Science of the Total Environment 493 (2014) 1122–1126
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Beta-blockers in the environment: Part II. Ecotoxicity study Joanna Maszkowska a, Stefan Stolte a,b, Jolanta Kumirska a,⁎, Paulina Łukaszewicz a, Katarzyna Mioduszewska a, Alan Puckowski a, Magda Caban a, Marta Wagil a, Piotr Stepnowski a, Anna Białk-Bielińska a,b a b
Department of Environmental Analysis, Institute for Environmental and Human Health Protection, Faculty of Chemistry, University of Gdańsk, ul. Wita Stwosza 63, 80-308 Gdańsk, Poland UFT — Center for Environmental Research and Sustainable Technology, University of Bremen, Leobener Straße, D-28359 Bremen, Germany
H I G H L I G H T S • Propranolol and metoprolol can be considered to be harmful to aquatic organisms. • Sorption explicitly inhibits the hazardous effects of beta-blockers. • The risks posed by these compounds for the environment are of minor importance.
a r t i c l e
i n f o
Available online 27 June 2014 Editor: Adrian Covaci Keywords: Green algae Duckweed Soil and marine bacteria
a b s t r a c t The increasing consumption of beta-blockers (BB) has caused their presence in the environment to become more noticeable. Even though BB are safe for human and veterinary usage, ecosystems may be exposed to these substances. In this study, three selected BB: propranolol, metoprolol and nadolol were subjected to ecotoxicity study. Ecotoxicity evaluation was based on a flexible ecotoxicological test battery including organisms, representing different trophic levels and complexity: marine bacteria (Vibrio fischeri), soil/sediment bacteria (Arthrobacter globiformis), green algae (Scenedesmus vacuolatus) and duckweed (Lemna minor). All the ecotoxicological studies were supported by instrumental analysis to measure deviation between nominal and real test concentrations. Based on toxicological data from the green algae test (S. vacuolatus) propranolol and metoprolol can be considered to be harmful to aquatic organisms. However, sorption explicitly inhibits the hazardous effects of BB, therefore the risks posed by these compounds for the environment are of minor importance. © 2014 Elsevier B.V. All rights reserved.
1. Introduction Beta-blockers (BB) are extensively used for the treatment of various cardiovascular disorders such as high blood pressure, ischaemic heart disease and heart rhythm disturbances (British Pharmacopoeia, 2005; The USA Pharmacopeial Convention, 2005). It is estimated that 100– 250 tons of these compounds are consumed in Germany per year (Cleuvers, 2005); several of them are in the top 200 prescribed medications in the United States, too (Huggett et al., 2003). They are also used in veterinary medicine and illegally as doping in many sports (Amendola et al., 2000; Huggett et al., 2002; Barbieri et al., 2012). BB work by blocking the effect of adrenaline on our body's receptors, thereby slowing the nerve impulses to the heart and reducing its workload (British Pharmacopoeia, 2005; The USA Pharmacopeial Convention, 2005). Their widespread use and often incomplete metabolism (Küster et al., 2009) mean that BB are commonly detected in sewage effluents and surface waters (e.g. Huggett et al., 2003; Kümmerer, 2004; Scheurer et al., 2010; Brooks and Huggett, 2012). For example, the ⁎ Corresponding author. Tel.: +48 58 5235212; fax: +48 58 5235454. E-mail address: [email protected] (J. Kumirska).
http://dx.doi.org/10.1016/j.scitotenv.2014.06.039 0048-9697/© 2014 Elsevier B.V. All rights reserved.
concentrations of metoprolol (MET), propranolol (PRO), and nadolol (NAD) in sewage treatment plant (STP) effluents vary between 25 and 2800 ng L−1 (Caban et al., 2013); they are found in rivers and streams, too (Lin and Tsai, 2009). MET, PRO and NAD (Fig. 1) are the most frequently analyzed beta-blockers in biological and environmental samples. Among these three compounds, PRO is the most hydrophobic and shows bioaccumulation potential (Kibbey et al., 2007; Maurer et al., 2007; Wang et al., 2012; Wilde et al., 2013). In our previous study (Maszkowska et al. Beta-blockers in the environment: Part I. Mobility and hydrolysis study. Sci. Total Environ. 2014) we confirmed that PRO, MET and NAD are highly resistant to hydrolysis, bioavailable and mobile in the environment. Hence, their common presence in the environment can lead to unexpected effects towards different organisms. Beta-adrenergic receptors have been characterized in fish and other aquatic animals, so it can be expected that physiological processes regulated by these receptors in wild animals may be affected by the presence of BB. Thus, even though BB are safe for human and veterinary usage, ecosystems may be exposed to these substances. There are two ways by which these compounds can influence other organisms living in the environment: by showing target effects on non-target organisms or by showing specific non-target effects on
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H
H O O
O
H
H
O
N
H
H
H
O
O
N
O
O
H N
O
Propranolol (PRO)
Metoprolol (MET)
Nadolol (NAD)
M = 259.34 g mol-1
M = 267.36 g mol-1
M = 309.40 g mol-1
pKa = 9.53
pKa = 9.52
pKa = 9.69
log Kow = 3.48
log Kow = 1.95
log Kow = 0.85
non-ß1-selective MSA (+)
ß1-selective MSA (-)
non-ß1-selective MSA (-)
Fig. 1. Chemical structures and selected properties of the analyzed beta-blockers.
non-target organisms. This is supported with the data presented by Santos et al. (2010), who have reviewed the available literature data on the ecotoxicity of BB to aquatic organisms which revealed that in fact these drugs may have deleterious effects on different organisms such as fish (Japanese medaka, rainbow trout), invertebrates (Dapnia magna, Hyalella azteca, Daphnia lumholtzi, Ceriodaphnia dubia), and green algae (Pseudokirchneriella subcapitata). Moreover, it must be also highlighted that BB belong to Endocrine Disruptive Compounds (ECDs) as it has already been proven that they affect both free and total testosterone levels in male organisms (Rosen et al., 1988; el-Sayed et al., 1998). Although some data is available, comprehensive studies are missing and most of the available data deal with only PRO and atenolol (Santos et al., 2010). Much less data is available for MET and NAD. Hence, their ecotoxicological potential has not been fully characterized yet. Therefore, it was decided to choose the same BB as in our previous study (Fig. 1) as a subject of ecotoxicity evaluation. A flexible (eco)toxicological test battery in order to enrich our hitherto limited knowledge of the potentially deleterious effects of these pharmaceuticals on the environment using marine bacteria (Vibrio fischeri), soil/sediment bacteria (Arthrobacter globiformis), limnic unicellular green algae (Scenedesmus vacuolatus) and duckweed (Lemna minor) has been selected. Additional experiments were performed in order to assess whether sorption diminishes deleterious effects of these pharmaceuticals on the soil bacteria A. globiformis. All the ecotoxicological studies were supported by instrumental analysis to measure deviation between nominal and real test concentrations. As a result, the reliable ERA data for all the tested BB were obtained.
2. Material and methods General scheme of ecotoxicity evaluation is presented in Fig. 1S (Appendix B data).
2.1. Chemicals Standards of (±) propranolol hydrochloride (PRO) [CAS No. 318-989], (±) metoprolol (+) tartrate (MET) [CAS No. 56392-17-7] and nadolol (NAD) [CAS No. 42200-33-9] as well as trifluoroacetic acid 99% (TFA) were purchased from Sigma-Aldrich (Steinheim, Germany). Methanol (MeOH) and acetonitrile (ACN) (both HPLC grade) used for the mobile phases were purchased from POCH (Gliwice, Poland). Salts used for the culturing media were purchased from Sigma–Aldrich (Steinheim, Germany). Deionized water was produced by the HYDROLAB System (Gdansk, Poland).
