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Chronic Effects Assessment and Plasma Concentrations of the β-blocker Propranolol
Title
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in Fathead Minnows (Pimephales promelas)
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Authors
Emma Giltrowa, Paul D. Ecclesab, Matthew J. Winterc, Paul J. McCormackc, Mariann
Rand-Weaverb, Thomas H. Hutchinsond, John P. Sumptera*
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a
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Kingdom
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b
Institute for the Environment, Brunel University, Kingston Lane, Uxbridge, Middlesex, UB8 3PH, United
Biosciences, School of Health Sciences and Social Care, Brunel University, Uxbridge, Middlesex, UB8
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3PH, United Kingdom
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c
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Brixham, Devon, TQ5 8BA, United Kingdom
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d
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Plymouth, PL1 3DH, United Kingdom
AstraZeneca Safety, Health and Environment, Brixham Environmental Laboratory, Freshwater Quarry,
Natural Environmental Research Council, Plymouth Marine Laboratory, Prospect Place, The Hoe,
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* Corresponding author. Tel: +44 1895 266303; Fax: +44 1895 269761; E-mail address;
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john.sumpter@brunel.ac.uk
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Abstract
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The presence of many human pharmaceuticals in the aquatic environment is now a worldwide
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concern, yet little is known of the chronic effects that these bioactive substances may be having on
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aquatic organisms. Propranolol, a non-specific beta adrenoreceptor blocker (β-blocker), is used to
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treat high blood pressure and heart disease in humans. Propranolol has been found in surface
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waters worldwide at concentrations ranging from 12 to 590 ng/L.
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ecologically relevant effects in fish in receiving waters, short-term (21d) adult reproduction studies
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were conducted, in which fathead minnows were exposed to concentrations propranolol
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hydrochloride [CAS number 318-98-9] ranging from nominal concentrations of free drug from
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0.001 to 10 mg/L (measured concentrations typically from 78-130%). Exposure of fish to 3.4 mg/L
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(measured) over 3 days caused 100% mortality or severe toxicity requiring euthanasia. The most
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sensitive endpoints from the studies were a decrease in hatchability (with regard to the number of
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days to hatch) and a dose-related increase in female gonadal-somatic index (GSI), giving LOEC
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hatchability
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observed, with LOEC
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Collectively these data do not suggest that propranolol was acting as a reproductive toxin.
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Propranolol plasma concentrations in male fish exposed to water concentrations of 0.1 and 1.0 mg/L
and LOEC
female GSI
To test the potential for
values of 0.1 mg/L. Dose related decreases in male weight were also
male wet weight value
of 1.0 mg/L, and LOEC
reproduction
values were 1.0 mg/L.
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were 0.34 and 15.00 mg/L respectively which constitutes 436 and 1546 % of measured water
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concentrations. These compare with predicted concentrations of 3.57 and 44.38 mg/L, and thus
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support the use of partition coefficient data models for predicting resultant fish plasma levels. In
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addition, propranolol plasma concentrations in fish exposed to water concentrations of 0.1 and 1.0
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mg/L were greater than the human therapeutic plasma concentration and hence these data very
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strongly support the fish plasma model proposed by Huggett et al. (2003). (Huggett, D.B., Cook,
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J.C., Ericson, J.F., Williams, R.T. 2003. A theoretical model for utilizing mammalian pharmacology
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and safety data to prioritize potential impacts of human pharmaceuticals to fish. Human and
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Ecological Risk Assessment, 9, 1789-1799).
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Keywords: β-blocker; Fathead minnow; Pharmaceutical; Propranolol; Fish plasma concentration
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1.
Introduction
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Pharmaceuticals in the environment has become a growing concern after the realisation that ethinyl
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estradiol could be found in the aquatic environment at concentrations exceeding 0.5ng/L, the
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concentration at which vitellogenin induction can occur in male fish (Purdom et al., 1994). Since
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then advancement of analytical techniques have shown that there are in fact many pharmaceuticals
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in our rivers and surface waters, including the human beta adrenoreceptor blocker (-blocker) (Fent
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et al., 2006; Ternes 2001). The question that follows therefore, is are the concentrations of
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pharmaceuticals found in the aquatic environment high enough to cause harm to the wildlife that are
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exposed to them?
