et al. Effects of selected tricyclic antidepressants on early-life stages of common carp
(Cyprinus carpio). Chemosphere 2017,185, 1072–1080. [CrossRef] [PubMed]
58.
Sehonova, P.; Zikova, A.; Blahova, J.; Svobodova, Z.; Chloupek, P.; Kloas, W. mRNA expression of antioxidant
andbiotransformationenzymesinzebrafish(Danio rerio)embryosafterexposuretothetricyclicantidepressant
amitriptyline. Chemosphere 2019,217, 516–521. [CrossRef] [PubMed]
59.
Huang, I.J.; Sirotkin, H.I.; McElroy, A.E. Varying the exposure period and duration of neuroactive
pharmaceuticals and their metabolites modulates effects on the visual motor response in zebrafish (Danio
rerio) larvae. Neurotoxicol. Teratol. 2019,72, 39–48. [CrossRef] [PubMed]
60.
Sehonova, P.; Hodkovicova, N.; Urbanova, M.; Örn, S.; Blahova, J.; Svobodova, Z.; Faldyna, M.; Chloupek, P.;
Briedikova, K.; Carlsson, G. Effects of antidepressants with different modes of action on early life stages of
fish and amphibians. Environ. Pollut. 2019,254, 112999. [CrossRef] [PubMed]
61.
Schmidt-Posthaus, H.; Bernet, D.; Wahli, T.; Burkhardt-Holm, P. Morphological organ alterations and
infectious diseases in brown trout Salmo trutta and rainbow trout Oncorhynchus mykiss exposed to polluted
river water. Dis. Aquat. Org. 2001,44, 161–170. [CrossRef] [PubMed]
62.
Klemetsen, A.; Amundsen, P.-A.; Dempson, J.B.; Jonsson, N.; Jonsson, B.; O’connell, M.; Mortensen, E.
Atlantic salmon Salmo salar L., brown trout Salmo trutta L. and Arctic charr Salvelinus alpinus (L.): A review
of aspects of their life histories. Ecol. Freshw. Fish. 2003,12, 1–59. [CrossRef]
63.
Cheng, J.; Flahaut, E.; Cheng, S.H. Effect of carbon nanotubes on developing zebrafish (Danio rerio) embryos.
Environ. Toxicol. Chem. 2007,26, 708–716. [CrossRef] [PubMed]
64. EU. Council Directive 2006/88/EC. Off. J. Eur. Union 2006,L 328, 14–56.
65.
Eitzen, L.; Paul, S.; Braun, U.; Altmann, K.; Jekel, M.; Ruhl, A.S. The challenge in preparing particle
suspensions for aquatic microplastic research. Environ. Res. 2019,168, 490–495. [CrossRef] [PubMed]
66.
Hansch, C.; Leo, A.; Hoekman, D. Exploring QSAR–Hydrophobic, Electronic, and Steric Constants; American
Chemical Society: Washington, DC, USA, 1995; p. 167.
67.
OECD. Test. No. 212: Fish, Short-term Toxicity Test on Embryo and Sac-fry Stages; OECD Publishing: Paris,
France, 1998. [CrossRef]
Water 2020,12, 2361 23 of 25
68.
Hermes-Lima, M.; Willmore, W.G.; Storey, K.B. Quantification of lipid peroxidation in tissue extracts based
on Fe(III)xylenol orange complex formation. Free Radic. Biol. Med. 1995,19, 271–280. [CrossRef]
69.
Monserrat, J.M.; Geracitano, L.A.; Pinho, G.L.L.; Vinagre, T.M.; Faleiros, M.; Alciati, J.C.; Bianchini, A.
Determination of Lipid Peroxides in Invertebrates Tissues Using the Fe(III) Xylenol Orange Complex
Formation. Arch. Environ. Contam. Toxicol. 2003,45, 177–183. [CrossRef] [PubMed]
70.
Lowry, O.H.; Rosebrough, N.J.; Farr, A.L.; Randall, R.J. Protein measurement with the Folin phenol reagent.
J. Biol. Chem. 1951,193, 265–275. [PubMed]
71.
Markwell, M.A.K.; Haas, S.M.; Bieber, L.L.; Tolbert, N.E. A modification of the Lowry procedure to simplify
protein determination in membrane and lipoprotein samples. Anal. Biochem. 1978,87, 206–210. [CrossRef]
72.