2.2. Ecotoxicological assessment 2.2.1. Preparation of standard stock and determination of soluble fraction of BB in tested media All the standard stock solutions of PRO, MET and NAD were prepared on the day the test was performed by balancing the proper mass of each BB and reconstituting in the respective test medium. In every test, the effects of each BB were first tested in a range of concentrations from 0.1 mg L−1 to 100 mg L−1 or from 0.1 mg kg−1 to 100 mg kg−1 in the case of sediment contact assay with soil bacteria (range finding tests). If a toxic effect was observed, the tests were repeated in the specified concentration range. In order to determine the soluble fraction of the investigated compounds in biological media (the difference between real and nominal concentrations) an HPLC-UV analysis was carried out using the conditions described in Section 2.3. The stock solution concentrations of PRO, MET, and NAD in media solutions used in four of the ecotoxicity tests (100 mg L−1 for L. minor, 112 mg L−1 for S. vacuolatus, 200 mg L−1 for V. fischeri) were estimated by matching the chromatographic peak areas to a calibration curve obtained for the standard solution of these compounds prepared in deionized water as well in the mixture of H2O: ACN (90:10, v/v) in order to check if there are any differences. Such investigation has not been made for A. globiformis as stock solutions have been prepared in deionized water.
2.2.2. Luminescent inhibition assay with marine bacteria The toxicity testing based on the luminescent bacterium V. fischeri was done using the LCK 482 test kit (Dr. Lange GmbH, Düsseldorf, Germany). The 30-min standard bioluminescence inhibition assay was carried out according to a modified DIN 38412-L34 protocol (1991). To exclude pH-effects, all the substances were prepared as phosphatebuffered solutions (0.02 M, pH 7.0, including 2% NaCl). The tests were carried out at least twice for each substance. Within each test, at least 4 controls (2% NaCl solution, phosphate-buffered) were used. Briefly, the tests were performed at 15 °C using thermostats (LUMIStherm, Dr Lange GmbH, Düsseldorf, Germany). The luminescence was measured with a luminometer (LUMIStox 300, Dr Lange GmbH, Düsseldorf, Germany). The freeze-dried bacteria were rehydrated according to the test protocol; then, 500 μL aliquots of the bacteria solution were pre-incubated for 15 min at 15 °C. After the initial luminescence had been measured, 500 μL of the diluted samples were added to the bacteria. The bioluminescence was measured again after an incubation time of 30 min. The relative toxicity of the samples was expressed as a percentage inhibition compared to the controls.
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2.2.3. Contact assay with soil bacteria The sediment contact assay with the soil bacterium A. globiformis (DIN 38412-L48) was applied in the version modified for liquid samples (Neumann-Hensel and Melbye, 2006). Bacterial enzyme inhibition is measured after 4 h by applying a colorimetric assay. Through the activity of the bacterial dehydrogenases, blue Resazurin is turned into pink Resorufin, indicating healthy cells. A. globiformis is one of the most common aerobic chemoheterotrophic bacteria in soils. In order to determine whether sorption inhibits the toxic effect of beta-blockers towards soil bacteria, a parallel test without soil was performed. 2.2.4. Reproduction inhibition assay with limnic green algae A synchronized unicellular limnic green algae S. vacuolatus (strain 211-15, SAG (Culture Collection of Algae), Universität Göttingen, Germany) culture was used for this assay. The stock culture was grown under photoautotrophic conditions at 28 °C (±0.5 °C) in an inorganic, sterilized medium (pH 6.4), with saturating white light (22 to 33 klx, Lumilux Daylight L 36 W-11 and Lumilux Interna L 36 W-41, Osram, Berlin, Germany). The cells were aerated with 1.5% vol CO2 and synchronized using a 14–10 h light–darkness cycle. The stock culture was diluted every day to a cell density of 5 × 105 cells mL−1. This test is a modified version of the assay described by Altenburger et al. (1990) and its sensitivity is comparable to the standardized 72 h test (ISO 8692,1989). The toxicity tests commenced with autospores (young algal cells at the beginning of the growth cycle). The algae were exposed to the test substances for one growth cycle (24 h). The endpoint of this assay was the inhibition of algal reproduction, measured as the inhibition of population growth. Cell numbers were determined with the Coulter Counter Z2 (Beckmann, Nürnberg, Germany). The tests were performed in sterilized glass tubes, the algae were stirred throughout the 24 h test period, and the test conditions were the same as for the stock culture except for the CO2 source. Here, 150 μL of a NaHCO3 solution was added to each test tube. The methods for stock culturing and testing are described in detail by Faust et al. (2001, 2003). All the substances were tested at least twice: first, a range finding was undertaken (4 concentrations, two replicates); then, the results were verified with 8 concentrations per substance in two replicates. This data was pooled for analysis. Growth inhibition was calculated using the cell counts of the treated samples in relation to the untreated controls. At least 6 controls were used for each assay. 2.2.5. Growth inhibition assay with duckweed The growth inhibition assay with L. minor was performed according to a modified version of the test protocol described in detail in Drost et al. (2007). The plants were grown in open Erlenmeyer flasks in a sterilized Steinberg medium (pH 5.5 ± 0.2) in a climate chamber at a constant temperature of 25 ± 2 °C. To exclude pH effects on plant growth, pH was checked at the beginning and the end of the test. The chamber was illuminated continuously with a maximum of 6000 lx. The assays were performed on six-well cell culture plates (Greiner Bio-One GmbH, Frickenhausen, Germany). All the BB were tested at least twice, with a minimum of 6 controls (pure Steinberg medium) in each test. The test started with one plant consisting of 3 fronds (leaves of the duckweed), and the measured endpoint was the inhibition of the growth rate determined by the frond area (mm2), which was calculated for the treated plants in relation to the untreated controls. The frond area was detected using a Scanalyzer from Lemnatec GmbH (Würselen, Germany). 2.2.6. Effect data modeling Dose–response curve parameters and plots were obtained using the drift package (version 2.15.1) for the R language and environment for statistical computing (www.r-project.org) (R Development Core Team, 2012).
2.3. Instrumental analysis In order to determine the soluble fraction of the investigated compounds in biological media the filtrate samples were analyzed by isocratic reversed phase HPLC using a Phenomenex Onyx C18-110A monolithic column, 150 mm × 4.6 mm (Torrance, USA). All compounds were detected at a wavelength of 220 nm. The mobile phase was MeOH: H2O (with 0.025% of TFA) (A:B) at a 1 mL min−1 flow rate and composition of A and B: 30:70, 40:60 and 48:52 for NAD, MET and PRO respectively. The injection volume was 50 μL. All chromatographic analyses were carried out on two replicates. Validation parameters of the applied HPLC-UV method are presented in Appendix B (Table 1S). The analytical system, Perkin Elmer Series 200, consisted of a chromatographic interface (Link 600), a binary pump, a UV/VIS detector, a vacuum degasser and a Rheodyne injection valve. 3. Results and discussion 3.1. Evaluation of soluble fraction of three BB in biological media Deviations between the nominal and real concentrations of the test compounds in the media of our test systems were tested using instrumental analysis to avoid misinterpretation of bioavailability and hence misinterpretation of the ecotoxicological data that was obtained. The results showing the percentage of soluble fraction of the three investigated BB in biological media are presented in Table 1. Almost no differences have been observed for the calibration based on water and mixture of water/ACN used as solvents for calibration solutions. These analyses indicate generally a low deviation (from − 1.4% up to +5.4%) between nominal and measured concentrations of stock solutions containing biological media. Therefore, the nominal concentration for all the tests can be considered to be the bioavailable fraction within the toxicity tests. Nevertheless, according to Owen et al. (2007) it must also be pointed out that accurate citation of CAS numbers is essential for pharmaceuticals in order to compare nominal concentration data in terms of either the drug free base or the drug salt complex. This is especially important for BB, since these compounds are available in different forms on the market. Moreover, the test solution pH should also be reported since most BB have acidity constants above 9, hence BB are often not present in aqueous solution as free base or salt solution, and at around pH 7 they are almost 99% in the positively charged form and not associated with any specific anion (Owen et al., 2007). To clarify this situation, in our study we have presented the ecotoxicological data referring to drug complex concentrations, which for the highest used concentrations expressed as 100 mg L−1 or 100 mg kg− 1 of each salt complex is equal to 0.3 mM of each Active Pharmaceutical Ingredient (API). 3.2. Ecotoxicity evaluation In Table 2, the EC50 values obtained in all the ecotoxicity tests — and for the reference substance: atrazine are presented (for detailed dose– response graphs see Figs. 2S and 3S, Appendix B). In general, these results revealed that PRO was the most toxic pharmaceutical to selected organisms, of which green algae (S. vacuolatus) were the most sensitive. The level of toxicity of the three selected BB towards Table 1 Percentage of soluble fraction of investigated compounds in biological media. Compound
Percentage of soluble fraction [%] Calibration H2O/calibration H2O:ACN
PRO MET NAD
L. minor
S. vacuolatus
V. fischeri
101.9/105.3 102.8/102.3 99.5/98.6
100.8/104.2 102.9/102.4 100.2/102.4
102.6/105.4 100.5/100.1 102.4/100.5
J. Maszkowska et al. / Science of the Total Environment 493 (2014) 1122–1126 Table 2 The EC50 values for three beta-blockers obtained during experiments and toxicity of referenced compounds. Compound
PRO (20–30) MET (66–87) NAD Atrazine a b c
EC50 (Confidence Interval) [mg L−1], *[mg kg−1] V. fischeri
A. globiformis with*/without soil
S. vacuolatus
N100 N100 N100 N100 N100 69.4a
N100/218 (192–244)
24
N100
75
N100
N100 0.039c
L. minor
N100 0.188b
Palma et al., 2008. Teodorović et al., 2012. Faust et al., 2001.