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Propranolol and other human and animal pharmaceuticals reach the aquatic environment through a
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variety of routes. However, the majority of pharmaceuticals are present in the aquatic environment
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due to incomplete removal at sewage treatment works (STWs). In Germany it was found that 96 %
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of propranolol was removed from the influent compared to the effluent, yet the 4 % remaining is
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largely why propranolol has been found ubiquitously in rivers and streams in America and Europe
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at concentrations in the ng/L range, with maximum and median concentrations reaching 590 ng/L
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and 12 ng/L, respectively (Ashton et al., 2004; Huggett et al., 2003; Ternes, 1998).
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There are many different classes of pharmaceuticals in the environment but this study and a
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companion study (Winter et al., 2008) focus solely on one group, the β-blockers. This is so that a
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comprehensive picture around one class of pharmaceuticals can be built up in order to try to
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establish some general principles that might be applicable to other classes of pharmaceuticals. A
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feature of the study was to apply the ‘fish plasma model’ as hypothesised by Huggett et al. (2003),
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which compares estimated or actual drug concentrations in fish plasma with human therapeutic
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plasma concentrations, in order to assess whether it is likely that environmental or experimental
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concentrations of a drug would produce therapeutic levels in fish. If the fish plasma model is found
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to be successful when applied to β-blockers, it might also be applicable to other groups of
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pharmaceuticals.
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β-blockers are grouped as one class of pharmaceuticals, and although they all act on beta-adrenergic
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receptors (β-ARs), they can differ greatly in their specificity and lipophilic properties (Owen et al.,
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2007). For example, propranolol has a relatively high log Kow of 3.48, whereas atenolol has a
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considerably lower log Kow of 0.23. A report by Winter et al. (2008) showed atenolol to have low
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chronic toxicity to fathead minnows. The results produced a 4 day embryo LOEC hatching value of >
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10 mg/L, a LOEC reproduction of >10 mg/L and the most sensitive endpoint was condition index,
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which value gave a LOEC condition index of 3.2 mg/L.
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Published data show propranolol, out of all the β-blockers investigated, to be the most toxic to
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aquatic organisms. For example, invertebrate LC50 values for metoprolol and propranolol range
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from 64 to >100 mg/L and 0.8 to 29.8 mg/L, respectively, showing that propranolol is harmful to
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invertebrates at much lower concentrations than metoprolol (Cleuvers, 2003; Huggett et al., 2002,
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Villegas-Navarro et al., 2003). In fish studies, the published data on propranolol are limited.
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However, a report by Huggett et al. (2002) showed propranolol to be relatively toxic with LOEC
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hatchability and egg production
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suggesting that propranolol is not particularly toxic to aquatic organisms but some suggesting this is
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not the case.
values of 0.0005 mg/L. Hence the message is mixed, with most studies
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2.
Materials and Methods
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2.1
Test design and concentration
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The experimental design incorporated two separate replicate experiments. From an ethical
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standpoint, to conduct two smaller experiments, rather than one large experiment, was preferable
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because it meant that if no effect whatsoever was found in the first experiment, the second
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experiment need not be undertaken, hence reducing the number of fish used. Secondly, if a
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particular concentration of the test substance evoked a toxic effect, that particular concentration
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need not be repeated in the second experiment, hence less fish are put under duress. For statistical
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reasons, conducting two separate experiments allows the results of the two experiments to be
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compared, and if any effects occur and are repeatable, the results will be more robust. In order to
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compare atenolol and propranolol the experimental protocol was based on that used by Winter et al.
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(2008). Hence, as in Winter et al. (2008), the first experiment in this study incorporated a 10 mg/L
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treatment group to replicate concentrations that may be used if conducting experimental work for a
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environmental risk assessment (ERA).
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2.2
Experimental procedure
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A 21 day adult reproduction study was carried out using the protocol set out in Ankley et al. (2001)
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and Harries et al. (2000), later modified by Winter et al. (2008).
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using a flow-through system in a room that contained two series of identical tanks.
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Thermostatically heated (25 ± 1 ºC) dechlorinated tap water from a header tank flowed through 6
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flow meters into 6 mixing chambers, via medical grade silicon tubing, and from these the water
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went to 6 fish tanks through silicone rubber tubing. In total there were 36 fish tanks which each had
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a working volume of 10.5 litres. Each tank was aerated and received roughly 9 tank renewals of
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water per day. During the exposure period, stock concentrations of propranolol were pumped from
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4L foil wrapped amber stock bottles through 0.16mm tubing to a peristaltic pump and out through
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peroxide cured silicon tubing to the mixing chambers.
