Effects of exposure to pharmaceuticals (diclofenac and carbamazepine) spiked sediments in the midge, Chironomus riparius (Diptera, Chironomidae)

Science of the Total Environment 609 (2017) 715–723

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Effects of exposure to pharmaceuticals (diclofenac and carbamazepine) spiked sediments in the midge, Chironomus riparius (Diptera, Chironomidae) Elena Nieto a, Carmen Corada-Fernández b, Miriam Hampel c, Pablo A. Lara-Martín c, Paloma Sánchez-Argüello d, Julián Blasco a,⁎ a

Instituto de Ciencias Marinas de Andalucía (ICMAN-CSIC), Campus Universitario Rio San Pedro, 11519 Puerto Real, Spain Facultad de Ciencias del Mar y Ambientales, Universidad de Cádiz (UCA), Campus Universitario Río San Pedro, 11510 Puerto Real, Spain c Instituto Universitario de Investigación Marina (INMAR), Campus de Excelencia Internacional del Mar (CEI•MAR), Universidad de Cádiz, Av. República Saharaui s/n, 11510 Puerto Real, Cádiz, Spain d Laboratorio de Ecotoxicologia, Departamento de Medio Ambiente, INIA, Ctra, A Coruña km 7, 28040 Madrid, Spain. b

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• The effect of sediment sorbed pharmaceuticals (diclofenac and carbamazepine) has been assessed • Chironomus riparius, first instar stage has been exposed –in a long-term experiment- to pharmaceutical spiked sediments • Growth, development, sex-ratio and emergence have been selected as endpoints • Adverse effects of both pharmaceuticals have been reported at the exposure conditions • Chronic exposure tests are useful to assess the environmental risk of these compounds

a r t i c l e

i n f o

Article history: Received 17 April 2017 Received in revised form 19 July 2017 Accepted 19 July 2017 Available online xxxx Keywords: Pharmaceuticals Diclofenac Carbamazepine Sediment toxicity Chronic toxicity Chironomus riparius

a b s t r a c t Human and veterinary pharmaceuticals and degradation products are continuously introduced into the environment. To date, there is a lack of information about the effects of pharmaceuticals in spiked toxicity tests with nontarget organisms. In this study, we have evaluated the effects of exposure to two common pharmaceuticals in the midge Chironomus riparius in spiked sediment experiments. The selected pharmaceuticals are the nonsteroidal anti-inflammatory drug (NSAID): diclofenac (DF) and the anti-depressant drug carbamazepine (CBZ). In order to assess the effects of the pharmaceuticals, a chronic toxicity test with the midge was carried out. The endpoints survival, growth and developmental stage by means of biomass, were measured after 10 days, and emergence rates and sexratio (male/female) were measured after 21 days of exposure. Significant mortality was observed in organisms at day 10 with a 40% of larvae surviving in the highest exposure concentration of CBZ. DF decreased the emergence ratio with respect to the controls in organisms exposed at concentrations of 34.0 μg·g−1 whereas CBZ reduced the growth of the midges (30,6% with respect to the control) and induced a significant change in sex-ratio at concentrations of 31.4 μg·g−1. The results obtained in the present study indicate possible adverse effects on aquatic invertebrates, which should be taken into account for environmental risk assessment of pharmaceutical compounds in sediments. © 2017 Elsevier B.V. All rights reserved.

⁎ Corresponding author at: Instituto de Ciencias Marinas de Andalucía (ICMAN-CSIC), Spain. E-mail address: [email protected] (J. Blasco).

http://dx.doi.org/10.1016/j.scitotenv.2017.07.171 0048-9697/© 2017 Elsevier B.V. All rights reserved.

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1. Introduction Due to globally increasing consumption, human and veterinary pharmaceuticals are being found in several environmental matrices in the majority of countries around the world (Daneshvar et al., 2010; Sim et al., 2010; Osorio et al., 2012; Writer et al., 2013; Valdés et al., 2014). The occurrence of pharmaceuticals in the environment has only relatively recently become an issue of concern, with the development of sensitive analytical tools that made possible the detection at low concentrations. This occurrence has been acknowledged as potentially hazardous to ecosystems (Kolpin et al., 2002; Fent et al., 2006). Although these compounds are being detected at concentrations in the range of ng·L−1 to μg·L−1 (Farré et al., 2001; Sacher et al., 2001; Gros et al., 2006; Martínez Bueno et al., 2007; Lara-Martín et al., 2014), the occurrence in the aquatic environment can affect water quality and potentially impact drinking water supplies, ecosystems and human health (Yuan et al., 2009; Sirés and Brillas, 2012), particularly since pharmaceutical compounds are designed to have a specific acute effect in the patient. However, little is known about the potential chronic effects on wildlife. In order to assess the potential effects that the occurrence of pharmaceuticals could have on wildlife, we have selected two of the most commonly found pharmaceuticals in water bodies: the non-steroidal anti-inflammatory drug (NSAID) diclofenac (DF), as well as the anti-epileptic and mood stabilizing drug carbamazepine (CBZ). Due to previously observed effects in non-target organisms (Ericson et al., 2010; Oviedo-Gómez et al., 2010; Lee et al., 2011; Schmidt et al., 2011), DF has already been included in the EU watch list of substances as possible candidate to be part of the priority list of substances (Directive 2013/39/EU, 2017). Likewise, CBZ has been proposed as marker for anthropogenic contamination (Clara et al., 2004) due to its poor elimination rate in waste water treatment plants (WWTPs) and its frequent occurrence in aquatic environments (Gros et al., 2009). Once discharged into the sewage system, mainly by excretion, these compounds generally reach WWTPs where they are partly or entirely removed. For CBZ, the average removal rate in WWTP is low and varies between 7 and 8% (Heberer, 2002a; Clara et al., 2004), whereas for DF the removal rate is around 22–39% (Paxéus, 2004; Bendz et al., 2005). Concentrations of these compounds have been analyzed in WWTP effluents with values comprised between 0.01 and 5.92 μg·L− 1 and 0.10 and 0.8 μg·L−1, for DF and CBZ, respectively (Öllers et al., 2001; Martínez Bueno et al., 2007; Gros et al., 2009). In freshwater rivers, DF and CBZ have been found at concentration ranges between 0.02 and 5.45 μg·L−1 and 0.001 and 3.7 μg·L−1, respectively (Öllers et al., 2001; Drillia et al., 2005; Gros et al., 2009). However, only little information is available about the concentration levels of these pharmaceuticals in sediments. Concentrations of 2.68 and 8 ng·g−1 have been reported for CBZ and DF, respectively in river sediments (Löffler and Terner, 2003; Silva et al., 2011), and Vazquez-Roig et al. (2010) reported CBZ concentrations from 1.81 to 6.85 ng·g−1 in sediments of marsh areas. Aquatic ecosystems near to municipal and industrial effluent discharges generally present sediments that contain a great variety of chemical contaminants (Liber et al., 1996). Sediment sorbed contaminants in general can affect benthic organisms via interstitial and overlying water (Gilroy et al., 2012) as well as through ingestion of sediment particles (Soares et al., 2005a, 2005b). Although not being lethal, an organism's ability to function normally in an ecosystem may be impaired at sublethal contaminant concentrations (Gerhardt et al., 2002). The freshwater macroinvertebrate Chironomus riparius, is a common midge in freshwater ecosystems and plays an important role in the aquatic food chain being a major food source for both fish and other macroinvertebrates (Lee and Choi, 2007). It is highly sensitive to the presence of environmental contaminants and is commonly employed in standard tests to assess the toxicity of sediment-associated contaminants (ASTM, 2005). C. riparius larvae pass through four instars, which are defined as developmental stages of arthropods between each moult, until sexual maturity is reached (Allaby, 2006). Thus, the juvenile