the selected organisms decreased in the order of PRO N MET N NAD, that corresponds well to their hydrophobicity as well as to their mode of action (Fig. 1). Nadolol was the least toxic compound. A No Observed Effect Level Concentration (NOEC) can be given for NAD as 100 mg L−1 for all the aquatic organisms and as 100 mg kg−1 to A. globiformis (in soil contact test). Also no adverse effect of NAD was observed on sediment bacteria in the test performed in parallel without soil. This is the first study reporting the toxicity of NAD to these aquatic organisms. No data is available in the literature concerning this compound apart from its toxicity to aquatic invertebrates as well as fish (Huggett et al., 2002). Nevertheless, available ecotoxicological data supports our observations and proves that NAD exhibits rather low acute toxicity to non-target aquatic organisms (Huggett et al., 2002; Fraysse and Garric, 2005). Even though for MET (similar to NAD) no adverse effects on V. fischeri, L. minor and A. globiformis were observed up to the highest tested concentrations (NOEC = 100 mg L−1 or 100 mg kg−1, respectively) this compound had a negative influence on the growth of green algae (S. vacuolatus). The EC50 value of MET on the green algae selected in our study was 76 mg L−1. Although this is the first time the toxicity of MET to these species of algae as well as to A. globiformis has been reported, our results are in agreement with those previously reported for marine bacteria, duckweed and other types of algae. While Escher et al. (2006) reported the toxicity of MET to V. fischeri with the EC50 values of 144 mg L−1, Cleuvers (2003) did not observe any effect of MET on L. minor after 7 days of exposure up to a concentration of 320 mg L− 1. This author has also evaluated the toxicity of MET to green algae Desmodesmus subcapitatus (formerly known as Scenedesmus subspicatus) with the EC50 value one order of magnitude lower than that obtained in our study (EC50 = 7.9 mg L−1 and 7.3 mg L−1 in 48 h and 3 days test, respectively) (Cleuvers, 2003, 2005). On the other hand, Escher et al. (2006) had an EC50 value for MET to these species of algae (D. subcapitatus) in a 24 h test in the same range as in our study (EC50 = 40 mg L−1). In contrast to MET and NAD, PRO had the strongest influence on the organisms we selected for ecotoxicological study, of which green algae were affected the most with an EC50 value of 24 mg L−1. No toxicity of PRO to L. minor up to 100 mg L−1 was observed. Only slight adverse effects were observed for marine bacteria (with LOEC — ang. Lowest Observed Effect Concentration, in the range of 10–100 mg L−1). These results are in agreement with Cleuvers (2003, 2005) who also observed the strongest toxicity of PRO to algae rather than to duckweed with the EC50 values of 0.7 to 5.8 mg L−1 for D. subcapitatus and from 113 to 114 mg L− 1 for L. minor. Although the same trend was observed as in our study, due to probable differences in the algae species, the EC50 values differ from each other in about one to two orders of magnitude depending on the compound. Liu et al. (2009) also supported the observations of Cleuvers (2003, 2005), reporting EC50 values of PRO to
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green algae (Pseudokirchneriella subcapitata) in the range of 0.77 to 7.40 mg L−1 depending on the test time and end point. Furthermore, the previously reported EC50s of PRO to V. fischeri in a range from 61 mg L−1 (Ferrari et al., 2004) up to 184 mg L−1 (Calleja et al., 1994) are also in agreement with our results. Nevertheless, very interesting results were obtained for A. globiformis. PRO exhibited no toxic effect at concentrations of up to 100 mg kg−1 in sediment contact assay (Fig. 3S, Appendix B). On the other hand, results obtained during a parallel test without soil, exhibited a toxic effect for PRO towards these bacteria. Thus, it can be concluded that sorption explicitly inhibits (by decreasing the bioavailable fraction of the pharmaceutical) the hazardous effects of beta-blockers. Nevertheless, the observed value of EC50 (218 mg L−1) is relatively high and therefore it can be assumed that these pharmaceuticals do not pose a risk for A. globiformis which ubiquitously occurs in soil. Strong differences in the toxicity of different BB to aquatic organisms have been previously observed by other scientists (Huggett et al., 2002; Cleuvers, 2003, 2005; Fraysse and Garric, 2005; Dzialowski et al., 2006). PRO was the most potent of the compounds tested and has the highest logKow; MET was the second most potent compound and has the second highest logKow. NAD — having the lowest logKow did not elicit any toxicological responses in this study (see Fig. 1). For this reason, taking bioconcentration factors into account, with resulting internal effect concentrations differing only slightly, Cleuvers (2005) suggested that the differences in the EC50s values depended mainly on the diverse logKow levels, causing narcosis via disruption of membrane integrity. Such a partitioning of pollutants into the membrane causes the membrane to expand or swell, increasing fluidity, lowering the phase transition temperature, and the ion permeability of the membrane. On the other hand, he emphasizes that baseline toxicants may interact with specific receptors in the hydrophobic region of membrane proteins. Additionally, disturbances of lipid–protein or lipid–lipid interactions or lateral pressure changes were thought to be responsible for the narcosis. Therefore, it is likely that in reality narcosis covers maybe dozens of unknown, but specific, toxic modes of action. Although neither daphnids nor plants have been reported to possess betareceptors, this is not an unimportant point. For as we know from humans, adverse drug reactions can be caused by previously unrecognized drug-receptor interactions, previously unidentified receptors and by a broad diversity in drug-metabolizing phenotypes. These variables are even more poorly characterized in aquatic biota. Hugget et al. (2002) have also suggested that other factors may be involved, as earlier studies with invertebrates showed that e.g. paracetamol (logKow = 0.49) was more toxic to D. magna than methotrexate, salicilic acid and clofibric acid (logKow = 2.28, 2.24 and 2.84, respectively). For mammals, including humans, BB are a competitive antagonist of the effects of catecholamines (adrenaline and noradrenaline) at ßadrenergic-receptor sites (ß1 and/or ß2). These compounds have different affinities for ß1 and ß2 receptors (see Fig. 1). Some of them, such as PRO, have also a membrane-stabilizing activity (MSA). PRO reduces membrane permeability for various ions (Na+, K+ and Ca2+) and has a local anesthetic activity. Thus, the effect on organism survival may in part be due to adverse effects on membrane stabilization (Fraysse and Garric, 2005; Dzialowski et al., 2006; Claessens et al., 2013). Moreover, as was previously highlighted (see Section 3.1), for accurate comparisons of the ecotoxicological data, citation of CAS numbers is essential not only to clarify whether the nominal concentration data refers to the drug free base or the drug salt complex but also whether it refers to enantiomer form, which unfortunately is not a common practice. For example, like many therapeutics and other aquatic contaminants, PRO is distributed as a recemix mixture (R,S)-propranolol hydrochloride, (S)-(−)-propranolol hydrochloride as well as (R)-(+)propranolol hydrochloride. While for example the (S)-enantiomer is the most active form in mammals (up to a 100-fold difference), only the latter is responsible for propranolol's membrane-stabilizing effect. When such compounds are present in an asymmetrical biological
