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Information of the Journal in which the present paper is
 Elsevier, Aquatic Toxicology, 2011, 101 (1), pp. 78-87.
Characterization of the multixenobiotic resistance (MXR)
mechanism in embryos and larvae of the zebra mussel
(Dreissena polymorpha) and studies on its role in tolerance to
single and mixture combinations of toxicants
Melissa Fariaa, Ana Navarroa, Till Luckenbachb, Benjamin Piñaa, Carlos
a Department of Environmental Chemistry, IDAEA-CSIC, Jordi Girona 18, 08034 Barcelona, Spain
b UFZ-Helmholtz Centre for Environmental Research, Permoserstr. 15, D-04318 Leipzig,
∗ Corresponding author. Tel.: +34 93 4006100x147; fax: +34 93 2045904. E-mail
address: (C. Barata).
The study of the cellular mechanisms of tolerance of organisms to pollution is a key
issue in aquatic environmental risk assessment. Recent evidence indicates that
multixenobiotic resistance (MXR) mechanisms represent a general biological defense of
many marine and freshwater organisms against environmental toxicants. In this work,
toxicologically relevant xenobiotic efflux transporters were studied in early life stages
of zebra mussels (Dreissena polymorpha). Expression of a P-gp1 (ABCB1) transporter
gene and its associated efflux activities during development were studied, using qRTPCR and the fluorescent transporter substrates rhodamine B and calcein-AM combined
with specific transporter inhibitors (chemosensitizers). Toxicity bioassays with the
model P-gp1 chemotherapeutic drug vinblastine applied singly and in combination with
different chemosensitizers were performed to elucidate the tolerance role of the P-gp1
efflux transporter. Results evidenced that the gene expression and associated efflux
activities of ABC transporters were low or absent in eggs and increased significantly in
1–3 d old trochophora and veliger larvae. Specific inhibitors of Pgp1 and/or MRP
transport activities including cyclosporine A, MK571, verapamil and reversin 205 and
the musk celestolide resulted in a concentration dependent inhibition of related transport
activities in zebra mussel veliger larvae, with IC50 values in the lower micromolar
range and similar to those reported for mammals, fish and mussels. Binary mixtures of
the tested transporter inhibitors except celestolide with the anticancer drug and P-gp1
substrate vinblastine increased the toxicity of the former compound more than
additively. These results indicate that MXR transporter activity is high in early lifestages of the zebra mussel and that may play an important role in the tolerance to
environmental contaminants.
Multixenobiotic resistance (MXR); MXR inhibitors; Larvae; Freshwater; Zebra mussel;
Musk fragrances; Mixture toxicity
1. Introduction
Aquatic organisms are constantly exposed to complex mixtures of structurally diverse
chemicals dissolved in the water and must tolerate their potential toxic impact. Multiple
evidence indicates that the cellular multixenobiotic resistance (MXR) system represents
a general biological defense mechanism for the protection of organisms against both
endogenous and environmental toxicants (Bard, 2000). MXR is mediated by
transmembrane transport proteins from the ABC (ATP binding cassette) protein family,
which recognize a wide variety of potential xenobiotics as substrates, pumping them out
of the cell in an energy dependent, ATP driven process (Epel, 1998 and Kurelec, 1992).
ABC transporters are a large protein family, sub-categorized into sub-families ABCAABCH with a variety of functions (Dean, 2005). The transporters with MXR-related
function are P-glycoproteins (P-gp, MDR1, ABCB1) belonging to the ABCB subfamily
(Germann, 1993), multidrug resistance associated proteins 1–5 (MRP1–5, ABCC1–5)
from the ABCC subfamily (Deeley et al., 2006) and breast cancer resistance protein
(BCRP, ABCG2) from the ABCG subfamily (Sarkadi et al., 2006).
Putative homologs of P-gp1 and of several MRPs have been identified in aquatic
invertebrates. Particularly high expression and activity of efflux transporters have been
reported for marine invertebrate larvae, such as innkeeper worm (Toomey et al., 1996),
mussel (McFadzen et al., 2000), sea urchin (Hamdoun et al., 2002) and sea star (Roepke
et al., 2006). However, MXR activity has not been investigated in larvae from any
freshwater invertebrate species yet.
Environmental pollutants may inhibit MXR transporter activity, increasing their
effective intracellular concentrations and enhancing their deleterious effects in exposed
cells and organisms. These compounds, able to block the function of transporters, and
therefore increasing the sensitivity of the cell to chemical agents, are termed
chemosensitizers (Epel et al., 2008 and Kurelec, 1997). They are classified into two
categories, based on their mechanisms of action. Competitive inhibitors are able, at
relatively large concentrations, to overwhelm the substrate-binding capacity of the
transporters, which typically present low substrate specificity. This is seen as a risk of
the presence of a vast variety of anthropogenic compounds in the environment
(Daughton and Ternes, 1999). Conversely, non-competitive chemosensitizers are
pollutants able to block the ATPase activity of the transporters, by allosteric interactions
and subsequently suppressing their detoxifying activity (Ramachandra et al., 1998).
Anthropogenic contaminants that have so far been found to act as chemosensitizers
include pesticides, perfluorochemicals, fragrances, pharmaceuticals and polyaromatic
hydrocarbons. The chemical diversity of these compounds indicates that
chemosensitization could be a common effect elicited by many environmental
chemicals. Research on environmental aspects of chemicals acting as chemosensitizers
is still at its early stage. The available literature indicates that the effect could be
important but is underappreciated. Its correct evaluation is particularly important for
risk assessment of mixtures, since it may provide a mechanistic explanation of cases
where combined effects of compounds are larger than predicted effects from single
chemicals (Altenburger et al., 2003 and Epel et al., 2008).
Bivalve species are excellent model organisms to study the physiological role of MXR
and its modulation by environmental chemosensitizers. They are able to successfully
reproduce, grow and develop in contaminated habitats and often show high MXR
related transporter activities (Smital et al., 2000).
The main objectives of this work are to study the transcription of MXR transporter
genes and functionally characterize toxicologically relevant xenobiotic efflux
transporters in early life stages of zebra mussel (Dreissena polymorpha). Zebra mussel
is an invasive species now widely distributed across North America and Europe. Due to
its abundance, the cellular toxic defense mechanisms of zebra mussel, including MXR
related gene and protein expression and MXR transporter efflux activities have been
actively studied ( Pain and Parant, 2007, Tutundjian and Minier, 2007 and Smital et al.,
2003). Information on MXR is so far limited to adult zebra mussels, being non-existing
for embryos and larvae.
