Content uploaded by Andrea Binelli
Author content
All content in this area was uploaded by Andrea Binelli on Feb 16, 2017
Content may be subject to copyright.
This article appeared in a journal published by Elsevier. The attached
copy is furnished to the author for internal non-commercial research
and education use, including for instruction at the authors institution
and sharing with colleagues.
Other uses, including reproduction and distribution, or selling or
licensing copies, or posting to personal, institutional or third party
websites are prohibited.
In most cases authors are permitted to post their version of the
article (e.g. in Word or Tex form) to their personal website or
institutional repository. Authors requiring further information
regarding Elsevier’s archiving and manuscript policies are
encouraged to visit:
http://www.elsevier.com/copyright
Author's personal copy
Multi-biomarker approach for the evaluation of the cyto-genotoxicity
of paracetamol on the zebra mussel (Dreissena polymorpha)
Marco Parolini
*
, Andrea Binelli, Daniele Cogni, Alfredo Provini
Department of Biology, University of Milan, Via Celoria 26, 20133 Milan, Italy
article info
Article history:
Received 10 November 2009
Received in revised form 16 February 2010
Accepted 22 February 2010
Available online 15 March 2010
Keywords:
NSAIDs
Paracetamol
Mussels
Cytotoxicity
Genotoxicity
abstract
Paracetamol (PCM; N-(4-hydroxyphenyl)acetamide) is a widely used analgesic and antipyretic agent that
is utilized in human medicine. Its use is so widespread that it is constantly being introduced into global
water bodies where it reaches concentrations up to several
l
gL
1
. A battery of eight biomarkers was
applied in the freshwater bivalve Dreissena polymorpha in order to evaluate its potential sub-lethal effect.
Mussels were exposed for 96 h to increasing environmental concentrations (1, 5, 10 nM) of PCM. Cyto-
genotoxicity was determined in mussel hemocytes by the lysosomal membrane stability (Neutral Red
Retention Assay), the single cell gel electrophoresis (SCGE) assay, the micronucleus test (MN test) and
assessments of the apoptotic frequency (DNA diffusion assay). Moreover, in order to evaluate the prob-
able alterations to the mussels’ oxidative status, measurements of the activity of superoxide dismutase
(SOD), catalase (CAT), glutathione peroxidase (GPx) and the detoxifying enzyme glutathione S-transferase
(GST) were performed using the cytosolic fraction extracted from a pool of entire mussels. The biomarker
battery demonstrated moderate cyto-genotoxicity in zebra mussel hemocytes since no primary DNA
fragmentation was measured by the SCGE assay and only a slight increase in fixed DNA damage was reg-
istered by apoptotic and MN frequencies. Significant destabilization of the lysosomal membrane from
baseline levels was evident at 5 and 10 nM at the end of the exposures, as was a high induction capacity
of the activities of CAT and GST.
Ó2010 Elsevier Ltd. All rights reserved.
1. Introduction
Pharmaceutical compounds are an emerging class of environ-
mental pollutants that are extensively and increasingly being used
in human and veterinary medicine (Fent et al., 2006). Due to their
continuous production, consumption and often abuse, many stud-
ies have shown worldwide measurable concentrations of about
100 of these drugs in the aquatic environment in the high ng L
1
to low
l
gL
1
range (Halling-Sorensen et al., 1998; Daughton and
Ternes, 1999; Kümmerer, 2001, 2004; Heberer, 2002). Since phar-
maceuticals have physico-chemical characteristics that are similar
to those of harmful xenobiotics (Sanderson et al., 2004), they could
be potentially dangerous to aquatic organisms (Fent et al., 2006)
that are particularly important non-target species, as they are ex-
posed to contaminants over their whole life-span. At present, only
the acute toxicity of a few pharmaceuticals has been tested on
organisms belonging to different trophic levels (Canesi et al.,
2007; Choi et al., 2008; Haap et al., 2008; Quinn et al., 2008; Yang
et al., 2008), but these data are not suitable for an accurate risk
assessment since chronic effects are much more probable (Fent,
2003; Crane et al., 2006). However, the environmental levels of
some pharmaceuticals are lower than the acute effect concentra-
tions measured by these assays (Ferrari et al., 2004; Bottoni and Fi-
dente, 2005). Currently, studies on chronic toxicity in aquatic
organisms are increasing to encompass a variety of different aqua-
tic species (Huggett et al., 2002; Pascoe et al., 2003; Quinn et al.,
2008), although these data are completely lacking for many phar-
maceuticals (Carlsson et al., 2006).
Among the many classes of pharmaceuticals, the non-steroidal-
anti-inflammatory drugs (NSAIDs) are one of the most important
groups. NSAIDs inhibit the synthesis and release of prostaglandins
acting as non-selective inhibitors of the enzyme cyclooxygenase,
inhibiting both the cyclooxygenase-1 (COX-1) and the cyclooxy-
genase-2 (COX-2) isoenzymes (Gagné et al., 2005). With an annual
production of several kilotons (Cleuvers, 2004), NSAIDs are the
sixth most sold drugs worldwide (Langman, 1999). Additionally,
since some of these pharmaceuticals can be purchased without
medical prescription, their consumption could be even higher.
Due to the continuous and increasing application, as well as their
pharmacokinetic properties (half-life, urinary and fecal excretion,
metabolism, etc.), NSAIDs can reach detectable concentrations
both in sewage and in surface water (Cleuvers, 2004). Many
authors have reported levels of these drugs exceeding 1
l
gL
1
in
wastewaters and in the effluents of sewage treatment plants
(STP), while lower concentrations have been found in surface
0045-6535/$ - see front matter Ó2010 Elsevier Ltd. All rights reserved.
doi:10.1016/j.chemosphere.2010.02.053
*Corresponding author. Tel.: +39 02 50314729; fax: +39 02 50314713.
E-mail address: marco.parolini@unimi.it (M. Parolini).
Chemosphere 79 (2010) 489–498
Contents lists available at ScienceDirect
Chemosphere
journal homepage: www.elsevier.com/locate/chemosphere
Author's personal copy
waters (Stumpf et al., 1996; Ternes et al., 1998; Farré et al., 2001;
Heberer, 2002; Metcalfe et al., 2003).
Although the analgesic and antipyretic agent paracetamol
(PCM; N-(4-hydroxyphenyl)acetamide) does not possess a real
anti-inflammatory action, it is usually considered an NSAID in tox-
icology due to its very similar mode of action (Misra et al., 1990).
Since it is considered a safe drug at therapeutic doses, it can be pur-
chased as an over-the-counter preparation in most countries, and
it is currently the most widely used drug worldwide (An et al.,
2009). Due to the huge production and quantity of use, it is re-
ported as one of the most frequently detected pharmaceuticals to
be found in surface waters, wastewaters and drinking water. In a
survey of 139 US streams, Kolpin et al. (2002) detected PCM in
24% of tested samples at a median concentration of 0.11
l
gL
1
,
with a maximum detection level up to 10
l
gL
1
, which is in per-
fect agreement with the PEC (Predicted Environmental Concentra-
tion) calculated by Kim et al. (2007) for Korean waters
(16.5
l
gL
1
). In worldwide surface waters, the median concentra-
tion of this compound is 0.055 ± 0.051
l
gL
1
(Bound and Voulvou-
lis, 2006; Gros et al., 2006), while in raw wastewaters was detected
at a higher median concentration of 48 ± 75
l
gL
1
(Gros et al.,
2006; Han et al., 2006).
Due to its widespread presence in aquatic ecosystems, PCM is
one of the possible dangerous compounds for the entire aquatic
biocoenosis (Crane et al., 2006; Schulte-Oehlmann et al., 2007).