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environment, such as the highly specific binding sites of many cellular receptors, enantiospecific differences in biological activity can occur. Enantiomers may also differ significantly in their rates of uptake and excretion, affinity for plasma proteins, structure of metabolites, potency, toxicity and bioavailability (Stanley et al., 2006), hence this aspect should be also considered in the future research. Nevertheless, the mechanism by which BB affect these organisms remains elusive and must be examined in greater depth. 4. Conclusion In this study the ecotoxicological evaluation of three selected BB: PRO, MET and NAD has been performed for better understanding of the possible effects that these compounds have towards non-target organisms. According to the chemical substance classification of the European Union Directive 93/67EEC, PRO and MET can be considered to be harmful (10 b EC50 b 100 mg L−1) to aquatic organisms (based on toxicological data from the green alga test (S. vacuolatus)). In addition, new ecotoxicological data for three BB to V. fischerii, S. vacuolatus and L. minor as well as to soil bacteria (A. globiformis) is presented. Taking all the results into account, as well as the environmental concentrations of these compounds, it can be concluded that the risks posed by these compounds are of minor importance. However, it must be stressed that in the future, chronic biotests on other aquatic organisms should be carried out, bearing in mind the contribution of each individual compound to the overall toxic potential of the mixture of substances present in the aquatic environment. Ultimately, it's possible that the classification of BB could change from “not harmful” to “very toxic” depending on whether the compounds are applied in single exposures or are part of a mixture of substances. This comprehensive study supports the Environmental Risk Assessment of these pharmaceuticals. Acknowledgments The authors would like to acknowledge the financial support of National Science Centre (Poland) under decision DEC-2011/03/B/NZ8/03010 and DEC-2011/03/B/NZ8/03009 and the German Academic Exchange Service (DAAD). The publication is financed from European Social Fund as a part of the project “Educators for the elite-integrated training program for PhD students, post-docs and professors as academic teachers at University of Gdansk” within the framework of Human Capital Operational Programme, Action IV. The authors are grateful to Stephanie Steudte and Katarzyna Wychodnik and Juliane Filser for their support. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.06.039. References Altenburger R, Bodeker W, Faust M, Grimme LH. Evaluation of the isobologram method for the assessment of mixtures of chemicals: combination effect studies with pesticides in algal biotests. Ecotoxicol Environ Saf 1990;20:98–114. Amendola L, Molaioni F, Botrè F. Detection of beta-blockers in human urine by GC–MS– MS–EI: perspectives for the antidoping control. J Pharm Biomed Anal 2000;23(1): 211–21. Barbieri M, Licha T, Nödler K, Carrera J, Ayora C, Sanchez-Vila X. Fate of β-blockers in aquifer material under nitrate reducing conditions: batch experiments. Chemosphere 2012;89(11):1272–7. British Pharmacopoeia Commission. British Pharmacopoeia. 34th ed. London: Pharmaceutical Press; 2005. Brooks B, Huggett D. Human pharmaceuticals in the environment. New York Heidelberg Dordrecht London: Springer; 2012. Caban M, Mioduszewska K, Stepnowski P, Kwiatkowski M, Kumirska J. Dimethyl(3,3,3trifluoropropyl)silyldiethylamine—a new silylating agent for the derivatization of βblockers and β-agonists in environmental samples. Anal Chim Acta 2013;782:78–88. Calleja MC, Persoone G, Geladi P. Comparative acute toxicity of the first 50 multicentre evaluation of in vitro cytotoxicity chemicals to aquatic non-vertebrates. Arch Environ Contam Toxicol 1994;26(1):69–78.
Claessens M, Vanhaecke L, Wille K, Janssen CR. Emerging contaminants in Belgian marine waters: single toxicant and mixture risks of pharmaceuticals. Mar Pollut Bull 2013;71(1–2):41–50. Cleuvers M. Aquatic ecotoxicity of pharmaceuticals including the assessment of combination effects. Toxicol Lett 2003;142(3):185–94. Cleuvers M. Initial risk assessment for three beta-blockers found in the aquatic environment. Chemosphere 2005;59(2):199–205. Drost W, Matzke M, Backhaus T. Heavy metal toxicity to Lemna minor: studies on the time dependence of growth inhibition and the recovery after exposure. Chemosphere 2007;67:36–43. Dzialowski EM, Turner PK, Brooks BW. Physiological and reproductive effects of beta adrenergic receptor antagonists in Daphnia magna. Arch Environ Contam Toxicol 2006;50(4):503–10. el-Sayed MG, el-Sayed MT, Elazab Abd el S, Hafeiz MH, el-Komy AA, Hassan E. Effects of some beta-adrenergic blockers on male fertility parameters in rates. Dtsch Tierarztl Wochenschr 1998;105(1):10–2. Escher BI, Bramaz N, Richter M, Lienert J. Comparative ecotoxicological hazard assessment of beta-blockers and their human metabolites using a mode-of-action-based test battery and a QSAR approach. Environ Sci Technol 2006;40(23):7402–8. Faust M, Altenburger R, Backhaus T, Blanck H, Boedeker W, Gramatica P, et al. Predicting the joint algal toxicity of multi-component s-triazine mixtures at low-effect concentrations of individual toxicants. Aquat Toxicol 2001;56:13–32. Faust M, Altenburger R, Backhaus T, Blanck H, Boedeker W, Gramatica P, et al. Joint algal toxicity of 16 dissimilarly acting chemicals is predictable by the concept of independent action. Aquat Toxicol 2003;63:43–64. Ferrari B, Mons R, Vollat B, Fraysse B, Paxéus N, Lo Giudice R, et al. Environmental risk assessment of six human pharmaceuticals: are the current environmental risk assessment procedures sufficient for the protection of the aquatic environment? Environ Toxicol Chem 2004;23(5):1344–54. Fraysse B, Garric J. Prediction and experimental validation of acute toxicity of ß-blockers in Ceriodaphnia dubia. Environ Toxicol Chem 2005;24:2470–6. Huggett DB, Brooks BW, Peterson B, Foran CM, Schlenk D. Toxicity of select beta adrenergic receptor-blocking pharmaceuticals (B-blockers) on aquatic organisms. Arch Environ Contam Toxicol 2002;43(2):229–35. Huggett DB, Khan I, Foran CM, Schlenk D. Determination of beta-adrenergic receptor blocking pharmaceuticals in United States wastewater effluent. Environ Pollut 2003;121(2):199–205. ISO Guideline 8692. Water quality — fresh water algal growth inhibition test with Scenedesmus subspicatus and Selenastrum capricornutum, G; 1989. Kibbey TCG, Paruchuri R, Sabatini DA, Chen L. Adsorption of beta blockers to environmental surfaces. Environ Sci Technol 2007;41(15):5349–56. Kümmerer K, editor. Pharmaceuticals in the environment. 