We present here relative expression levels of ABCB1 transporter mRNA in different
larval stages and efflux transporter activities determined by measuring uptake of
fluorescent dyes by the larvae in the presence of transporter inhibitors. We then address
the question in toxicity assays with a toxic transporter substrate and various inhibitors
whether chemosensitization explains synergic joint toxic effects in mussel embryos.
2. Material and methods
2.1. Chemicals
The five tested compounds, cyclosporine A (CsA), verapamil (VER), reversin 205
(REV205), MK571, and celestolide (Cel), and vinblastine, rhodamine B (RhB),
calcein=AM (Ca=AM) and serotonin creatinine sulphate monohydrate were purchased
from Sigma–Aldrich (Steinheim, Germany). Ethanol and DMSO (analytical grade) were
obtained from Merck (Darmstadt, Germany).
2.2. Experimental animals
Sexually mature zebra mussels (D. polymorpha) were collected in shallow water (0.5–1
m deep) from the Ebro River in the Ribaroja reservoir (NE, Spain) during the May–
August period in 2009 and 2010. Within 3 h of collection, animals were transported in
local water in aerated 10 L plastic containers to the lab and acclimated to 20 °C and
ASTM hard water ( ASTM, 1995) for at least 24 h. Spawning of the zebra mussels was
induced by exposure to 10−3 M serotonin creatinine sulphate monohydrate for 15 min.
The mussels were then transferred to clean ASTM hard water. Spawning occurred
within 15–30 min in males and within 1–3 h in females. A composite of about 106 eggs
from at least 5 females was pooled in 200 mL ASTM hard water and fertilized with 1
mL of sperm from at least 3 males. Only gametes obtained from clean ASTM hard
water (no serotonin) were used ( Faria et al., 2010).
For studying the transcription levels of P-glycoprotein (P-gp1) gene and efflux MXR
transporter activity in 1 d old trochophora and 2 d and 3 d D shape veliger larvae, about
105 of fertilized eggs were either collected (1 h of age) or cultured in a 10 L ASTM
hard water under constant oxygenation (>90% air saturation), temperature (20 °C) and
photoperiod (14 h:10 h; light:dark) without food. Culture of zebra mussel larvae for
more than 3 d requires addition of food and handling greater numbers of individuals to
account for high mortality rates (Vanderploeg et al., 1996). Eggs (1 h after being
fertilized) and larvae were collected and concentrated by filtration into 50 mL centrifuge
tubes and then used in efflux MXR transporter assays. For determining gene
transcription profiles eggs and larvae were further concentrated by gentle centrifugation
(1000 rpm, 10 min). Pellets were resuspended in 5 mL and equally aliquoted in 1.5 mL
eppendorf tubes. After a brief centrifugation the supernatant was discharged and the
pellet, containing larvae, was resuspended in RNAlater® (Sigma–Aldrich) and stored at
−80 °C.
2.3. RNA extraction and qRT-PCR analysis
Total RNA was isolated from the tissues using Trizol reagent® (Invitrogen). The RNA
concentration was measured by spectrophotometric absorption at 260 nm with a
NanoDrop ND-1000 spectrophotometer (NanoDrop Technologies, Delaware, DE) and
the RNA quality was checked with an Agilent 2100 Bioanalyzer (Agilent Technologies,
Santa Clara, CA). RNA was treated with DNAseI to remove genomic DNA
contamination. Quantities from 1 μg to 100 ng of DNAseI-treated RNA were
retrotranscribed to cDNA using First Strand cDNA Synthesis Kit (Roche®, Germany)
and stored at −20 °C. Aliquots of 10 ng of total RNA were used to quantify specific
transcripts in LightCycler® 480 Real-Time PCR System (Roche) using SYBR® Green
Mix (Takara Bio Inc., Siga, Japan) and specific primers designed from reference genes
AJ517687 (encoding D. polymorpha S3 ribosomal protein) and AJ506742 (encoding
putative multixenobiotic resistance protein with high degree of identity with pglycoprotein, P-gp1; Tutundjian and Minier, 2007), using the Primer Express 2.0
software (Applied Biosystems, Foster City, CA). PCR efficiency values for both genes
were calculated using at least five serial 1:3 dilutions for each gene ( Pfaffl, 2001);
primer sequences were refined until obtaining efficiency values better than 95%. Primer
sequences: S3 forward: 5′-CAGTGTGAGTCCCTGAGATACAAG-3′; S3 reverse: 5′AACTTCATGGACTTGGCTCTCT-3′; P-gp1 forward: 5′CACCTGGACGTTACCAAAGAAGATATA-3′; P-gp1 reverse: 5′TCACCAACCAGCGTCTCATATTT-3′. The corresponding amplicons were
sequenced in an Applied Biosystems 3730 DNA Analyzer (Applied Biosystems) and
compared to the corresponding references in GenBank by the BLAST algorithm at
NCBI server ( Relative mRNA
abundances of both reference (S3) and test (P-gp1) genes were calculated from the
second derivative maximum of their respective amplification curves (Cp, calculated by
triplicates). Ribosomal protein S3 was selected as reference gene after examination of
its variability among all available samples (not shown). Variations on P-gp1 transcript
abundance were calculated as ΔCp values (ΔCp = CpS3 − CpP-gp1). To facilitate the
interpretation of results, these values were expressed as mRNA copies of target gene per
1000 copies of the reference gene mRNA (‰ of reference gene, 1000 × 2ΔCp).