Although very few studies have been carried out to evaluate its
environmental risk, Henschel et al. (1997) classified this drug as
harmful to aquatic organisms on the basis of some ecotoxicological
tests in different biological models including bacteria, algae, Daph-
nia spp., and fish embryos. According to this assumption, on the ba-
sis of the Predicted No Effect Concentration (PNEC) value of
9.2
l
gL
1
,Kim et al. (2007) calculated a PCM hazard quotient that
correspond to 1.8, suggesting potential adverse ecological conse-
quences. In order to increase the knowledge about its possible envi-
ronmental effects, the purpose of this study was to assess the cyto-
genotoxicity of PCM on the freshwater bivalve zebra mussel (Dreis-
sena polymorpha) by a multi-biomarker approach. This mussel was
chosen as a biological model because mollusks plays an important
role in freshwater ecosystems and are particularly susceptible to
environmental stressors (Gagné et al., 2006). Additionally, previous
studies have revealed that D. polymorpha is an useful and sensible
organism capable of highlighting sub-lethal effects when exposed
to synthetic chemicals, like persistent organic pollutants (Riva
et al., 2007; Binelli et al., 2008b) and pharmaceuticals (Binelli
et al., 2008a, 2009). We measured the end-points of eight individual
biomarkers, whose integrated response can be helpful for detecting
the sub-lethal effects caused by PCM on zebra mussel specimens
exposed to three different environmental concentrations of this
drug. Its genotoxic potential was evaluated by the single cell gel
electrophoresis (SCGE) assay, the micronucleus test (MN test) and
the measure of apoptotic frequency (DNA Diffusion assay), while
the cytotoxicity was measured by a lysosomal membrane stability
test (NRRA – Neutral Red Retention Assay), a classical parameter
of generic cellular stress in bivalves (Lowe et al., 1995). Addition-
ally, we also measured the activity of three antioxidant phase I en-
zymes, catalase (CAT), superoxide dismutase (SOD) and glutathione
peroxidase (GPx), as well as the phase II detoxifying enzyme gluta-
thione S-transferase (GST) to study in-depth the toxicity of PCM and
to investigate its possible mechanism of action.
2. Materials and methods
Standard PCM (CAS number 103-90-2) was obtained from Sig-
ma–Aldrich (Steinheim, Germany), as well as all other chemicals
used for biomarker determination. Dimethylsulfoxide (DMSO;
CAS number 67-68-5; purity = 99.5%) was obtained from VWR
International (Milan, Italy).
2.1. Mussel acclimation and maintenance conditions
Several hundred specimens of D. polymorpha tied by the byssus
to the rocks were collected by a scuba diver at a depth of 4–6 m in
Lake Lugano (Northern Italy), which is considered a reference site
due to its low xenobiotic pollution (Binelli et al., 2005). The mus-
sels were rapidly transferred in laboratory in bags filled with lake
water, the rocks were rinsed and introduced into 100-L glass aqua-
ria filled with tap water, which was maintained at a natural photo-
period, constant temperature (20 ± 1 °C), pH (7.5) and oxygenation
(>90% of saturation). Bivalves were fed daily with an algae replace-
ment-substitute-enrichment medium (AlgaMac-2000
Ò
, Bio-Mar-
ine Inc., Hawthorne, USA), and the water was changed regularly
every two days for at least two weeks to gradually purify the mol-
lusks of the pollutants that had accumulated in their soft tissues.
Several specimens (n= 300 for each aquarium) with the same shell
length (about 20 mm) were chosen for in vivo tests, including a
control assay. They were gently cut from the rocks and placed on
glass sheets suspended in 15-L aquaria filled with tap and de-chlo-
rinated water (1:1 v/v) and maintained at the same conditions de-
scribed above. Only specimens able to re-form their byssus and
reattach themselves to the glass sheet were used in the experi-
ments. Mussels were used for the subsequent in vivo exposures
only when target biomarkers showed values comparable with
baseline levels previously checked.
2.2. Concentration choice
Doses to be tested under laboratory conditions must be care-
fully selected in order to give information with maximum utility
in the real world. Since the purpose of this study was to investigate
particular sub-lethal effects, and bearing in mind that no data on
the toxicity of PCM was available for D. polymorpha, drug doses
were chosen according to the concentration currently found in
freshwaters worldwide. We selected 1 nM (0.154
l
gL
1
), 5 nM
(0.75
l
gL
1
) and 10 nM (1.51
l
gL
1
) concentrations of PCM. The
first and second doses are comparable with most of the measured
environmental concentrations (MECs; Bound and Voulvoulis,
2006; Grujic
´et al., 2009). The highest one is comparable with the
maximum level revealed currently in aquatic ecosystems
(1.95
l
gL
1
;Kolpin et al., 2004), but much lower than the PEC
value calculated by Kim et al. (2007) for Korean water (10
l
gL
1
)
and by Stuer-Lauridsen et al. (2000) for the European Union
(65.4
l
gL
1
).
2.3. Exposure assays
Exposure assays were conducted in semi-static conditions. The
entire water volume was changed daily and PCM was added up to
the selected concentration. Given the hydrophilic nature of PCM
(log K
ow
= 0.46; Lorphensri et al., 2007), a PCM working solution
(10 mg L
1
) was prepared by using deionized water. Exact volumes
of this working solution were added daily to each aquarium, until
the desired exposure concentrations were reached. The complete
water and chemical changes should guarantee a constant solution
concentration of PCM over each 24-h period and prevent losses of
contaminant as well as the transformation of the parental com-
pound into its metabolites (Binelli et al., 2009). Mussels were fed
daily with AlgaMac-2000
Ò
, which was added 2 h before each water
and chemical change. Temperature, oxygenation and pH were
checked daily. Several specimens (n= 33) were collected each
day from the control and exposure aquaria to measure cyto- and
genotoxic biomarkers in the hemocytes, whose viability was
490 M. Parolini et al. /Chemosphere 79 (2010) 489–498
Author's personal copy
checked by the Trypan blue exclusion method. The entire soft tis-
sue of the other 25 specimens was frozen in liquid nitrogen and
maintained at 80 °C until the enzymatic activity was measured.
2.4. SCGE assay
The SCGE assay was basically performed according to the
alkaline (pH > 13) version of the assay developed by Singh
et al. (1988), with the subsequent optimizations for the zebra
mussel detailed by Buschini et al. (2003). A total of 100
l
Lof
hemolymph from ten specimens was withdrawn from the sinus
near to the posterior adductor muscle with a hypodermic syringe
containing 200
l
L of phosphate-buffered saline (PBS; pH 7.4).
Ten aliquots of 10
l
L of cell suspension mixed with 85
l
Lof
low-melting agarose (LMA-0.7%) in PBS (37 °C) were spread onto
ten coated slides (previously dipped in 1% normal-melting aga-
rose; NMA), then covered by a coverslip and kept at 4 °C for
40 min until the agarose layer was hardened. A third layer of
85
l
L was added to the slides in the same way. After agarose
solidification, the coverslip was removed and slides were placed
in a lysing solution (2.5 M NaCl, 100 mM Na
2
EDTA, 8 mM Tris–
HCl, 1% Triton X-100 and 10% DMSO, pH 10) in a Coplin jar at
4°C in the dark for 1 h. Alkaline DNA unwinding was carried
out for 5 min in a gel electrophoresis chamber filled with freshly
prepared buffer (1 mM Na
2
EDTA, 300 mM NaOH, pH 13) and
then in an ice-water bath (4 °C). Electrophoresis was then per-
formed at 0.78 V cm
1
(25 V) and 300 mA for 10 min. After the
electrophoresis, slides were washed with a neutralization buffer
(0.4 M Tris–HCl, pH 7.5) for 5 min and fixed in absolute ethanol
for other 5 min. After staining with DAPI (4
0
,6-diamidino-2-phen-
ylindole) DNA dye (Sigma–Aldrich), a coverslip was placed over
the slides. The observations were carried out under a fluores-
cence microscope (Leitz DMR, Germany) equipped with a FITC
filter. All steps were performed in the dark in order to minimize
additional UV-induced DNA damage. Positive controls were car-
ried out exposing hemocytes to H
2
O
2
to check the effectiveness
of the electrophoresis conditions. All samples were blindly coded
and evaluated; 50 cells for each slide were analyzed using an
image analysis system (Comet Score
Ò
) for a total of 500 analyzed
hemocytes per sample. Two DNA damage end-points were eval-
uated: the ratio between migration length and comet head
diameter (LDR) and the percentage of tail DNA. The first one
was chosen to represent DNA damage data, the second since
the working group on genetic toxicology testing from the 4th
IWGT (International Workshop on Genotoxicity Test Procedures)
recently agreed that the percentage of tail DNA is the measure
most linearly related to dose and the easiest to understand
(Kirkland et al., 2007).