and pupal stages are spent in the sediments feeding on organic detritus making them a relevant stage for assessing sediment contamination. Towards the end of the fourth instar, the larvae cease feeding and pupation begins (Benoit et al., 1997). The emergence occurs immediately after pupae swim to the water surface (Environment Canada, 1997) being this development stage an useful endpoint to determine effects in the development of the organisms. Test procedures with C. riparius are widely standardized (ASTM, 1997, 2005; OECD, 2004a, 2004b, 2010) and the organism has been widely used in short-term assessments of contaminated sediments, mainly due to its easy culture and handling under laboratory conditions and the fact that it remains in close contact with the sediment during larvae development (Benoit et al., 1997). In the past years, the number of studies that have reported the negative effects of pharmaceutical compounds on benthic macroinvertebrates from contaminated sediments employing midges has increased significantly (Oetken et al., 2005; Nentwig, 2007; Sánchez-Argüello et al., 2009; Gilroy et al., 2012; López-Doval et al., 2012; Heye et al., 2016) underlining the rising concern about the presence of this kind of compounds in the environment by the general population and regulatory organisms. Among the analyzed responses in these studies, the authors observed decrease of emergence, reduced growth, increase of the biomass and increase of the female/male ratio in spiked sediment exposure experiments at low concentrations with different pharmaceuticals. On the background of the current information on the effects of sediment associated pharmaceuticals, this study aims to address current knowledge gaps related to the sublethal effects of sediment sorbed pharmaceuticals on non-target organisms. For this purpose, we have selected the freshwater midge, C. riparius, due to its high sensitivity and the fact that it undergoes part of its life-cycle within the sediment and two model pharmaceutical compounds (DF and CBZ) on the basis of the high consumption rates and presence in WWTP effluents (Kolpin et al., 2002; Roberts and Thomas, 2006; Carballa et al., 2008). The exposure route was through spiked sediments and the analyzed responses included: survival, growth (biomass) and development stage after 10 days of exposure, as well as emergence rate, cumulative emergence and sex-ratio (male/female) at days 15–21, of the surviving organisms. This study aims to contribute to fill the gap of information about the impact of these two pharmaceuticals sorbed on sediments for their consideration in sediment quality guidelines. 2. Material and methods 2.1. Chemicals DF (CAS no. 15307-86-05) and CBZ (CAS no. 298-46-4) were purchased from Sigma-Aldrich (St. Louis, USA). 2.2. Organisms C. riparius were obtained from our laboratory culture at the facilities of the Institute of Marine Sciences of Andalusia following the recommendations described by OECD guidelines 218 and 219 (OECD, 2004a, 2004b) and Sánchez-Argüello et al. (2005). Briefly, C. riparius were allocated in glass aquaria with dechlorinated water, under static flow and gentle aeration, a photoperiod of 16:8 h (light:dark) and a temperature of 20° ± 1 °C. Egg masses were placed in a glass beaker and green algae (Chlorella vulgaris) were added ad libitum. After hatching (2–3 days), b48 h old larvae were transferred to 25 L glass aquaria containing sandy sediment. On the 7th day, larvae were removed and placed into a clean rearing container which was covered with plastic mesh to avoid escape of the emerging adults. Commercial fish food, Tetramin®, was added to the culture chamber three times a week. Water was replaced weekly. Under these conditions adult emergence occurred on the 13–15th day. Emerging adults were left in the rearing aquaria to

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permit mating and oviposition (2–3 days). After removal of adults by aspiration, egg ropes were collected and then used to initiate the next generation.

2.3. Sediment spiking Sediments were collected from a pristine site in the Guadalete river (36°47′54.93″N-5°19′53.74″W). Sediment was transported to the laboratory and stored at 4 °C in the dark until spiking. Before spiking, the sediment was sieved through a 2 mm mesh and homogenized. A portion of the sediment was removed to be used in the physicochemical characterization (humidity 13.3%; loss of ignition 3.2% and organic carbon 1.10%). Humidity was measured on basis of the UNE 77311 procedure (UNE, 77311, 2000), organic carbon according to Gaudette et al. (1974) modified by El Rayis (1985) and loss of ignition was estimated in a muffle furnace at 550 °C and gravimetric determination. Sediment was then frozen in liquid nitrogen over 24 h to remove indigenous macrofauna. Sediment was spiked adapting the method described by Gilroy et al. (2012) with 0.2% (v/v) methanol or DMSO in the case of DF and CBZ, respectively. A varying volume of each solution was added in increasing concentrations of each compound to achieve the targeted sediment exposure concentrations (0.12; 1.2; 12.0; 30.0 and 60 μg·g−1). Sediments were mixed three times a day during 72 h, allowed to rest for further 7 days at room temperature and stored at 4 °C until the beginning of the experiment, in order to provide an adequate mixing while minimizing the biodegradation of the compounds (ASTM, 2004).

2.4. Experimental design Bioassays were carried out adapting the standard procedure ASTM E 1706-05 for long-term sediment toxicity tests. After 10 days survival, growth and developmental stage, and during day 15 to day 21, adult emergence and sex-ratio were assessed. For each compound, seven conditions (sediment control + solvent control + five concentrations), and six replicates (three for survival, growth and development stage and three for adult emergence and sex-ratio) were used in each test. On the day before the start of the experiment, 80 g wet weight (w/w) of spiked sediment and subsequently 320 mL of dechlorinated water were added to each beaker. After a period of 24 h to allow the settling of the sediment, twenty larvae of C. riparius (b3-d old) were added to each replicate. Overlying water was renewed every day substituting 90% of the total volume with dechlorinated water. The tests were performed at 20° ± 1 °C with a photoperiod of 16:8 (light: darkness). Midges in each beaker were fed daily with 1.5 mL of a 4 g·L− 1 fish food flakes suspension. Dissolved oxygen, pH and conductivity of the overlying water were measured at the beginning and end of the experiment and twice a week in at least two replicates per treatment.

2.5. Evaluation of selected toxicity endpoints Larval survival, growth and developmental stage were determined on day 10. Sediments of three replicates were passed through a 300 μm sieve to retain the larvae. Surviving animals were removed, counted and collected for estimation of developmental stage and growth as dry weight (DW). For determination of DW, the larvae were dried at 60 °C for 24 h and weight estimated with a precision balance Mettler AE166 (± 0.1 mg). Developmental stage was determined by measuring the width of the cephalic capsule and the first segment with a Nikon SMA 800 stereomicroscope. From day 15 onwards, adult emergence and sex-ratio (defined as n° males/n° females) were recorded daily in the remaining three replicates according to the procedures described by OECD (2010).