2.4. Efflux transporter assays
Efflux transporter activities were determined using the fluorescent dyes RhB and CaAM in the absence and presence of mammalian transporter inhibitors VER and
REV205, specific for P-gp1; MK571, specific for MRPs; CsA, specific of both P-gp1
and MRP transporters; and with the musk Cel, which according to previous studies
strongly inhibits P-gp1 transporter activities in mussels (Luckenbach and Epel, 2005). If
transporters are active little dye accumulates inside the cells and fluorescence is weak. If
transporter activity is blocked by inhibitors accumulation of dye increases in the cells
and fluorescence levels are higher (Neyfakh, 1988). Test solutions were prepared in
ASTM hard water with 5 μM RhB or 1 μM of Ca-AM. Preliminary experiments
performed with and without the model transporter inhibitor CsA at 10 μM showed that
these selected substrate dye concentrations provided the greatest fold increase of
accumulated dye relative to controls (supplementary material Fig. 1). Stocks of the test
compounds or dyes dissolved in water (RhB), DMSO (Ca-AM, MK571, REV205) or
ethanol (CsA, VER, Cel) were added to ASTM water to the desired concentration, and
all solutions were adjusted to 1 mL/L ethanol or DMSO along with a carrier control.
Five to seven exposure treatments plus control and solvent controls were carried out for
each model inhibitor. Treatments were replicated at least four times. Assays were
initiated with 6000 larvae incubated in 50 mL glass vials in a volume of 40 mL at 20 °C
in the dark under gentle shaking and lasted 1 h 30 min. Following exposure, larvae were
filtered using a 20 μm nylon mesh, washed and resuspended in 5 mL of clean ASTM.
The fluorescence of approximately 1000 larvae was then measured in a microplate
fluorescence reader (Synergy 2, BioTek, USA) using an excitation/emission wavelength
of 535/590 and 480/530 for RhB and Ca-AM, respectively. Measurements of each
replicate were run in triplicate and were corrected for background fluorescence levels of
resuspended ASTM water. Final results were reported as fluorescent units per larva,
which were counted using a 1 mL Sedgewick rafter chamber (Pyser-Sgi, UK) and a 10×
magnification microscopy (Nikon Abbe Labphot – 2 bright field microscope, Japan).
Additionally, the fluorescence of larvae was visually evaluated using epifluorescence
microscopy (Zeiss Axiophot, West Germany) including a red and green range filter
(BP-546 and LP-590) and an Olympus DP70 digital camera.
2.5. Toxicity assays
Two independent sets of experiments were performed. For single substances, 6–8
concentrations of vinblastine as toxic ABC transporter substrate and VER, Cel, CsA,
REV205 and MK571 as ABC transporter inhibitors were used to obtain accurate
concentration–response curves that were used to define the studied combinations and to
predict mixture toxicity responses. In a second set of experiments, binary toxicity
analyses were carried out with the model MXR toxicant vinblastine and each of the
above mentioned inhibitor compounds. The toxicity of the mixtures was determined
using a fixed ratio design and using from 6 to 8 exposure levels. While keeping the
mixture ratio constant, the total concentration of the mixture was varied so that a
complete concentration–response relationship of the mixture could be described
experimentally. The components of the mixtures were mixed in a ratio of the EC50s of
the individual compounds. This is an often used design that leads to so-called equitoxic
mixtures, and permitted confrontation of observed and predicted responses following
the (CA) and (IA) conceptual models (Altenburger et al., 2003 and Barata et al., 2006).
Furthermore, at a mixture concentration that equals the sum of 1/n of the EC50s of the
individual components, the CA model predicts exactly 50% effect of the mixture (see
Eq. (3)).
Within each experiment, additions of the test substance or mixture were done in
triplicate. Embryonic development tests were initiated with fertilized eggs (50 eggs/mL)
until larvae reached the D-veliger stage (approximately 48 h after fertilization at 20 °C;
Ackerman, 1995). After the incubation period, embryos were fixed with lugol, allowed
to settle for 24 h at 4 °C and then counted using a 1 mL Sedgewick rafter chamber
(Pyser-Sgi, UK). The total number of normal D shape veliger larvae was then recorded
and compared to that of control treatments using a ×10 magnification microscopy
(Nikon Abbe Labphot – 2 bright field microscope, Japan) to determine the median
effective concentration (His et al., 1997).
2.6. Data analyses
One way ANOVAs of relative mRNA expression were performed using ΔCp values, as
this parameter followed normal distributions, as assessed by the Kolmogorov–Smirnov
test. Transporter efflux activity across embryos and larval stages was evaluated
comparing the levels of fluorescence in the absence and presence of 10 μM CsA plus
controls using one way ANOVAs. Prior to analyses data were log transformed to meet
ANOVA assumptions of normality and variance homoscedasticity. Concentration–
response RhB, Ca-AM accumulation curves in presence of the studied chemosensitizers
were analysed using ANOVA and regression analyses. Firstly, one way ANOVAs
followed by one-side post hoc Dunnet's test were used to identify the lowest
concentration of model inhibition increasing significantly the accumulation of dye
relative to controls (LOEC). Secondly, IC50 values were obtained by fitting data to
classical sigmoidal four parameter dose–response model (Eq. (1))
where R is the response, b represents minimum of the response, a represents maximum
of the response, k is shape parameter, ci is logarithm of inhibitor concentration and
IC50 is the concentration of chemosensitizers that corresponds to 50% of the maximal
In toxicity assays, EC50s and their confidence intervals were obtained by fitting
responses relative to control treatments (R) to the non-linear allosteric decay regression
model depicted in Eq. (1):
where R(ci) – proportional biological response at concentration ci relative to controls; ci
– concentration of the toxic substance (i); EC50 – the half saturation constant (i.e. the
concentration that caused an inhibition of 50% in the biological process); k – decay
The allosteric decay model was selected to fit the data obtained since it can describe non
linear type responses (Barata et al., 2006). In Eq. (2), the EC50 and its 95% confidence
intervals are regression parameters and hence can be calculated by the least square
method (Barata et al., 2006). Model accuracy was assessed by using the adjusted
coefficient of determination (r2) and by analysing residual distribution ( Zar, 1996).
Significance of the entire regression and regression coefficients were determined by
analysis of variance (ANOVA) and Student t-tests, respectively ( Zar, 1996).
Predicted values for mixture combinations considering the CA and IA model were
determined following a previous established procedure (Eqs. (3) and (4); Barata et al.,
2006). Following the CA model and considering a mixture of n chemicals, where each
chemical contributes to the overall toxicity proportionally to its EC50i (expressed as
TUi), expected responses of Eq. (1) can be calculated as:
calculating k′ as the geometric mean of the ki obtained for each chemical of the mixture
weighted by its relative toxicity (TUi). This approach assumes that the response in a
mixture of n chemicals is proportional to the addition of the equi-effective individual
concentrations of its constituents. Thus by solving Eq. (3) it is possible to obtain
expected mixture responses based on the CA model considering nonlinear allosteric
decay biological responses of the n mixture constituents.
