Science of the Total Environment 518–519 (2015) 225–237

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Acute toxicological effects on the earthworm Eisenia fetida of 18 common pharmaceuticals in artificial soil Mª. Rosa Pino a,⁎, Jonatan Val a, Ana Mª. Mainar b, Estefanía Zuriaga a, Cecilia Español a, Elisa Langa a a b

Universidad San Jorge, Instituto de Medio Ambiente, Facultad de Ciencias de la Salud, GIMACÉS, Villanueva de Gállego, Zaragoza 50830, Spain Universidad de Zaragoza, Instituto de Investigación en Ingeniería de Aragón (I3A), GATHERS, Calle Mariano Esquillor, s/n, Edificio de Institutos, I3A, Bloque 5, 2.ª planta, 20018 Zaragoza, Spain

H I G H L I G H T S • • • •

Acute toxicity of 18 pharmaceuticals on the earthworm Eisenia fetida was measured. 4 NSAIDs and 3 blood lipid-lowering agents showed acute toxicity on E. fetida. β-Blockers and antibiotics showed no detectable lethality on E. fetida. NSAIDs, antibiotics, timolol and bezafibrate were degraded in artificial soil.

a r t i c l e

i n f o

Article history: Received 3 November 2014 Received in revised form 18 February 2015 Accepted 22 February 2015 Available online 9 March 2015 Editor: E. Capri Keywords: Antibiotic β-Blocker Blood lipid-lowering agent Earthworm NSAID Soil

a b s t r a c t Following soil applications of recycled water and biosolids, pharmaceutical residues can eventually enter the terrestrial environment. In vitro and in vivo assays have largely focused on the acute ecotoxicity of these compounds in aquatic systems. However, studies on the ecotoxicological effects of pharmaceuticals in soil biota are especially scarce. The aim of this study was to investigate the acute toxicity of 18 pharmaceuticals (4 NSAIDs, 5 blood lipidlowering agents, 6 β-blockers and 3 antibiotics) that are usually found in the environment by using an Eisenia fetida bioassay. In addition, the presence of these pharmaceuticals in artificial soil was verified at the end of the test. Our results indicate that seven of the studied drugs cause acute adverse effects in E. fetida, in particular, the NSAIDs and the blood lipid-lowering agents. Ibuprofen (LC50 = 64.80 mg/kg) caused the highest acute toxicity for all tested compounds, followed by diclofenac (LC50 = 90.49 mg/kg) and simvastatin (LC50 = 92.70 mg/kg). Other tested pharmaceuticals from NSAIDs and blood lipid-lowering families have toxicity effects, from a LC50 = 140.87 mg/kg for gemfibrozil to 795.07 mg/kg for lovastatin. Atorvastatin, bezafibrate, β-blockers and antibiotics showed no detectable lethality in E. fetida. The four NSAIDs showed evidence of modification of their original chemical structure after 14 days so the detected toxicity may be due to the original product as well as their degradation products. The three blood lipid-lowering agents seem to be more stable in soil. From an environmental perspective, the lethal concentrations of the tested drugs are much greater than those reported in wastewater and biosolids, therefore acute toxic effects may be improbable. However, little is known about the accumulation of these substances in soils after regular applications, so accumulative and chronic effects cannot be excluded. Moreover, more studies are needed to determine the role of the degradation products of these pharmaceuticals on terrestrial toxicity. © 2015 Elsevier B.V. All rights reserved.

1. Introduction ⁎ Corresponding author at: Universidad San Jorge, Campus Universitario Villanueva de Gállego, Autovía A-23 Zaragoza-Huesca, km. 510, 50830 Villanueva de Gállego, Zaragoza, Spain. E-mail addresses: [email protected] (M.ªR. Pino), [email protected] (J. Val), [email protected] (A.M.ª Mainar), [email protected] (E. Zuriaga), [email protected] (C. Español), [email protected] (E. Langa). URL: http://www.usj.ess (M.ªR. Pino).

http://dx.doi.org/10.1016/j.scitotenv.2015.02.080 0048-9697/© 2015 Elsevier B.V. All rights reserved.

Pharmaceutical innovation has provided an extraordinary variety of drug families that have undoubtedly helped to improve our health and social welfare, but as a result many new active ingredients are released into the environment. The metabolic and environmental degradation of these compounds produces a huge variety of metabolites and degradation products that increase the complexity of the problem.

226

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

The consumption of pharmaceuticals is substantial. Approximately 3000 different substances are used for human medicine in the European Union (EU). Pharmaceutical compounds such as non-steroidal antiinflammatory drugs (NSAIDs) and antibiotics are used extensively worldwide and their consumption, predominantly in developed countries, is assumed to be higher than several hundred tons per year (Daughton and Ternes, 1999). Approximately 70% of administered drugs are estimated to be released into the environment (Jacobsen and Berglind, 1988), which would explain the increasing evidence of pharmaceutical ubiquity in the environment. Pharmaceuticals can be introduced into the environment via a number of routes but are primarily introduced by both untreated and treated sewage (Ternes et al., 2004; Yu et al., 2011). Treatment Plants (STPs) are not designed to remove these contaminants so, they can be found in the effluent of STPs and in their sewage sludge. Many pharmaceuticals have limited biodegradability, resulting in only partial removal from the water phase. As a result, considerable amounts of pharmaceuticals remain in treated sewage sludge, and they are commonly called biosolids (Heidler et al., 2006; Martin et al., 2012a,b; Miege et al., 2009; Ternes et al., 2004). Pharmaceutical residues can enter the terrestrial environment following soil applications of wastewater (recycled water) and biosolids (Kinney et al., 2006; Xia et al., 2005). Even if pharmaceutical sorption is quite small relative to biodegradation, it cannot be completely neglected because each drug could become significantly more concentrated during sludge dewatering. Pharmaceutical sorption into the sludge/soil matrix in biosolid-amended soils reduces the bioavailability needed for microbial biodegradation (Drillia et al., 2005; Williams et al., 2006) and could lead to the environmental persistence of these substances. Available data show levels of human pharmaceuticals in biosolids and wastewater in the range of ng/kg to μg/kg (Lishman et al., 2006; Martin et al., 2012b; Radjenovic et al., 2009; Rosal et al., 2010; Ternes, 1998) so they are usually known as “micro-pollutants”. However, some antibiotics seem to reach higher ranges. The antibiotic ciprofloxacin and norfloxacin were detected in the range from 3.1 to 6.9 mg/kg respectively in digested sludge (Heidler and Halden, 2008) and tetracycline can be detected at a maximum concentration of 2.7 mg/kg in U.S. biosolids (Hamscher et al., 2002; McClellan and Halden, 2010). Once in upper soil layers, the active ingredients might either accumulate in soil or be absorbed by crops, or they may be readily available for transport into surface and groundwater through leaching and over surface flow runoff (Jongbloed and Lenis, 1998), contaminating both the surrounding surface and groundwater (Kemper, 2008; Topp et al., 2008). However, little is known about the specific behavior and concentrations of each class of drugs in soils. Analyses of soil environmental samples are especially scarce (Kim and Carlson, 2005; Loffler and Ternes, 2003), primarily because of the difficulty of extracting this type of compound from this matrix. Sometimes the analytical methods employed for soil samples are merely slight variations of those developed for aqueous ones. In addition, the global impact of pharmaceuticals in terrestrial ecosystems still remains unclear. Although these substances are generally less persistent than classical persistent organic pollutants, their continuous release and their concentration in the soil matrix makes them semi-persistent (Daughton and Ternes, 1999) so these substances deserve special attention as environmental pollutants. Furthermore, they are already designed to interfere with biological processes at low concentrations (Ankley et al., 2007). Some of the negative impacts of these products in terrestrial ecosystems have been preliminarily revealed as the inhibition of seed germination and crop growth and phytotoxicity (Kong et al., 2007; Xie et al., 2011), inhibited microbial activity and soil enzymatic activity (Kong et al., 2006; Liu et al., 2009) and product accumulation in plants and crop biomass (Boxall et al., 2006; Dolliver et al., 2007; Kumar et al., 2005; Wu et al., 2010b). Earthworms have been found to accumulate drugs directly from sewage effluent (Markman et al.,