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2.6. Sediment extractions For the determination of effective exposure concentrations, a sediment sample from each concentration was collected at the beginning, after 10 days of exposure and at the end of the experiment, and stored at −20 °C until determination. Analysis of pharmaceuticals in sediment samples was performed by pressurized liquid extraction (PLE) followed by solid-phase extraction (SPE) and ultra-performance liquid chromatography-triple quadrupole mass spectrometry (UPLC-QqQMS/MS). Briefly, pharmaceuticals were extracted from sediments by PLE using water at 100 °C and 1500 psi. Subsequently, the aqueous extracts were purified and pre-concentrated according to the method reported by Baena-Nogueras et al. (2016). SPE was carried out using 200 mg Oasis HLB cartridges activated with 8 mL of metanol and 8 mL of Milli-Q water at a flow rate of 1 mL min−1. Subsequently, 30 mL of aqueous extracts were passed through the cartridges at the same flow and the cartridges were rinsed with 8 mL of Milli-Q water and dried using a vacuum for 20 min in order to remove the excess of water. The analytes retained were eluted from the HLB sorbent with 10 mL of methanol at a flow rate of 1 mL min−1. Methanolic extracts were evaporated to dryness under a nitrogen current and finally, samples were reconstituted with 1 mL methanol-water (25:75, v/v). Prior to injection, extracts were filtered using PTFE filters (0.45 μm from Teknokroma Corporation, Spain). Separation, identification and quantification of the compounds were carried out using UPLC-QqQ-MS/MS (Bruker EVOQ Elite system) with autosampler and C18 analytical column of 100 mm × 2.1 mm and 2 μm particle size. Calibration curves were constructed for target compounds in the range of 0.1–100 μg·L−1. The compounds were identified by comparison of their retention times of two transitions of each analyte (one for quantification and one for confirmation) and their ion ratio. Internal standards were added to the vials before injection to correct possible fluctuations in the MS signal. The method limits of detection (mLOD) were below 1 ng·L−1 for the target compounds. All the data were processed using peak areas with Bruker MS Workstation. Recoveries were 86 and 101%, for DF and CBZ, respectively. 2.7. Risk assessment of pharmaceuticals Environmental risk for sediment sorbed pharmaceuticals (diclofenac and carbamazepine) was assessed using the Risk Quotient (RQ) based on the ratio between predicted environmental concentration (PEC) and predicted no effect concentration (PNEC). RQ ¼

PEC PNEC

We have employed the most sensitive measured endpoint of Chironomus riparius test to estimate PNEC (Trombini et al., 2016). The RQ for a mixture of both pharmaceuticals can also be calculated as the sum of PEC/PNEC. We applied an assessment factor (100) in the RQ quantification as chronic toxicity endpoints (NOEC) were used (ECB, 2003) RQmixture ¼

n X PECi PNECi i¼1

For mixtures, risk levels were established as: low (0.01–0.1), medium (0.1–1.0) and high (N 1) (Hernando et al., 2006). 2.8. Statistical analysis Data analysis was carried out by means of the statistical software package IBM SPSS Statistics Package 21.0. The Levene test was first applied to evaluate the homogeneity of variances of the results. Significant differences were determined by one-way ANOVA (with treatment as a

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factor) at a significance level lower than 0.05 (p b 0.05). When significant differences were found for the factor treatment, the Dunnett's test (significance, p b 0.05) was used as post hoc comparison between controls and treatment. 3. Results 3.1. Exposure assessment Nominally targeted sediment concentrations for both DF and CBZ were 0.12; 1.2; 12.0; 30 and 60 μg·g−1. Measured concentrations at the start of the experiment are displayed in Table 1. Initial exposure concentrations of the spiked sediments ranged between 0.1 and 34.04 and 0.2 and 31.4 μg·g−1 for DF and CBZ, respectively. During the experiment, sediment concentrations decreased with exposure time, reaching final concentrations ranges between (n.d and 2.4) for DF and (n.d and 6.3 μg·g−1) for CBZ. 3.2. Effect assessment Due to the lack of significant differences in the selected endpoints between sediment control and solvent control, only the sediment control was considered for statistical analysis. In both sediment and solvent controls, survival was N 70%, as recommended by ASTM E 1706-05 for test acceptability in all assays. At day 10, 92 and 40% of larvae had survived in the highest exposure concentration of DF and CBZ, respectively (Fig. 1). Growth was significantly reduced after exposure to the highest concentration of CBZ in 30.65% with respect to the control (Fig. 2). With regard to the developmental endpoint instar measured by means of mean head capsule width ± SD, all surviving organisms exposed to DF and CBZ were found within the range of the fourth instar. However, although not statistically significant, a decrease in head capsule and 1st segment widths could be observed with exposure to increasing concentrations of DF (Table 1). Changes in population dynamics could take more relevance with the results obtained for CBZ in this study with respect to the endpoint sexratio. After exposure to high concentrations of this compound, the proportion of males observed increased (Fig. 3). On the contrary, DF did not change the sex-ratio in exposed organisms significantly (Fig. 3). Regarding the results obtained after 21 days of exposure, emergence rates were significantly lower in organisms exposed at 34.0 μg·g−1 DF (p = 0.01) compared with the control (Fig. 3). Emergence started on day 15 in sediments spiked with DF and CBZ. Percentages of cumulative emergence varied between 23% and 87% in the case of DF where at 34.0 μg·g−1 a significant difference was found with respect to the control (p = 0.003) (Fig. 4). The sex-ratio was statistically greater in organisms Table 1 Mean values (±SD) of head capsule width and 1st segment width (mm) of Chironomus riparius after 10 days of exposure to diclofenac (DF) and carbamazepine (CBZ). Measured concentrationa (μg g−1)

Instar

Head capsule width (mm)

1st segment width (mm)

DF Control 0.1 0.7 11.1 27.8 34.0

4th 4th 4th 4th 4th 4th

0.54 (±0.01) 0.54 (±0.07) 0.52 (±0.06) 0.51 (±0.06) 0.45 (±0.07) 0.44 (±0.06)

0.80 (±0.06) 0.77 (±0.06) 0.75 (±0.06) 0.77 (±0.05) 0.72 (±0.07) 0.73 (±0.07)

CBZ Control 0.2 1.6 13.6 38.7 31.4

4th 4th 4th 4th 4th 4th

0.42 (±0.05) 0.43 (±0.05) 0.42 (±0.05) 0.42 (±0.05) 0.43 (±0.05) 0.41 (±0.05)

0.63 (±0.09) 0.62 (±0.09) 0.62 (±0.09) 0.63 (±0.10) 0.61 (±0.10) 0.62 (±0.10)

a

Data correspond 0 day.