Alternatively, expected responses (Rmix) of an n-compound mixture by the IA model
can be obtained directly by Eq. (3), sensu Bliss, 1939 and many others ( Altenburger et
al., 2003 and Barata et al., 2006).
where ci is again the concentration of the ith component and R(ci) is the proportional
response relative to control treatments of the compound applied alone.
Using Eqs. (3) and (4), then it is possible to plot predicted combine toxicity responses
for similarly and dissimilarly acting chemicals according to the CA and IA concepts,
respectively, and hence to compare them with the observed responses.
The adequacy of CA and IA models to predict combined toxicity of the studied
mixtures was tested applying regression model between predicted and observed
toxicities. A good agreement would result in a linear regression line with slope of one
and intercept of zero.
Regression curves were obtained with the aid of SigmaPlot 2002 (1086-2001 SPSS Inc.)
and the rest of analyses with the statistical software SPSS 17 (SPSS Inc., 2002,Wacker
Drive, Chicago, IL, USA) using the least square method. Wherever the inspection of
residuals evidenced variance heteroscedasticity, the estimates were corrected using the
weighted – least-square method, where the weight was proportional to the reciprocal of
the variance.
3. Results
3.1. mRNA expression of a P-gp1 like gene in mussel embryos and larvae
P-gp1 mRNA was barely detected in D. polymorpha eggs, but its abundance increased
by a factor of 10 at the first day after fertilization ( Fig. 1). These levels decreased
moderately in the two consecutive days, although maintaining values significantly
higher (3–4 fold) than egg levels ( Fig. 1).
3.2. Accumulation rates of RhB and Ca-AM dye as measures of MXR transport
activity in mussel embryos and larvae
Differences in accumulation rates of RhB and Ca-AM in mussel larval stages in the
absence and presence of CsA were visible with fluorescence microscopy (Fig. 2). In the
presence of CsA, trochophora (1 d) and veliger (2, 3 d) larvae appeared brighter
indicating (Fig. 2 E,F,J) increased accumulation of RhB or Ca-AM compared to larvae
treated with the studied dyes only (Fig. 2 B,C,H). These differences of RhB and Ca-AM
accumulation due to CsA were not seen in the unfertilized eggs (Fig. 2 D,I). Images K,
L, M of Fig. 2 illustrate the typical shape of the studied eggs and larval stages taken
with bright field.
RhB and Ca-AM accumulation in zebra mussel larvae was quantified using a
fluorescence microplate reader. Whereas RhB and Ca-AM fluorescence levels of
unfertilized eggs were unchanged in the presence of 10 μM CsA (P < 0.05, Fig. 3 A,B),
trochophora (1 d old) and veliger (2, 3 d old) larvae showed a statistically significant (P
< 0.05) increase in fluorescence compared to the controls ( Fig. 3 A,B).
RhB and Ca-AM accumulation was further quantified in veliger larvae that were
exposed to increasing levels of transporter inhibitors VER and REV205 (specific for Pgp), MK571 (specific for MRPs), CsA (specific inhibitor of both transporter systems),
or with the musk Cel. The effects of model transport inhibitors on the
accumulation/retention of RhB and Ca-AM are shown in Fig. 4 A and B and estimated
IC50 values, maximal fold inhibitions and LOEC are presented in Table 1. One way
ANOVAs performed on relative fluorescence units indicated that all inhibitors led to the
significant (P < 0.01) increase in the accumulation of RhB or Ca-AM relative to control
treatments. The lowest concentration effect on dye accumulation levels relative to
controls (LOEC) for most inhibitors occurred at 5 μM except for Cel that increased CaAM accumulation at 10 μM ( Table 1). Regression analyses denoted that all tested
inhibitors were able to accumulate Ca-AM in a concentration-dependent manner but
only CsA and REV205 did so for RhB ( Table 1). For RhB both CsA and REV205
showed equivalent levels of inhibitory potential (their 95% CI of IC50 overlap, Table 1)
and fold increase in accumulated dye. For Ca-AM, REV205 and VER were found to be
the most potent inhibitors with IC50 values of 4.2 and 4.7 μM, respectively, whereas
Cel, MK571 and CsA had the lowest potency (IC50s of 7.8, 8.2 and 9.8 μM,
respectively). The greatest maximal fold inhibition of Ca-AM occurred under MK571
exposure (23.4 fold), followed by CsA (6.1 fold), REV205 (5.1 fold), Cel (2.1 fold) and
VER (1.7 fold).
3.3. Toxicity
3.3.1. Bioassay performance
The percentage (mean ± SE) of fertilized eggs of D. polymorpha that reached the D
shape veliger stage in control and solvent control treatments after 48 h of incubation
were 79.8 ± 8.8 and 74.6 ± 6.5%, respectively. Student's t tests did not reveal significant
(P < 0.05) differences between the treatments (t tests = 1.45; df = 16; p = 0.052).
REV205 and CsA were only marginally toxic to D. polymorpha embryos at exposure
concentrations below 10 and 20 μM, respectively ( Fig. 5B). Concentration response
curves depicted in Fig. 5 denoted substantial differences in sensitivities of exposed
embryos to the tested drugs and transporter inhibitors. In all cases regression models
were significant (P < 0.05) and were able to predict accurately the observed responses
(r2 > 0.8; Table 2). The toxic model P-gp1 substrate vinblastine was 10- to 30-fold more
toxic than the studied transporter inhibitors ( Fig. 5A). Note also that vinblastine
toxicity to zebra mussel embryos varied slightly in the two studied consecutive years.