2.5. DNA diffusion assay
The evaluation of the frequency of apoptotic cells was carried
out by the method described by Singh (2000), based on the same
protocol used for the SCGE assay. An aliquot of 10
l
L was shared
among five different slides for each sample and then processed
with the same method described above for the SCGE assay. The
only methodological difference was that, after lysing the hemo-
cytes, slides were subjected to 5 min of alkaline DNA unwinding
(pH 13) without the subsequent electrophoresis. Slides were then
washed in the neutralization buffer and fixed in absolute ethanol.
Finally, after DAPI dying, they were observed under a fluorescence
microscope. Two hundred cells per slide were analyzed for a total
of 1000 cells per sample. Necrotic cells were eliminated from the
count.
2.6. Micronucleus test
The MN test was performed according to the method developed
by Pavlica et al. (2000). Hemolymph (100
l
L) was withdrawn by
the sinus near the posterior adductor muscle with a syringe filled
with 100
l
L of PBS and ethylenediaminetetraacetic acid (EDTA,
10 mM) solution (1:1 v/v). The cell suspension was spread on a
slide and left for 15 min in a humidity chamber at room tempera-
ture to allow the hemocytes to settle. The hemocytes were then
fixed with glutaraldehyde (25% solution, diluted to 1% in PBS) for
5 min. After rinsing the excess fixative with PBS, the slides were
stained with bisbenzimide 33258 (Hoechst; CAS number 23491-
45-4) at a concentration of 1 mg mL
1
for 5 min, prior to being
washed and mounted in glycerol-McIlvane buffer (1:1 v/v). Slides
were kept in the dark at 4 °C until examination under a Leitz
DCM fluorescence microscope that was equipped with a sub-
merged lens at 100magnification. All samples were coded and
evaluated by a blinded observer. Four hundred cells were counted
for each slide, for a total of 4000 cells per sample. Only intact and
non-overlapping hemocyte nuclei were scored. Micronuclei were
identified by the criteria proposed by Kirsch-Volders et al.
(2000), and the MN frequency (MN‰) was calculated.
2.7. Neutral Red Retention Assay (NRRA)
The NRRA method followed the protocol proposed by Lowe and
Pipe (1994) and International Council for the Exploration of the Sea
(ICES, 2004). The Neutral Red stock solution was prepared by dis-
solving 20 mg of dye in 1 mL of dimethylsulfoxide (DMSO), while
the working solution was made by diluting 5
l
L of stock solution
in 2.5 mL of PBS. Microscope slides were previously coated with
2
l
L of polylysine with the help of a coverslip. Five slides were
used for each sample. Hemolymph was withdrawn using a hypo-
dermic syringe containing 100
l
L of PBS/EDTA solution (1:1 v/v),
as described above. The entire withdrawal was spread carefully
on each slide. Slides were suspended on a rack in a light-proof
humidity chamber for 20 min, and excess solution was carefully
tipped off. Lastly, 40
l
L of Neutral Red working solution was
added. After 20 min of incubation in the humidity chamber, slides
were observed under an optical microscope. Slides were examined
systematically thereafter at 15 min intervals to determine at what
point in time there was evidence of dye loss from the lysosomes to
the cytosol. Tests were terminated when dye loss was evident in at
least 50% of the hemocytes. The mean retention time (NRRT) was
then calculated from the five replicates.
2.8. Enzymatic activity
The enzymatic activities were measured in the entire organism
according to the observations made by Osman et al. (2007) and Os-
man and van Noort (2007) that CAT and GST activities in the whole
soft tissue were much higher than in a single gill. Enzymatic activ-
ities were determined spectrophometrically as described by Orbea
et al. (2002). Measurements were carried out in triplicate using the
cytosolic fraction extracted from a pool of 6–8 entire mussels (1g
fresh weight) homogenized in 100 nM phosphate buffer (pH 7.4;
KCl 100 mM, EDTA 1 mM) using a Potter homogenizer. Specific
protease inhibitors (1:10) were also added to the buffer: dithio-
threitol (DTT, 100 mM), phenanthroline (Phe, 10 mM) and trypsin
inhibitor (Try, 10 mg mL
1
). The homogenate was centrifuged at
500gfor 15 min at 4 °C. The supernatant was subsequently trans-
ferred into clean tubes and centrifuged again at 2000gfor 30 min
at 4 °C. Finally, the supernatant was ultra-centrifuged at
100 000gfor 90 min at 4 °C. The cytosolic fraction was held in ice
and immediately processed for the determination of protein and
enzymatic activities. The total protein content of all samples was
M. Parolini et al. /Chemosphere 79 (2010) 489–498 491
Author's personal copy
determined according to the Bradford method (1976) using bovine
serum albumin (BSA) as the standard. The activity of each enzyme
(CAT, SOD, GPx and GST) was measured in the cytosolic fraction.
CAT activity was determined by measuring the consumption of
H
2
O
2
at 240 nm using 50 mM of H
2
O
2
substrate in 80 mM potas-
sium phosphate buffer (pH 7). SOD activity was determined by
measuring the degree of inhibition of cytochrome c reduction at
550 nm by superoxide anion generated by the xanthine oxidase/
hypoxanthine reaction. The concentrations of the reagents used
during these reactions were as follows: potassium phosphate buf-
fer (50 mM, pH 7.8), hypoxanthine (50
l
M), xanthine oxidase
(1.87 mU mL
1
) and cytochrome c (10
l
M). The activity is given
in SOD units (1 SOD unit = 50% inhibition of the xanthine oxidase
reaction).
GPx activity was measured by monitoring the consumption of
NADPH at 340 nm during the formation of reduced glutathione
by glutathione reductase. The reaction medium consisted of the
following: 0.2 mM H
2
O
2
substrate in 100 mM potassium phos-
phate buffer (pH 7), containing additional glutathione (2 mM), so-
dium azide (NaN
3
; 0.5 mM), glutathione reductase (2 U mL
1
) and
NADPH (120
l
M). GST activity was measured by adding reduced
glutathione (20 mM) and 1-chloro-2,4-dinitrobenzene (CDNB) in
phosphate buffer (pH 7.4) to the cytosolic fraction. The resulting
reaction was monitored for 1 min at 340 nm.
2.9. Statistical analysis
Data normality and homoscedasticity were verified using the
Shapiro–Wilk and Levene’s tests, respectively. To identify dose/ef-
fect and time/effect relationships a two-way analysis of variance
(ANOVA) was performed using time and PCM concentrations as
variables, while biomarker end-points served as cases. The ANOVA
was followed by a Bonferroni post hoc test to evaluate eventual sig-
nificant differences (p< 0.05) between treated samples and related
controls (time to time) as well as among exposures. The Pearson’s
correlation test was carried out on all measured variables in the
three exposure assays to investigate possible correlations between
Fig. 1. Mean values (±SEM) of the SCGE assay expressed by the length/diameter ratio (LDR; a) and the mean of the mean percentages of DNA in tail (b). Significant differences
(two-way ANOVA, Bonferroni post hoc test,
p< 0.01) are referred to the comparison between treated mussels and the correspondent control (time to time).