2nd ed. Berlin Heidelberg, New York: Springer; 2004. Küster A, Alder AC, Escher BI, Duis K, Fenner K, Garric J, et al. Environmental risk assessment of human pharmaceuticals in the European Union: a case study with the β-blocker atenolol. Integr Environ Assess Manag 2009;6:514–23. Lin Yu-Chen A, Tsai YT. Occurrence of pharmaceuticals in Taiwan's surface waters: impact of waste streams from hospitals and pharmaceutical production facilities. Sci Total Environ 2009;407:3793–802. Liu QT, Williams T, Cumming RI, Holm G, Hetheridge MJ. Comparative aquatic toxicity of propranolol and its photodegradated mixtures: algae and rotifer screening. Environ Toxicol Chem 2009;28:2622–31. Maurer M, Escher BI, Richle P, Schaffner C, Alder AC. Elimination of beta-blockers in sewage treatment plants. Water Res 2007;41(7):1614–22. Neumann-Hensel H, Melbye K. Optimisation of the solid-contact test with Arthrobacter globiformis. J. Soils Sediments 2006;6(4):201–7. Owen SF, Giltrow E, Huggett DB, Hutchinson TH, Saye J, Winter MJ, et al. Comparative physiology, pharmacology and toxicology of beta-blockers: mammals versus fish. Aquat Toxicol 2007;82(3):145–62. Palma P, Palma VL, Fernandes RM, Soares AMVM, Barbosa IR. Acute toxicity of atrazine, endosulfan sulphate and chlorpyrifos to Vibrio fischeri, Thamnocephalus platyurus and Daphnia magna, Relative to their concentrations in surface waters from the Alentejo Region of Portugal. Bull Environ Contam Toxicol 2008;81:485–9. R Development Core Team. Writing R extensions. Vienna, Austria: R Foundation for Statistical Computing; 2012. URL http://CRAN.R-Project.org/doc/manuals/R-exts.html, Accessed March 15 2014. Rosen RC, Kostis JB, Jekelis AW. Beta-blocker effects on sexual function in normal males. Arch Sex Behav 1988;17(3):241–55. Santos LHMLM, Araújo AN, Fachini A, Pena A, Delerue-Matos C, Montenegro MCBSM. Ecotoxicological aspects related to the presence of pharmaceuticals in the aquatic environment. J Hazard Mater 2010;175(1-3):45–95. Scheurer M, Ramil M, Metcalfe CD, Groh S, Ternes TA. The challenge of analyzing betablocker drugs in sludge and wastewater. Anal Bioanal Chem 2010;396(2):845–56. Stanley JK, Ramirez AJ, Mottaleb M, Chambliss CK, Brooks BW. Enantiospecific toxicity of the beta-blocker propranolol to Daphnia magna and Pimephales promelas. Environ Toxicol Chem 2006;25(7):1780–6. Teodorović I, Knežević V, Tunić T, Cučak M, Lečić JN, Leovac A, et al. Myriophyllum aquaticum versus Lemna minor: sensitivity and recovery potential after exposure to atrazine. Environ Toxicol Chem 2012;31(2):417–26. The United States Pharmacopeial Convention. USP28-N-NF23; 2005. p. 1281. Wang L, Xu H, Cooper WJ, Song W. Photochemical fate of beta-blockers in NOM enriched waters. Sci Total Environ 2012;426:289–95. Wilde ML, Mahmoud WMM, Kümmerer K, Martins AF. Oxidation–coagulation of β-blockers by K2FeVIO4 in hospital wastewater: assessment of degradation products and biodegradability. Sci Total Environ 2013;452–453:137–47.
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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv
Beta-blockers in the environment: Part II. Ecotoxicity study Joanna Maszkowska a, Stefan Stolte a,b, Jolanta Kumirska a,⁎, Paulina Łukaszewicz a, Katarzyna Mioduszewska a, Alan Puckowski a, Magda Caban a, Marta Wagil a, Piotr Stepnowski a, Anna Białk-Bielińska a,b a b
Department of Environmental Analysis, Institute for Environmental and Human Health Protection, Faculty of Chemistry, University of Gdańsk, ul. Wita Stwosza 63, 80-308 Gdańsk, Poland UFT — Center for Environmental Research and Sustainable Technology, University of Bremen, Leobener Straße, D-28359 Bremen, Germany
H I G H L I G H T S • Propranolol and metoprolol can be considered to be harmful to aquatic organisms. • Sorption explicitly inhibits the hazardous effects of beta-blockers. • The risks posed by these compounds for the environment are of minor importance.
a r t i c l e
i n f o
Available online 27 June 2014 Editor: Adrian Covaci Keywords: Green algae Duckweed Soil and marine bacteria
a b s t r a c t The increasing consumption of beta-blockers (BB) has caused their presence in the environment to become more noticeable. Even though BB are safe for human and veterinary usage, ecosystems may be exposed to these substances. In this study, three selected BB: propranolol, metoprolol and nadolol were subjected to ecotoxicity study. Ecotoxicity evaluation was based on a flexible ecotoxicological test battery including organisms, representing different trophic levels and complexity: marine bacteria (Vibrio fischeri), soil/sediment bacteria (Arthrobacter globiformis), green algae (Scenedesmus vacuolatus) and duckweed (Lemna minor). All the ecotoxicological studies were supported by instrumental analysis to measure deviation between nominal and real test concentrations. Based on toxicological data from the green algae test (S. vacuolatus) propranolol and metoprolol can be considered to be harmful to aquatic organisms. However, sorption explicitly inhibits the hazardous effects of BB, therefore the risks posed by these compounds for the environment are of minor importance. © 2014 Elsevier B.V. All rights reserved.
1. Introduction Beta-blockers (BB) are extensively used for the treatment of various cardiovascular disorders such as high blood pressure, ischaemic heart disease and heart rhythm disturbances (British Pharmacopoeia, 2005; The USA Pharmacopeial Convention, 2005). It is estimated that 100– 250 tons of these compounds are consumed in Germany per year (Cleuvers, 2005); several of them are in the top 200 prescribed medications in the United States, too (Huggett et al., 2003). They are also used in veterinary medicine and illegally as doping in many sports (Amendola et al., 2000; Huggett et al., 2002; Barbieri et al., 2012). BB work by blocking the effect of adrenaline on our body's receptors, thereby slowing the nerve impulses to the heart and reducing its workload (British Pharmacopoeia, 2005; The USA Pharmacopeial Convention, 2005). Their widespread use and often incomplete metabolism (Küster et al., 2009) mean that BB are commonly detected in sewage effluents and surface waters (e.g. Huggett et al., 2003; Kümmerer, 2004; Scheurer et al., 2010; Brooks and Huggett, 2012). For example, the ⁎ Corresponding author. Tel.: +48 58 5235212; fax: +48 58 5235454. E-mail address: [email protected] (J. Kumirska).
http://dx.doi.org/10.1016/j.scitotenv.2014.06.039 0048-9697/© 2014 Elsevier B.V. All rights reserved.
concentrations of metoprolol (MET), propranolol (PRO), and nadolol (NAD) in sewage treatment plant (STP) effluents vary between 25 and 2800 ng L−1 (Caban et al., 2013); they are found in rivers and streams, too (Lin and Tsai, 2009). MET, PRO and NAD (Fig. 1) are the most frequently analyzed beta-blockers in biological and environmental samples. Among these three compounds, PRO is the most hydrophobic and shows bioaccumulation potential (Kibbey et al., 2007; Maurer et al., 2007; Wang et al., 2012; Wilde et al., 2013). In our previous study (Maszkowska et al. Beta-blockers in the environment: Part I. Mobility and hydrolysis study. Sci. Total Environ. 2014) we confirmed that PRO, MET and NAD are highly resistant to hydrolysis, bioavailable and mobile in the environment. Hence, their common presence in the environment can lead to unexpected effects towards different organisms. Beta-adrenergic receptors have been characterized in fish and other aquatic animals, so it can be expected that physiological processes regulated by these receptors in wild animals may be affected by the presence of BB. Thus, even though BB are safe for human and veterinary usage, ecosystems may be exposed to these substances. There are two ways by which these compounds can influence other organisms living in the environment: by showing target effects on non-target organisms or by showing specific non-target effects on
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H
H O O
O
H
H
O
N
H
H
H
O
O
N
O
O
H N
O
Propranolol (PRO)
Metoprolol (MET)
Nadolol (NAD)
M = 259.34 g mol-1
M = 267.36 g mol-1
M = 309.40 g mol-1
pKa = 9.53
pKa = 9.52
pKa = 9.69
log Kow = 3.48
log Kow = 1.95
log Kow = 0.85
non-ß1-selective MSA (+)
ß1-selective MSA (-)
non-ß1-selective MSA (-)
Fig. 1. Chemical structures and selected properties of the analyzed beta-blockers.
