Chronic toxicity of ibuprofen to Daphnia magna: Effects on life history traits and population dynamics

Toxicology Letters 172 (2007) 137–145

Chronic toxicity of ibuprofen to Daphnia magna: Effects on life history traits and population dynamics Lars-Henrik Heckmann a,∗ , Amanda Callaghan a , Helen L. Hooper a , Richard Connon a,1 , Thomas H. Hutchinson b , Steve J. Maund c , Richard M. Sibly a a

b

University of Reading, School of Biological Sciences, Environmental Biology, Philip Lyle Building, Reading RG6 6BX, United Kingdom AstraZeneca Global Safety, Health & Environment, Brixham Environmental Laboratory, Devon TQ5 8BA, United Kingdom c Syngenta Crop Protection AG, 4002 Basel, Switzerland Received 8 April 2007; received in revised form 4 June 2007; accepted 5 June 2007 Available online 14 June 2007

Abstract The non-steroidal anti-inflammatory drug (NSAID) ibuprofen (IB) is a widely used pharmaceutical that can be found in several freshwater ecosystems. Acute toxicity studies with Daphnia magna suggest that the 48 h EC50 (immobilisation) is 10–100 mg IB l−1 . However, there are currently no chronic IB toxicity data on arthropod populations, and the aquatic life impacts of such analgesic drugs are still undefined. We performed a 14-day exposure of D. magna to IB as a model compound (concentration range: 0, 20, 40 and 80 mg IB l−1 ) measuring chronic effects on life history traits and population performance. Population growth rate was significantly reduced at all IB concentrations, although survival was only affected at 80 mg IB l−1 . Reproduction, however, was affected at lower concentrations of IB (14-day EC50 of 13.4 mg IB l−1 ), and was completely inhibited at the highest test concentration. The results from this study indicate that the long-term crustacean population consequences of a chronic IB exposure at environmentally realistic concentrations (ng l−1 to ␮g l−1 ) would most likely be of minor importance. We discuss our results in relation to recent genomic studies, which suggest that the potential mechanism of toxicity in Daphnia is similar to the mode of action in mammals, where IB inhibits eicosanoid biosynthesis. © 2007 Elsevier Ireland Ltd. All rights reserved. Keywords: Invertebrate; Stress response; Fecundity; Reproduction; Mode of action; NSAID

1. Introduction The release of human and veterinary pharmaceuticals into the environment has been the focus of recent ∗

Corresponding author. Tel.: +44 118 378 4426; fax: +44 118 931 0180. E-mail address: [email protected] (L.-H. Heckmann). 1 Present address: University of California, School of Veterinary Medicine, Department of Anatomy, Physiology and Cell Biology, Davis, CA 95616, USA.

research and review programmes (K¨ummerer, 2004). In Europe, several over-the-counter drugs like acetylsalicylic acid, ibuprofen (IB) and paracetamol are consumed in amounts of over 100 t per year (Bound and Voulvoulis, 2005; Zwiener and Frimmel, 2000). IB is not fully metabolised by humans and may, therefore, enter the sewage system as the parent compound or metabolites (Buser et al., 1999). Treatment in modern sewage treatment plants appears to eliminate the vast majority of IB and its metabolites with degradation by >95% reported in the literature (Buser et al., 1999). In surface water

0378-4274/$ – see front matter © 2007 Elsevier Ireland Ltd. All rights reserved. doi:10.1016/j.toxlet.2007.06.001

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monitoring programmes, IB has been detected at the ng l−1 to ␮g l−1 range in various parts of the world (Andreozzi et al., 2003; Han et al., 2006; Kolpin et al., 2002; Metcalfe et al., 2003; Stumpf et al., 1999). IB is a non-steroidal anti-inflammatory drug (NSAID) that has relatively high mobility in aquatic environments, but a low persistence compared to other pharmaceuticals (Buser et al., 1999). The half-life of IB has been estimated to be 32 days in the field (Tixier et al., 2003). Maximum concentrations of IB in UK surface waters have been found to be 5 ␮g l−1 with an estimated risk quotient (RQ) of 0.01 (Ashton et al., 2004). Previously, Stuer-Lauridsen et al. (2000) calculated a rather conservative RQ above 1 that was based on no elimination of the active compound, and a predicted environmental concentration of IB. Both of these RQ-values are based on data from the few acute studies that have been published on the cladoceran crustacean Daphnia magna Straus, which report that the 48 h EC50 for immobility is in the region of 10–100 mg IB l−1 (Cleuvers, 2003, 2004; Halling-Sørensen et al., 1998; Han et al., 2006; Heckmann et al., 2005). However, there are no published data on the chronic population effects of IB on D. magna or other arthropod species. In one of the more extensive studies investigating the effects of pharmaceuticals on planktonic communities, Richards et al. (2004) revealed that a mixture of IB (0.6 mg l−1 ), fluoxetine (1.0 mg l−1 ) and ciprofloxacin (1.0 mg l−1 ) decreased the diversity of a zooplankton community in a 35day microcosm experiment, concurrently increasing the overall abundance of a few species. Other non-arthropod studies show that IB, acetylsalicylic acid and paracetamol had no effect on the cnidarian Hydra vulgaris at concentrations up to 1 mg l−1 following 7 days exposure (Pascoe et al., 2003). However, 1 mg IB l−1 was sufficient to reduce the growth of duckweed Lemna minor by 25%, with a 7-day EC50 of 4 mg IB l−1 (Pomati et al., 2004). Conversely, growth of the cyanobacteria Synechocystis sp. was stimulated at 10 ␮g IB l−1 (Pomati et al., 2004). At present, there is limited ecotoxicological information available in the literature regarding the potential chronic impact of pharmaceuticals on aquatic organisms and ecosystems (Fent et al., 2006; K¨ummerer, 2004); although the few reported chronic-effect studies indicate that IB probably has very little impact on aquatic environments (Han et al., 2006; Pascoe et al., 2003; Pomati et al., 2004; Pounds et al., submitted for publication; Richards et al., 2004). Here, we report the chronic effects of IB on the life history traits and population dynamics of D. magna following a 14-day exposure. Furthermore, we suggest a potential mecha-