Each experiment was carried out
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Inside each fish tank, a half piece of PVC gutter was placed over a stainless steel mesh and glass
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base (Thorpe et al., 2007). Most eggs were collected from the underside of the tile, to which they
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had been attached by the fish. However, in the event that some eggs did not adhere to the tile, they
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would fall onto the mesh or into the glass base.
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After a 21 day pre-exposure period followed by 21 days of exposure to propranolol-HCl, the fish
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were humanely terminated using an overdose of ethyl 3-aminobenzoate methanesulfonate (MS-
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222). Plasma samples were pooled from each sex from each concentration for analysis of
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propranolol concentration in the plasma. Wet weight, fork length, gonad weight and fatpad weight
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were recorded. The number of tubercles and the prominence of tubercles were determined using the
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qualitative scheme set out by Smith (1978).
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2.3
Hatchability trails
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To assess whether the eggs produced by each pair were viable (i.e. successfully produced live
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young), hatchability trials were undertaken. After the first successful spawning in the baseline
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period, fifty eggs were left on the tile from the next spawn and placed in a wire meshed cage that
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was submersed in a separate tank receiving the same concentration of propranolol (i.e. the
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developing eggs were exposed to the same concentration of propranolol as were the adult fish that
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laid the eggs). The eggs were checked and counted daily and any abnormalities recorded and dead
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eggs removed.
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2.4
Test substance
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The test substance, pharmaceutical grade propranolol (DL-Propranolol hydrochloride, 1–
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isopropylamino-3-(1-naphthyloxy) proprano-2-ol hydrochloride, 99 % pure, racemic mixture, CAS
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318-98-9, to be referred to as propranolol), was obtained from AstraZeneca, Brixham (UK). In the
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first experiment, five concentrations of propranolol namely 0.001, 0.01, 0.1, 1.0 and 10 mg/L were
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used, together with a dilution water control (DWC). In the second experiment this was reduced to 4
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concentrations of propranolol which were 0.001, 0.01, 0.1, 1.0 mg/L and DWC. In both
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experiments, there were six tanks for each treatment group, i.e. four pairs of fathead minnow and 2
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tanks for hatchability trials at each concentration. Because unexpected results occurred at 10 mg/L
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in the first experiment (see later), and because it was not intended to collect acute toxicity data for
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propranolol, the 10 mg/L concentration was not repeated in experiment 2, and so for the second
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experiment it was possible to collect data from five fathead minnow pairs in the control tanks.
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2.5
Test species
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The fish in both experiments were one year old fathead minnows, Pimephales promelas. They were
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virginal stock that had been bred at Brunel University. One male and one female were placed in
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each fish tank. All male fish used in the pair-breeding experiments exhibited secondary sexual
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characteristics prior to the start of the experiments. The fish were fed twice a day with frozen adult
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brine shrimp (Artemia salina) during the week and once a day during the weekend, and in addition,
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fish flakes were also fed once a day. The fish received a photoperiod of 16 hours of light followed
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by 8 hours of darkness, with a 20 minute dawn: dusk transition.
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2.6
Analysis of water samples
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Liquid chromatography was used for analyte separation with 0.1 % ammonia in water and 0.1 %
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ammonia in methanol over an elution gradient. Electrospray ionisation-tandem mass spectrometry
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(ESI-MS/MS) was then used to quantify the analytes, which was carried out in the positive
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ionization mode (+) ESI. For increased selectivity, MS/MS with selected reaction monitoring
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(SRM) was used for detection of protonated propranolol ([M + H]+, 260.2 Da) and XcaliburTM
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software was used for data acquisition and processing.
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2.7
Analysis of plasma samples
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All samples and standards were prepared by protein precipitation (PPT) by which 50 µL of plasma
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was added to acetonitrile (200µL) containing metoprolol internal standard (0.25 µg L-1) in a
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filtration vial (0.45µm). After vortexing the filtrate transferred to a vial and diluted with 100 µL,
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30:70 acetonitrile/water. Propranolol concentration was then measured using online solid-phase
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extraction/liquid chromatography (SPE/LC).
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2.8
Statistical analyses
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Data were checked for normal distribution and equal variance using the Kolmogorov-Smirnov test
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and Levene Median test, respectively. Statistical analyses of normally distributed data of equal
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variance were carried out either using a t-test, or a One Way Repeated Measures ANOVA to
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compare two sets of data or more than two data sets, respectively. If a statistical difference was
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found in the latter case, a Holm-Sidak post-hoc test was carried out to show which treatment groups
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were different. If the data were not of a normal distribution, a non-parametric Kruskal-Wallis One
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Way Analysis of Variance on Ranks test was performed, followed by a Dunn’s post hoc test. The
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data for 1.0 mg/L hatchability trials could not be analysed as there were fewer than two data-points
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from the exposure period
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3.