Ellman, G.L.; Courtney, K.D.; Andres, V.; Featherstone, R.M. A new and rapid colorimetric determination of
acetylcholinesterase activity. Biochem. Pharmacol. 1961,7, 88–95. [CrossRef]
73.
Rault, M.; Collange, B.; Mazzia, C.; Capowiez, Y. Dynamics of acetylcholinesterase activity recovery in two
earthworm species following exposure to ethyl-parathion. Soil Biol. Biochem.
2008
,40, 3086–3091. [CrossRef]
74.
Sanchez-Hernandez, J.C.; Mazzia, C.; Capowiez, Y.; Rault, M. Carboxylesterase activity in earthworm gut
contents: Potential (eco)toxicological implications. Comp. Biochem. Physiol. Part C Toxicol. Pharmacol.
2009
,
150, 503–511. [CrossRef] [PubMed]
75.
Benjamini, Y.; Hochberg, Y. Controlling the false discovery rate: A practical and powerful approach to
multiple testing. J. R. Stat. Soc. Ser. B 1995,57, 289–300. [CrossRef]
76.
Moermond, C.T.A.; Kase, R.; Korkaric, M.; Ågerstrand, M. CRED: Criteria for reporting and evaluating
ecotoxicity data. Environ. Toxicol. Chem. 2016,35, 1297–1309. [CrossRef] [PubMed]
77.
Baena-Nogueras, R.M.; Gonz
á
lez-Mazo, E.; Lara-Mart
í
n, P.A. Degradation kinetics of pharmaceuticals
and personal care products in surface waters: Photolysis vs biodegradation. Sci. Total Environ.
2017
,
590–591, 643–654. [CrossRef] [PubMed]
78.
Chang, P.-H.; Jiang, W.-T.; Li, Z.; Kuo, C.-Y.; Jean, J.-S.; Chen, W.-R.; Lv, G. Mechanism of amitriptyline
adsorption on Ca-montmorillonite (SAz-2). J. Hazard. Mater. 2014,277, 44–52. [CrossRef] [PubMed]
79.
Lv, G.; Stockwell, C.; Niles, J.; Minegar, S.; Li, Z.; Jiang, W.-T. Uptake and retention of amitriptyline by
kaolinite. J. Colloid Interface Sci. 2013,411, 198–203. [CrossRef] [PubMed]
80.
David, A.; Lange, A.; Tyler, C.R.; Hill, E.M. Concentrating mixtures of neuroactive pharmaceuticals and
altered neurotransmitter levels in the brain of fish exposed to a wastewater effluent. Sci. Total Environ.
2018
,
621, 782–790. [CrossRef] [PubMed]
81.
Lajeunesse, A.; Gagnon, C.; Gagn
é
, F.; Louis, S.; ˇ
Cejka, P.; Sauv
é
, S. Distribution of antidepressants and their
metabolites in brook trout exposed to municipal wastewaters before and after ozone treatment—Evidence of
biological effects. Chemosphere 2011,83, 564–571. [CrossRef] [PubMed]
82.
Muir, D.; Simmons, D.; Wang, X.; Peart, T.; Villella, M.; Miller, J.; Sherry, J. Bioaccumulation of pharmaceuticals
and personal care product chemicals in fish exposed to wastewater effluent in an urban wetland. Sci. Rep.
2017,7, 16999. [CrossRef] [PubMed]
83.
Ziarrusta, H.; Mijangos, L.; Izagirre, U.; Plassmann, M.M.; Benskin, J.P.; Anakabe, E.; Olivares, M.; Zuloaga, O.
Bioconcentration and biotransformation of amitriptyline in gilt-head bream. Environ. Sci. Technol.
2017
,
51, 2464–2471. [CrossRef] [PubMed]
84.
Grierson, J.P.; Neville, A.C. Helicoidal architecture of fish eggshell. Tissue Cell
1981
,13, 819–830. [CrossRef]
85.
Killeen, J.; McLay, H.A.; Johnston, I.A. Development in Salmo trutta at different temperatures, with a
quantitative scoring method for intraspecific comparisons. J. Fish Biol. 1999,55, 382–404. [CrossRef]
86.