2007) and from the biosolids applied to agricultural soils (Kinney et al., 2008). Studies about pharmaceutical toxicity in soil systems have focused primarily on plants (D'Abrosca et al., 2008; Feito et al., 2013; Furtula et al., 2012; Hillis et al., 2008; Migliore et al., 2003; Schmitt-Jansen et al., 2007) and much less on soil invertebrates, despite earthworms having been used as a model organism for soil toxicity studies of many classical pollutants. Earthworms, are common invertebrates in a wide range of soils and may represent 60–80% of the total soil biomass. Because of their impact on soil aeration, moisture content, nutrient cycling, and overall soil structure, earthworms are closely linked to the physical and chemical dynamics of the terrestrial environment and they are highly affected by soil pollutants (Spurgeon et al., 2003). The Oligochaeta Eisenia fetida, (family Lumbricidae), is considered to be a model organism for environmental surveillance (Cortez and Bouche, 1992) and this organism is a recommended test species (OECD 207, 1984). Nonetheless, standard acute toxicity data of human pharmaceuticals in the E. fetida have only been reported, to the best of our knowledge, for the oxytetracycline (Baguer et al., 2000; Boleas et al., 2005). Therefore, there is an urgent need to fill the gaps about the ecotoxicological effects of pharmaceuticals in soil ecosystems to provide an accurate risk assessment and proper risk management. Moreover, knowledge about the environmental fate and toxicity to non-target organisms for most APIs (active pharmaceutical ingredients) currently on the market remains scarce and the European Medicines Agency (EMA) requires (and has since 2005) that an environmental risk assessment be included when applying for new marketing authorisations for human use pharmaceuticals (EMEA, 2006). The aim of this study was to measure the acute toxicity of 18 pharmaceuticals with a widespread presence in the environment by using E. fetida as a terrestrial model organism to provide toxicological information from the soil environment. In addition, we also evaluated to what extent the soil used for E. fetida experiments affected the initial chemical structure of our compounds over the assay time. The 18 selected drugs belong to four therapeutic families on which the scientific community has focused its attention because of the drugs' expected environmental impact from their extensive consumption and their systematic presence in the environment, namely non-steroidal anti-inflammatory drugs (NSAIDs), blood lipid-lowering agents, β-blockers and antibiotics. 2. Material and methods 2.1. Chemicals We studied four NSAIDs (ibuprofen, diclofenac, paracetamol and salicylic acid); five blood lipid-lowering agents (bezafibrate, gemfibrozil, atorvastatin, simvastatin and lovastatin); six β-blockers (propranolol, atenolol, acebutolol, metoprolol, timolol and nadolol); and three antibiotics (sulphamethoxazole, trimethoprim and tetracycline). Table 1 summarizes the relevant information about the 18 pharmaceuticals and other chemicals and solvents used in this study. 2.2. E. fetida acute toxicity test Adult E. fetida earthworms were obtained from the company TODOVERDE (Ourense, Spain). The tests were performed in accordance with the OECD 207 (1984) methodology. Adult earthworms, at least 2 months old with clitella with individual weights from 300 to 600 mg were selected and used in the experiments. All organisms allowed 15 days of acclimatization prior to testing and maintained at a controlled temperature of 18–25 °C, pH 7.5 to 8 and 80%–85% humidity in a sphagnum peat substrate from the company FLOWER (Tarragona, Spain). The tests were conducted in 1 L polypropylene rectangular jars with lids to minimize water evaporation and animal escape. The jar

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237 Table 1 Relevant information of the 18 pharmaceuticals and other chemicals and solvents used in this study. Compound NSAIDs Ibuprofen Diclofenac Paracetamol Salicylic acid

CAS number

Laboratory

Purity

Table 2 Summary results for the Eisenia fetida toxicity test of the 18 drugs expressed as LC50 with a 95% confidence interval, CI (in parentheses). Values are based on nominal concentrations (mg·kg−1 dry weight soil). All derived toxicity values were significant (P b 0.0001). Chemical

15687-27-1 15307-79-6 103-90-2 69-72-7

Blood lipid lowering agents Bezafibrate 41859-67-0 Gemfibrozil 25812-30-0 Atorvastatin 134523-00-5 Simvastatin 79902-63-9 Lovastatin 81739-26-6 β-Blockers Propranolol Atenolol Timolol Acebutolol Metoprolol Nadolol

525-66-6 29122-68-7 26921-17-5 34-381-68-5 83-43-2 42200-33-9

Antibiotics Sulfamethoxazole Trimethoprim Tetracycline

723-46-6 738-70-5 60-54-8

Solvents and other chemicals Acetonitrile 75-05-8 Methanol 67-56-1 KH2PO4 7778-77-0 K2HPO4 7758-11-4 Ethanol 64-17-5

Acofarma

100% 100% 100% 99.9%

NSAIDs Ibuprofen Diclofenac

Sigma Teva Pharma Ercros S.A.

Acofarma

Sigma-Aldrich Fagron Iberica

Acofarma

Fluka

Sigma-Aldrich

≥98% 100% 99% 99% 99% N99% N98% N98% 99.5% N98% 99.8%

100% 90% 98% ≥99.9% (GC) ≥99.9% (GC) ≥99.5% ≥99.0% ≥99.8% (GC)

lids were punched with holes to allow ventilation and oxygen supply. 500 g (wet weight) of standardized OECD soil substrate consisting of industrial fine sand, sphagnum peat, and kaolin clay in a 7:1:2 ratio, respectively was placed into the jars. Sphagnum peat was purchased from Verdecora vivarium (Spain) whereas; kaolin and industrial sand were obtained from Imerys Ceramics España, S.A. The water content of the mixture was determined by weighing the sample and drying it to a constant mass at 105 °C for 24 h. Deionized water was added to adjust the overall moisture content in an amount equivalent to 35% of dry weight of soil, and the medium was thoroughly mixed. The pH was measured with a pH meter and a 1 M KCl solution. Finally, we add 10 adult earthworms to each of the jars. Tests were performed with 3 replicates for each concentration. In order to determine the range of pharmaceutical concentrations for use in the final tests, preliminary range-finding assays (0–10–200– 1000 mg/kg) were performed. Several test concentrations in a geometric series (Table 2) were then prepared. The highest dose tested was 2000 mg/kg except for those classes of drugs in which no toxicity was observed. In these cases, higher doses were tested (Table 2). We performed 3 replicates of negative controls for each experiment by using the same procedure and standardized soil without applying any drug. Furthermore, in order to assess the sensitivity of the test organisms chloroacetamide (OECD 207, 1984) was used as positive control in these experiments at a concentration of 24 mg/kg after a preliminary range-finding test (0–20–30–40–50 mg/kg). The jars were stored at 20 ± 2 °C under 80–85% relative humidity and 400–800 lx of constant light. Mortality was assessed at 7 and 14 days after treatment. After 14 days, the EC50 value for E. fetida was determined by log-Probit analysis (Bliss, 1934). 2.3. Stability tests To reproduce the experimental conditions used for the E. fetida toxicological tests with sufficient accuracy, 0.0100 g of each drug was

227

Paracetamol Salicylic acid

Concentrations tested (mg/kg)

LC50 (95% CI) (mg/kg)

R2 of the regression

0, 15, 30, 50, 60, 80, 100, 150 0, 30, 60, 100, 150, 180, 200, 500 0, 200, 400, 500, 600, 700, 850, 1000 0, 40, 100, 200, 350, 500, 800, 1000

64.80 (58.49–71.78) 90.49 (78.43–101.78) 693.50 (662.13–727.88) 162.68 (134.02–193.72)

0.961

140.87 (133.56–147.99) 92.70 (68.86–126.28) 795.07 (631.01–1017.50) N2000

0.977

N2000

–

Blood lipid lowering agents Gemfibrozil 0, 60, 100, 125, 150, 175, 200, 300 Simvastatin 0, 10, 20, 50, 75, 100, 200, 400 Lovastatin 0, 100, 200, 500, 750, 1000, 1500, 2000 Atorvastatin 0, 200, 500, 1000, 1500, 2000 Bezafibrate 0, 150, 850, 2000 β-Blockers Propranolol Atenolol Acebutolol Metoprolol Timolol Nadolol