Results
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3.1
Propranolol water concentration
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Propranolol concentrations from the water chemistry analyses showed that the measured
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propranolol concentration for each treatment was in the expected range, apart from the 10 mg/L
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tanks in experiment 1 (Table 1). This showed that there was no contamination of control tanks with
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propranolol. At a nominal concentration of 10 mg/L, only 34 % of the nominal dose was measured
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in the tanks. This may have been due to solubility issues, and/or uptake of the drug by these fish in
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the tanks. In all other treatments, the mean propranolol concentrations in experiments 1 and 2
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ranged from 53 to 112 % and from 78 to 130 % of the nominal concentrations, respectively,
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demonstrating that the measured concentrations were relatively close to the nominal concentrations.
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Statistical analysis of the water chemistry results showed that there were no statistically significant
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differences in the mean results at each concentration between the two experiments, and so data from
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each experiment were combined to make one data set.
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3.2
Fish mortality
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During the three day transition period in experiment 1, during which the propranolol concentrations
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were expected to rise steadily towards the normal values, the fish in the 10 mg/L tanks (actual
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concentration 3.4 mg/L) died or had to be euthanized before the 21 day exposure period began. In
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all pairs the males became obviously ill before the females, and before termination showed a lack of
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orientation and were noted to be swimming on one side and nosing the bottom of the tank. At the
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1.0 mg/L concentration on day 5 in experiment 1, two males started to show similar behaviour to
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the 10 mg/L fish and were euthanized. During experiment 2, the fish at the 1.0 mg/L concentration
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were closely observed during the exposure period. It was noted on day 2 that all the fish exposed to
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1.0 mg/L propranolol had reduced appetites, as a noticeable quantity of uneaten food could be seen
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on top and around the spawning tiles. By day 4, two males were euthanized as they had started to
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swim on one side with an apparent loss of orientation and their females were terminated on day 5 as
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they also showed the same symptoms. The remaining fish at this concentration, despite having
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reduced appetites, showed no obvious signs of distress. On day 11 during experiment 2 of the
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exposure period, one of the males at the 1.0 mg/L concentration was found dead outside of the tank.
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This was most unexpected, as no signs of distress had been observed in the days prior to this
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incident. The remaining fish being maintained at 1.0 mg/L were then culled after this event and
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sampling data and tissues collected. Hence, egg data were collected for only 11 days in experiment
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2 from the 1.0 mg/L treatment group. Four other deaths occurred during the two experiments and
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in all cases it was because the female had become egg-bound. Data from fish that died during the
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exposure period were excluded from the baseline data.
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3.3
Cumulative egg production
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Table 2 shows a summary of egg production data during the baseline and exposure period from
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each experiment. During the baseline period of 21 days across both experiments, the total number
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of eggs per female ranged from 590 to 1074, and the mean number of eggs per female per
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reproductive day ranged from 28 to 51. These data compare favourably to pair breeding assay data
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reported by Winter et al. (2008), where the number of eggs per female per reproductive day ranged
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from 27 to 50, but are lower than those reported by Thorpe et al. (2007) (80 to 93 eggs/female/day).
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The mean number of eggs per spawning over the 21 days was between 147 and 238 in the baseline
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period. This range is higher than data reported by Winter et al. (2008) (103 to 161 eggs/spawn) and
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lower than that reported by Thorpe et al. (2007) (358 ± 14 eggs/spawn). In the 1.0 mg/L treatment
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group the data collected over 11 days during the baseline period show the number of eggs produced
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per day and per spawn in experiment 1 to be reduced compared to other treatment groups. Egg
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production in the exposure period was similar to that in the baseline period in the DWC group, and
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also in the groups exposed to 0.001, 0.01 and 0.1 mg propranolol/L. There was evidence of reduced
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egg production in response to exposure to 1 mg/L. This apparent effect was not significant in
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experiment 1 (in which low survival of the adult fish, and unexpected low egg production in the
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baseline period complicated any analysis) but was significant (p<0.05) in experiment 2. There was
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no statistically significant difference between baseline and exposure periods for any of the other
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treatments.