Schmieg, H.; Huppertsberg, S.; Knepper, T.P.; Krais, S.; Reitter, K.; Rezbach, F.; Ruhl, A.S.; Köhler, H.-R.;
Triebskorn, R. Polystyrene microplastics do not affect juvenile brown trout (Salmo trutta f. fario) or modulate
effects of the pesticide methiocarb. Environ. Sci. Eur. 2020,32, 49. [CrossRef]
87.
Fonte, E.; Ferreira, P.; Guilhermino, L. Temperature rise and microplastics interact with the toxicity of the
antibiotic cefalexin to juveniles of the common goby (Pomatoschistus microps): Post-exposure predatory
behaviour, acetylcholinesterase activity and lipid peroxidation. Aquat. Toxicol.
2016
,180, 173–185. [CrossRef]
[PubMed]
88.
Lu
í
s, L.G.; Ferreira, P.; Fonte, E.; Oliveira, M.; Guilhermino, L. Does the presence of microplastics influence
the acute toxicity of chromium(VI) to early juveniles of the common goby (Pomatoschistus microps)? A study
with juveniles from two wild estuarine populations. Aquat. Toxicol.
2015
,164, 163–174. [CrossRef] [PubMed]
Water 2020,12, 2361 24 of 25
89.
Miranda, T.; Vieira,L.R.; Guilhermino, L.Neurotoxicity, behavior, and lethal effectsof cadmium, microplastics,
and their mixtures on Pomatoschistus microps juveniles from two wild populations exposed under laboratory
conditions—Implications to environmental and human risk assessment. Int. J. Environ. Res. Public Health
2019,16, 2857. [CrossRef] [PubMed]
90.
Ferreira, P.; Fonte, E.; Soares, M.E.; Carvalho, F.; Guilhermino, L. Effects of multi-stressors on juveniles of the
marine fish Pomatoschistus microps: Gold nanoparticles, microplastics and temperature. Aquat. Toxicol.
2016
,
170, 89–103. [CrossRef] [PubMed]
91.
Qiao, R.; Sheng, C.; Lu, Y.; Zhang, Y.; Ren, H.; Lemos, B. Microplastics induce intestinal inflammation,
oxidative stress, and disorders of metabolome and microbiome in zebrafish. Sci. Total Environ.
2019
,
662, 246–253. [CrossRef] [PubMed]
92.
Tosetto, L.; Williamson, J.E.; Brown, C. Trophic transfer of microplastics does not affect fish personality.
Anim. Behav. 2017,123, 159–167. [CrossRef]
93.
Critchell, K.; Hoogenboom, M.O. Effects of microplastic exposure on the body condition and behaviour of
planktivorous reef fish (Hoogen). PLoS ONE 2018,13, e0193308. [CrossRef] [PubMed]
94.
Jacob, H.; Gilson, A.; Lanct
ô
t, C.; Besson, M.; Metian, M.; Lecchini, D. No Effect of Polystyrene Microplastics
on Foraging Activity and Survival in a Post-larvae Coral-Reef Fish, Acanthurus triostegus.Bull. Environ.
Contam. Toxicol. 2019,102, 457–461. [CrossRef] [PubMed]
95.
Yin, L.; Chen, B.; Xia, B.; Shi, X.; Qu, K. Polystyrene microplastics alter the behavior, energy reserve and
nutritional composition of marine jacopever (Sebastes schlegelii). J. Hazard. Mater.
2018
,360, 97–105. [CrossRef]
[PubMed]
96.
Yin, L.; Liu, H.; Cui, H.; Chen, B.; Li, L.; Wu, F. Impacts of polystyrene microplastics on the behavior and
metabolism in a marine demersal teleost, black rockfish (Sebastes schlegelii). J. Hazard. Mater.
2019
,380, 120861.
[CrossRef] [PubMed]
97.
Barboza, L.G.A.; Vieira, L.R.; Guilhermino, L. Single and combined effects of microplastics and mercury on
juveniles of the European seabass (Dicentrarchus labrax): Changes in behavioural responses and reduction
of swimming velocity and resistance time. Environ. Pollut. 2018,236, 1014–1019. [CrossRef] [PubMed]
98.
De S
á
, L.C.; Lu
í
s, L.G.; Guilhermino, L. Effects of microplastics on juveniles of the common goby
(Pomatoschistus microps): Confusion with prey, reduction of the predatory performance and efficiency,
and possible influence of developmental conditions. Environ. Pollut.