0, 50, 500, 1000, 2000, 5000 3298.63 (2817.03–3969.48) 0, 500, 800, 1000, 1500, N2000 2000 0, 1000, 2000 N2000 0, 1000, 2000 N2000 0, 1000, 2000 N2000 0, 1000, 2000 N2000

Antibiotics Sulfamethoxazole 0, 1500, 2000, 4000 Trimethoprim 0, 100, 600, 2000 Tetracycline 0, 60, 600, 2000

N4000 N2000 N2000

0.968 0.970 0.944

0.824 0.962 –

0.899 – – – – – – – –

placed into a 10 mL glass vial together with 5 g of soil (the same soil used for the E. fetida worms) and the proportional amount of water to reproduce the E. fetida experimental conditions. No worms were placed in these vials. The presence of the test drugs was evaluated after 14 days. Drug extractions from soil samples were performed with ethanol except for atorvastatin and trimethoprim, for which methanol was used (because of the insolubility of both drugs in ethanol). The control vials (for methanol and for ethanol) were also analyzed, and they only contained water and soil according to the cited proportion. The soil drug extraction procedure consisted of washing each sample in the vials with 3 mL of solvent aliquots, sonicating for 2 min, centrifuging for 5 min at 5000 rpm and collecting the supernatant. This step was repeated 10 times for each compound. All supernatants were collected in a unique sample and kept at −20 °C in darkness for subsequent analysis. The detection of the studied pharmaceuticals in the collected supernatants was performed in a Shimadzu-LC 20AD HPLC equipped with a diode-array detector (SPD-M20A) and a Waters Nova-PakC18 column (15 cm × 3.9 cm and a 4 μm particle size). Samples were eluted with two different mobile phases using an isocratic method, namely 50% A (acetonitrile) + 50% B (phosphate buffer solution pH = 3) for mobile phase 1 and 70% A + 30% B for mobile phase 2 at a 1 mL/min flow rate and a column temperature of 25 °C for all the tests. The absorbance was registered at all UV–VIS wavelengths. Calibration curves were created with known concentrations (100–2000 ppm) of each pharmaceutical for quantification. Total time for analysis was 30 min.

228

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

2.4. Statistical analysis In order to examine the relationship between mortality and pharmaceuticals concentration, we calculate the Probit logistic regression using XLSTAT (2014.5.03) software:   pffiffiffiffiffiffiZ 2 p ¼ 1= 2π −∞…βX exp −x =2 ∂x where βX represents the linear combination of variables (including constants). The knowledge of the distribution of the event being studied gives the likelihood of the sample. To estimate the β parameters of the model (the coefficients of the linear function), the software tries to maximize the likelihood function by using a Newton–Raphson algorithm. All values are presented with their confidence intervals (95%), dose–response curves by logistic regression are represented and LC50 is calculated. Model results were statistically tested by using a Chisquare test on the logarithm of the likelihood ratio that evaluates if the variables bring significant information by comparing the model as it is defined with a simpler model with only one constant. Pearson correlation coefficients (R2) were also calculated for each analysis using the SPSS 18.00 software (IBM SPSS software). 3. Results The dose–response data with modeled dose–response curves by logistic regression are presented in Figs. 1 and 2. The calculated LC50 with their confidence intervals are shown in Table 2 together with the Pearson's correlation coefficient values (R2). Regarding the validity of tests, the mortality in the negative controls never exceeds 10% at the end of each test. All the earthworms exposed to the positive control of chloroacetamide died within the exposure time. As shown, all of the R2 values were above 0.950 with the exception of simvastatin and propranolol, which had an R2 of 0.824 and 0,899 respectively. All of the derived toxicity values analyzed with the Chi-square test were highly significant (P b 0.0001). The toxicity in earthworms of the 18 selected drugs varied according to the chemical category. 3.1. NSAIDs All of the tested non-steroidal anti-inflammatory drugs caused toxicity in E. fetida with LC50 values from 693.50 to 64.80 mg/kg. The most toxic compound for this earthworm is ibuprofen, followed by diclofenac, salicylic acid and paracetamol (Table 2). The shapes of the dose–response curves (Fig. 1) were different for the four NSAIDs. Ibuprofen and diclofenac exhibit steep dose–response curves at concentrations over 50–100 mg/kg, and salicylic acid has a

smoother curve that starts at the same concentrations but increases more slowly until reaching higher concentrations. Paracetamol shows a different pattern, that is, it has no toxicity at concentrations below 700 mg/kg and then has a very sharp slope dose–response curve. It is noteworthy that diclofenac and ibuprofen can be toxic at concentrations lower than 50 mg/kg. As cited in the Materials and methods section, pharmaceuticals were placed in soil samples without worms. After 14 days, the drugs were extracted from the soil and their chemical structure and quantity were evaluated, if possible. HPLC analyses for a pure drug solution at 1000 ppm and for the extracted drug after 14 days are shown in Figs. 3 to 6. Table 3 shows information related to the mobile phase and wavelength for each chromatography test, together with the removal efficiency for the non-degraded compounds. Although the absorbance was registered at all wavelengths, the results are given for the wavelength at which the greatest differences between the original compound and the structure were observed after 14 days (Table 3). Fig. 3a–h shows chromatograms for ibuprofen (a and b), paracetamol (Fig. 3c, and d), salicylic acid (Fig. 3e and f) and diclofenac (Fig. 3g and h). Clear evidence of degradation can be observed for the four NSAIDs because the chromatograms of their original chemical structures (Fig. 3a, c, e and g) are not equal to their chromatograms after 14 days (Fig. 3b, d, f and h). For ibuprofen and diclofenac, the differences are related to another growing peak that occurs practically at the same retention time as the original one (Fig. 3a and b, and g and h, respectively). However, new peaks for paracetamol and salicylic acid are observed some minutes before the original peaks (Fig. 3c, d, e and f, respectively). 3.2. Blood lipid-lowering agents The tested blood lipid-lowering agents exhibited different toxicity effects in E. fetida with LC50 values from 92.70 to 795.05 mg/kg. The most toxic compound was simvastatin, followed by gemfibrozil and lovastatin (Fig. 2). Bezafibrate does not exhibit lethal effects below 850 mg/kg. A single dead worm was observed in only one case of the replicates at the maximum tested concentration (2000 mg/kg), and thus the bezafibrate LC50 is indicated as higher than 2000 mg/kg in Table 2. Atorvastatin does not present detectable lethal effects below 2000 mg/kg. Simvastatin and lovastatin present a smoothly sloped dose–response curve that increases progressively, but simvastatin is still more toxic at lower concentrations. In fact, although simvastatin has a similar LC50 to that of diclofenac, it is the most toxic drug at low concentrations of all of the tested chemicals. Gemfibrozil has a more steeply sloping dose–response curve that rapidly increases from 100 mg/kg.

Fig. 1. Dose–response curves for non-steroidal anti-inflammatory drugs studied for Eisenia fetida: ibuprofen, diclofenac, salicylic acid and paracetamol. Pale gray lines indicate the confidence limits. Each point is the average value of three replicates.

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

229

Fig. 2. Dose–response curves for acute toxic blood lipid lowering agent drugs studied for Eisenia fetida: simvastatin, gemfibrozil and lovastatin. Pale gray lines indicate the confidence limits. Each point is the average value of three replicates.