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3.4
Body and organ size
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Analyses of male and female wet weight data obtained at the end of the experiments showed that
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weight ranged from 2.433 to 5.680 g in male fish and from 0.811 to 2.490 g in female fish. The
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male wet weight data showed a dose-related decrease, in which there was a statistically significant
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difference between the DWC group and the 1.0 mg/L treatment group (p < 0.05). This trend was
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not seen in the female wet weight data, and there were no statistically significant differences
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between treatment groups. Male fork length ranged from 52 to 71 mm and female fork length
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ranged from 42 to 65 mm. No statistically significant differences were found between treatment
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groups.
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Gonad weight in males ranged from 0.01 to 0.074 g and gonadal somatic index (GSI) ranged from
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0.22 to 1.39 %. Figure 1A shows male GSI data, in which a statistically significant difference was
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found between the DWC and 0.01 mg/L treatment groups. Both the 0.001 mg/L and the 0.01 mg/L
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groups also had higher mean GSI values than the DWC group, however there was not a significant
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trend of increasing GSI with increasing propranolol concentration. Female gonad weight ranged
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from 0.01 to 0.357 g and female GSI ranged from 1.32 to 17.14 %. In contrast, female GSI data
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exhibited a dose-related effect, for as the concentration of propranolol increased, so did the GSI, as
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shown in Figure 1B. These female GSI data produced statistically significant differences between
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treatment groups DWC and 0.1 mg/L, DWC and 1.0 mg/L, and between 0.001 and 1.0 mg/L.
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Figure 2 shows the weight of the fatpads from the male fish at each concentration of propranolol.
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The male fatpad weight ranged from 0.090 to 0.477 g. As the concentration of propranolol
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increased from 0.01 mg/L, there was a dose-related decrease in the weight of the fatpad, and a
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statistically significant difference was found between treatment groups 0.01mg/L and 1.0 mg/L and
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also between groups 0.1 mg/L and 1.0 mg/L (p < 0.05). However, no treatment group showed a
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statistically significant difference when compared to the control group. The number of tubercles on
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the males ranged from 12 to 16, and the tubercle prominence score from 1 to 4 (Smith, 1978).
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There were no statistically significant differences between treatment groups in either of these sets of
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data. No secondary sexual characteristics were observed in any female fish sampled.
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Figure 3 shows the condition index (CI) of male and female fish in each treatment group. These
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data range from 1.21 to 1.99 in male fish and between 1.00 and 1.58 in female fish. A statistically
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significant difference exits between the male fish CI of the 0.001 mg/L and 1.0mg/L groups and
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between the CI of the 0.01 mg/L and 1.0 mg/L treatment groups. However, there were no treatment
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groups in which the CI was statistically significantly different from that of the control group. No
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statistically significant differences were found between treatment groups in the female data.
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3.5
Hatchability
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The data from the hatchability trials are summarized in Table 3. Data from the 1.0 mg/L group could not
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be statistically analysed because there were fewer than 2 data points. A high proportion of eggs (between
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75 and 95 %) hatched. There was no evidence that exposure of eggs to propranolol affected the proportion
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of eggs that hatched. Time to hatch was unaffected by 0.001 and 0.01 mg propranolol/L (Table 3), but was
289
significantly delayed (by one day) in eggs exposed to 0.1 mg propranolol/L. Time to hatch also appeared
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to be delayed (by 3 days) when eggs were exposed to 1 mg propranolol/L, but low sample number
291
compromised this apparent effect. In the 0.1mg/L treatment group there was an increase in the time to
292
hatch between the baseline and exposure periods, which was statistically significant.
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3.6
NOEC and LOEC values
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Table 4 summarises the lowest observed effect concentrations (LOEC) and the no observed effect
296
concentration (NOECs) for each of the endpoints that were measured. In summary, propranolol
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affected hatchability (with regard to the number of days it takes for eggs to hatch) and female GSI
298
at 0.1 mg/L. These were the most sensitive endpoints. At 1.0 mg/L, propranolol significantly
299
affected male survival and male wet weight, and at 3.4 mg/L propranolol significantly affected
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female survival.
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3.7
Plasma propranolol concentration
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Fish in the control tanks (DWC) had no measurable propranolol in their plasma, but as the
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concentration of propranolol increased in the water, the fish plasma propranolol concentration
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increased dramatically, as shown in Table 5. At exposure concentrations of 0.1 and 1.0 mg/L, the
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fish bioaccumulated propranolol in their blood, and this was especially obvious in male fish. At
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low propranolol concentrations (0.001 and 0.01 mg/L), fish plasma concentrations did not reach, or
308
only just reached water concentrations, whereas at higher water concentrations (0.1 and 1.0 mg/L),
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plasma concentrations were higher than water concentrations.