The toxicity of vinblastine co-administered with the different transporter inhibitors was
used to study the role of P-gp1 and MRP like transporters in protecting embryos to
environmental toxicants (Fig. 6). For those model inhibitors that were also toxic to
embryos (VER, MK571 and Cel) combined toxicity of equitoxic mixtures of vinblastine
with VER (Fig. 6A), MK571 (Fig. 6B) or Cel (Fig. 6C) were used. Observed joint
toxicity was greater than additive as predicted by the CA and IA concepts in the binary
combinations of vinblastine with VER and MK571 but not with celestolide (Fig. 6A–
C). A thorough analysis of the equitoxic mixtures can be obtained by plotting and
analysing together predicted versus observed responses according to the CA and IA
models. Additivity and a good agreement between observed and predicted joint toxicity
response against the CA and IA concepts would result in a linear regression line with
slope of one and intercept of zero (for further details see Fig. 2 and explanation
provided in supplementary material). For binary mixtures involving vinblastine with
VER or MK571, IA or CA models could not accurately predict joint toxicity of binary
mixtures (elevations and/or slopes were different than 0 and/or 1, respectively, Table 3).
Joint toxicity of vinblastine with Cel was additive and could be accurately predicted by
both IA and CA concepts (elevations and/or slopes were not significantly different than
0 and/or 1, respectively, Table 3, Fig. 6C).
For inhibitors showing only marginal toxicity to embryos, vinblastine was coadministered with a concentration known to inhibit substantially inhibit MXR
transporters (i.e. 5 μM REV205 and 10 μM CsA). Combined toxicity of vinblastine
coadministered with CsA or REV205 was also greater than that of vinblastine alone
(Fig. 6 D,E).
The degree of synergism of the studied compounds was further analysed comparing the
EC50 of the mixture with that of the predicted toxicity by the CA model (for VER,
MK571, Cel) or the toxicity of vinblastine alone (REV205, CsA). The results depicted
in Table 2 indicated that CsA and REV205 showed the greatest synergistic effects on
vinblastine toxicity, followed by MK571 and VER.
4. Discussion
Reported studies indicated that MXR efflux transporters constitute a general mechanism
of tolerance to environmental pollutants in adult stages in many aquatic species. There
is, however, less information on the type and role that different MXR transporters may
have in early life stages of aquatic organisms and no information exists on freshwater
species. The results presented in this study indicate that the expression of a putative Pgp1 (ABCB1) gene was low in fertilized eggs increasing shortly after in 1 d old
trochophora and 2–3 d old veliger larvae (Fig. 1). To confirm the functional expression
of a P-gp1 gene, transport activity studies using different fluorescent substrates and a
wide range of inhibitors were conducted. Accumulation rates of RhB and Ca-AM, well
established substrates of mammalian P-gp1 (ABCB1), were low in control larvae stages
but not in eggs indicating high efflux rates (Fig. 2). When the previous model inhibitors
were added, an increase in substrate accumulation only occurred in 1 d old trochophora
and 2–3 d old veliger (Fig. 3), thus confirming gene transcription results. In 2 d veliger
larvae, VER, a well characterized competitive substrate of human and rat P-gp1,
inhibited the efflux of both substrates at concentrations higher than 5 μM but only in a
sigmoid concentration dependent manner with Ca-AM (Table 1, Fig. 4). VER, is known
to be a potent MXR transporter inhibitor in in vitro studies performed in mussel gills
(Luckenbach and Epel, 2005). Conversely, and in agreement with our results Zaja et al.,
2007 and Zaja et al., 2008 reported a low inhibitory potential of VER in fish hepatoma
cell lines, probably reflecting group and tissue specific differences concerning VER
sensitivity between mussel, fish, mammalian, gill and embryonic P-gp1. CsA and VER
are not fully specific inhibitors of P-pg1 and at higher micromolar concentrations they
are also known to inhibit transport proteins from the MRP (ABCC) subfamily (Haimeur
et al., 2004). Therefore, REV205, a specific inhibitor of mammalian P-gp1 (Sharom et
al., 1999) was used to additionally prove that functional P-gp1 is expressed in zebra
mussel embryos. REV205 was found to be the most potent inhibitor (the lowest IC50)
for both substrates (Table 1).
It has been reported that RhB is solely a P-gp1 substrate, but Ca-AM is most likely
substrate for MRP1 and MRP2 (ABCC2) as well (Hollo et al., 1994). These findings
from mammalian models were confirmed in our experiments in mussel embryos.
Inhibitors of MRP mediated efflux in mammalian systems such as CsA and MK571 also
led to a marked increase in the accumulation of Ca-AM in mussel embryos. It is
noticeable that CsA and REV205 (all P-gp1 inhibitors) did not change their inhibitory
properties depending on the substrate used. When the maximal level of inhibition was
considered differences were also observed between RhB and Ca-AM. REV205, CsA
and MK571 resulted in five, six and 23 fold maximal level of inhibition when Ca-AM
was used as a substrate (Table 1). The described differences in IC50 values and levels of
maximal inhibition probably reflect differences in affinities of the P-gp1 and/or MRPs
binding sites between Ca-AM and RhB.
Hamdoun et al. (2002) using calcein-AM and a model MRP inhibitor (MK571) found
that MRP transporters appeared shortly after fertilization in sear urchin embryos.
McFadzen et al. (2000) reported that accumulation of RhB in embryos of Mytilus edulis
was increased in the presence of a competitive P-gp1 transporter inhibitor (VER).
Toomey and Epel (1993) found that embryos of the marine worm Urechis campo
expressed P-gp1 like proteins and had a VER sensitive accumulation of RhB, whereas
the sea urchin larvae species (Lytechinus anamesus) did not appear to express P-gp1
proteins nor showed a RhB VER sensitive dye efflux capacity. Therefore, the study here
is in line with other studies that showed that MXR activity is in fact quite high in
embryo and larval stages of different aquatic animals.
The best line of evidence of the role of MXR in protecting embryos against pollutants
comes from the analysis of the toxicological consequences of MXR inhibition during
embryo development. Previous results have shown that inhibitors of P-gp1 efflux
transporters increased the toxicity of pollutants to embryos, which may or may not be
also substrates of MXR system (Toomey and Epel, 1993, Britvic and Kurelec, 1999 and
Smital et al., 2004). These observations confirm that various MXR efflux transporters
(P-gp1, MRP and probably others) contributed simultaneously to resistance towards
many pollutants. However, the previous studies did not test rigorously until which
extent inhibition of different types of MXR transporters may contribute to ameliorate
the toxicity of chemicals that are detoxified by MXR.