Fig. 1. Percentage of survival (mean ± SD) of Chironomus riparius after 10 days of exposure to diclofenac (DF) and carbamazepine (CBZ) in spiked sediments. Measured concentrations at the beginning of the test are presented.

exposed to 31.4 μg·g−1 CBZ compared to the controls indicating a relative increase in the abundance of male over female individuals (Fig. 5). 3.3. Risk assessment To calculate the RQ for DF and CBZ, the selected concentration corresponding to the highest concentration where no significant differences were reported to respect to the control was used. For DF and CBZ, these values are close to 30.0 μg·g−1 (the second highest tested concentration). Instead of use PEC, we have employed the measured environmental concentration (MEC) reported by Ebele et al. (2017) where the current state of knowledge about global pharmaceutical levels has been reviewed. The concentrations selected for DF and CBZ were 57 and 33 ng·g−1, respectively, corresponding to worst case scenarios reported in Ebele et al. (2017). The RQ for DF and CBZ were 0.19 and 0.11, respectively and for the mixture 0.30, indicating medium risk for both compounds or its mixture. 4. Discussion The freshwater midge C. riparius has been frequently employed to evaluate the effects of sediment sorbed contaminants in non-target organisms. Also in the case of pharmaceuticals as environmental contaminants, the number of studies reporting defined endpoints in spiked sediments using benthic invertebrates (Oviedo-Gómez et al., 2010; Méndez et al., 2013; Maranho et al., 2014, 2015) and more specifically midges (Oetken et al., 2005; Sánchez-Argüello et al., 2005, 2009;

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Fig. 2. Growth of larvae measured as mean dry weight (d.w ± SD) and mean values of percentage of increase with respect to sediment control (Δd.w.%) after 10 days of exposure to diclofenac (DF) and carbamazepine (CBZ). Asterisks indicate significant differences between means compared with sediment control (p b 0.05). Measured concentrations at the beginning of the test are presented.

Nentwig, 2007; Péry et al., 2008; Gilroy et al., 2012; López-Doval et al., 2012) as non-target organisms, are increasing. However, there is still an important knowledge gap about the effects of hundreds of pharmaceuticals and derivates on aquatic and benthic organisms which have

to be addressed urgently to prevent potential adverse effects of this kind of compounds at community level. For this reason, we performed a spiked sediment exposure test in order to evaluate the possible effects of two pharmaceutical compounds in the midge C. riparius. The selected

Fig. 3. Mean values (±SD) of emergence ratio for Chironomus riparius after 21 days of exposure at diclofenac (DF) and carbamazepine (CBZ). Asterisks indicate significant differences between means compared with sediment control (p b 0.05).

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sex-ratio (male/female). The obtained results were compared, when data were available, with those obtained for other benthic invertebrates. 4.1. Exposure

Fig. 4. Cumulative emergence curves (%) of midges exposed at diclofenac (DF) and carbamazepine (CBZ) in spiked sediment. Asterisks indicate significant differences between means compared with sediment control (p b 0.05).

compounds belong to two important groups of prescription drugs: antiepileptics and mood stabilizing drugs (carbamazepine) and nonsteroidal anti-inflammatory drugs, NSAIDs (diclofenac). The selected endpoints included mortality as well as several developmental endpoints such as growth (biomass), developmental instar, emergence rate and

For all concentrations, except the highest, targeted and measured values are satisfactorily similar. However, the highest targeted spiked exposure concentration was not achieved in both cases, DF and CBZ, presenting instead values similar to the next lower concentration. Differences in targeted and measured concentration in spiking procedures can be due to binding of the compound to glass surfaces, evaporation processes, biological and chemical degradation or sub-optimal extraction efficiency as it has been also observed by other authors (Ramil et al., 2010; López-Doval et al., 2012) as well as to the different sorption behavior of the selected compounds onto the sediment and its characteristics to achieve the steady-state of the water-sediment system, as for example in a study carried out by Buser et al. (1998), who showed negligible adsorption of DF onto sediment particles in laboratory experiments. A clear continuous decrease of the sediment exposure concentrations was reported, reaching at the end of the experiment values between 2.4 for DF and 6.3 μg·g−1 for CBZ in the highest exposure concentrations. Biological or chemical degradation of the compounds as explanation for the decrease in exposure concentrations can rather certainly be ruled out: CBZ has been found to be persistent in soils, biosolids and soil-biosolid mixtures with no degradation observed over 60 d (Monteiro and Boxall, 2009) and Löffler et al. (2005) reported high persistence of CBZ in sediment with values of N365 d (DT90). In the case of DF, a half-live of 20.44 d was reported by Qin et al. (2015) in agricultural soils irrigated with reclaimed water. These residence times, although not determined in freshwater sediments, suggest that the concentration of these compounds decreased along the experiment due to other reasons than degradation and can be rather contributed to their polar or moderately polar character, evidenced by the octanolwater partition coefficients of 1.51 and 1.90 for CBZ and DF, respectively (Scheytt et al., 2005) being desorbed to the overlying water. In our experiment, the exposure water was renewed daily, leading probably to the leaching of the compounds from the sediment into the water column where a further degradation could also be possible. Our experiments aimed to simulate the effects of punctually contaminated sediment under the conditions of non-continuous contaminant input. For the assessment of continuously contaminated sediments, water change frequency has to be reconsidered in the case of polar compounds envisaging a compromise between avoiding the leaching of the

Fig. 5. Mean values (±SD) of sex-ratio (n° males/n° females) for Chironomus riparius after 21 days of exposure at diclofenac (DF) and carbamazepine (CBZ). Asterisks indicate significant differences between means compared with sediment control (p b 0.05).