Chromatograms for blood lipid-lowering agents are shown in Fig. 4a–j for both a 1000 ppm solution of the drug in its original structure and the drug that was collected after 14 days. These chemicals seem to be much more stable than NSAIDs, with bezafibrate (Fig. 4g–h) being the only degraded drug (a new peak appears on the chromatogram before the standard retention time) at the end of the test. Generally speaking, good removal efficiencies (between 60 and 70%) were found for the compounds that were not degraded, namely gemfibrozil, simvastatin and lovastatin. Only atorvastatin had a removal efficiency below 50% (Table 3). 3.3. β-Blockers None of the six tested β-blockers (propranolol, atenolol, acebutolol, metoprolol, timolol and nadolol) caused lethal effects in E. fetida below the highest dose (2000 mg/kg). When a higher dose of 5000 mg/kg was assayed in the case of propranolol a high LC50 value (3298.62 mg/kg) was obtained. All these data would seem to indicate that these drugs lead to low toxicity in E. fetida. Only the β-blocker timolol seems to be degraded after 14 days, as shown in Fig. 5g and h. Metoprolol (Fig. 5a and b), nadolol (Fig. 5c and d), propranolol (Fig. 5e and f), acebutolol (Fig. 5i and j) and atenolol (Fig. 5k and l) apparently maintain the original structure during the assay time because no additional peaks or peak deformations are observed over the wavelength range. The removal efficiencies are above 50% for these drugs, except for atenolol and acebutolol, which have values of 15 and 39%, respectively (Table 3). 3.4. Antibiotics Finally, three antibiotics (sulphamethoxazole, trimethoprim and tetracycline) were tested and, similar to the β-blockers, none of them were toxic to E. fetida below 2000 mg/kg. In addition, sulphamethoxazole was tested at 4000 mg/kg, and all of the earthworms survived in the replicates at this concentration. Thus, the initial screening of these products exerted low toxicity over E. fetida. Fig. 6a and b shows the chromatograms for the sulphamethoxazole 1000 ppm standard solution and sulphamethoxazole supernatant collected after 14 days, respectively. Although the shapes of the plotted peaks are very similar and the retention times are not modified, a longer peak tail can be observed in Fig. 6b. This finding might indicate that sulphamethoxazole degradation could have taken place. The trimethoprim chromatogram for the original compound shows a double peak, with the first maximum higher than the second maximum

(Fig. 6c). After 14 days, we observed the opposite maximum order, that is, the first peak is lower than the second one (Fig. 6d). The scale for the OX axis was plotted differently from Fig. 6c to d to detect this subtle change in the chromatograms. Tetracycline was completely degraded in 14 days, as seen in Fig. 6e and f. Among the entire group of tested drugs, tetracycline is the only one with no presence at all at the end of the experimental period. 4. Discussion Our results show that four NSAIDs (ibuprofen, paracetamol, salicylic acid and diclofenac) and three blood lipid-lowering agents (gemfibrozil, simvastatin and lovastatin) exhibited acute toxicity in E. fetida, with ibuprofen being the most toxic (LC50 = 64.80 mg/kg) and lovastatin the least toxic (LC50 = 795.07 mg/kg). Nevertheless, none of the drugs from the tested β-blockers families caused lethality below 2000 mg/kg. Propranolol shows a LC50 value above 3000 mg/kg. None of the three antibiotics assayed caused detectable lethality. Toxicity from degradation products rather than the original compounds cannot be dismissed because the four tested NSAIDs were found to be degraded during the test time. In addition, the other studied pharmaceuticals, such as bezafibrate, timolol, sulphamethoxazole, trimethoprim and tetracycline, were also degraded. The other nine drugs were apparently not degraded, with the following order from easy to difficult removal from tested soil: metoprolol and gemfibrozil N propranolol, simvastatin and nadolol N acebutolol N atorvastatin N atenolol. The results presented here demonstrate that pharmaceuticals are likely to cause acute adverse effects in terrestrial organism as E. fetida (three of them below 100 mg/kg) but these data must be put in perspective if an environmental risk assessment is made as discussed for each family of drugs below. 4.1. NSAIDs The two most toxic drugs to E. fetida, ibuprofen and diclofenac, are two of the most highly consumed drugs among human populations (Carballa et al., 2008; Ortiz de Garcia et al., 2013), and their presence has been widely reported in the environment. These two pharmaceuticals were detected in the final products, digested sludge and compost, that are applied onto soils. However, LC50 values between 60 and 90 mg/kg are far higher than any concentration of these drugs in the biosolids or recycled water. In addition, these pharmaceutical concentrations tend to be lower in the sludge applied to soils. For example, diclofenac has been detected in wastewater treatment plant effluent

230

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

mAU

a

t (min)

mAU

b

t (min)

mAU

c

t (min) mAU

d

t (min)

mAU

e

t (min)

mAU

f

t (min)

AU

g

t (min)

mAU

h

t (min) Fig. 3. Absorbance units (AU) vs time (t) for HPLC analysis at the cited wavelengths in Table 3 for 1000 ppm concentration of NSAIDs ibuprofen (a), paracetamol (c), salicylic acid (e) and diclofenac (g); and for the supernatants collected from soil samples after 14 days for sample with ibuprofen (b), sample with paracetamol (d), sample with salicylic acid (f) and sample with diclofenac (h).

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

231

mAU

a

t (min) mAU

b

t (min)

mAU

c

t (min)

mAU

d

t (min)

mAU

e

t (min)

mAU

f

t (min)

AU

g

t (min)

mAU

h

t (min) mAU

i

t (min) mAU

j

t (min) Fig. 4. Absorbance units (AU) vs time (t) for HPLC analysis at the cited wavelengths in Table 3 for 1000 ppm concentration of blood lipid lowering agents simvastatin (a), atorvastatin (c), gemfibrozil (e), bezafibrate (g) and lovastatin (i); and for the supernatants collected from soil samples after 14 days for sample with simvastatin (b), sample with atorvastatin (d), sample with gemfibrozil (f), sample with bezafibrate (h) and lovastatin (j).

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

a

mAU

232

t (min) mAU

b

t (min)

AU

c

t (min) mAU

d

t (min)

AU

e

t (min) AU

f

mAU

t (min)

g

mAU

t (min)

h

t (min) mAU

i

mAU

t (min)

j

t (min) AU

k

mAU

t (min)

l

t (min) Fig. 5. Absorbance units (AU) vs time (t) for HPLC analysis at the cited wavelengths in Table 3 for 1000 ppm concentration of β-blockers metoprolol (a), nadolol (c), propranolol (e) timolol (g), acebutolol (i) and atenolol (k); and for the supernatants collected from soil samples after 14 days for sample with metoprolol (b), sample with nadolol (d), sample with propranolol (f), sample with timolol (h), acebutolol (j) and atenolol (l).

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

233

AU

a

t (min)

mAU

b

t (min)

mAU

c

t (min)

mAU

d

t (min)

mAU

e

t (min)

mAU

f

t (min) Fig. 6. Absorbance units (AU) vs time (t) for HPLC analysis at the cited wavelengths in Table 3 for 1000 ppm concentration of antibiotics sulfamethoxazole (a), trimethoprim (c) and tetracycline (e); and for the supernatants collected from soil samples after 14 days for sample with sulfamethoxazole (b), sample with trimethoprim (d), sample with tetracycline (f).

in Germany (Stuelten et al., 2008; Ternes, 1998), Spain (Rosal et al., 2010) and Canada (Lishman et al., 2006) having been reported the average diclofenac concentrations in wastewater from 0.2 μg/L to 220 ng/L. Ibuprofen was also found in wastewater at values ranged from 8.45 μg/L to 135 ng/L. Concerning the presence of those two

drugs in sewage sludge, ibuprofen concentration levels were up to 741.1 ng/g dry weight (Martin et al., 2012a; Nieto et al., 2010) and up to 380,7 ng/g dry weight for diclofenac (Radjenovic et al., 2009). Some biosolids also contained ibuprofen at concentrations of 246 μg/kg dry weight (Albero et al., 2014; McClellan and Halden,

234

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

Table 3 Chemicals used for ecotoxicity tests, mobile phase and wavelength selected for their HPLC and % removal efficiency from soil samples. Chemical

Mobile phase

Selected wavelength

% removal efficiency

1

197 nm 197 nm 220 nm 268 nm

Degraded Degraded Degraded Degraded

Blood lipid lowering agents Gemfibrozil 2 Simvastatin Lovastatin Atorvastatin 1 Bezafibrate