In the present study the physiological role of P-gp1 efflux transporters protecting zebra
mussel embryos against toxic chemicals was further tested assessing joint toxicity of
binary combinations of model P-gp1 efflux and MRP inhibitors with the anti-cancer
drug vinblastine, which acts by inhibiting mitosis and is also a substrate of P-gp1
transporters (Chan et al., 2004). For transporter inhibitors that were also toxic to zebra
mussel embryos (i.e. VER, MK571), joint toxicity can only be properly analysed using
established mixture toxicity protocols that allow prediction of combined toxicity of a
given mixture from that of its individual constituents (Altenburger et al., 2003 and
Barata et al., 2006). A priori the mode of action of VER and MK571 to zebra mussel
embryos is not known and thus two predictive models were used: the CA and IA ones
that allow prediction of joint toxicity of mixture constituents having similar and
dissimilar modes of action, respectively. Observed results indicate that the binary
combinations of VER and MK571 with vinblastine had a toxicity greater than additivity
according to the predictions estimated by CA and IA models (Barata et al., 2006). The
remaining model inhibitors (REV205 and CsA) were only marginally toxic to zebra
mussel embryos and thus it was possible to study synergism comparing the toxicity of
vinblastine alone with that co-administered with a single dose of the inhibitor known to
inhibit MXR transporters. The results obtained denoted a high degree of synergism of
both compounds when combined with vinblastine. Indeed CsA and REV205 increased
the toxicity of vinblastine 4.2 and 2.8 fold followed closely by MK571 (2.6 fold
increase). The rationale behind the observed synergism was that these model inhibitors
inhibited the efflux of vinblastine by P-gp1 transporters and hence the former drug
accumulated inside the embryo cells inhibiting mitosis to a much greater extent than
with vinblastine alone. Surprisingly, MK571, which is not a specific P-gp1 inhibitor,
had elevated levels of synergism (2.6 fold). These last results suggest that despite the
fact that MK571 is considered a specific inhibitor of MRP transporters, at micromolar
concentrations if may also inhibit P-gp1 as it was indicated in efflux transporter assays
with RhB at 5 μM MK571 (Table 1).
The above mentioned results and previous studies support the general hypothesis that
MXR inhibitors are a possible threat for both early and adult life stages of aquatic
organisms. From a regulatory perspective it is especially important to find out whether
or not powerful MXR inhibitors of anthropogenic origin exist in the environment.
Emerging pollutants are of special concern due to their continuous release into the
environment and, most importantly, their unknown mode of toxicity to aquatic biota. As
Luckenbach and Epel (2005) we also found that the synthetic musk fragrance Cel had
an impact on the MXR system as indicated by dye uptake assays (Table 1). However,
the inhibitory potential on the MXR system of zebra mussel embryos was
comparatively low. On the other hand Cel showed considerable toxicity to mussel
embryos, which is along the lines with findings of Wollenberger et al. (2003), who also
found particular sensitivity of early life stages of a marine copepod to musks, including
Cel. Little inhibitory potential together with high toxicity would explain why the joint
toxicity of Cel and vinblastine was not greater than additive (Fig. 6, Table 3), indicating
that Cel in our experiments had no chemosensitizing effect on mussel embryos.
In summary the results presented in this study indicated that gene expression and
associated MXR transporter activity were already significant in 1 d old trochophora
larvae in D. polymorpha. Additionally, functional and toxicity assays showed that both
P-gp1 and MRP transporters may be present in zebra mussel embryos shortly after the
fecundation of eggs and that P-gp1 transporters contributed actively to detoxify the
cytotoxic drug vinblastine. These results agree with recent studies that showed that the
MXR system of aquatic invertebrates may be based on many different transporter types.
For instance, the genome of the sea urchin Strongylocentrotus purpuratus contains 65
ABC transporters, with many being homologous to known mammalian xenobiotic
transporters ( Sodergren et al., 2006). Transcripts and activities of different xenobiotic
transporter types were also identified in tissues from bivalves ( Ludeking and Kohler,
2002 and Luckenbach and Epel, 2008), including full-length P-gp1 and MRP type
transcripts with 40–49% amino acid identity with human P-gp1 and 33–45% identity
with human MRP transporters ( Luckenbach and Epel, 2008). Although limited to a
single drug (vinblastine) and one component of the MXR transporter system (P-gp1),
the results reported in this study also indicate that predictive joint toxicity modelling
approaches can be successfully used in zebra mussel embryos to assess the toxicological
consequences of chemosensitization in an in vivo system.
This study was funded by the Ministerio de Ciencia e Inovación and Ministrerio de
Medio Ambiente, Medio Rural y Marino projects CGL2008-01898,
041/SGTB/2007/1.1 and 042/RN08/03.4 and by the Spanish and German Integrated
action project DE2009-0089. We thank the aid of two ammoniums referees to improve
this manuscript.
Ackerman, J.D., 1995. Zebra mussel life history. In: Fifth international zebra mussel
and other aquatic nuisance organisms conference, pp. 1–8.
Altenburger, R., Nendza, M., Schüürmann, G., 2003. Mixture toxicity and its
mod¬elling by quantitative structure – activity relationships. Environ. Toxicol. Chem.
22, 1900–1925.
ASTM, 1995. Standard Methods for the Examination of Water and Wastewater, 19th
edition. American Public Association/America Water Works Association/Water
Environment Federation, Washington, DC.
Barata, C., Baird, D.J., Nogueira, A.J.A., Soares, A.M.V., Riva, M.C., 2006. Toxicity of
binary mixtures of metals and pyrethroid insecticides to Daphnia magna Straus.
Implications for multi-substance risks assessment. Aquat. Toxicol. 78, 1–14.
Bard, S.M., 2000. Multixenobiotic resistance as a cellular defense mechanism in aquatic
organisms. Aquat. Toxicol. 48, 357–389.
Britvic, S., Kurelec, B., 1999. The effect of inhibitors of multixenobiotic resistance
mechanism on the production of mutagens by Dreissena polymorpha in waters spiked
with premutagens. Aquat. Toxicol. 47, 107–116.
Chan, L.M.S., Lowes, S., Hirst, B.H., 2004. The ABCs of drug transport in intestine and
liver: efflux proteins limiting drug absorption and bioavailability. Eur. J. Pharm. Sci. 21,
Daughton, C.G., Ternes, T.A., 1999. Pharmaceuticals and personal care products in the
environment: agents of subtle change? Environ. Health Perspect. 107, 907–938. Dean,
M., 2005. The genetics of ATP-binding cassette transporters. In: Methods in
Enzymology, 409–429.