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compound and the wellbeing of the exposed population in the test. Within this context, the biologic activity of benthic organisms (e.g. burrowing) in the sediments can further alter certain processes such as sorption or desorption of compounds (Timmermann et al., 2011; Chen et al., 2016). In the case of more polar compounds, bioturbation could increase desorption from the sediment facilitating the transfer into the overlying water and thus increasing bioavailability and the risk of water contamination (Goedkoop and Peterson, 2003). 4.2. Effect assessment Gilroy et al. (2012) reported a survival N 72.5% in Chironomus dilutus exposed to CBZ in a range of concentrations between 0.1 and 100 μg·g−1, being the exposure concentrations slightly higher than the reported in our study. Maranho et al. (2014) reported significant mortalities in the marine polychaete Hediste diversicolor exposed to 0.05–0.5 μg·g−1 CBZ even if the exposure concentrations were much lower. Previously, Oetken et al. (2005) reported no acute toxicity in C. riparius exposed to CBZ spiked sediments in a series of tests at CBZ concentrations of 0.1; 0.2; 0.7 and 2.9 μg·g−1 indicating that CBZ seems not to be acutely toxic at environmentally relevant concentrations. These different responses of invertebrates to stress agents are related to the taxonomy and phylogeny (Holan et al., 2016; Malaj et al., 2016). With respect to our results obtained for DF, López-Doval et al. (2012) observed a survival between 70 and 90% in C. riparius larvae exposed during 10 days to 0.3–300 μg·g−1 to indomethacin, belonging like DF also to the group of NSAIDs. On the other hand, Oviedo-Gómez et al. (2010) observed that sediment sorbed DF produced oxidative stress in H. azteca exposed to 46.7 μg·kg−1 however without being significantly lethal. Maranho et al. (2015) did not observe any significant mortality in the amphipod A. brevicornis exposed to ibuprofen, another NSAID for 10 days. The endpoint growth is an environmentally very relevant and easily measurable indicator of changes due to sediment contamination. Changes in growth are ecologically very relevant since they can manifest at population level through effects on reproductive output on the longer term (Sibley et al., 1997). In our experiments, we observed significant changes in growth measured by biomass (somatic growth) (dry weight) in those organisms exposed to the highest tested CBZ concentration with respect to the control (Fig. 2). Although not significant, there is a continuous decreasing trend in growth of the organisms with increasing exposure concentrations of both of the pharmaceuticals, CBZ and DF. The difference in the case of CBZ, against organisms from the control experiment is the biggest, with those from the increasing exposure concentrations progressively being smaller. For DF, all exposed organisms are bigger than the control organisms suggesting the stimulation of the larval growth, with the highest relative increase in growth occurring at the lowest DF exposure concentration and successively decreasing with the concentration (Fig. 2). A similar observation was made by Gilroy et al. (2012) who found a growth occasionally greater than controls in Chironomus dilutus but exposed in this case to 56.6 μg·g−1 d.w. of CBZ. Sibley et al. (1997) suggested that in larvae of C. tentans, a closely related species of the family Chironomidae, the organisms should have a minimum weight of approx. 0.5 mg dry weight/individual for emergence to take place. According with this suggestion, in our study, the organisms exposed to the selected pharmaceutical compounds reach that estimated weight being comprised between 1.4 and 2.0 mg and 1.7 and 2.4 mg for DF and CBZ, respectively. With regard to the developmental stage, morphometry is a widely used tool in the determination of insect instars (Daly, 1985). The dimensions most commonly used to determine the number of larval instars is the width (Ecole, 1999; Silva et al., 2008) from ventral view. In our experiments, the organisms exposed to DF and CBZ had reached the fourth instar independently from the compound and the exposure concentration applied. In this context, the emergence ratio for CBZ was comprised around 0.5 for all treatments including the control, whereas for DF all

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treatments resulted in an emergence rate of about 0.9 except for the highest concentration where the emergence rate decreased significantly to 0.2 (Fig. 3) In agreement, the total emergence of adults (%) decreased with the highest DF concentration (Fig. 3). Oetken et al. (2005) observed a significant and concentration dependent decrease of emergence with a negative impact of CBZ on emergence of midges at sediment concentrations N 0.07 μg·g−1. These authors also observed a blockade of pupation in CBZ-exposed C. riparius larvae such that they survived up to 4 weeks without undergoing pupation until they finally died. The suppression of pupation in C. riparius under CBZ exposure refers to a specific mode of action, most probably some interference with a physiologic pathway first activated in this life stage or to some modulation of endocrine functions (Oetken et al., 2005). Similar findings have been made by Oropesa et al. (2016) who observed that chronic exposure to CBZ significantly decreased the reproductive output and the number of molts of D. magna at 200 μg/L. CBZ induced the production of male offspring (12 ± 1.7%), in a concentration independent manner, acting as a weak juvenile hormone analog. Further, CBZ showed to be toxic to daphnid embryos through maternal exposure interfering with their normal gastrulation and organogenesis stages but not producing direct embryo toxicity. No information on effects of emergence rate in C. riparius has been found for treatment with DF. Other authors have previously associated effects between the relationships growth-emergency. At ecological level, larval growth retardation and low emergence rate, could, in a longer term, lead to reduced reproductive success and consequently reduction in abundance in natural populations (Liber et al., 1996; Benoit et al., 1997). Changes in growth induced by stress can be used to make meaningful predictions regarding reproduction and population dynamics (Sibley et al., 1997). These authors observed that a reduction in the growth during larval stages in C. tentans is associated with a proportional decline in reproductive output of adult females. Effects on sex-ratio may affect reproduction and therefore, population dynamics, even more in the context of chronic exposure that affects successive generations of organisms (Sánchez-Argüello et al., 2009). Recent experiments showed reduction of development rate and fecundity (egg ropes/female) when C. riparius was exposed to 100 μg/L of CBZ fortifications of reclaimed water (Sánchez-Argüello et al. personal communication). In our experiments, we have assessed several endpoints throughout the life cycle of the midge, C. riparius exposed to sediments spiked with two different pharmaceutical compounds. At our assayed concentrations, we observed an increase of mortality for CBZ and a reduction in the emergence rate for DF. CBZ also induced a significant change in the sex-ratio, increasing the number of males with respect to females. CBZ, as well as other antidepressants have been previously related with endocrine disrupting activity in human (Isojärvi et al., 2004; Löfgren et al., 2006; Leśkiewicz et al., 2008) as well as environmentally exposed non-target organisms (Hampel et al., 2014; Fong and Ford, 2014; Hazelton et al., 2014). This fact is of great environmental relevance, as changes in natural proportions between male and female can have important consequences in population in the longer term. 4.3. Risk assessment To establish the risk associated to the occurrence of pharmaceuticals in the sediment involve the need to have a wide database about presence and effect. However, our results can be considered only as a preliminary assessment in relation to the risk associated to DF and CBZ in sediments. One of the limitations is that only one species C. riparius has been considered for the calculation and, on the other hand, the database about quantification of these compounds in sediments is limited. Nevertheless, the use of chronic instead of acute toxicity data gives additional value from the point of view of ecological relevance (ECB, 2003). Our results showed higher risk than those found by Trombini et al. (2016) for the species Tisbe battagliai using acute toxicity test