492 M. Parolini et al. /Chemosphere 79 (2010) 489–498
Author's personal copy
various biological responses. All statistical analyses were per-
formed using the STATISTICA 7.0 software package.
3. Results
3.1. Baseline levels
No mortality or changes in hemocyte viability were recorded in
the control aquarium. In comparison to the corresponding control
value, however, significant viability differences (p< 0.01) were re-
corded in bivalves exposed to the higher PCM dose, beginning from
72 h of exposure (data not showed). The viability of the hemocytes
was always higher than 78%, according to recommendations made
by the 4th International Workshop on Genotoxicity Test Proce-
dures (IWGTP), which suggested a viability >70% for the overall
genotoxicity assays (Kirkland et al., 2007). All control data from
the tested cyto- and genotoxic biomarkers agreed those obtained
by our research group in previous studies (Riva et al., 2007; Binelli
et al., 2008a,b; Binelli et al., 2009). Moreover, also the baseline lev-
els obtained for enzymatic activities were similar to those obtained
in zebra mussel specimens by Osman and van Noort (2007), Binelli
et al. (2009) and Faria et al. (2009).
3.2. Cyto-genotoxicity assay results
PCM did not induce primary genetic damage in zebra mussel
hemocytes at each tested concentration, since no significant differ-
ences (p> 0.05) were noticed about the LDR values between con-
trols and PCM-treated specimens at each exposure time and dose
(Fig. 1A). Neither time-dependent (two-way ANOVA, Bonferroni
post hoc test; F= 0.79; p> 0.05) nor dose-dependent (F= 1.86;
p> 0.05) relationships were noticed for the LDR end-point. By con-
trast, considering the mean of the percentage of DNA in the tail,
both time/effect (F= 9.148; p< 0.01) and dose/effect (F= 9.166;
p< 0.01) relationships were evident (Fig. 1B). Moreover, we ob-
served an overall significant difference (p< 0.01) between each
PCM treatment and control. In addition, at the end of the exposure,
5 nM and 10 nM PCM increased significantly (p< 0.01) the levels of
the percentage of DNA in the tail.
According to the tail DNA data, the MN assay (Fig. 2) showed an
overall significant (p< 0.05) difference not only between each
treatment and the controls, but also between 10 nM and the other
two tested doses. Moreover, we noticed significant overall dose/ef-
fect (F= 20.48; p< 0.01) and time/effect (F= 19.32; p< 0.01) rela-
tionships. At the end of the exposure at 5 nM and 10 nM, the
data indicated a significant (p< 0.01) frequency of micronuclei
about 3.7 and 5 times higher than baseline levels, respectively.
By contrast, the measured apoptotic cell frequencies (Fig. 3)
showed neither a time/effect (F= 1.82; p> 0.05) nor a dose/effect
(F= 1.6; p>0.05) relationship and no differences (p>0.05) were
observed between each treatment and the controls or among each
PCM dose. The only significant difference (p< 0.05) in the apoptotic
frequency was noticed at the end of exposure at 10 nM, with about
a fourfold increase of apoptosis compared to the correspondent
baseline level.
Finally, data obtained by the NRRA (Fig. 4) showed that PCM
was able to induce a significant destabilization of lysosomal mem-
branes in an overall dose-dependent (F= 33.43; p< 0.01) and time-
dependent (F= 23.29; p< 0.01) manner. The statistical approach
highlighted an overall significant difference (p< 0.05) not only be-
tween each treatment and baseline levels, but also among each
other. A decreasing temporal trend was recorded at each dose
and a significant (p< 0.01) increase of cellular stress in bivalves
was noticed at the highest dose after only 48 h of exposure.
3.3. Enzymatic activity results
CAT, GPx and GST showed a very similar trend with a clear and
significant (p< 0.05) induction of their enzymatic activity starting
from 5 nM, while only SOD exhibited a more complex enzymatic
trend (Fig. 5). Notwithstanding, we notice a clear dose/effect
(two-way ANOVA, p< 0.01) and time/effect (p< 0.01) relationship
for each single enzyme. The CAT activity reached values about two-
fold higher than the correspondent controls at the end of the expo-
sure at the higher dose. Moreover, the higher treatment differed
significantly (p< 0.01) from the others. The GPx, like CAT, showed
a significant (p< 0.01) increase of 63% compared to controls after
96 h at 5 nM and after 24 h at 10 nM. Each dose differed signifi-
cantly from controls (p< 0.01) and 5 nM and 10 nM as well as from
the lowest dose (p< 0.01). Each single PCM dose induced signifi-
cant (p< 0.05) changes in the GST activity that exhibited an
increasing trend similar to CAT and GPx at 5 nM and 10 nM, with
significant differences (p< 0.01) compared to the correspondent
controls after 72 h of exposure at each concentration. By contrast,
SOD was characterized by a drastic increase (p< 0.01) in activity
after only 24 h at 1 nM, followed by a return to baseline levels.
The middle dose showed a similar behavior, but data were not sig-
nificantly different from controls. PCM showed a slight, but non-
significant increasing trend at 10 nM. Notwithstanding this high
variability of responses, each dose trend differed significantly from
the controls (p< 0.05), and the lowest concentration differed
(p< 0.01) also from 5 nM and 10 nM.
4. Discussion
The first approach to characterize the possible adverse effects
due to a specific chemical could be to utilize in vitro assays that
provide basic information concerning the toxicity of xenobiotics
and often help elucidate the probable mechanism of action of pol-
lutants. In previous in vitro screening studies on zebra mussel
hemocytes, we pointed out the high capability of PCM to induce
both cytotoxic and genotoxic damage (Parolini et al., 2009). Results
obtained by this previous research showed that PCM has a high
tendency to induce DNA damage, as highlighted by the significant
increase in LDR values and apoptotic cell frequency. Moreover,
data obtained by NRRA revealed that PCM was able to significantly
reduce the lysosomal membrane stability. These results, obtained
at administered doses (30, 150 and 450
l
gL
1
) much higher than
environmental levels, revealed the possible adverse effects that
PCM can cause to D. polymorpha but may not completely reflect
the real risk for the organism and consequently for the aquatic bio-
coenosis. The in vivo assays yield additional toxicological informa-
tion for the selected compound and help to thoroughly complete
and analyze the topic of the environmental risk assessment. In vivo
results gave quite different responses than in vitro experiments,
confirming that a tiered approach can give a more exhaustive pic-
ture of the pollutants’ toxicity (Hartmann et al., 2004). While the
in vitro assays showed clear primary DNA damage highlighted by
the Comet Test, PCM tested in vivo at environmental concentra-
tions was not able to produce significant DNA fragmentation, ex-
cept for two values obtained by the use of the percentage of DNA
in the tail, considered by the IWGTP as the measure that is most
linearly related to dose (Kirkland et al., 2007). The lack of correla-
tion (p> 0.05) between the Comet Test end-points, which evaluate
primary repairable genetic damage, and the other biomarkers of
genotoxicity (DNA diffusion assay and MN test) that detect fixed
DNA damage, seems to indicate that the PCM genotoxicity ap-
peared in D. polymorpha through great genetic injuries, such as
double strand breaks (dsb), translocations, and inversions, without
a previous increase in DNA fragmentation. Repairing systems are
M. Parolini et al. /Chemosphere 79 (2010) 489–498 493
Author's personal copy
capable of repairing minor DNA damage caused by PCM, while
hemocytes activate the apoptotic processes or produce micronuclei
only when broad genetic damage is evident. As suggested by Binelli
et al. (2009) for the exposure of zebra mussel specimens to the
antibacterial trimethoprim, the significant increase of micronuclei
(p< 0.01) observed at the end of the exposure at 5 nM and 10 nM
(Fig. 2) should indicate that one of the mechanism of PCM cytotox-
icity involves the aneuploidogen pathway, instead of clastogenic
effects or that the latter are produced by broad genetic damage.