Deeley, R.G., Westlake, C., Cole, S.P.C., 2006. Transmembrane transport ofendo- and
xenobiotics by mammalian ATP-binding cassette multidrug resistance proteins. Physiol.
Rev. 86, 849–899.
Epel, D., 1998. Use of multidrug transporters as first lines of defense against toxins in
aquatic organisms. Comp. Biochem. Physiol. 120 A, 23–28.
Epel, D., Luckenbach, T., Stevenson, C.N., MacManus-Spencer, L.A., Hamdoun,A.,
Smi¬tal, T., 2008. Efflux transporters: newly appreciated roles in protection against
pollutants. Environ. Sci. Technol. 42, 3914–3920.
Faria, M., López, M., Fernández-Sanjuán, M., Lacorte, L., Barata, C., 2010. Comparative toxicity of single and combined mixtures of selected pollutants among larval
stages of the native freshwater mussels (Unio elongatulus) and the invasive zebra
mussel (Dreissena polymorpha). Sci. Total Environ. 408, 2452–2458.
Germann, U.A., 1993. Molecular analysis of the multidrug transporter.
Cytotechnol¬ogy 12, 33–62.
Haimeur, A., Conseil, G., Deeley, R.G., Cole, S.P.C., 2004. The MRP-related and
BCRP/ABCG2 multidrug resistance proteins: biology, substrate specificity and
regulation. Curr. Drug Metabol. 5, 21–53.
Hamdoun, A.M., Griffin, F.J., Cherr, G.N., 2002. Tolerance to biodegraded crude oil in
marine invertebrate embryos and larvae is associated with expression of a
multixenobiotic resistance transporter. Aquat. Toxicol. 61, 127–140.
His, E., Seaman, M.N.L., Beiras, R., 1997. A simplification the bivalve embryogenesis
and larval development bioassay method for water quality assessment. Water Res. 31,
Hollo, Z., Homolya, L., Davis, C.W., Sarkadi, B., 1994. Calcein accumulation as a
fluo¬rometric functional assay of the multidrug transporter. Biochim. Biophys. Acta
11191, 384–388.
Kurelec, B., 1992. The multixenobiotic resistance mechanism in aquatic organisms.
Crit. Rev. Toxicol. 22, 23–43.
Kurelec, B.A., 1997. A new type of hazardous chemical: the chemosensitizers of
multixenobiotic resistance. Environ. Health Perspect. 105, 855–860.
Luckenbach, T., Epel, D., 2005. Nitromusk and polycyclic musk compounds as
long¬term inhibitors of cellular xenobiotic defense systems mediated by multidrug
transporters. Environ. Health Perspect. 113, 17–24.
Luckenbach, T., Epel, D., 2008. ABCB- and ABCC-type transporters confer
multixeno¬biotic resistance and form an environment–tissue barrier in bivalve gills.
Am. J. Physiol. – Regul. Integr. Comp. Physiol. 294, R1919–R1929.
Ludeking, A., Kohler, A., 2002. Identification of six mRNA sequences of genes related
to multixenobiotic resistance (MXR) and biotransformation in Mytilus edulis. Mar.
Ecol. Prog. Ser. 238, 115–124.
McFadzen, I., Eufemia, N., Heath, C., Epel, D., Moore, M., Lowe, D., 2000. Multidrug
resistance in the embryos and larvae of the mussel Mytilus edulis. Mar. Environ. Res.
50, 319–323.
Neyfakh, A.A., 1988. Use of fluorescent dyes as molecular probes for the study of
multidrug resistance. Exp. Cell Res. 174, 168–176.
Pain, S., Parant, M., 2007. Identification of multixenobiotic defence mechanism (MXR)
background activities in the freshwater bivalve Dreissena polymorpha as reference
values for its use as biomarker in contaminated ecosystems. Chemo-sphere 67, 1258–
Pfaffl, M.W., 2001. A new mathematical model forrelative quantification in real-time
RT-PCR. Nucl. Acids Res. 29, e45.
Ramachandra, M., Ambudkar, S.V., Chen, D., Hrycyna, C.A., Dey, S., Gottesman,
M.M., Pastan, I., 1998. Human P-glycoprotein exhibits reduced affinity for substrates
during a catalytic transition state. Biochemistry 37, 5010–5019.
Roepke, T.A., Hamdoun, A.M., Cherr, G.N., 2006. Increase in multidrug transport
activity is associated with oocyte maturation in sea stars. Dev. Growth Differ. 48, 559–
Sarkadi, B., Homolya, L., Szakács, G., Váradi, A., 2006. Human multidrug resistance
ABCB and ABCG transporters: participation in a chemoimmunity defense sys¬tem.
Physiol. Rev. 86, 1179–1236.
Sharom, F.J., Yu, X., Lu, P., Liu, R., Chu, J.W., Szabó, K., Müller, M., Hose, C.D.,
Monks, A., Váradi, A., Seprôdi, J., Sarkadi, B., 1999. Interaction of the P-glycoprotein
multidrug transporter (MDR1) with high affinity peptide chemosensitizers in isolated
membranes, reconstituted systems, and intact cells. Biochem. Pharma¬col. 58, 571–586.
Smital, T., Luckenbach, T., Sauerborn, R., Hamdoun, A.A., Vega, R.L., Epel, D., 2004.
Emerging contaminants – pesticides, PPCPs, microbial degradation products and natural
substances as inhibitors of multixenobiotic defense in aquatic organ¬isms. Mutat. Res. –
Fund. Mol. Mech. 552, 101–117.
Smital, T., Sauerborn, R., Hackenberger, B.K., 2003. Inducibility of the P-glycoprotein
transport activity in the marine mussel Mytilus galloprovincialis and the fresh¬water
mussel Dreissena polymorpha. Aquat. Toxicol. 65, 443–465.
Smital, T., Sauerborn, R., Picevic, B., Kurelec, B., 2000. Inhibition of multixenobiotic
resistance mechanism in aquatic organisms by commercially used pesticides. Mar.
Environ. Res. 50, 334–335.
Sodergren, E., Weinstock, G.M., Davidson, E.H., Cameron, R.A., Gibbs, R.A., Angerer,
R.C., et al., 2006. The genome of the sea urchin Strongylocentrotus purpuratus. Science
314, 941–952.