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and an assessment factor (AF) = 1000, according to ECB (2003) guidelines. The use of RQ can be employed also for priorization of emerging organic contaminants (EOCs) individually or such as mixture of compounds, because exposure and effect are taken in account together. Although there is still an urgent need to establish links between the observed effects in laboratory exposure experiments and their impact on ecosystem level, the present study contributes to the understanding on the possible chronic effects of the selected pharmaceuticals, DF and CBZ at low concentrations. Our results indicate that their presence in sediments may produce changes in the survival and development of C. riparius which in a long term could have important ecological implications in natural populations. 5. Conclusions Our spiked sediment experiments with DF and CBZ employing the midge C. riparius have shown that exposure to these compounds may produce negative effects in survival and growth and provoke certain changes in emergence and sex-ratio. The C. riparius bioassay is suitable for the quality assessment of sediments contaminated by pharmaceuticals compounds. The employed endpoints are considered useful to indicate changes in population structure and size. Therefore, the obtained results should be taken into account for threshold settings and quality guidelines. Nevertheless, further research should be done to assess the environmental risk of these emerging contaminants in benthic organisms that inhabit in the sediments at population level to guarantee the protection of ecosystems. Acknowledgements Authors would like to thank to Dr. A.M. Arias for the Chironomus larvae image employed in the graphical abstract and the Spanish Ministry of Economy and Competitiveness for its financial support through the project SCARCE (Consolider-Ingenio 2010 CSD2009-00065), NETSCARCE (CTM2015-69780-REDC) and the PIE 20143E072 (CSIC). M.H. is supported by a Ramón y Cajal contract (RYC-2012-12217) from the Spanish Ministry of Economy and Competitiveness (MINECO). Ministry of Economy and Competitiveness and European Regional Development Fund (ERDF). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2017.07.171. References Allaby, M., 2006. A Dictionary of Ecology. Oxford University Press, USA, p. 234. ASTM, 1997. Standard test methods for measuring the toxicity of sediment-associated contaminants with freshwater invertebrates. ASTM Annu. B. Stand, pp. 1138–1220. ASTM, 2004. Standard guide for collection, storage, characterization, and manipulation of sediments for toxicological testing and for selection of samplers used to collect benthic invertebrates. Annual Book of ASTM Standards. ASTM International, West Conshohocken, PA, pp. 698–790. ASTM, 2005. E 1706-05 Standard Test Method for Measuring the Toxicity of Sedimentassociated Contaminants with Freshwater Invertebrates. (West Conshohocken, PA). www.astm.org. Baena-Nogueras, R.M., Pintado-Herrera, M.G., González-Mazo, E., Lara-Martín, P.A., 2016. Determination of pharmaceuticals in coastal systems using solid phase extraction (SPE) followed by ultraperformance liquid chromatography-tandem mass spectrometry (UPLC-MS/MS). Curr. Anal. Chem. 12, 183–201. Bendz, D., Paxéus, N.A., Ginn, T.R., Loge, F.J., 2005. Occurrence and fate of pharmaceutically active compounds in the environment, a case study: Höje River in Sweden. J. Hazard. Mater. 122, 195–204. Benoit, D.A., Sibley, P.K., Juenemann, J.L., Ankley, G.T., 1997. Chironomus tentans life-cycle test: design and evaluation for use in assessing toxicity of contaminated sediments. Environ. Toxicol. Chem. 16, 1165–1176. Buser, H.-R., Poiger, T., Müller, M.D., 1998. Occurrence and fate of the pharmaceutical drug diclofenac in surface waters: rapid photodegradation in a lake. Environ. Sci. Technol. 32, 3449–3456. Carballa, M., Fink, G., Omil, F., Lema, J.M., Ternes, T., 2008. Determination of the solid–water distribution coefficient (Kd) for pharmaceuticals, estrogens and musk fragrances in digested sludge. Water Res. 42, 287–295.

Chen, M., Ding, S., Liu, L., Xu, D., Gong, M., Tang, H., Zhang, C., 2016. Kinetics of phosphorus release from sediments and its relationship with iron speciation influenced by the mussels (Corbicula fluminea) bioturbation. Sci. Total Environ. 542, 833–840. Clara, M., Strenn, B., Kreuzinger, N., 2004. Carbamazepine as a possible anthropogenic marker in the aquatic environment: investigations on the behaviour of carbamazepine in wastewater treatment and during groundwater infiltration. Water Res. 38, 947–954. Daly, H.V., 1985. Insect morphometrics. Annu. Rev. Entomol. 30, 415–438. Daneshvar, A., Svanfelt, J., Kronberg, L., Prévost, M., Weyhenmeyer, G.A., 2010. Seasonal variations in the occurrence and fate of basic and neutral pharmaceuticals in a Swedish river–lake system. Chemosphere 80, 301–309. Directive 2013/39/EU of the European parlamiament and of the council of 12 August 2013 amending directives 2000/60/EC and 2008/105/EC as regards priority substances in the field of water policy, 2017. Drillia, P., Stamatelatou, K., Lyberatos, G., 2005. Fate and mobility of pharmaceuticals in solid matrices. Chemosphere 60, 1034–1044. Ebele, A.J., Abdallah, M.A.-E., Harrad, S., 2017. Pharmaceuticals and personal care products in the freshwater aquatic environment. Emerg. Contam. 3, 1–16. Ecole, C.C., 1999. Determinação do número de ínstares larvais em Plutella xylostella (L.) (Lepidoptera: Yponomeutidae). Acta Sci. 21, 331–335. El Rayis, O.A., 1985. Reassessment of the titration method for determination organic carbon in recent sediments. Rapp. Comm. Int. Mer. Medit. 29, 45–47. Environment Canada, 1997. Biological Test Method: Test for Survival and Growth in Sediment Using the Larvae of Freshwater Midges (Chironomus tentans or Chironomus riparius). EPS 1/RM/32. Environmental Protection Series (Ottawa, ON, Canada). Ericson, H., Thorsén, G., Kumblad, L., 2010. Physiological effects of diclofenac, ibuprofen and propranolol on Baltic Sea blue mussels. Aquat. Toxicol. 99, 223–231. European Chemical Bureau, 2003. Technical Guidance Document on Risk Assessment in Support of Commission Directive 93/67/EEC on Risk Assessment for New Notified Substances, Commission Regulation (EC) No 1488/94 on Risk Assessment for Existing Substances Directive 98/8/EC of the European Parliament and the Council Concerning the Biocidal Products on the Market. Environmental Chemicals Bureau, Ispra, Italy. Farré, M., Ferrer, I., Ginebreda, A., Figueras, M., Olivella, L., Tirpau, L., Vilanova, M., Barcelò, D., 2001. Determination of drugs in surface water and wastewater samples by liquid chromatography: methods and preliminary results including toxicity studies with Vibrio fischeri. J. Chromatogr. A 938, 187–197. Fent, K., Weston, A.A., Caminada, D., 2006. Ecotoxicology of human pharmaceuticals. Aquat. Toxicol. 76, 122–159. Fong, P.P., Ford, A.T., 2014. The biological effects of antidepressants on the molluscs and crustaceans: a review. Aquat. Toxicol. 151, 4–13. Gaudette, H., Flight, W., Toner, L., Folger, D., 1974. An inexpensive titration method for the determination of organic carbon in recent sediments. J. Sediment. Petrol. 44, 249–253. Gerhardt, A., Janssens de Bisthoven, L., Mo, Z., Wang, C., Yang, M., Wang, Z., 2002. Shortterm responses of Oryzias latipes (Pisces: Adrianichthyidae) and Macrobrachium nipponense (Crustacea: Palaemonidae) to municipal and pharmaceutical waste water in Beijing, China: survival, behaviour, biochemical biomarkers. Chemosphere 47, 35–47. Gilroy, È.A.M., Balakrishnan, V.K., Solomon, K.R., Sverko, E., Sibley, P.K., 2012. Behaviour of pharmaceuticals in spiked lake sediments – effects and interactions with benthic invertebrates. Chemosphere 86, 578–584. Goedkoop, W., Peterson, M., 2003. The fate, distribution, a toxicity of lindane in tests with Chironomus riparius: effects of bioturbation and sediment organic matter content. Environ. Toxicol. Chem. 22, 67–76. Gros, M., Petrovic, M., Barceló, D., 2006. Multi-reside analytical methods using LC-tandem MS for the determination of pharmaceuticals in environmental and wastewater samples: a review. Anal. Bioanal. Chem. 386, 941–952. Gros, M., Petrović, M., Barceló, D., 2009. Tracing pharmaceutical residues of different therapeutic classes in environmental waters by using liquid chromatography/quadrupole-linear ion trap mass spectrometry and automated library searching. Anal. Chem. 81, 898–912. Hampel, M., Bron, J.E., Taggart, J.B., Leaver, M.J., 2014. The antidepressant drug carbamazepine induces differential transcriptome expression in the brain of Atlantic salmon, Salmo salar. Aquat. Toxicol. 151, 114–123. Hazelton, P.D., Du, B., Haddad, S.P., Fritts Andrea, K., Chambliss, C.K., Brooks, B.W., Bringolf, R.B., 2014. Chronic fluoxetine exposure alters movement and burrowing in adult freshwater mussels. Aquat. Toxicol. 151, 27–35. 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. Hernando, M.D., Mezcua, M., Fernández-Alba, A.R., Barceló, D., 2006. Environmental risk assessment of pharmaceutical residues in wastewater effluents, surface waters and sediments. Talanta 69, 334–342. Heye, K., Becker, D., Lütke Eversloh, C., Durmaz, V., Ternes, T.A., Oetken, M., Oehlmann, J., 2016. Effects of carbamazepine and two of its metabolites on the non-biting midge Chironomus riparius in a sediment full life cycle toxicity test. Water Res. 98, 19–27. Holan, J.R., King, C.R., Dabis, A.R., 2016. Sensitivity of six subantartic marine invertebrates to common metal contaminants. Environ. Toxicol. Chem. 35 (9), 2245–2251. Isojärvi, J.I., Löfgren, E., Juntunen, K.S., Pakarinen, A.J., Päivänsalo, M., Rautakorpi, I., Tuomivaara, L., 2004. Effect of epilepsy and antiepileptic drugs on male reproductive health. Neurology 27 (62), 247–253. Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber, L.B., Buxton, H.T., 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. Streams, 1999–2000: a national reconnaissance. Environ. Sci. Technol. 36, 1202–1211. Lara-Martín, P.A., Gonzalez-Mazo, E., Petrovic, M., Barcelo, D., Brownawell, B.J., 2014. Occurrence, distribution and partitioning of nonionic surfactants and pharmaceuticals in the urbanized Long Island Sound Estuary (NY). Mar. Pollut. Bull. 85, 710–719. Lee, S.-B., Choi, J., 2007. Effects of bisphenol A and ethynyl estradiol exposure on enzyme activities, growth and development in the fourth instar larvae of Chironomus riparius (Diptera, Chironomidae). Ecotoxicol. Environ. Saf. 68, 84–90.