Additionally, the significant (p< 0.05) increase of apoptotic cells
(Fig. 3) measured at the end of exposure at 10 nM, the strong cor-
relation with micronuclei frequency noticed already at 5 nM and
the lack of correlation with SCGE end-points can confirm the
hypothesis that hemocytes chose a programmed cell death only
in presence of great DNA injuries, without transitioning through
the DNA fragmentation.
Since the genotoxicity potential of PCM was registered only at
the end of exposure (Figs. 2 and 3), we can suppose that this hydro-
philic compound can be readily eliminated by the detoxification
system through feces at least at the beginning of exposure. After-
ward, when the levels of PCM increase in tissues, broad genetic
damage appears at the highest doses and cells can choose either
to follow the programmed death or the micronuclei production.
However, this hypothesis does not explain the possible mechanism
of action by which PCM damages macromolecules. Many authors
(Hazai et al., 2002) have suggested that the metabolism of PCM
leads to ROS production since the parental molecule is biotrans-
formed in the toxic NAPQI (N-acetyl-p-benzoquinoneimine), which
Fig. 2. Frequency of micronucleated hemocytes (mean‰± SEM) measured in zebra mussel specimens. Significant differences (two-way ANOVA, Bonferroni post hoc test,
p< 0.01) are referred to the comparison between treated mussels and the correspondent control (time to time).
Fig. 3. Percentages of apoptotic hemocytes (mean values ± SEM) measured by the DNA diffusion assay in zebra mussel specimens. Significant differences (two-way ANOVA,
Bonferroni post hoc test,
p< 0.05) are referred to the comparison between treated mussels and the correspondent control (time to time).
494 M. Parolini et al. /Chemosphere 79 (2010) 489–498
Author's personal copy
besides being a reactive metabolite that interacts with proteins
and nucleic acids (Huber et al., 2009), it is an electrophilic interme-
diate and can increase the formation of reactive oxygen species
(ROS) and reactive nitrogen species (RNS), such as superoxide an-
ion, hydroxyl radical, hydrogen peroxide, nitro oxide and peroxyni-
trite (Yen et al., 2006). Furthermore, results obtained measuring
the bivalve enzymatic activity explain the onset of genetic damage
with the increase in oxidative stress through the production of
ROS.
SOD activity in D. polymorpha exposed to PCM showed a very
particular response since it significantly (p< 0.01) increased only
at the lowest concentration after 24 h of exposure (Fig. 5). The lack
Fig. 4. Neutral Red Retention Time (mean NRRT values ± SEM) measured in zebra mussel specimens. Significant differences (two-way ANOVA, Bonferroni post hoc test,
p< 0.01) referred to the comparison between treated mussels and the correspondent control (time to time).
Fig. 5. Effects of increasing doses of PCM on the activity of catalase (CAT), glutathione peroxidase (GPx), superoxide dismutase (SOD) and glutathione S-transferase (GST)
measured in the body of zebra mussels. Significant differences (two-way ANOVA, Bonferroni post hoc test,
p< 0.05,
p< 0.01) are referred to the comparison between treated
mussels and the correspondent control (time to time).
M. Parolini et al. /Chemosphere 79 (2010) 489–498 495
Author's personal copy
of a significant SOD response at 5 nM and 10 nM could indicate
that sub-lethal damage caused by PCM at higher doses compro-
mised the ability of the zebra mussel to respond in an adaptive
manner (Regoli et al., 2003). On the other hand, the antioxidant
chain reaction is not blocked by the lack of SOD induction because
of several other cellular enzymes, such as those contained in per-
oxisomes, are able to generate H
2
O
2
(Khessiba et al., 2005) that
represents the selective substrate for the subsequent antioxidant
enzymes. The significant induction of CAT and GPx obtained in
our study was also observed in several previous studies carried
out in mussels exposed to different xenobiotics. A significant in-
crease of CAT was observed in the digestive gland of a ribbed mus-
sel (Geukensia demissa) exposed to paraquat (Wenning et al., 1988),
in M. edulis exposed to menadione (Livingstone et al., 1990) and in
a green-lipped mussel (Perna viridis) exposed to polycyclic aro-
matic hydrocarbons (PAHs) and organochlorine pesticides (OCs)
(Richardson et al., 2008). Furthermore, Pan et al. (2006) have found
a similar GPx increasing trend in the Chlamys ferrari hemolymph
exposed to benzo(
a
)pyrene, benzo(k)fluoranthene and their mix-
ture. The significant increase (p< 0.01) of GST activity was already
found by many authors (Gagné et al., 2004; Canesi et al., 2007;
Binelli et al., 2009) both in the entire body and in the hepatic
glands of some mussel species exposed to different xenobiotics,
and it seems to confirm the possible increase in oxidative stress
due to an excretion mechanism. On the other hand, PCM is a very
polar chemical (log K
ow
= 0.46; Lorphensri et al., 2007) and can be
easily conjugated to glutathione by GST through weakly active
phase I biotransformation enzymes. This pathway was verified in
mammals where high PCM concentrations are oxidized by en-
zymes of the CYP 450 mixed-function oxidase (MFO) system to
NAPQI (Corcoran et al., 1980; Bowles et al., 2005), which quickly
reacts with glutathione, resulting in the formation of a glutathione
conjugate mediated by GST (Huber et al., 2009). On the basis of
these findings we can hypothesize a possible mechanism of action
of PCM in the zebra mussel (Fig. 6): PCM is oxidized by CYP 450 to
the highly reactive NAPQI that can directly interact with proteins
and nucleic acids. In addition, this electrophilic intermediate can
increase the formation of ROS and reactive nitrogen species
(RNS) that can react and create broad injuries to DNA. Finally,
GST detoxifies NAPQI, catalyzing its conjugation with glutathione.
To verify this suggested mechanism of action it will be necessary to
investigate the formation of NAPQI and/or its glutathione-conju-
gated metabolite. Also the significant decrease of the NRRT
(Fig. 4), which is a common parameter for the assessment of gen-
eral stress in bivalves and an indicator of the induction of oxidative
stress (Lowe et al., 1995), along with the significant negative corre-
lation with apoptosis and micronuclei end-points obtained at 5 nM
and 10 nM confirm that the oxidative stress could be the main
pathway for the formation of DNA damage.
The correlation analysis (Table 1) performed on all measured
biomarkers could be an useful tool to support the proposed mech-
anisms of action. The genotoxicity revealed mainly for the highest
administered dose is most likely due simply to the increase in oxi-
dative stress, since overall significant (p< 0.05) correlations were
found between genotoxic end-points and all the antioxidant en-
zymes. Moreover, another confirmation was made by the signifi-
cant negative (p< 0.05) correlation between NRRT, genotoxic
biomarkers and antioxidants, which was in accordance with work
by many authors who reported good correlations between lyso-
somal membrane destabilization, impairment of the antioxidant
defense system and DNA injuries (Regoli et al., 2002; Mamaca
et al., 2005; Binelli et al., 2009). In addition, other authors have
suggested that reactive free-radicals contribute to the destabiliza-
tion of lysosomal membranes and that the intralysosomal environ-
ment is already an oxyradical production site (Kirchin et al., 1992;
Winston et al., 1996). Finally, the significant correlation (p< 0.05)
between GST and phase I enzymes confirms the activation of the
entire enzymatic defense chain of the zebra mussel. Moreover, it
supports the thesis that GST play a fundamental role in the protec-
tion against PCM, but at the same time that it is also the responsi-
ble for the insurgence of oxidative stress, through the PCM
conjugation with glutathione.