non-target organisms. This is supported with the data presented by Santos et al. (2010), who have reviewed the available literature data on the ecotoxicity of BB to aquatic organisms which revealed that in fact these drugs may have deleterious effects on different organisms such as fish (Japanese medaka, rainbow trout), invertebrates (Dapnia magna, Hyalella azteca, Daphnia lumholtzi, Ceriodaphnia dubia), and green algae (Pseudokirchneriella subcapitata). Moreover, it must be also highlighted that BB belong to Endocrine Disruptive Compounds (ECDs) as it has already been proven that they affect both free and total testosterone levels in male organisms (Rosen et al., 1988; el-Sayed et al., 1998). Although some data is available, comprehensive studies are missing and most of the available data deal with only PRO and atenolol (Santos et al., 2010). Much less data is available for MET and NAD. Hence, their ecotoxicological potential has not been fully characterized yet. Therefore, it was decided to choose the same BB as in our previous study (Fig. 1) as a subject of ecotoxicity evaluation. A flexible (eco)toxicological test battery in order to enrich our hitherto limited knowledge of the potentially deleterious effects of these pharmaceuticals on the environment using marine bacteria (Vibrio fischeri), soil/sediment bacteria (Arthrobacter globiformis), limnic unicellular green algae (Scenedesmus vacuolatus) and duckweed (Lemna minor) has been selected. Additional experiments were performed in order to assess whether sorption diminishes deleterious effects of these pharmaceuticals on the soil bacteria A. globiformis. All the ecotoxicological studies were supported by instrumental analysis to measure deviation between nominal and real test concentrations. As a result, the reliable ERA data for all the tested BB were obtained.
2. Material and methods General scheme of ecotoxicity evaluation is presented in Fig. 1S (Appendix B data).
2.1. Chemicals Standards of (±) propranolol hydrochloride (PRO) [CAS No. 318-989], (±) metoprolol (+) tartrate (MET) [CAS No. 56392-17-7] and nadolol (NAD) [CAS No. 42200-33-9] as well as trifluoroacetic acid 99% (TFA) were purchased from Sigma-Aldrich (Steinheim, Germany). Methanol (MeOH) and acetonitrile (ACN) (both HPLC grade) used for the mobile phases were purchased from POCH (Gliwice, Poland). Salts used for the culturing media were purchased from Sigma–Aldrich (Steinheim, Germany). Deionized water was produced by the HYDROLAB System (Gdansk, Poland).
2.2. Ecotoxicological assessment 2.2.1. Preparation of standard stock and determination of soluble fraction of BB in tested media All the standard stock solutions of PRO, MET and NAD were prepared on the day the test was performed by balancing the proper mass of each BB and reconstituting in the respective test medium. In every test, the effects of each BB were first tested in a range of concentrations from 0.1 mg L−1 to 100 mg L−1 or from 0.1 mg kg−1 to 100 mg kg−1 in the case of sediment contact assay with soil bacteria (range finding tests). If a toxic effect was observed, the tests were repeated in the specified concentration range. In order to determine the soluble fraction of the investigated compounds in biological media (the difference between real and nominal concentrations) an HPLC-UV analysis was carried out using the conditions described in Section 2.3. The stock solution concentrations of PRO, MET, and NAD in media solutions used in four of the ecotoxicity tests (100 mg L−1 for L. minor, 112 mg L−1 for S. vacuolatus, 200 mg L−1 for V. fischeri) were estimated by matching the chromatographic peak areas to a calibration curve obtained for the standard solution of these compounds prepared in deionized water as well in the mixture of H2O: ACN (90:10, v/v) in order to check if there are any differences. Such investigation has not been made for A. globiformis as stock solutions have been prepared in deionized water.
2.2.2. Luminescent inhibition assay with marine bacteria The toxicity testing based on the luminescent bacterium V. fischeri was done using the LCK 482 test kit (Dr. Lange GmbH, Düsseldorf, Germany). The 30-min standard bioluminescence inhibition assay was carried out according to a modified DIN 38412-L34 protocol (1991). To exclude pH-effects, all the substances were prepared as phosphatebuffered solutions (0.02 M, pH 7.0, including 2% NaCl). The tests were carried out at least twice for each substance. Within each test, at least 4 controls (2% NaCl solution, phosphate-buffered) were used. Briefly, the tests were performed at 15 °C using thermostats (LUMIStherm, Dr Lange GmbH, Düsseldorf, Germany). The luminescence was measured with a luminometer (LUMIStox 300, Dr Lange GmbH, Düsseldorf, Germany). The freeze-dried bacteria were rehydrated according to the test protocol; then, 500 μL aliquots of the bacteria solution were pre-incubated for 15 min at 15 °C. After the initial luminescence had been measured, 500 μL of the diluted samples were added to the bacteria. The bioluminescence was measured again after an incubation time of 30 min. The relative toxicity of the samples was expressed as a percentage inhibition compared to the controls.
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2.2.3. Contact assay with soil bacteria The sediment contact assay with the soil bacterium A. globiformis (DIN 38412-L48) was applied in the version modified for liquid samples (Neumann-Hensel and Melbye, 2006). Bacterial enzyme inhibition is measured after 4 h by applying a colorimetric assay. Through the activity of the bacterial dehydrogenases, blue Resazurin is turned into pink Resorufin, indicating healthy cells. A. globiformis is one of the most common aerobic chemoheterotrophic bacteria in soils. In order to determine whether sorption inhibits the toxic effect of beta-blockers towards soil bacteria, a parallel test without soil was performed. 2.2.4. Reproduction inhibition assay with limnic green algae A synchronized unicellular limnic green algae S. vacuolatus (strain 211-15, SAG (Culture Collection of Algae), Universität Göttingen, Germany) culture was used for this assay. The stock culture was grown under photoautotrophic conditions at 28 °C (±0.5 °C) in an inorganic, sterilized medium (pH 6.4), with saturating white light (22 to 33 klx, Lumilux Daylight L 36 W-11 and Lumilux Interna L 36 W-41, Osram, Berlin, Germany). The cells were aerated with 1.5% vol CO2 and synchronized using a 14–10 h light–darkness cycle. The stock culture was diluted every day to a cell density of 5 × 105 cells mL−1. This test is a modified version of the assay described by Altenburger et al. (1990) and its sensitivity is comparable to the standardized 72 h test (ISO 8692,1989). The toxicity tests commenced with autospores (young algal cells at the beginning of the growth cycle). The algae were exposed to the test substances for one growth cycle (24 h). The endpoint of this assay was the inhibition of algal reproduction, measured as the inhibition of population growth. Cell numbers were determined with the Coulter Counter Z2 (Beckmann, Nürnberg, Germany). The tests were performed in sterilized glass tubes, the algae were stirred throughout the 24 h test period, and the test conditions were the same as for the stock culture except for the CO2 source. Here, 150 μL of a NaHCO3 solution was added to each test tube. The methods for stock culturing and testing are described in detail by Faust et al. (2001, 2003). All the substances were tested at least twice: first, a range finding was undertaken (4 concentrations, two replicates); then, the results were verified with 8 concentrations per substance in two replicates. This data was pooled for analysis. Growth inhibition was calculated using the cell counts of the treated samples in relation to the untreated controls. At least 6 controls were used for each assay. 2.2.5. Growth inhibition assay with duckweed The growth inhibition assay with L. minor was performed according to a modified version of the test protocol described in detail in Drost et al. (2007). The plants were grown in open Erlenmeyer flasks in a sterilized Steinberg medium (pH 5.5 ± 0.2) in a climate chamber at a constant temperature of 25 ± 2 °C. To exclude pH effects on plant growth, pH was checked at the beginning and the end of the test. The chamber was illuminated continuously with a maximum of 6000 lx. The assays were performed on six-well cell culture plates (Greiner Bio-One GmbH, Frickenhausen, Germany). All the BB were tested at least twice, with a minimum of 6 controls (pure Steinberg medium) in each test. The test started with one plant consisting of 3 fronds (leaves of the duckweed), and the measured endpoint was the inhibition of the growth rate determined by the frond area (mm2), which was calculated for the treated plants in relation to the untreated controls. The frond area was detected using a Scanalyzer from Lemnatec GmbH (Würselen, Germany). 2.2.6. Effect data modeling Dose–response curve parameters and plots were obtained using the drift package (version 2.15.1) for the R language and environment for statistical computing (www.r-project.org) (R Development Core Team, 2012).