199 nm 242 nm 235 nm 243 nm 197 nm

76 67 61 27 Degraded

β-Blockers Propranolol Atenolol Acebutolol Metoprolol Timolol Nadolol

1

197 nm 197 nm 241 nm 222 nm 215 nm 200 nm

68 15 39 78 Degraded 63

1

197 nm 237 nm 235 nm

Degraded Degraded Degraded

NSAIDs Ibuprofen Diclofenac Paracetamol Salicylic acid

Antibiotics Sulfamethoxazole Trimethoprim Tetracycline

2010) and paracetamol at average concentrations of 130 μg/kg approximately (Albero et al., 2014). Some authors reported the presence of paracetamol in STP effluents at concentrations between 20 ng/L (Roberts and Thomas, 2006) and 246 μg/L (Gomez et al., 2007) and salicylic acid at average concentrations of 0.106 μg/L (Lishman et al., 2006), in primary sludge (561 μg/kg dm), in digested sludge (32.9 μg/kg dm) (Martin et al., 2012a) and in biosolids at average concentrations of 547 μg/kg (Albero et al., 2014). These concentrations are far lower than toxic concentrations detected in our results with E. fetida. Environmental risk assessment of pharmaceutical compounds found in sludge showed potential risks in the case of the NSAID ibuprofen (Martin et al., 2012b) in digested sludge (RQ value: 3.70) as well as in compost (RQ value: 1.57). However, the usual dosage of sludge applied to soil leads to a decrease in pharmaceutical concentrations and, consequently, a disappearance of ecotoxicological risk in the case of compost (Martin et al., 2012b). Concerning the stability of the studied drugs, the toxic effect of the NSAIDs may be due to their original chemical structure as well as to any of their degradation products, because the four NSAIDs were degraded after 14 days. These degradation data are consistent with other authors for ibuprofen or diclofenac (Grossberger et al., 2014; Lin and Tsai, 2009; Lin and Gan, 2011; Xu et al., 2009). For paracetamol (Loffler et al., 2005) and salicylic acid (Martin et al., 2012a), rapid degradation processes were also reported. However, very divergent degradation data for a given compound can be found probably due to different experimental conditions and the soil type. For ibuprofen or diclofenac, the studied degradation time (t1/2) diverges from 0.91 days to 6.09 days and from 3.07 to 20.44 days, respectively (Grossberger et al., 2014; Xu et al., 2009). Other authors (Lin and Gan, 2011) found a t1/2 for ibuprofen of between 10 and 50 days. Very little data about ecotoxicity of degradation products of these drugs are available in the literature. For ibuprofen (Illes et al., 2013) and diclofenac (Coelho et al., 2009) the degradation products are generated due to degradation induced in treatments for the removal of these compounds from effluents of STP, and considers how their ecotoxicity is modified, which is a very different situation to the conditions of this study.

4.2. Blood lipid-lowering agents Although simvastatin is one of the most commonly used blood lipidlowering agents in the world, its related data are very scarce. Observed effluent concentrations of this drug are quite low, roughly 100–300 ng/L (Ottmar et al., 2012), much lower than LC50 values for E. fetida. Gemfibrozil and lovastatin have been also detected in wastewater treatment plant effluents at concentrations below 1 μg/L (Lishman et al., 2006; Rosal et al., 2010) and at concentrations below 1 μg/L (Conley et al., 2008) respectively. No data available for the presence of simvastatin and lovastatin was found in biosolids. However traces of gemfibrozil were found in biosolids in the range of ng/L (Albero et al., 2014). Martin et al. (2012b) found environmental risks in digested sludge for gemfibrozil (RQ value: 4.63) but ecotoxicological risk cannot be detected in the case of compost. Gemfibrozil, simvastatin and lovastatin showed a lethal effect on E. fetida and their original chemical structure remained for the time assay, so this toxic effect is due to these drugs and not because of a possible degradation product. Only bezafibrate was found to be degraded after 14 days but no toxic effect was observed when tested on E. fetida so neither this drug nor any of its degraded products are toxic under experimental conditions. Consistency related to drug recovery from soil and degradation was found when comparing our experimental results with reported values for bezafibrate (Grossberger et al., 2014; Maeng et al., 2011, 2012) and gemfibrozil (Fang et al., 2012; Grossberger et al., 2014; Maeng et al., 2011, 2012). The estimated degradation time for gemfibrozil was 18 days in the study performed by Fang et al. (2012), between 27 and 94 days in the study performed by Lin and Gan (2011), and between 10 and 68 days in the study performed by Grossberger et al. (2014). These authors also reported a t1/2 from 7 to 44 days for bezafibrate. The degradation time for gemfibrozil was estimated to be 18 days in the study performed by Fang et al. (2012), between 27 and 94 days for Lin and Gan (2011) and between 10 and 68 days for Grossberger et al. (2014). These authors also reported a t1/2 from 7 to 44 days for bezafibrate. 4.3. β-Blockers None of the tested β-blockers cause lethal effects on E. fetida because they could be removed from soil in their original chemical structures. Toxicity effects can be detected for propranolol but only at very high concentrations. Our results show that atenolol and acebutolol were the most persistent drugs in the soil because their removal efficiencies (15% and 39%, respectively) are low and no degradation was observed after 14 days. Propranolol, metoprolol and nadolol were all within the same range for removal efficiency (60–70%) and were not degraded, in contrast to information from other authors (Lin et al., 2010). Lin et al. performed sorption and combined sorption–biodegradation experiments for several pharmaceuticals, including propranolol and acebutolol, for 35 days. Propranolol and acebutolol concentrations decreased by 83% and 90%, respectively, after 14 days, and their half-lives were 2.2 and 2.4 days, respectively. Finally, timolol seems to be degraded after 14 days, so in this case, it cannot be guaranteed that the non-toxic effect is due to its original chemical structure but because of any of its degradation products. 4.4. Antibiotics None of the tested antibiotics are likely to cause acute toxic effects on E. fetida worms but all of them were degraded after 14 days, so two situations may occur: a) these drugs are so quickly degraded that their effect on E. fetida can't be observed in such a long time and their degraded products are non-toxic, b) neither the original drugs nor the degraded products are toxic on E. fetida.

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

Otherwise, the presence of these drugs in biosolids has not been reported. Tetracycline is the only antibiotic occasionally detected at concentrations in the range of mg in biosolids (Hamscher et al., 2002; McClellan and Halden, 2010). Degradation of sulphamethoxazole is consistent with data offered by Lin and Gan (2011), in which the t1/2 for this substance is estimated to be between 9 and 60 days. Kim et al. (2005) found no evidence of tetracycline biodegradation over the test duration. Bansal (2012) studied tetracycline behavior in different soil types, and this author found t1/2 values ranging from 27 to 52 days, depending on the soil type and water/soil proportion. Once again, the information from different authors is divergent, primarily because of the heterogeneity of experimental conditions. The tetracycline in our tests was completely degraded after 14 days. Lin and Gan (2011) reported completely different data in the same work for trimethoprim when its behavior is compared to two soils. This drug is well adsorbed in one type of soil, but 100% of the initial drug is recovered from the other. Degradation exhibits the same trend, and this compound is barely degraded in one tested soil, whereas there is a t1/2 value equal to 26 days for the other. Summarizing, it appears that the concentrations at which these drugs may produce toxic effects are much higher than those found in wastewater or biosolids, the main input channels of these substances to the soil. Therefore, we could discard acute toxicity effects. These results are consistent with other studies of veterinary pharmaceuticals products on E. fetida in which no toxic effects were found or they occurred only at very high concentrations. Boleas et al. (2005) studied the effects of oxytetracycline on soil organisms using a multispecies-soil system and no mortality was observed for E. fetida at the highest concentration 100 mg/kg soil. Similar results were obtained by Qu et al. (2005). Baguer et al. (2000) tested the effects of antibiotics, tylosin and oxytetracycline, on three species of soil fauna: earthworms (Aporrectodea caliginosa), springtails and enchytraeids and the lowest observed effect concentration was 3000 mg/kg and in many cases no effect was seen even at the highest test concentration of 5000 mg/kg. However, data about concentrations of these drugs in soils are very scarce and we could not find specific information about the concentrations at which these 18 drugs can occur in soils fertilized with wastewater or biosolids, Literature suggests that generally, the concentrations found in the field are 2–3° of magnitude lower than in the biosolids, which is likely to be due to dilution, degradation and leaching processes (Wu et al., 2010a). But it is noteworthy that application of municipal biosolids and manure is a common practice worldwide so the transfer of pharmaceuticals to soil is globally relevant (EPA, 2006; Chang et al., 2002). Once in soil, compounds with strong sorption and recalcitrant to degradation could remain on the surface and accumulate after repeated applications of biosolids or irrigation. Therefore, although these concentrations are not immediately toxic, accumulative and long-term effects cannot be ruled out. Kinney et al. (2008) described that earthworms in common agricultural soil environments can accumulate a wide range of chemically diverse organic contaminants originated from biosolids or manure applied to terrestrial ecosystems. Moreover, through predation of earthworms, these compounds could be further dispersed beyond the point of application in terrestrial ecosystems. In any case, it seems clear that the presence of some pharmaceuticals in the soil must be sufficiently high and persistent because different field studies have found that pharmaceutical products can also be taken up by plants that are grown in soils fertilized with biosolids (Aryal and Reinhold, 2011; Gottschall et al., 2012; Holling et al., 2012). Therefore, acute laboratory tests presented in this study provide a valuable data in the absence of prior information about the effect of these drugs in E. fetida but it should be considered as a preliminary step in assessments of environmental risk that can guide future studies. It might be appropriate for future research to consider chronic effects of these substances on earthworm behavior, growth, and reproduction