E. Nieto et al. / Science of the Total Environment 609 (2017) 715–723 Lee, J., Ji, K., Lim Kho, Y., Kim, P., Choi, K., 2011. Chronic exposure to diclofenac on two freshwater cladocerans and Japanese medaka. Ecotoxicol. Environ. Saf. 74, 1216–1225. Leśkiewicz, M., Budziszewska, B., Lasoń, W., 2008. Endocrine effects of antiepileptic drugs. Przegl. Lek. 65, 795–798. Liber, K., Call, D.J., Dawson, T.D., Whiteman, F.W., Dillon, T.M., 1996. Effects of Chironomus tentans larval growth retardation on adult emergence and ovipositing success: implications for interpreting freshwater sediment bioassays. Hydrobiologia 323, 155–167. Löffler, D., Terner, T.A., 2003. Determination of acidic pharmaceuticals, antibiotics and ivermectin in river sediment using liquid chromatography-tandem mass spectrometry. J. Chromatogr. A 102, 133–144. Löffler, D., Römbke, J., Meller, M., Ternes, T.A., 2005. Environmental fate of pharmaceuticals in water/sediment systems. Environ. Sci. Technol. 39, 5209–5218. Löfgren, E., Tapanainen, J.S., Koivunen, R., Pakarinen, A., Isojärvi, J.I., 2006. Effects of carbamazepine and oxcarbazepine on the reproductive endocrine function in women with epilepsy. Epilepsia 47 (9), 1441–1446. López-Doval, J.C., Kukkonen, J.V.K., Rodrigo, P., Muñoz, I., 2012. Effects of indomethacin and propranolol on Chironomus riparius and Physella (Costatella) acuta. Ecotoxicol. Environ. Saf. 78, 110–115. Malaj, E., Guenard, G., Schaffer, R.B., Ohe von der, P.C., 2016. Evolutionary patterns and physicochemical properties explain macroinvertebrate sensitivity to heavy metals. Ecol. Appl. 26, 1249–1259. Maranho, L.A., Baena-Nogueras, R.M., Lara-Martín, P.A., DelValls, T.A., Martín-Díaz, M.L., 2014. A candidate short term toxicity test using Ampelisca brevicornis to assess sublethal responses. Arch. Environ. Contam. Toxicol. 68, 237–258. Maranho, L.A., Garrido-Pérez, M.C., DelValls, T.A., Martín-Díaz, M.L., 2015. Suitability of standardized acute toxicity tests for marine sediment assessment: pharmaceutical contamination. Water Air Soil Pollut. 226, 65. Martínez Bueno, M.J., Agüera, A., Gómez, M.J., Hernando, M.D., García-Reyes, J.F., Fernández-Alba, A.R., 2007. Application of liquid chromatography/quadrupolelinear ion trap mass spectrometry and time-of-flight mass spectrometry to the determination of pharmaceuticals and related contaminants in wastewater. Anal. Chem. 79, 9372–9384. Méndez, N., Lacorte, S., Barata, C., 2013. Effects of the pharmaceutical fluoxetine in spikedsediments on feeding activity and growth of the polychaete Capitella teleta. Mar. Environ. Res. 89, 76–82. Monteiro, S.C., Boxall, A.B., 2009. Factors affecting the degradation of pharmaceuticals in agriculture soils. Environ. Toxicol. Chem. 28, 2546–2554. Nentwig, G., 2007. Effects of pharmaceuticals on aquatic invertebrates. Part II: the antidepressant drug fluoxetine. Arch. Environ. Contam. Toxicol. 52, 163–170. OECD, 2004a. Test No. 218: Sediment-water Chironomid Toxicity Using Spiked Sediment, OECD Guidelines for the Testing of Chemicals, Section 2. OECD Publ., pp. 1–21. OECD, 2004b. Test No. 219: Sediment-water Chironomid Toxicity Using Spiked Water, OECD Guidelines for the Testing of Chemicals, Section 2. OECD Publ., pp. 1–21. OECD, 2010. Test No. 233: Sediment-water Chironomid Life-cycle Toxicity Test Using Spiked Water or Spiked Sediment. OECD Guidelines for the Testing of Chemicals, Section 2. OECD Publ., pp. 1–29. Oetken, M., Nentwig, G., Löffler, D., Ternes, T., Oehlmann, J., 2005. Effects of pharmaceuticals on aquatic invertebrates. Part I. The antiepileptic drug carbamazepine. Arch. Environ. Contam. Toxicol. 49, 353–361. Öllers, S., Singer, H.P., Fässler, P., Müller, S.R., 2001. Simultaneous quantification of neutral and acidic pharmaceuticals and pesticides at the low-ng/l level in surface and waste water. J. Chromatogr. A 911, 225–234. Oropesa, A.L., Floro, A.M., Palma, P., 2016. Assessment of the effects of the carbamazepine on the endogenous endocrine system of Daphnia magna. Environ. Sci. Pollut. Res. 23, 17311. Osorio, V., Marcé, R., Pérez, S., Ginebreda, A., Cortina, J.L., Barceló, D., 2012. Occurrence and modeling of pharmaceuticals on a sewage-impacted Mediterranean river and their dynamics under different hydrological conditions. Integr. Model. Monit. Differ. river basin scales under Glob. Chang. 440, pp. 3–13. Oviedo-Gómez, D.G.C., Galar-Martínez, M., García-Medina, S., Razo-Estrada, C., GómezOliván, L.M., 2010. Diclofenac-enriched artificial sediment induces oxidative stress in Hyalella azteca. Environ. Toxicol. Pharmacol. 29, 39–43. Paxéus, N., 2004. Removal of selected non-steroidal anti-inflammatory drugs (NSAIDs), gemfibrozil, carbamazepine, beta-blockers, trimethoprim and triclosan in conventional wastewater treatment plants in five EU countries and their discharge to the aquatic environment. Water Sci. Technol. 50, 253–260.