2.3. Instrumental analysis In order to determine the soluble fraction of the investigated compounds in biological media the filtrate samples were analyzed by isocratic reversed phase HPLC using a Phenomenex Onyx C18-110A monolithic column, 150 mm × 4.6 mm (Torrance, USA). All compounds were detected at a wavelength of 220 nm. The mobile phase was MeOH: H2O (with 0.025% of TFA) (A:B) at a 1 mL min−1 flow rate and composition of A and B: 30:70, 40:60 and 48:52 for NAD, MET and PRO respectively. The injection volume was 50 μL. All chromatographic analyses were carried out on two replicates. Validation parameters of the applied HPLC-UV method are presented in Appendix B (Table 1S). The analytical system, Perkin Elmer Series 200, consisted of a chromatographic interface (Link 600), a binary pump, a UV/VIS detector, a vacuum degasser and a Rheodyne injection valve. 3. Results and discussion 3.1. Evaluation of soluble fraction of three BB in biological media Deviations between the nominal and real concentrations of the test compounds in the media of our test systems were tested using instrumental analysis to avoid misinterpretation of bioavailability and hence misinterpretation of the ecotoxicological data that was obtained. The results showing the percentage of soluble fraction of the three investigated BB in biological media are presented in Table 1. Almost no differences have been observed for the calibration based on water and mixture of water/ACN used as solvents for calibration solutions. These analyses indicate generally a low deviation (from − 1.4% up to +5.4%) between nominal and measured concentrations of stock solutions containing biological media. Therefore, the nominal concentration for all the tests can be considered to be the bioavailable fraction within the toxicity tests. Nevertheless, according to Owen et al. (2007) it must also be pointed out that accurate citation of CAS numbers is essential for pharmaceuticals in order to compare nominal concentration data in terms of either the drug free base or the drug salt complex. This is especially important for BB, since these compounds are available in different forms on the market. Moreover, the test solution pH should also be reported since most BB have acidity constants above 9, hence BB are often not present in aqueous solution as free base or salt solution, and at around pH 7 they are almost 99% in the positively charged form and not associated with any specific anion (Owen et al., 2007). To clarify this situation, in our study we have presented the ecotoxicological data referring to drug complex concentrations, which for the highest used concentrations expressed as 100 mg L−1 or 100 mg kg− 1 of each salt complex is equal to 0.3 mM of each Active Pharmaceutical Ingredient (API). 3.2. Ecotoxicity evaluation In Table 2, the EC50 values obtained in all the ecotoxicity tests — and for the reference substance: atrazine are presented (for detailed dose– response graphs see Figs. 2S and 3S, Appendix B). In general, these results revealed that PRO was the most toxic pharmaceutical to selected organisms, of which green algae (S. vacuolatus) were the most sensitive. The level of toxicity of the three selected BB towards Table 1 Percentage of soluble fraction of investigated compounds in biological media. Compound
Percentage of soluble fraction [%] Calibration H2O/calibration H2O:ACN
PRO MET NAD
L. minor
S. vacuolatus
V. fischeri
101.9/105.3 102.8/102.3 99.5/98.6
100.8/104.2 102.9/102.4 100.2/102.4
102.6/105.4 100.5/100.1 102.4/100.5
J. Maszkowska et al. / Science of the Total Environment 493 (2014) 1122–1126 Table 2 The EC50 values for three beta-blockers obtained during experiments and toxicity of referenced compounds. Compound
PRO (20–30) MET (66–87) NAD Atrazine a b c
EC50 (Confidence Interval) [mg L−1], *[mg kg−1] V. fischeri
A. globiformis with*/without soil
S. vacuolatus
N100 N100 N100 N100 N100 69.4a
N100/218 (192–244)
24
N100
75
N100
N100 0.039c
L. minor
N100 0.188b
Palma et al., 2008. Teodorović et al., 2012. Faust et al., 2001.
the selected organisms decreased in the order of PRO N MET N NAD, that corresponds well to their hydrophobicity as well as to their mode of action (Fig. 1). Nadolol was the least toxic compound. A No Observed Effect Level Concentration (NOEC) can be given for NAD as 100 mg L−1 for all the aquatic organisms and as 100 mg kg−1 to A. globiformis (in soil contact test). Also no adverse effect of NAD was observed on sediment bacteria in the test performed in parallel without soil. This is the first study reporting the toxicity of NAD to these aquatic organisms. No data is available in the literature concerning this compound apart from its toxicity to aquatic invertebrates as well as fish (Huggett et al., 2002). Nevertheless, available ecotoxicological data supports our observations and proves that NAD exhibits rather low acute toxicity to non-target aquatic organisms (Huggett et al., 2002; Fraysse and Garric, 2005). Even though for MET (similar to NAD) no adverse effects on V. fischeri, L. minor and A. globiformis were observed up to the highest tested concentrations (NOEC = 100 mg L−1 or 100 mg kg−1, respectively) this compound had a negative influence on the growth of green algae (S. vacuolatus). The EC50 value of MET on the green algae selected in our study was 76 mg L−1. Although this is the first time the toxicity of MET to these species of algae as well as to A. globiformis has been reported, our results are in agreement with those previously reported for marine bacteria, duckweed and other types of algae. While Escher et al. (2006) reported the toxicity of MET to V. fischeri with the EC50 values of 144 mg L−1, Cleuvers (2003) did not observe any effect of MET on L. minor after 7 days of exposure up to a concentration of 320 mg L− 1. This author has also evaluated the toxicity of MET to green algae Desmodesmus subcapitatus (formerly known as Scenedesmus subspicatus) with the EC50 value one order of magnitude lower than that obtained in our study (EC50 = 7.9 mg L−1 and 7.3 mg L−1 in 48 h and 3 days test, respectively) (Cleuvers, 2003, 2005). On the other hand, Escher et al. (2006) had an EC50 value for MET to these species of algae (D. subcapitatus) in a 24 h test in the same range as in our study (EC50 = 40 mg L−1). In contrast to MET and NAD, PRO had the strongest influence on the organisms we selected for ecotoxicological study, of which green algae were affected the most with an EC50 value of 24 mg L−1. No toxicity of PRO to L. minor up to 100 mg L−1 was observed. Only slight adverse effects were observed for marine bacteria (with LOEC — ang. Lowest Observed Effect Concentration, in the range of 10–100 mg L−1). These results are in agreement with Cleuvers (2003, 2005) who also observed the strongest toxicity of PRO to algae rather than to duckweed with the EC50 values of 0.7 to 5.8 mg L−1 for D. subcapitatus and from 113 to 114 mg L− 1 for L. minor. Although the same trend was observed as in our study, due to probable differences in the algae species, the EC50 values differ from each other in about one to two orders of magnitude depending on the compound. Liu et al. (2009) also supported the observations of Cleuvers (2003, 2005), reporting EC50 values of PRO to