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concentrations of 0.1 and 1.0 mg/L for 21 days had plasma propranolol concentrations higher than
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the human therapeutic plasma concentration (HTPC).
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313
4
Discussion
Fish exposed to propranolol
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The toxicity data obtained from the propranolol 21-day adult reproduction study reported here
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differed at least ten-fold from the comparable data obtained with atenolol (Winter et al., 2008).
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Fathead minnows were able to tolerate high concentrations of atenolol (LOEC survival >10 mg/L),
317
unlike in this study, where propranolol caused acute toxicity at lower concentrations (LOEC survival
318
3.4 mg/L). The main reason for this could be due to differences in lipophilicity between the two β-
319
blockers. Atenolol has a much lower log Kow, compared to propranolol, and is expected therefore
320
to less readily bioconcentrate. Indeed, plasma concentrations of atenolol in relation to the
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corresponding measured water levels only ranged from 1.8 to 12.2 %, unlike plasma concentrations
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of propranolol, which reached up to 1550 % of the surrounding water concentration.
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Before the fish in the 10 mg/L treatment group (actual concentration = 3.4 mg/L) were euthanized,
325
they were observed to be very disorientated and swimming on one side or nosing the bottom of the
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tank. The loss of appetite together with the disorientation of the fish and lack of sexual behaviour
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indicates that propranolol may have been acting as a CNS toxin. The relatively high lipophilic
328
nature of propranolol means it can easily move across the blood-brain barrier in organisms (Vu &
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Beckwith 2007). Studies have shown that propranolol can be found in human brain tissue at
330
concentrations ten to twenty six times higher than in plasma (Cruickshank & Neil-Dwyer, 1985;
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Westerlund, 1985). This is a high degree of bioconcentration, especially when compared to brain
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concentrations of the hydrophilic β-blocker atenolol, which has a brain: plasma ratio of 0.2
333
(Cruickshank & Neil-Dwyer, 1985). Similar accumulation can also be found in the brains of rabbits
334
and monkeys that were administered propranolol (Srivastava & Kapoor, 1983). Thus it is possible
335
that in our experiments the fish exposed to 1.0 and 10 mg propranolol/L bioaccumulated enough
336
propranolol within the CNS to cause the acute toxicity that occurred.
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338
When Huggett et al. (2002) exposed Japanese medaka to propranolol, egg production was found to
339
be decreased at 0.0005 mg/L, suggesting that propranolol was acting as a reproductive toxin. The
340
effects found by Huggett et al. (2002), although statistically significant when compared with the
341
controls, did not however exhibit a dose-response relationship. The data from this study do not
342
agree with those reported in Huggett et al. (2002), as egg production was only affected at 1.0 mg/L,
343
a 2000-fold higher concentration than that which appeared to inhibit egg production in medaka.
344
Surprisingly, fathead minnows (LOEC male survival = 1.0 mg/L, LOEC female survival = 3.4 mg/L) were
345
much more acutely sensitive to propranolol than medaka (LC50 = 24.3 mg/L) (Huggett et al., 2002).
346
The reason for this inter-species difference is presently unclear. The data from this study show that
347
propranolol concentrations in the aquatic environment (maximum values = 590 ng/L, Ternes, 1998)
348
and would not be expected to cause reproductive effects.
349
350
When pharmaceuticals are taken by patients, the most relevant concentration that describes the
351
quantity of drug needed to initiate a physiological response is the human plasma concentration of
352
the drug. This is because it is the plasma containing the drug which comes into contact with the
353
target site. Since a certain plasma concentration of pharmaceutical is required to affect target
354
receptors and enzymes in humans, it has been hypothesised that approximately the same
355
concentration would be required to affect another species sharing the same target (Brown et al.,
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2007). This assumption has been used to describe concentrations of pharmaceuticals in the
357
environment that may induce effects in receiving biota. One such model, the ‘fish plasma model’
358
proposed by Huggett et al. (2003) compares estimated (from the partition coefficient) or actual drug
359
concentrations in fish plasma with human therapeutic plasma concentrations, in order to assess
360
whether it is likely that environmental or experimental concentrations of a drug would produce
361
therapeutic levels in fish (Brown et al., 2007). The effect ratio (ER) is calculated as shown below:
362
363
ER = Human therapeutic plasma concentration (HTPC)/ Fish steady state plasma
364
concentration (FSSPC)
365
366
The lower the ER (i.e. ≤ 1), the greater the potential there is for a pharmacological response to
367
occur in fish (Huggett et al., 2003).