235

that might indirectly affect soil fertility or modify the quality of earthworms as a food source (Kelsey et al., 2005; Klok et al., 2006), especially in families such as NSAIDs and blood lipid-lowering agents where it seems that the toxicity is greater. In addition, although E. fetida is a recommended species for standardized acute toxicity tests (OECD 207, 1984; Fourie et al., 2007) and still regarded as suitable organism for ecological risk assessment in terrestrial ecosystems in recent studies (Chen et al., 2014), is an epigenic species being located in organic matter on the soil surface and has a distribution that is mainly restricted to woodlands and accumulations of organic matter. Therefore the use of more ecologically relevant soil dwelling species as a model for environmental assessment is necessary for extrapolation of laboratory tests to field conditions. A complementary ecotoxicity studies should be also carried out in species from different trophic chains such as plants or microorganisms. On the other hand, the absence of detectable acute toxicity data for certain compounds, such as some β-blockers or antibiotics, does not mean that they are innocuous. There is a general consensus that the toxic effects of a certain contaminants are first triggered at the molecular level. For that reason, molecular biology techniques are becoming more popular for studying drug consequences at the genetic level (De Felice et al., 2012; Lin et al., 2012; Ragugnetti et al., 2011). For instance, DNA damage and biochemical toxicity of soil tetracyclines or oxytetracyclines in the earthworm E. fetida (Dong et al., 2012; Lin et al., 2012) have been verified, although our results demonstrate that they are not toxic. Finally, more studies are needed to determine the role of the degradation products of these pharmaceuticals on toxicity. Only after filling these gaps can more reliable environmental risk assessments with much lower uncertainty be performed.

5. Conclusions Our results reveal that four NSAIDs and three blood lipid-lowering agents show acute toxicity in the terrestrial non-target organism, E. fetida under laboratory conditions. From an environmental perspective, the lethal concentrations of the tested drugs are much greater than those reported in the terrestrial environment, so acute toxic effects may be improbable. However, accumulation and chronic effects cannot be excluded because of the regular practice of land application of biosolids and wastewater containing these substances. Therefore, the toxicity data of this study provide essential baseline information and represent an important first step that should guide further research in this area as long-term studies of chronic effects. In addition, the four NSAIDs analyzed in soil, show clear evidence of modification of their original chemical structure after 14 days but blood lipid-lowering agents seem to be more stable. These results indicate that the degradation products can play a key role in the toxicity of these products.

Conflict of interest There is not any actual or potential conflict of interest including any financial, personal or other relationships with other people or organizations within the three years from the beginning of the submitted work that could inappropriately influence, or be perceived to influence the work.

Acknowledgments The authors thank TEVA PHARMA for providing the atorvastatin and simvastatin used in this study.

236

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237

References Albero, B., Sanchez-Brunete, C., Miguel, E., Aznar, R., Tadeo, J.L., 2014. Determination of selected pharmaceutical compounds in biosolids by supported liquid extraction and gas chromatography–tandem mass spectrometry. J. Chromatogr. A 1336, 52–58. Ankley, G.T., Brooks, B.W., Huggett, D.B., Sumpter, J.P., 2007. Repeating history: pharmaceuticals in the environment. Environ. Sci. Technol. 41, 8211–8217. Aryal, N., Reinhold, D.M., 2011. Phytoaccumulation of antimicrobials from biosolids: impacts on environmental fate and relevance to human exposure. Water Res. 45, 5545–5552. Baguer, A.J., Jensen, J., Krogh, P.H., 2000. Effects of the antibiotics oxytetracycline and tylosin on soil fauna. Chemosphere 40, 751–757. Bansal, O.P., 2012. A laboratory study on degradation studies of tetracycline and chlortetracycline in soils of Aligarh district as influenced by temperature, water content, concentration of farm yield manure, nitrogen and tetracyclines. Proc. Natl. Acad. Sci. India Sect. B Biol. Sci. 82, 503–509. Bliss, C.I., 1934. The method of probits. Science 79, 38–39. Boleas, S., Alonso, C., Pro, J., Fernandez, C., Carbonell, G., Tarazona, J.V., 2005. Toxicity of the antimicrobial oxytetracycline to soil organisms in a multi-species-soil system (MS center dot 3) and influence of manure co-addition. J. Hazard. Mater. 122, 233–241. Boxall, A.B.A., Johnson, P., Smith, E.J., Sinclair, C.J., Stutt, E., Levy, L.S., 2006. Uptake of veterinary medicines from soils into plants. J. Agric. Food Chem. 54, 2288–2297. Carballa, M., Omil, F., Lema, J.M., 2008. Comparison of predicted and measured concentrations of selected pharmaceuticals, fragrances and hormones in Spanish sewage. Chemosphere 72, 1118–1123. Chang, A.C., Pan, G., Page, A.L., Asano, T., 2002. Developing Human Health-related Chemical Guidelines for Reclaimed Water and Sewage Sludge Applications in Agriculture. World Health Organization, Geneva. Chen, C., Wang, Y., Zhao, X., Qian, Y., Wang, Q., 2014. Combined toxicity of butachlor, atrazine and lambda-cyhalothrin on the earthworm Eisenia fetida by combination index (CI)-isobologram method. Chemosphere 112, 393–401. Coelho, A.D., Sans, C., Aguera, A., Gomez, M.J., Esplugas, S., Dezotti, M., 2009. Effects of ozone pre-treatment on diclofenac: intermediates, biodegradability and toxicity assessment. Sci. Total Environ. 407, 3572–3578. Conley, J.M., Symes, S.J., Schorr, M.S., Richards, S.M., 2008. Spatial and temporal analysis of pharmaceutical concentrations in the upper Tennessee River basin. Chemosphere 73, 1178–1187. Cortez, J., Bouche, M.B., 1992. Do earthworms eat living roots. Soil Biol. Biochem. 24, 913–915. D'Abrosca, B., Fiorentino, A., Izzo, A., Cefarelli, G., Pascarella, M.T., Uzzo, P., et al., 2008. Phytotoxicity evaluation of five pharmaceutical pollutants detected in surface water on germination and growth of cultivated and spontaneous plants. J. Environ. Sci. Health A Tox. Hazard Subst. Environ. Eng. 43, 285–294. 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. De Felice, B., Copia, L., Guida, M., 2012. Gene expression profiling in zebrafish embryos exposed to diclofenac, an environmental toxicant. Mol. Biol. Rep. 39, 2119–2128. Dolliver, H., Kumar, K., Gupta, S., 2007. Sulfamethazine uptake by plants from manureamended soil. J. Environ. Qual. 36, 1224–1230. Dong, L., Gao, J., Xie, X., Zhou, Q., 2012. DNA damage and biochemical toxicity of antibiotics in soil on the earthworm Eisenia fetida. Chemosphere 89, 44–51. Drillia, P., Dokianakis, S.N., Fountoulakis, M.S., Kornaros, M., Stamatelatou, K., Lyberatos, G., 2005. On the occasional biodegradation of pharmaceuticals in the activated sludge process: the example of the antibiotic sulfamethoxazole. J. Hazard. Mater. 122, 259–265. EMEA (European Medicines Agency Inspections), 2006. Guideline on the Requirements to the Chemical and Pharmaceutical Quality Documentation Concerning Investigational Medicinal Products in Clinical Trials. EMEA editions, London (CHMP/QWP/185401/2004 final). Fang, Y., Karnjanapiboonwong, A., Chase, D.A., Wang, J., Morse, A.N., Anderson, T.A., 2012. Occurrence, fate, and persistence of gemfibrozil in water and soil. Environ. Toxicol. Chem. 31, 550–555. Feito, R., Valcarcel, Y., Catala, M., 2013. Preliminary data suggest that venlafaxine environmental concentrations could be toxic to plants. Chemosphere 90, 2065–2069. Fourie, F., Reinecke, S.A., Reinecke, A.J., 2007. The determination of earthworm species sensitivity differences to cadmium genotoxicity using the comet assay. Ecotoxicol. Environ. Saf. 67, 361–368. Furtula, V., Stephenson, G.L., Olaveson, K.M., Chambers, P.A., 2012. Effects of the veterinary pharmaceutical salinomycin and its formulation on the plant Brassica rapa. Arch. Environ. Contam. Toxicol. 63, 513–522. Gomez, M.J., Bueno, M.J.M., Lacorte, S., Fernandez-Alba, A.R., Aguera, A., 2007. Pilot survey monitoring pharmaceuticals and related compounds in a sewage treatment plant located on the Mediterranean coast. Chemosphere 66, 993–1002. Gottschall, N., Topp, E., Metcalfe, C., Edwards, M., Payne, M., Kleywegt, S., et al., 2012. Pharmaceutical and personal care products in groundwater, subsurface drainage, soil, and wheat grain, following a high single application of municipal biosolids to a field. Chemosphere 87, 194–203. Grossberger, A., Hadar, Y., Borch, T., Chefetz, B., 2014. Biodegradability of pharmaceutical compounds in agricultural soils irrigated with treated wastewater. Environ. Pollut. 185, 168–177. Hamscher, G., Sczesny, S., Hoper, H., Nau, H., 2002. Determination of persistent tetracycline residues in soil fertilized with liquid manure by high-performance liquid chromatography with electrospray ionization tandem mass spectrometry. Anal. Chem. 74, 1509–1518. Heidler, J., Halden, R.U., 2008. Meta-analysis of mass balances examining chemical fate during wastewater treatment. Environ. Sci. Technol. 42, 6324–6332.