5. Conclusions
Since this work is only partially in accordance with results ob-
tained from our in vitro study on the evaluation of cyto-genotoxi-
city of PCM, it confirms that the tiered approach can give a more
exhaustive and detailed report of the pollutant toxicity. The new
Fig. 6. Supposed mechanism of action of PCM in zebra mussel.
Table 1
Pearson’s correlation matrix obtained by using all measured end-points (pool of 18–
24 specimens) at 1 nM, 5 nM and 10 nM. Significant correlations (
p< 0.05) are
indicated in bold.
LDR %DNA MN Apo NRRT CAT SOD GPx
1nM
%DNA 0.79
MN 0.03 0.01
Apo 0.07 0.04 0.02
NRRT 0.08 0.01 0.19 0.09
CAT 0.32 0.21 0.05 0.16 0.22
SOD 0.06 0.35 0.26 0.17 0.29 0.17
GPx 0.09 0.24 0.27 0.24 0.65 0.06 0.40
GST 0.09 0.05 0.03 0.18 0.24 0.35 0.09 0.13
5nM
%DNA 0.78
MN 0.16 0.27
Apo 0.05 0.17 0.33
NRRT 0.08 0.21 0.50 0.23
CAT 0.39 0.37 0.21 0.21 0.25
SOD 0.04 0.17 0.18 0.32 0.10 0.02
GPx 0.26 0.32 0.60 0.34 0.72 0.55 0.01
GST 0.18 0.26 0.55 0.26 0.77 0.21 0.08 0.79
10 nM
%DNA 0.81
MN 0.06 0.12
Apo 0.04 0.10 0.45
NRRT 0.00 0.11 0.69 0.51
CAT 0.04 0.16 0.62 0.69 0.70
SOD 0.05 0.14 0.49 0.50 0.60 0.75
GPx 0.19 0.19 0.36 0.28 0.54 0.59 0.70
GST 0.15 0.01 0.49 0.43 0.64 0.78 0.85 0.85
LDR = length/diameter ratio; % DNA = percentage of tail DNA; Apo = frequency of
apoptotic cells; MN = frequency of micronuclei; NRRT = Neutral Red Retention
Time; CAT = catalase; GPx = glutathione peroxidase; SOD = superoxide dismutase;
GST = glutathione S-transferase.
496 M. Parolini et al. /Chemosphere 79 (2010) 489–498
Author's personal copy
data revealed the capacity of this drug to induce moderate geno-
toxicity in bivalves exposed to environmental concentrations, but
they do not show any primary DNA damage, unlike the in vitro re-
sults. This genetic damage was probably due both to the increase in
oxidative stress and/or to a direct interaction between its metabo-
lite NAPQI with DNA. These data confirm the hypothesis that PCM
could be a problematic compound for aquatic organisms, although
the current environmental concentrations can seemingly cause
only low and moderate adverse effects. On the other hand, the sig-
nificant DNA injuries found at the end of exposure can show a pos-
sible delay of PCM genotoxic action that can be confirmed by
longer assays. Nonetheless, considering the increasing production
and use of this drug, its environmental presence and hazardous im-
pact on aquatic organisms could quickly increase.
References
An, J., Zhou, Q.X., Sun, F.H., Zhang, L., 2009. Ecotoxicological effects of paracetamol
on seed germination and seedling development of wheat (Triticum aestivum L.).
J. Hazard. Mater. 169, 751–757.
Binelli, A., Ricciardi, F., Riva, C., Provini, A., 2005. Screening of POP pollution by AChE
and EROD activities in zebra mussels from the Italian great lakes. Chemosphere
61 (8), 1074–1082.
Binelli, A., Cogni, D., Parolini, M., Riva, C., Provini, A., 2008a. In vivo experiments for
the evaluation of genotoxic and cytotoxic effects of Triclosan in zebra mussel
hemocytes. Aquat. Toxicol. 91, 238–244.
Binelli, A., Riva, C., Cogni, D., Provini, A., 2008b. Genotoxic effects of p,p
0
-DDT (1,1,1-
trichloro-2,2-bis-(chlorophenyl)ethane) and its metabolites in zebra mussel (D.
Polymorpha) by SCGE assay and micronucleus test. Environ. Mol. Mutagen. 49,
406–415.
Binelli, A., Parolini, M., Cogni, D., Pedriali, A., Provini, A., 2009. A multi-biomarker
assessment of the impact of the antibacterial trimethoprim on the non-target
organism zebra mussel (Dreissena polymorpha). Comp. Biochem. Physiol. C 150,
329–336.
Bottoni, P., Fidente, R., 2005. Un primo contributo alla problematica dei farmaci
come contaminanti delle acque. Ann. Ist. Super San 41 (3), 333–342.
Bound, J.P., Voulvoulis, N., 2006. Predicted and measured concentrations for
selected pharmaceuticals in UK rivers: implications for risk assessment.
Water Res. 40, 2885–2892.
Bowles, D., Isayenkova, J., Lim, E.K., Poppenberger, B., 2005. Glycosyltransferases:
managers of small molecules. Curr. Opin. Plant Biol. 8, 254–263.
Bradford, M.M., 1976. A rapid and sensitive method for the quantification of
microgram quantities of protein using the principle of protein–dye binding.
Anal. Biochem. 72, 248–254.
Buschini, A., Carboni, P., Martino, A., Poli, P., Rossi, C., 2003. Effects of temperature
on baseline and genotoxicant-induced DNA damage in haemocytes of Dreissena
polymorpha. Mutat. Res. 537, 81–92.
Canesi, L., Ciacci, C., Lorusso, L.C., Betti, M., Gallo, G., Pojana, G., Marcomini, A., 2007.
Effects of Triclosan on Mytilus galloprovincialis hemocyte function and digestive
gland enzyme activities: possible modes of action on non target organisms.
Comp. Biochem. Physiol. C 145, 464–472.
Carlsson, C., Johansson, A.K., Alyan, G., Bergman, K., Kühler, T., 2006. Are
pharmaceuticals potent environmental pollutants? Part I: environmental risk
assessments of selected active pharmaceutical ingredients. Sci. Total Environ.
364, 67–87.
Choi, K., Kim, Y., Jung, J., Kim, M.H., Kim, C.S., Kim, N.H., Park, J., 2008. Occurrence
and ecological risks of roxithromycin, trimethoprim, and chloramphenicol in
the Han River, Korea. Environ. Toxicol. Chem. 27 (3), 711–719.
Cleuvers, M., 2004. Mixture toxicity of the anti-inflammatory drugs diclofenac,
ibuprofen, naproxen, and acetylsalicylic acid. Ecotoxicol. Environ. Saf. 59, 309–
315.
Corcoran, G.B., Mitchell, J.R., Vaishnav, Y.N., Horning, E.C., 1980. Evidence that
acetaminophen and N-hydroxyacetaminophen form a common arylating
intermediate, N-acetyl-p-benzoquinoneimine. Mol. Pharmacol. 18, 536–542.
Crane, M., Watts, C., Boucard, T., 2006. Chronic aquatic environmental risks from
exposure to human pharmaceuticals. Sci. Total Environ. 367, 23–41.
Daughton, C.G., Ternes, T.A., 1999. Pharmaceuticals and personal care products in
the environment: agents of subtle change? Environ. Health Perspect. 107
(Suppl. 6), 907–938.
Faria, M., Carrasco, L., Diez, S., Riva, M.C., Bayona, J.M., Barata, C., 2009. Multi-
biomarker responses in the freshwater mussel Dreissena polymorpha exposed to
polychlorobiphenyls and metals. Comp. Biochem. Physiol. C 149, 281–288.
Farré, M.L., Ferrer, I., Ginebreda, A., Figueras, M., Olivella, L., Tirapu, L., Vilanova, M.,
Barcelo, D., 2001. Determination of drugs in surface water and wastewater
samples by liquid chromatography–mass spectrometry: methods and
preliminary results including toxicity studies with Vibrio fischeri.J.