368
369
Due to the different physical properties of pharmaceuticals, some drugs are more likely to reach
370
human plasma therapeutic levels from waterborne exposure of aquatic organisms than others.
371
Propranolol has a log Kow of 3.48, and since it is relatively lipophilic it is likely to bioaccumulate in
372
fish from exposure via water. The plasma propranolol concentrations in fathead minnows showed a
373
dose-related relationship, in that as the concentration of propranolol in the water increased,
374
propranolol plasma concentrations also increased (Table 5). The measured FSSPC concentration in
375
the 0.1 and 1.0 mg/L treatment groups exceeded the water concentration of propranolol, showing
376
bioaccumulation of propranolol. These data also agree with data by Wong et al. (1979), who
377
established that propranolol plasma concentration and daily dose were significantly correlated. By
378
using the ‘fish plasma model’, the predicted FSSPC, as calculated by using log Kow (Table 5) are
379
within 1 order of magnitude to the measured FSSPC, and hence the model does quite well in
380
predicting the plasma concentrations of propranolol.
381
382
The ER values of male and female fathead minnows exposed to 0.1 and 1.0 mg/L of propranolol
383
were less than 1; hence, the measured FSSPC in the fish maintained at these concentrations was
384
greater than the drug concentration in human plasma which is needed to elicit a therapeutic effect.
385
This could explain why fish exposed to concentrations of 0.1mg/L of propranolol, and higher,
386
showed statistically significant effects from the control in some endpoints, and why propranolol at
387
3.4 mg/L was so toxic. These data support the fish plasma model and show that the read-across of
388
information from humans to fish is plausible in that existing human pharmaco and toxicodynamic
389
data could be used as a starting point to predict effects on organisms in the environment (at least
390
those sharing targets with humans).
391
392
In summary, these data do not suggest that propranolol is acting as a reproductive toxin and
393
therefore do not agree with the findings of Huggett et al., 2002. The environmental concentrations
394
of propranolol are not greater, or equal to, 0.1mg/L, which was the lowest concentration at which
395
effects were observed, and so adult fathead minnows are not considered to be at risk from
396
propranolol in the aquatic environment. However, these data clearly support the fish plasma model,
397
as effects were seen in treatment groups where the internal plasma concentration of propranolol
398
exceeded the human therapeutic plasma concentration required to treat patients for hypertension
399
(Huggett et al., 2003; Wong et al., 1979).
400
401
Acknowledgements
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The authors would like to thank the European Union for funding this work as part of the ERAPharm
403
project, contract no. 511135 (Knacker et al., 2005), and to NERC for providing studentship funding
404
to PDE
405
406
References
407
Ankley, G.T., Jensen, K.M., Kahl, M.D., Korte, J.J., Makynen, E.A. 2001. Description and
408
evaluation of a short-term reproduction test with the fathead minnow (Pimephales promelas).
409
Environ. Toxicol. Chem. 20, 1276-1290.
410
411
Ashton, D., Hilton, M., Thomas, K.V. 2004. Investigating the environmental transport of human
412
pharmaceuticals to streams in the United Kingdom. Sci. Total Environ. 333, 167-184.
413
414
Brown, J.N., Paxeus, N., Forlin, L., Larsson, D.G.J. 2007. Variations in bioconcentration of human
415
pharmaceuticals from sewage effluents into fish blood plasma. Environ. Toxicol. Pharmacol. 24,
416
267-274.
417
418
Cleuvers, M. 2003. Aquatic ecotoxicity of pharmaceuticals including the assessment of combination
419
effects. Toxicol. Lett. 142, 185-194.
420
421
Cruickshank, J.M., Neil-Dwyer, G. 1985. Beta-blocker brain concentrations in man. Eur. J. Clin.
422
Pharmacol. 28, 21-23.
423
424
Fent, K., Weston, A.A., Caminada, D. 2006. Ecotoxicology of human pharmaceuticals.
425
Aquat. Toxicol. 76, 122-159.
426
427
Harries, J.E., Runnalls, T., Hill, E., Harris, C.A., Maddix, S., Sumpter, J.P., Tyler, C.R. 2000.