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green algae (Pseudokirchneriella subcapitata) in the range of 0.77 to 7.40 mg L−1 depending on the test time and end point. Furthermore, the previously reported EC50s of PRO to V. fischeri in a range from 61 mg L−1 (Ferrari et al., 2004) up to 184 mg L−1 (Calleja et al., 1994) are also in agreement with our results. Nevertheless, very interesting results were obtained for A. globiformis. PRO exhibited no toxic effect at concentrations of up to 100 mg kg−1 in sediment contact assay (Fig. 3S, Appendix B). On the other hand, results obtained during a parallel test without soil, exhibited a toxic effect for PRO towards these bacteria. Thus, it can be concluded that sorption explicitly inhibits (by decreasing the bioavailable fraction of the pharmaceutical) the hazardous effects of beta-blockers. Nevertheless, the observed value of EC50 (218 mg L−1) is relatively high and therefore it can be assumed that these pharmaceuticals do not pose a risk for A. globiformis which ubiquitously occurs in soil. Strong differences in the toxicity of different BB to aquatic organisms have been previously observed by other scientists (Huggett et al., 2002; Cleuvers, 2003, 2005; Fraysse and Garric, 2005; Dzialowski et al., 2006). PRO was the most potent of the compounds tested and has the highest logKow; MET was the second most potent compound and has the second highest logKow. NAD — having the lowest logKow did not elicit any toxicological responses in this study (see Fig. 1). For this reason, taking bioconcentration factors into account, with resulting internal effect concentrations differing only slightly, Cleuvers (2005) suggested that the differences in the EC50s values depended mainly on the diverse logKow levels, causing narcosis via disruption of membrane integrity. Such a partitioning of pollutants into the membrane causes the membrane to expand or swell, increasing fluidity, lowering the phase transition temperature, and the ion permeability of the membrane. On the other hand, he emphasizes that baseline toxicants may interact with specific receptors in the hydrophobic region of membrane proteins. Additionally, disturbances of lipid–protein or lipid–lipid interactions or lateral pressure changes were thought to be responsible for the narcosis. Therefore, it is likely that in reality narcosis covers maybe dozens of unknown, but specific, toxic modes of action. Although neither daphnids nor plants have been reported to possess betareceptors, this is not an unimportant point. For as we know from humans, adverse drug reactions can be caused by previously unrecognized drug-receptor interactions, previously unidentified receptors and by a broad diversity in drug-metabolizing phenotypes. These variables are even more poorly characterized in aquatic biota. Hugget et al. (2002) have also suggested that other factors may be involved, as earlier studies with invertebrates showed that e.g. paracetamol (logKow = 0.49) was more toxic to D. magna than methotrexate, salicilic acid and clofibric acid (logKow = 2.28, 2.24 and 2.84, respectively). For mammals, including humans, BB are a competitive antagonist of the effects of catecholamines (adrenaline and noradrenaline) at ßadrenergic-receptor sites (ß1 and/or ß2). These compounds have different affinities for ß1 and ß2 receptors (see Fig. 1). Some of them, such as PRO, have also a membrane-stabilizing activity (MSA). PRO reduces membrane permeability for various ions (Na+, K+ and Ca2+) and has a local anesthetic activity. Thus, the effect on organism survival may in part be due to adverse effects on membrane stabilization (Fraysse and Garric, 2005; Dzialowski et al., 2006; Claessens et al., 2013). Moreover, as was previously highlighted (see Section 3.1), for accurate comparisons of the ecotoxicological data, citation of CAS numbers is essential not only to clarify whether the nominal concentration data refers to the drug free base or the drug salt complex but also whether it refers to enantiomer form, which unfortunately is not a common practice. For example, like many therapeutics and other aquatic contaminants, PRO is distributed as a recemix mixture (R,S)-propranolol hydrochloride, (S)-(−)-propranolol hydrochloride as well as (R)-(+)propranolol hydrochloride. While for example the (S)-enantiomer is the most active form in mammals (up to a 100-fold difference), only the latter is responsible for propranolol's membrane-stabilizing effect. When such compounds are present in an asymmetrical biological
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environment, such as the highly specific binding sites of many cellular receptors, enantiospecific differences in biological activity can occur. Enantiomers may also differ significantly in their rates of uptake and excretion, affinity for plasma proteins, structure of metabolites, potency, toxicity and bioavailability (Stanley et al., 2006), hence this aspect should be also considered in the future research. Nevertheless, the mechanism by which BB affect these organisms remains elusive and must be examined in greater depth. 4. Conclusion In this study the ecotoxicological evaluation of three selected BB: PRO, MET and NAD has been performed for better understanding of the possible effects that these compounds have towards non-target organisms. According to the chemical substance classification of the European Union Directive 93/67EEC, PRO and MET can be considered to be harmful (10 b EC50 b 100 mg L−1) to aquatic organisms (based on toxicological data from the green alga test (S. vacuolatus)). In addition, new ecotoxicological data for three BB to V. fischerii, S. vacuolatus and L. minor as well as to soil bacteria (A. globiformis) is presented. Taking all the results into account, as well as the environmental concentrations of these compounds, it can be concluded that the risks posed by these compounds are of minor importance. However, it must be stressed that in the future, chronic biotests on other aquatic organisms should be carried out, bearing in mind the contribution of each individual compound to the overall toxic potential of the mixture of substances present in the aquatic environment. Ultimately, it's possible that the classification of BB could change from “not harmful” to “very toxic” depending on whether the compounds are applied in single exposures or are part of a mixture of substances. This comprehensive study supports the Environmental Risk Assessment of these pharmaceuticals. Acknowledgments The authors would like to acknowledge the financial support of National Science Centre (Poland) under decision DEC-2011/03/B/NZ8/03010 and DEC-2011/03/B/NZ8/03009 and the German Academic Exchange Service (DAAD). The publication is financed from European Social Fund as a part of the project “Educators for the elite-integrated training program for PhD students, post-docs and professors as academic teachers at University of Gdansk” within the framework of Human Capital Operational Programme, Action IV. The authors are grateful to Stephanie Steudte and Katarzyna Wychodnik and Juliane Filser for their support. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.06.039. References Altenburger R, Bodeker W, Faust M, Grimme LH. Evaluation of the isobologram method for the assessment of mixtures of chemicals: combination effect studies with pesticides in algal biotests. Ecotoxicol Environ Saf 1990;20:98–114. Amendola L, Molaioni F, Botrè F. Detection of beta-blockers in human urine by GC–MS– MS–EI: perspectives for the antidoping control. J Pharm Biomed Anal 2000;23(1): 211–21. Barbieri M, Licha T, Nödler K, Carrera J, Ayora C, Sanchez-Vila X. Fate of β-blockers in aquifer material under nitrate reducing conditions: batch experiments. Chemosphere 2012;89(11):1272–7. British Pharmacopoeia Commission. British Pharmacopoeia. 34th ed. London: Pharmaceutical Press; 2005. Brooks B, Huggett D. Human pharmaceuticals in the environment. New York Heidelberg Dordrecht London: Springer; 2012. Caban M, Mioduszewska K, Stepnowski P, Kwiatkowski M, Kumirska J. Dimethyl(3,3,3trifluoropropyl)silyldiethylamine—a new silylating agent for the derivatization of βblockers and β-agonists in environmental samples. Anal Chim Acta 2013;782:78–88. Calleja MC, Persoone G, Geladi P. Comparative acute toxicity of the first 50 multicentre evaluation of in vitro cytotoxicity chemicals to aquatic non-vertebrates. Arch Environ Contam Toxicol 1994;26(1):69–78.
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