2015
,196, 359–362. [CrossRef]
[PubMed]
99.
Guven, O.; Bach, L.; Munk, P.; Dinh, K.V.; Mariani, P.; Nielsen, T.G. Microplastic does not magnify the acute
effect of PAH pyrene on predatory performance of a tropical fish (Lates calcarifer). Aquat. Toxicol.
2018
,
198, 287–293. [CrossRef] [PubMed]
100.
Wu, M.; Liu, S.; Hu, L.; Qu, H.; Pan, C.; Lei, P.; Shen, Y.; Yang, M. Global transcriptomic analysis of zebrafish
in response to embryonic exposure to three antidepressants, amitriptyline, fluoxetine and mianserin.
Aquat. Toxicol. 2017,192, 274–283. [CrossRef] [PubMed]
101.
Mannerström, M.; Toimela, T.; Ylikomi, T.; Tähti, H. The combined use of human neural and liver cell lines
and mouse hepatocytes improves the predictability of the neurotoxicity of selected drugs. Toxicol. Lett.
2006
,
165, 195–202. [CrossRef] [PubMed]
102.
Müller, T.C.; Rocha, J.B.T.; Morsch, V.M.; Neis, R.T.; Schetinger, M.R.C. Antidepressants inhibit human
acetylcholinesterase and butyrylcholinesterase activity. Biochim. Biophys. Acta Mol. Basis Dis.
2002
,
1587, 92–98. [CrossRef]
103.
Ziegler, M.; Knoll, S.; Köhler, H.R.; Tisler, S.; Huhn, C.; Zwiener, C.; Triebskorn, R. Impact of the antidepressant
citalopram on the behaviour of two different life stages of brown trout. PeerJ
2020
,8, e8765. [CrossRef]
[PubMed]
104.
Kellner, M.; Porseryd, T.; Porsch-Hällström, I.; Borg, B.; Roufidou, C.; Ols
é
n, K.H. Developmental exposure
to the SSRI citalopram causes long-lasting behavioural effects in the three-spined stickleback (Gasterosteus
aculeatus). Ecotoxicology 2018,27, 12–22. [CrossRef] [PubMed]
105.
Meijide, F.J.; Da Cuña, R.H.; Prieto, J.P.; Dorelle, L.S.; Babay, P.A.; Lo Nostro, F.L. Effects of waterborne
exposure to the antidepressant fluoxetine on swimming, shoaling and anxiety behaviours of the mosquitofish
Gambusia holbrooki. Ecotoxicol. Environ. Saf. 2018,163, 646–655. [CrossRef] [PubMed]
Water 2020,12, 2361 25 of 25
106.
Painter,M.M.; Buerkley, M.A.; Julius,M.L.; Vajda,A.M.; Norris,D.O.; Barber, L.B.; Furlong,E.T.; Schultz, M.M.;
Schoenfuss, H.L. Antidepressants at environmentally relevant concentrations affect predator avoidance
behavior of larval fathead minnows (Pimephales promelas). Environ. Toxicol. Chem.
2009
,28, 2677–2684.
[CrossRef] [PubMed]
107.
Maulvault, A.L.; Santos, L.H.M.L.M.; Paula, J.R.; Camacho, C.; Pissarra, V.; Fogaça, F.; Barbosa, V.; Alves, R.;
Ferreira, P.P.; Barcel
ó
, D.; et al. Differential behavioural responses to venlafaxine exposure route, warming
and acidification in juvenile fish (Argyrosomus regius). Sci. Total Environ.
2018
,634, 1136–1147. [CrossRef]
[PubMed]
108.
Heinrich, P.; Hanslik, L.; Kämmer, N.; Braunbeck, T. The tox is in the detail: Technical fundamentals for
designing, performing, and interpreting experiments on toxicity of microplastics and associated substances.
Environ. Sci. Pollut. Res. 2020,27, 22292–22318. [CrossRef] [PubMed]
109.
Lu, K.; Qiao, R.; An, H.; Zhang, Y. Influence of microplastics on the accumulation and chronic toxic effects of
cadmium in zebrafish (Danio rerio). Chemosphere 2018,202, 514–520. [CrossRef] [PubMed]
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