428
Development of a reproductive performance test for endocrine disrupting chemicals using pair-
429
breeding fathead minnow (Pimephales promelas). Environ. Sci. Technol. 34, 3003-3011.
430
431
Huggett, D.B., Brooks B.W., Peterson B., Foran, C.M., Schlenk, D. 2002. Toxicity of select beta
432
adrenergic receptor-blocking pharmaceuticals (β-blockers) on aquatic organisms. Arch. Environ.
433
Contam. Toxicol. 43, 229-235.
434
435
Huggett, D.B., Khan, I.A., Foran, C.M., Schlenk, D. 2003. Determination of beta-adrenergic
436
receptor blocking pharmaceuticals in United States wastewater effluent. Environ. Pollut. 121, 199-
437
205.
438
439
Huggett, D.B., Cook, J.C., Ericson, J.F., Williams, R.T. 2003. A theoretical model for utilizing
440
mammalian pharmacology and safety data to prioritize potential impacts of human pharmaceuticals
441
to fish. Human and Ecological Risk Assessment. 9, 1789-1799.
442
443
Knacker, T., Duis, K., Ternes, T., Fenner, K., Escher, B., Schmitt, H., Rombke, J., Garric, J.,
444
Hutchinson, T.H., Boxall, A.B. 2005. The EU-project ERAPharm, Incentives for the further
445
development of guidance documents? Envrion. Sci. Pollut. Res. Int. 12, 62-65.
446
447
Owen, S.F., Giltrow, E., Huggett, D.B., Hutchinson, T.H., Saye, J., Winter, M.J., Sumpter, J.P.
448
2007. Comparative physiology, pharmacology and toxicology of beta-blockers: Mammals versus
449
fish. Aquat. Toxicol. 82, 145-162..
450
451
Purdom C.E., Hardiman P.A., Bye V.J., Eno N.C., Tyler C.R., Sumpter J.P. 1994. Estrogenic
452
effects of effluents from sewage treatment works. Chem. Ecol. 8, 75-285.
453
454
Smith, R.J.F. 1978. Seasonal changes in the histology of the gonads and dorsal skin of the fathead
455
minnow, Pimephales promelas. Can. J. Zool. 56, 2103-2109.
456
457
Srivastava, M., Kapoor, N.K. 1983. The effect of propranolol on rat-brain catecholamine
458
biosynthesis. J. Biosci. 5, 261-266.
459
460
Ternes, T.A. 1998. Occurrence of drugs in German sewage treatment plants and rivers. Water Res.
461
32, 3245-3260.
462
463
Ternes, T., Bonerz, M., Schmidt, T. 2001. Determination of neutral pharmaceuticals in wastewater
464
and
465
J. Chromatogr. A. 938, 175-185.
rivers
by
liquid
chromatography-electrospray
tandem
mass
spectrometry.
466
467
Thorpe, K.L., Benstead, R., Hutchinson, T.H., Tyler, C.R. 2007. An optimised experimental test
468
procedure for measuring chemical effects on reproduction in the fathead minnow, Pimephales
469
promelas. Aquat. Toxicol. 81, 90-98.
470
471
Villegas-Navarro, A., Rosas-L, E, Reyes, J.L. 2003. The heart of Daphnia magna: effects of four
472
cardioactive drugs. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 136, 127-134.
473
474
Vu, J., Beckwith, M.C. 2007. Central nervous system side effects of beta-adrenergic blocking
475
agents. http://health.utah.gov/medicaid/pharmacy.
476
477
Westerlund, A. 1985. Central nervous-system side-effects with hydrophilic and lipophilic beta-
478
blockers. Eur. J. Clin. Pharmacol. 28, 73-76.
479
480
Winter, M.J., Lillicrap, A.D., Caunter, J.E., Schaffner, C., Alder, A.C., Ramil, M., Ternes, T.A.,
481
Giltrow, E., Sumpter, J.P., Hutchinson, T.H. 2008. Defining the chronic impacts of atenolol on
482
embryo-larval development and reproduction in the fathead minnow (Pimephales promelas). Aquat.
483
Toxicol. 86, 361-369.
484
485
Wong, L., Nation, R.L., Chiou, W.L., Mehta, P.K., 1979. Plasma concentrations of propranolol and
486
4-hydroxypropranolol during chronic oral propranolol therapy. Br. J. Clin. Pharmacol. 8, 163-167.
487
488