Heidler, J., Sapkota, A., Halden, R.U., 2006. Partitioning, persistence, and accumulation in digested sludge of the topical antiseptic triclocarban during wastewater treatment. Environ. Sci. Technol. 40, 3634–3639. Hillis, D.G., Antunes, P., Sibley, P.K., Klironomos, J.N., Solomon, K.R., 2008. Structural responses of Daucus carota root-organ cultures and the arbuscular mycorrhizal fungus, Glomus intraradices, to 12 pharmaceuticals. Chemosphere 73, 344–352. Holling, C.S., Bailey, J.L., Heuvel, B.V., Kinney, C.A., 2012. Uptake of human pharmaceuticals and personal care products by cabbage (Brassica campestris) from fortified and biosolids-amended soils. J. Environ. Monit. 14, 3029–3036. Illes, E., Takacs, E., Dombi, A., Gajda-Schrantz, K., Racz, G., Gonter, K., et al., 2013. Hydroxyl radical induced degradation of ibuprofen. Sci. Total Environ. 447, 286–292. Jacobsen, P., Berglind, L., 1988. Persistence of oxytetracycline in sediments from fish farms. Aquaculture 70, 365–370. Jongbloed, A.W., Lenis, N.P., 1998. Environmental concerns about animal manure. J. Anim. Sci. 76, 2641–2648. Kelsey, J.W., Colino, A., White, J.C., 2005. Effect of species differences, pollutant concentration, and residence time in soil on the bioaccumulation of 2,2-bis(p-chlorophenyl)1,1-dichloroethylene by three earthworm species. Environ. Toxicol. Chem. 24, 703–708. Kemper, N., 2008. Veterinary antibiotics in the aquatic and terrestrial environment. Ecol. Indic. 8, 1–13. Kim, S.C., Carlson, K., 2005. LC–MS2 for quantifying trace amounts of pharmaceutical compounds in soil and sediment matrices. TrAC Trends Anal. Chem. 24, 635–644. Kim, S., Eichhorn, P., Jensen, J.N., Weber, A.S., Aga, D.S., 2005. Removal of antibiotics in wastewater: effect of hydraulic and solid retention times on the fate of tetracycline in the activated sludge process. Environ. Sci. Technol. 39, 5816–5823. Kinney, C.A., Furlong, E.T., Werner, S.L., Cahill, J.D., 2006. Presence and distribution of wastewater-derived pharmaceuticals in soil irrigated with reclaimed water. Environ. Toxicol. Chem. 25, 317–326. Kinney, C.A., Furlong, E.T., Kolpin, D.W., Burkhardt, M.R., Zaugg, S.D., Werner, S.L., et al., 2008. Bioaccumulation of pharmaceuticals and other anthropogenic waste indicators in earthworms from agricultural soil amended with biosolid or swine manure. Environ. Sci. Technol. 42, 1863–1870. Klok, C., Van der Hout, A., Bodt, J., 2006. Population growth and development of the earthworm Lumbricus rubellus in a polluted field soil: possible consequences for the godwit (Limosa limosa). Environ. Toxicol. Chem. 25, 213–219. Kong, W.D., Zhu, Y.G., Fu, B.J., Marschner, P., He, J.Z., 2006. The veterinary antibiotic oxytetracycline and Cu influence functional diversity of the soil microbial community. Environ. Pollut. 143, 129–137. Kong, W.D., Zhu, Y.G., Liang, Y.C., Zhang, J., Smith, F.A., Yang, A., 2007. Uptake of oxytetracycline and its phytotoxicity to alfalfa (Medicago sativa L.). Environ. Pollut. 147, 187–193. Kumar, K., Gupta, S.C., Baidoo, S.K., Chander, Y., Rosen, C.J., 2005. Antibiotic uptake by plants from soil fertilized with animal manure. J. Environ. Qual. 34, 2082–2085. Lin, K., Gan, J., 2011. Sorption and degradation of wastewater-associated non-steroidal anti-inflammatory drugs and antibiotics in soils. Chemosphere 83, 240–246. Lin, A.Y.-C., Tsai, Y.-T., 2009. Occurrence of pharmaceuticals in Taiwan's surface waters: impact of waste streams from hospitals and pharmaceutical production facilities. Sci. Total Environ. 407, 3793–3802. Lin, A.Y.-C., Lin, C.-A., Tung, H.-H., Chary, N.S., 2010. Potential for biodegradation and sorption of acetaminophen, caffeine, propranolol and acebutolol in lab-scale aqueous environments. J. Hazard. Mater. 183, 242–250. Lin, D., Zhou, Q., Xu, Y., Chen, C., Li, Y., 2012. Physiological and molecular responses of the earthworm (Eisenia fetida) to soil chlortetracycline contamination. Environ. Pollut. 171, 46–51. Lishman, L., Smyth, S.A., Sarafin, K., Kleywegt, S., Toito, J., Peart, T., et al., 2006. Occurrence and reductions of pharmaceuticals and personal care products and estrogens by municipal wastewater treatment plants in Ontario, Canada. Sci. Total Environ. 367, 544–558. Liu, F., Ying, G.-G., Tao, R., Zhao, J.-L., Yang, J.-F., Zhao, L.-F., 2009. Effects of six selected antibiotics on plant growth and soil microbial and enzymatic activities. Environ. Pollut. 157, 1636–1642. Loffler, D., Ternes, T.A., 2003. Determination of acidic pharmaceuticals, antibiotics and ivermectin in river sediment using liquid chromatography–tandem mass spectrometry. J. Chromatogr. A 1021, 133–144. Loffler, D., Rombke, J., Meller, M., Ternes, T.A., 2005. Environmental fate of pharmaceuticals in water/sediment systems. Environ. Sci. Technol. 39, 5209–5218. Maeng, S.K., Sharma, S.K., Abel, C.D.T., Magic-Knezev, A., Amy, G.L., 2011. Role of biodegradation in the removal of pharmaceutically active compounds with different bulk organic matter characteristics through managed aquifer recharge: batch and column studies. Water Res. 45, 4722–4736. Maeng, S.K., Sharma, S.K., Abel, C.D.T., Magic-Knezev, A., Song, K.-G., Amy, G.L., 2012. Effects of effluent organic matter characteristics on the removal of bulk organic matter and selected pharmaceutically active compounds during managed aquifer recharge: column study. J. Contam. Hydrol. 140, 139–149. Markman, S., Guschina, I.A., Barnsley, S., Buchanan, K.L., Pascoe, D., Mueller, C.T., 2007. Endocrine disrupting chemicals accumulate in earthworms exposed to sewage effluent. Chemosphere 70, 119–125. Martin, J., Camacho-Munoz, D., Santos, J.L., Aparicio, I., Alonso, E., 2012a. Occurrence of pharmaceutical compounds in wastewater and sludge from wastewater treatment plants: removal and ecotoxicological impact of wastewater discharges and sludge disposal. J. Hazard. Mater. 239, 40–47. Martin, J., Dolores Camacho-Munoz, M., Luis Santos, J., Aparicio, I., Alonso, E., 2012b. Distribution and temporal evolution of pharmaceutically active compounds alongside sewage sludge treatment. Risk assessment of sludge application onto soils. J. Environ. Manag. 102, 18–25.