Toomey, B.H., Epel, D., 1993. Multixenobiotic resistance in Urechis-caupo embryos –
protection from environmental toxins. Biol. Bull. 185, 355–364.
Toomey, B.H., Kaufman, M.R., Epel, D., 1996. Marine bacteria produce compounds
that modulate multixenobiotic transport activity in Urechis caupo embryos. Mar.
Environ. Res. 42, 393–397.
Tutundjian, R., Minier, C., 2007. Effect of temperature on the expression of Pglycoprotein in the zebra mussel, Dreissena polymorpha. J. Therm. Biol. 32, 171–177.
Vanderploeg, H.A., Liebig, J.R., Gluck, A.A., 1996. Evaluation of different
phytoplank¬ton for supporting development of zebra mussel larvae (Dreissena
polymorpha): the importance of size and polyunsaturated fatty acid content. J. Great
Lakes Res. 22, 36–45.
Wollenberger, L., Breitholtz, M., Ole Kusk, K., Bengtsson, B.E., 2003. Inhibition of
larval development of the marine copepod Acartia tonsa by four synthetic musk
substances. Sci. Total Environ. 305, 53–64.
Zaja, R., Caminada, D., Lonncar, J., Fent, K., Smital, T., 2008. Development and
char¬acterization of P-glycoprotein 1 (Pgp1, ABCB1)-mediated doxorubicin-resistant
PLHC-1 hepatoma fish cell line. Toxicol. Appl. Pharmacol. 227, 207–218.
Zaja, R., Klobucar, R.S., Smital, T., 2007. Detection and functional characterization of
Pgp1 (ABCB1) and MRP3 (ABCC3) efflux transporters in the PLHC-1 fish hep¬atoma
cell line. Aquat. Toxicol. 81, 365–376.
Zar, J.H., 1996. Biostatistical Analysis. Prentice-Hall International, Inc., New Jersey.
Figure captions
Figure 1. Relative expression of P-pgl levels depicted as ‰ of S3 ribosomal mRNA
abundance (mean ± SE, n = 5) in D. polymorpha eggs, trochophora larvae of 1 d and D
shape veliger larvae of 2 and 3 d old. Low case letters indicate significantly (P < 0.05)
different sets of data after ANOVA and Tukey's post hoc tests.
Figure 2. Fluorescent microscopy of RhB (red colour) and calcein (green colour) in
eggs (1 h post-fertilization), trochophora (1 d; Tph) and D shape veliger (2–3 d;Vel)
larvae. Images were taken after 1.5 h incubation period with 5 μM RhA and 1 μM CaAM in absence (Ct; fist and third panel line of images) and with the presence of 10 μM
of CsA (second and fourth line of images). Images of eggs, trochophora and D shape
larvae (K, L, M, respectively) taken with bright field.
Figure 3. Efflux transporter activity for RhB (A) and Ca-AM (B) (fluorescence units per
larvae) of eggs, trochophora (1 d old) and D shape veliger larvae of 2 and 3 d. Data are
expressed as mean SE (n = 5–8). Asterisks indicate significant (P < 0.05) differences
from solvent controls within each age group following ANOVA and post hoc Dunnet's
and Tukey's multiple comparison tests.
Figure 4. Effect of transported MXR inhibitors on the accumulation of RhB (A) and CaAM (B) in D-veliger larvae. Results are expressed in fluorescence units per embryo.
Symbols and error bars are mean ± SE (n = 4). Data were fitted to a sigmoidal four
parameter dose–response model. Solvent controls (C) are also included. Horizontal axis
is depicted in log scale.
Figure 5. Percentage of D shape D. polymorpha larvae after 48 h single exposure to
vinblastine (A) and the remaining studied compounds (B). Each symbol corresponds to
a single replicate. Responses have been fitted to the allosteric decay function of
Eq. (2) (see text). Horizontal axis is depicted in log scale.
Figure 6. Percentage of D shape D. polymorpha larvae after 48 h exposures to equitoxic
binary mixtures of vinblastine with VER (A), MK571 (B) and Cel (C) and of
vinblastine co-administered with 10 μM of CsA (D) and 5 μM of REV205. For
comparison purposes the responses of larvae exposed to vinblastine alone are depicted
in graphs D and E. Each symbol corresponds to a single replicate. Responses have been
fitted to the allosteric decay function (solid line). Concentration–responses curved
predicted by the CA and IA concepts are also depicted in graphs A, B, C. Horizontal
axis is depicted in log scale.
Figure 1
Figure 2
Figure 3
Figure 4
Figure 5
Figure 6
Table 1. IC50 values (95% CI), maximal accumulation (fold increase) of RhB and Ca-AM with the model inhibitors used. The lowest significant
(P < 0.01) concentration effect in dye accumulation levels relative to controls (LOEC) is also depicted. nd, not determined by the regression
model used.
Table 2. Estimated EC50s and 95% confidence intervals (μM) of the studied compounds in zebra mussel (D. polymorpha) embryonic
development. Results from binary mixture exposures are also depicted. In binary mixtures, fold increased levels of observed EC50s relative to
those predicted according to the CA (VER, MK571, Cel) or to vinblastine alone (REV205, CsA) are depicted in brackets. r2, coefficient of
determination; N, sample size.
a.Toxicity response obtained in 2009.
b.Toxicity response obtained in 2010.
c.EC50s for REV205 and CsA were not depicted since only marginal effects were detected at 10 and 20 μM, respectively.
d.EC50 of vinblastine co-administered with 5 and 10 μM REV205 and CsA, respectively.
Table 3. Regression models fitted to the predicted versus observed embryonic development relationships obtained for the binary mixtures
following the concentration addition (CA) and independent action (IA) concepts. a, b are the intercepts and slopes of the regression models. SE:
associated standard errors; N: sample size, r2: coefficient of determination. Probability levels of t-tests (t) testing for departures of a = 0 and b = 1
are depicted. Within each mixture, different letters indicate significant differences among slopes and elevations between IA and CA concepts
following ANCOVA. No significant (ns) P ≥ 0.05; * 0.01 < P < 0.05; ** 0.001 < P < 0.01; *** 0.001 < P.
a.Linear model y = a + bX.
b.Power model y = aXb.