Chromatogr. A 938 (1/2), 187–197.
Fent, K., 2003. Ökotoxikologie. Georg Thieme Verlag, Stuttgart.
Fent, K., Weston, A.A., Caminada, D., 2006. Ecotoxicology of human
pharmaceuticals. Aquat. Toxicol. 76, 122–159.
Ferrari, B., Mons, R., Vollat, B., Fraysse, B., Paxeus, N., Lo Giudice, R., Pollio, A., Garric,
J., 2004. Environmental risk assessment of six human pharmaceuticals: are the
current environmental risk assessment procedures sufficient for the protection
of the aquatic environment? Environ. Toxicol. Chem. 23 (5), 1344–1354.
Gagné, F., Blaise, C., Hellou, J., 2004. Endocrine disruption and health effects of caged
mussels, Elliptio complanata, placed downstream from a primary treated
municipal effluent plume for 1 year. Comp. Biochem. Physiol. C 138, 33–44.
Gagné, F., Bérubé, E., Fournier, M., Blaise, C., 2005. Inflammatory properties of
municipal effluents to Elliptio complanata mussels – lack of effects from anti-
inflammatory drugs. Comp. Biochem. Physiol. C 141, 332–337.
Gagné, F., Blaise, C., Fournier, M., Hansen, P.D., 2006. Effects of selected
pharmaceutical products on phagocytic activity in Elliptio complanata
mussels. Comp. Biochem. Physiol. C 143, 179–186.
Gros, M., Petrovic
´, M., Barceló, D., 2006. Development of a multi-residue analytical
methodology based on liquid chromatography–tandem mass spectrometry
(LC–MS/MS) for screening and trace level determination of pharmaceuticals in
surface and wastewaters. Talanta 70 (4), 678–690.
Grujic
´, S., Vasiljevic
´, T., Lauševic
´, M., 2009. Determination of multiple
pharmaceutical classes in surface and ground waters by liquid
chromatography–ion trap–tandem mass spectrometry. J. Chromatogr. A 1216,
4989–5000.
Haap, T., Triebskorn, R., Hohler, H.-R., 2008. Acute effects of diclofenac and DMSO to
Daphnia magna: immobilisation and hsp70-induction. Chemosphere 73, 353–
359.
Halling-Sorensen, B., Nors Nielsen, S., Lanzky, P.F., Ingerslev, F., Holten Lutzhoft,
H.C., Jorgensen, S.E., 1998. Occurrence, fate and effects of pharmaceutical
substances in the environment – a review. Chemosphere 36 (2), 357–393.
Han, G.H., Hur, H.G., Kim, S.D., 2006. Ecotoxicological risk of pharmaceuticals from
wastewater treatment plants in Korea: occurrence and toxicity to Daphnia
magna. Environ. Toxicol. Chem. 25, 265–271.
Hartmann, A., Schumacher, M., Plappert-Helbig, U., Lowe, P., Suter, W., Mueller, L.,
2004. Use of the alkaline in vivo comet assay for mechanistic genotoxicity
investigations. Mutagenesis 19 (1), 51–59.
Hazai, E., Vereczkey, L., Monostory, K., 2002. Reduction of toxic metabolite
formation of acetaminophen. Biochem. Biophys. Res. Commun. 291, 1089–
1094.
Heberer, T., 2002. Occurrence, fate, and removal of pharmaceutical residues in the
aquatic environment: a review of recent research data. Toxicol. Lett. 131, 5–17.
Henschel, K.P., Wenzel, A., Diedrich, M., Fliedner, A., 1997. Environmental hazard
assessment of pharmaceuticals. Regul. Toxicol. Pharmacol. 25 (3), 220–225.
Huber, C., Bartha, B., Harpaintner, R., Schröder, P., 2009. Metabolism of
acetaminophen (paracetamol) in plants – two independent pathways result in
the formation of a glutathione and a glucose conjugate. Environ. Sci. Pollut. Res.
16, 206–213.
Huggett, D.B., Brooks, B.W., Peterson, B., Foran, C.M., Schlenk, D., 2002. Toxicity of
selected beta adrenergic receptor-blocking pharmaceuticals (B-blockers) on
aquatic organisms. Arch. Environ. Contam. Toxicol. 43 (2), 229–235.
ICES, 2004. In: Moore, M.N., Lowe, D. (Eds.), Biological Effects of Contaminants:
Measurement of Lysosomal Membrane Stability: ICES Techniques in Marine
Environmental Sciences, Copenhagen, Denmark, vol. 36, p. 39.
Khessiba, A., Roméo, M., A, P., 2005. Effects of some environmental parameters on
catalase activity measured in the mussel (Mytilus galloprovincialis) exposed to
lindane. Environ. Pollut. 133, 275–281.
Kim, Y., Choi, K., Jung, J., Park, S., Kim, P.G., Park, J., 2007. Aquatic toxicity of
acetaminophen, carbamazepine, cimetidine, diltiazem and six major
sulfonamides, and their potential ecological risks in Korea. Environ. Int. 33,
370–375.
Kirchin, M.A., Moore, M.N., Dean, R.T., Winston, G.W., 1992. The role of oxyradicals
in intracellular proteolysis and toxicity in mussels. Mar. Environ. Res. 34, 315–
320.
Kirkland, D.J., Hayashi, M., Jacobson-Kram, D., Kasper, P., MacGregor, J.T., Müller, L.,
Uno, Y., 2007. Summary of major conclusions from the 4th IWGT, San Francisco,
9–10 September, 2005. Mutat. Res. 627, 5–9.
Kirsch-Volders, M., Sofuni, T., Aaderma, M., Albertini, S., Eastmond, D., Fenech, M.,
Ishidate, M., Lorge, E., Norppa, H., Surallés, J., Von der Hude, W., Wakata, A.,
2000. Report from the in vitro micronucleus assay working group. Environ. Mol.
Mutagen. 35, 167–172.
Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., et al., 2002.
Response to comment on, ‘Pharmaceuticals, hormones, and other organic
wastewater contaminants in US streams, 1999–2000: a national
reconnaissance’. Environ. Sci. Technol. 36, 4007–4008.
Kolpin, D.W., Skopec, M., Meyer, M.T., Furlong, E.T., Zaugg, S.D., 2004. Urban
contribution of pharmaceuticals and other organic wastewater contaminants to
streams during differing flow conditions. Sci. Total Environ. 328, 119–130.
Kümmerer, K., 2001. Drugs in the environment: emission of drugs, diagnostic aids
and disinfectants into wastewater by hospitals in relation to other sources – a
review. Chemosphere 45 (6/7), 957–969.
Kümmerer, K., 2004. Pharmaceuticals in the environment – scope of the book and
introduction. In: Kümmerer, K. (Ed.), Pharmaceuticals in the Environment,
second ed. Springer-Verlag, Berlin, pp. 3–11.
Langman, M.J.S., 1999. Epidemiology of non-steroidal anti-inflammatory drug
damage to stomach and duodenum. Ital. J. Gastroenterol. Hepatol. 31 (S1),
2–5.
Livingstone, D.R., Garcia Martinez, P., Michel, X., Ribera, D., Winston, G.W., 1990.
Oxyradical production as a pollution-mediated mechanism of toxicity in the
common mussel Mytilus edulis L., and other mollusks. Funct. Ecol. 4, 415–424.
M. Parolini et al. /Chemosphere 79 (2010) 489–498 497
Author's personal copy
Lorphensri, O., Sabatini, D.A., Kibbey, T.C.G., Osathaphan, K., Saiwan, C., 2007.
Sorption and transport of acetaminophen, 17a-ethynylestradiol, nalidixic acid
with low organic contentaquifer sand. Water Res. 41, 2180–2188.