M.ªR. Pino et al. / Science of the Total Environment 518–519 (2015) 225–237 McClellan, K., Halden, R.U., 2010. Pharmaceuticals and personal care products in archived US biosolids from the 2001 EPA national sewage sludge survey. Water Res. 44, 658–668. Miege, C., Choubert, J.M., Ribeiro, L., Eusebe, M., Coquery, M., 2009. Fate of pharmaceuticals and personal care products in wastewater treatment plants — conception of a database and first results. Environ. Pollut. 157, 1721–1726. Migliore, L., Cozzolino, S., Fiori, M., 2003. Phytotoxicity to and uptake of enrofloxacin in crop plants. Chemosphere 52, 1233–1244. Nieto, A., Borrull, F., Pocurull, E., Marce, R.M., 2010. Occurrence of pharmaceuticals and hormones in sewage sludge. Environ. Toxicol. Chem. 29, 1484–1489. OECD (Organisation for Economic Cooperation and Development), 1984. Earthworm, acute toxicity tests. OECD-Guideline for Testing Chemicals vol. 207. OECD publications, Paris. Ortiz de Garcia, S., Pinto Pinto, G., Garcia Encina, P., Irusta Mata, R., 2013. Consumption and occurrence of pharmaceutical and personal care products in the aquatic environment in Spain. Sci. Total Environ. 444, 451–465. Ottmar, K.J., Colosi, L.M., Smith, J.A., 2012. Fate and transport of atorvastatin and simvastatin drugs during conventional wastewater treatment. Chemosphere 88, 1184–1189. Qu, M., Xu, Y., Chen, H., Li, Z., Sun, L., Xu, D., et al., 2005. Toxicological study of three veterinary drugs on Eisenia fetida. Ying Yong Sheng Tai Xue Bao 16, 1108–1111. Radjenovic, J., Petrovic, M., Barcelo, D., 2009. Fate and distribution of pharmaceuticals in wastewater and sewage sludge of the conventional activated sludge (CAS) and advanced membrane bioreactor (MBR) treatment. Water Res. 43, 831–841. Ragugnetti, M., Adams, M.L., Guimaraes, A.T.B., Sponchiado, G., de Vasconcelos, E.C., Ribas de Oliveira, C.M., 2011. Ibuprofen genotoxicity in aquatic environment: an experimental model using Oreochromis niloticus. Water Air Soil Pollut. 218, 361–364. 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. Rosal, R., Rodriguez, A., Antonio Perdigon-Melon, J., Petre, A., Garcia-Calvo, E., Jose Gomez, M., et al., 2010. Occurrence of emerging pollutants in urban wastewater and their removal through biological treatment followed by ozonation. Water Res. 44, 578–588. Schmitt-Jansen, M., Bartels, P., Adler, N., Altenburger, R., 2007. Phytotoxicity assessment of diclofenac and its phototransformation products. Anal. Bioanal. Chem. 387, 1389–1396.

237

Spurgeon, D.J., Weeks, J.M., Van Gestel, C.A.M., 2003. A summary of eleven years progress in earthworm ecotoxicology. Pedobiologia 47, 588–606. Stuelten, D., Zuehlke, S., Lamshoeft, M., Spiteller, M., 2008. Occurrence of diclofenac and selected metabolites in sewage effluents. Sci. Total Environ. 405, 310–316. Ternes, T.A., 1998. Occurrence of drugs in German sewage treatment plants and rivers. Water Res. 32, 3245–3260. Ternes, T.A., Joss, A., Siegrist, H., 2004. Scrutinizing pharmaceuticals and personal care products in wastewater treatment. Environ. Sci. Technol. 38, 392A–399A. Topp, E., Monteiro, S.C., Beck, A., Coelho, B.B., Boxall, A.B.A., Duenk, P.W., et al., 2008. Runoff of pharmaceuticals and personal care products following application of biosolids to an agricultural field. Sci. Total Environ. 396, 52–59. U.S. Environmental Protection Agency, 2006. Emerging Technologies for Biosolid Management. U.S. Environmental Protection Agency, Washington, DC (EPA 832-R-06-005). Williams, C.F., Williams, C.E., Adamsen, E.J., 2006. Sorption–desorption of carbamazepine from irrigated soils. J. Environ. Qual. 35, 1779–1783. Wu, C., Spongberg, A.L., Witter, J.D., Fang, M., Ames, A., Czajkowski, K.P., 2010a. Detection of pharmaceuticals and personal care products in agricultural soils receiving biosolids application. Clean-Soil, Air, Water 38, 230–237. Wu, C., Spongberg, A.L., Witter, J.D., Fang, M., Czajkowski, K.P., 2010b. Uptake of pharmaceutical and personal care products by soybean plants from soils applied with biosolids and irrigated with contaminated water. Environ. Sci. Technol. 44, 6157–6161. Xia, K., Bhandari, A., Das, K., Pillar, G., 2005. Occurrence and fate of pharmaceuticals and personal care products (PPCPs) in biosolids. J. Environ. Qual. 34, 91–104. Xie, X., Zhou, Q., Lin, D., Guo, J., Bao, Y., 2011. Toxic effect of tetracycline exposure on growth, antioxidative and genetic indices of wheat (Triticum aestivum L.). Environ. Sci. Pollut. Res. 18, 566–575. Xu, J., Wu, L., Chang, A.C., 2009. Degradation and adsorption of selected pharmaceuticals and personal care products (PPCPs) in agricultural soils. Chemosphere 77, 1299–1305. Yu, Y., Huang, Q., Wang, Z., Zhang, K., Tang, C., Cui, J., et al., 2011. Occurrence and behavior of pharmaceuticals, steroid hormones, and endocrine-disrupting personal care products in wastewater and the recipient river water of the Pearl River Delta, South China. J. Environ. Monit. 13, 871–878.