723

Péry, a.R.R., Gust, M., Vollat, B., Mons, R., Ramil, M., Fink, G., Ternes, T., Garric, J., 2008. Fluoxetine effects assessment on the life cycle of aquatic invertebrates. Chemosphere 73, 300–304. Qin, Q., Chen, X., Zhuang, J., 2015. The fate and impact of pharmaceuticals and personal care products in agricultural soils irrigated with reclaimed water. Crit. Rev. Environ. Sci. Technol. 45, 1379–1408. Ramil, M., El Aref, T., Fink, G., Scheurer, M., Ternes, T.A., 2010. Fate of beta blockers in aquatic-sediment systems: sorption and biotransformation. Environ. Sci. Technol. 44, 962–970. Roberts, P.H., Thomas, K.V., 2006. The occurrence of selected pharmaceuticals in wastewater effluent and surface waters of the lower Tyne catchment. Sci. Total Environ. 356, 143–153. Sacher, F., Lange, F.K., Brauch, H.J., Blankenhorn, I., 2001. Pharmaceuticals in groundwaters. Analytical methods and results of a monitoring program in BadenWürttemberg, Germany. J. Chromatogr. A 938, 199–210. Sánchez-Argüello, P., Alonso, C., Fernández, C., Vega, M.M., García, M.P., Tarazona, J.V., 2005. Evaluation of a Multi-species test System for assessing acute and chronic toxicity of sediments and water to aquatic invertebrates: effects of pentachlorophenol on Dapnia magna and Chironomus prasinus. J. Soils Sediments 5, 53–58. Sánchez-Argüello, P., Fernández, C., Tarazona, J.V., 2009. Assessing the effects of fluoxetine on Physa acuta (Gastropoda, Pulmonata) and Chironomus riparius (Insecta, Diptera) using a two-species water-sediment test. Sci. Total Environ. 407, 1937–1946. Scheytt, T., Mersmann, P., Lindstädt, R., Heberer, T., 2005. 1-Octanol/water partition coefficients of 5 pharmaceuticals from human medical care: carbamazepine, clofibric acid, diclofenac, ibuprofen, and propyphenazone. Water Air Soil Pollut. 165, 3–11. Schmidt, W., O'Rourke, K., Hernan, R., Quinn, B., 2011. Effects of the pharmaceuticals gemfibrozil and diclofenac on the marine mussel (Mytilus spp.) and their comparison with standardized toxicity tests. Mar. Pollut. Bull. 62, 1389–1395. Sibley, P.K., Benoit, D.A., Ankley, G.T., 1997. The significance of growth in Chironomus tentans sediment toxicity tests: relationship to reproduction and demographic endpoints. Environ. Toxicol. Chem. 16, 336–345. Silva, D.S., Kaminski, L.A., Dell'Erba, R., Moreira, G.R.P., 2008. Morfologia externa dos estágios imaturos de heliconíneos neotropicais: VII. Dryadula phaetusa (Linnaeus) (Lepidoptera, Nymphalidae, Heliconiinae). Rev. Bras. Entomol. 52, 500–509. Silva, B.F., Jelic, A., López-Serna, R., Mozeto, A.A., Petrovic, M., Barceló, D., 2011. Occurrence and distribution of pharmaceuticals in surface water, suspended solids and sediments of the Ebro river basin, Spain. Chemosphere 85, 1331–1339. Sim, W.J., Lee, J.W., Oh, J.E., 2010. Occurrence and fate of pharmaceuticals in wastewater treatment plants and rivers in Korea. Environ. Pollut. 158, 1938–1947. Sirés, I., Brillas, E., 2012. Remediation of water pollution caused by pharmaceutical residues based on electrochemical separation and degradation technologies: a review. Environ. Int. 40, 2012–2229. Soares, S., Cativa, I., Moreira-Santos, M., Soares, A.M.V.M., Ribeiro, R., 2005a. A short-term sublethal in situ sediment assay with Chironomus riparius based on postexposure feeding. Arch. Environ. Contam. Toxicol. 49, 163–172. Soares, S., Moreira, S., Andrade, S., Guilhermino, L., Ribeiro, R., 2005b. Freshwatersaltwater interface and estuarine sediment in situ assays based on post-exposure feeding of chironomids and polychaetes. Estuaries 28, 314–319. Timmermann, K., Banta, G.T., Klinge, L., Andersen, O., 2011. Effects of bioturbation on the fate of oil in coastal sandy sediments—an in situ experiment. Chemosphere 82, 1358–1366. Trombini, C., Hampel, M., Blasco, J., 2016. Evaluation of acute effects of four pharmaceuticals and their mixtures on the copepod Tisbe battagliai. Chemosphere 155, 319–328. UNE, 77311, 2000. Calidad del suelo, Determinación del contenido en materia seca y agua en base al peso. Método gravimétrico (ISO 11465:1993). Valdés, M.E., Amé, M.V., de los A Bistoni, M., Wunderlin, D.A., 2014. Occurrence and bioaccumulation of pharmaceuticals in a fish species inhabiting the Suquía River basin (Córdoba, Argentina). Sci. Total Environ. 472, 389–396. Vazquez-Roig, P., Segarra, R., Blasco, C., Andreu, V., Picó, Y., 2010. Determination of pharmaceuticals in soils and sediments by pressurized liquid extraction and liquid chromatography tandem mass spectrometry. J. Chromatogr. A 1217, 2471–2483. Writer, J.H., Ferrer, I., Barber, L.B., Thurman, E.M., 2013. Widespread occurrence of neuroactive pharmaceuticals and metabolites in 24 Minnesota rivers and wastewaters. Sci. Total Environ. 461–462, 519–527. Yuan, F., Hu, C., Hu, X., Qu, J., Yang, M., 2009. Degradation of selected pharmaceuticals in aqueous solution with UV and UV/H2O2. Water Res. 43, 1766–1774.