Lowe, D.M., Pipe, R.K., 1994. Contaminant induced lysosomal membrane damage
in marine mussel digestive cells: an in vitro study. Aquat. Toxicol. 30,
357–365.
Lowe, D.M., Fossato, V.U., Depledge, M.H., 1995. Contaminant-induced lysosomal
membrane damage in blood cells of mussels Mytilus galloprovincialis from the
Venice Lagoon: an in vitro study. Mar. Ecol. Prog. 129, 189–196.
Mamaca, E., Bechmann, R.K., Torgrimsen, S., Aas, E., Bjørnstad, A., Baussant, T., Le
Floch, S., 2005. The neutral red lysosomal retention assay and comet assay on
haemolymph cells from mussels (Mytilus edulis) and fish (Symphodus melops)
exposed to styrene. Aquat. Toxicol. 75, 191–201.
Metcalfe, C.D., Koenig, B.G., Bennie, D.T., Servos, M., Ternes, T.A., Hirsch, R., 2003.
Occurrence of neutral and acidic drugs in the effluents of Canadian sewage
treatment plants. Environ. Toxicol. Chem. 22 (12), 2872–2880.
Misra, R., Pandey, H., Chandra, M., Agarwal, P.K., Pandeya, S.N., 1990. Effects of
commonly used non steroidal anti-inflammatory drugs on gastric mucosa. A
clinical, endoscopic and histopathological study. J. Assoc. Phys. India 38 (9),
636–638.
Orbea, A., Ortiz-Zarragoitia, M., Solé, M., Porte, C., Cajaraville, M.P., 2002.
Antioxidant enzymes and peroxisome proliferation in relation to
contaminant body burdens of PAHs and PCBs in bivalve molluscs, crabs
and fish from the Urdaibai and Plentzia estuaries (Bay of Biscay). Aquat.
Toxicol. 58, 75–98.
Osman, A.M., van Noort, P.C.M., 2007. Comparison of key enzymes in the zebra
mussel, Dreissena polymorpha, the earthworm Allolobophora chlorotica and
Chironomus riparius larvae. Ecotoxicol. Environ. Saf. 67, 212–217.
Osman, A.M., van den Heuvel, H., van Noort, P.C.M., 2007. Differential responses of
biomarkers in tissues of a freshwater mussel, Dreissena polymorpha, to the
exposure of sediment extracts with different levels of contamination. J. Appl.
Toxicol. 27, 51–59.
Pan, L.Q., Ren, J., Liu, J., 2006. Responses of antioxidant systems and LPO level to
benzo(
a
)pyrene and benzo(k)fluoranthene in the haemolymph of the scallop
Chlamys ferrari. Environ. Pollut. 141, 443–451.
Parolini, M., Binelli, A., Cogni, D., Riva, C., Provini, A., 2009. An in vitro biomarker
approach for the evaluation of the ecotoxicity of non-steroidal anti-
inflammatory drugs (NSAIDs). Toxicol. in Vitro 23, 935–942.
Pascoe, D., Karntanut, W., Muller, C.T., 2003. Do pharmaceuticals affect freshwater
invertebrates? A study with the cnidarian Hydra vulgaris. Chemosphere 51 (6),
521–528.
Pavlica, M., Klobucar, G.I.V., Vetma, N., Erben, R., Papeš, D., 2000. Detection of
micronuclei in haemocytes of zebra mussel and ramshorn snail exposed to
pentachlorophenol. Mutat. Res. 465, 145–150.
Quinn, B., Gagnè, F., Blaise, C., 2008. An investigation into the acute and
chronic toxicity of eleven pharmaceuticals (and their solvents) found in
wastewater effluent on the cnidarian, Hydra attenuata. Sci. Total Environ.
389, 306–314.
Regoli, F., Gorbi, S., Frenzilli, G., Nigro, M., Corsi, I., Focardi, S., Winston, G.W., 2002.
Oxidative stress in ecotoxicology: from the analysis of individual antioxidants
to a more integrated approach. Mar. Environ. Res. 54, 419–423.
Regoli, F., Winston, G.W., Gorbi, S., Frenzilli, G., Nigro, M., Corsi, I., Focardi, S., 2003.
Integrating enzymatic responses to organic chemical exposure with total
oxyradical absorbing capacity and DNA damage in the European eel Anguilla
anguilla. Environ. Toxicol. Chem. 22, 2120–2129.
Richardson, B.J., Mak, E., De Luca-Abbott, S.B., Martin, M., McClellan, K., Lam, P.K.S.,
2008. Antioxidant responses to polycyclic aromatic hydrocarbons and
organochlorine pesticides in green-lipped mussels (Perna viridis): do mussels
‘‘integrate” biomarker responses? Mar. Pollut. Bull. 57, 503–514.
Riva, C., Binelli, A., Cogni, D., Provini, A., 2007. Evaluation of DNA damages induced
by BDE-209 in haemocytes of Dreissena polymorpha using SCGE assay and MN
test. Environ. Mol. Mutagen. 48, 735–743.
Sanderson, H., Johnson, D.J., Reitsma, T., Brain, R.A., Wilson, C.J., Solomon, K.R., 2004.
Ranking and prioritization of environmental risks of pharmaceuticals in
surfacewaters. Regul. Toxicol. Pharm. 39 (2), 158–183.
Schulte-Oehlmann, U., Oehlmann, J., Püttmann, W., 2007.
Humanpharmakawirkstoffe in der Umwelt: Einträge, Vorkommen und der
Versuch einer Bestandsaufnahme. Umweltwiss Schadst Forsch 19 (3), 68–179.
Singh, N.P., 2000. Microgels for estimation of DNA strand breaks, DNA protein
crosslinks and apoptosis. Mutat. Res. 455, 111–127.
Singh, N.P., McCoy, M.T., Tice, R.R., Schneider, E., 1988. A simple technique for
quantification of low levels of DNA damage in individual cells. Exp. Cell Res.
175, 184–191.
Stuer-Lauridsen, F., Birkved, M., Hansen, L.P., Holten Lützhøft, H.C., Halling-
Sørensen, B., 2000. Environmental risk assessment of human pharmaceuticals
in Denmark after normal therapeutic use. Chemosphere 40, 783–793.
Stumpf, M., Ternes, T., Haberer, K., Seel, P., Baumann, W., 1996. Nachweis von
Arzneimittelrückständen in Kläranlagen und Fliessgewässern. Vom Wasser 86,
291–303.
Ternes, T., Hirsch, R., Mueller, J., Haberer, K., 1998. Methods for the determination of
neutral drugs as well as betablockers and alpha2-sympathomimetics in
aqueous matrices using GC/MS and LC/MS/MS. Fresenius. J. Anal. Chem. 362
(3), 329–340.
Wenning, R.J., Di Giulio, R.T., Gallagher, E.P., 1988. Oxidant-mediated biochemical
effects of paraquat in the ribbed mussel, Guekensia demissa. Aquat. Toxicol. 12,
157–170.
Winston, G.W., Moore, M.N., Kirchin, M.A., Soverchia, C., 1996. Production of
reactive oxygen species (ROS) by hemocytes from the marine mussel, Mytilus
edulis. Comp. Biochem. Physiol. C 113, 221–229.
Yang, L.H., Ying, G.G., Su, H.C., Stauber, J.L., Adams, M.S., Binet, M.T., 2008. Growth
inhibiting effects of 12 antibacterial agents and their mixtures on the
freshwater microalga Pseudokirchneriella subcapitata. Environ. Toxicol. Chem.
27 (5), 1201–1208.
Yen, F.L., Wu, T.H., Lin, L.T., Lin, C.C., 2006. Hepatoprotective and antioxidant effects
of Cuscuta chinensis against acetaminophen-induced hepatotoxicity in rats. J.
Ethnopharmacol. 111 (1), 123–128.
498 M. Parolini et al. /Chemosphere 79 (2010) 489–498