nism of toxicity for this NSAID based on recent genomic studies. 2. Materials and methods 2.1. Test species D. magna (clone type 5-IRCHA), originally obtained from the Water Research Centre, Medmenham, UK, was used for this study. For an outline of the life history of D. magna and full information on culturing conditions see Hooper et al. (2006). 2.2. Test chemical The NSAID IB is widely used in the treatment of rheumatic disorders, pain and fever due to its analgesic, antipyretic and anti-inflammatory properties. IB was obtained as a sodium salt (C13 H17 NaO2 ) from Sigma–Aldrich (CAS no. 31121-93-4: batch no. 64K0892). In principle, IB is a fatty acid and, thus, virtually insoluble at low pH, but fully ionized at pH 7 and above (Hadgraft and Valenta, 2000). In the case of IB sodium (IB-Na), the parent compound dissociates to form C13 H17 O2 − and Na+ at pH 7 and above. Here, we refer to concentrations of the NSAID based on the molecular weight of the pharmacological active ingredient IB. 2.3. Experimental design Concentrations of IB were selected based on preliminary acute and chronic toxicity test results. Initially, a standard acute test method (OECD, 2004) was used to determine the toxicity of IB to our laboratory clone. The 48 h EC50 for immobilisation was estimated as 108 mg IB l−1 (Heckmann et al., 2005). Subsequently, populations were exposed to eight IB concentrations ranging from 10 to 160 mg IB l−1 for 12 days (Heckmann et al., 2005), using an experimental design similar to the one described below. This pilot study revealed that reproduction was markedly reduced above 10 mg IB l−1 , and that survival was unaffected at concentrations up to 40 mg IB l−1 with no survivors above 80 mg l−1 . Accordingly, we chose the concentrations 20, 40 and 80 mg IB l−1 and a dilution water control for the current study. Each treatment was replicated four times and assigned to a randomised block design. During the first 24 h of the test, each replicate consisted of 310 fourth brood neonates (<24 h old) that were held in a glass cylinder (height 13 cm; internal diameter 8.5 cm) with a 200 ␮m mesh covering the bottom. This cylinder was contained within a larger glass vessel (height 22 cm; internal diameter 18.5 cm; thickness 5 mm) (Harzkristall GmbH, Derenburg, Germany) with a clear plastic lid, containing 5 l of reconstituted water (Hooper et al., 2006) with or without the addition of IB-Na. After 24 h exposure, 10 neonates were transferred to the larger vessel, and the cylinder containing the remaining 300 neonates was removed. These daphnids were analysed for differences in gene expression (not reported here). The 10 transferred neonates were observed over the following 13 days to assess the effects of IB on popula-

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tion growth rate (PGR), reproduction, survival, somatic growth and population size structure. The latter two parameters were derived from measurements of individual’s body surface area, captured by digital imaging using the method and settings of Hooper et al. (2006). There was no feeding during the first day (24 h), but, subsequently, all populations were fed daily with equal amounts of green algae Chlorella vulgaris var viridis (equivalent to 0.25 mg carbon per day on day 2; 0.50 mg carbon per day from days 3–9; 0.75 mg carbon per day on days 10 and 11; and 1.00 mg carbon per day on days 12 and 13). The test vessels were kept in a 20 ± 1 ◦ C temperature-controlled room with a light:dark regime of 16:8 h with lighting provided by a 70 W, cool white fluorescent tube, situated 10 cm directly above the test vessels. 2.4. Physicochemical conditions During the test, the water was aerated regularly (2 h per day) via a glass tube (length 20 cm, internal diameter 3 mm) submerged 3 cm below the water surface. Temperature was monitored daily having a mean of 19.5 ± 0.2 ◦ C S.E. (n = 60) throughout the 14-day exposure. Dissolved oxygen (DO), conductivity and pH were monitored on days 2, 4, 8 and 14 (Table 1). Conductivity increased with IB concentration (P < 0.05) following the dissociation of sodium ions from IB-Na. However, the difference in conductivity between the control and 80 mg IB l−1 was less than 5% throughout the experiment. There was a consistent increase in both DO and pH in all treatments until day 8, most likely the result of

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increased photosynthesis in the test medium from the daily addition of C. vulgaris. By day 14, DO and pH levels had decreased in the control and the 20 and 40 mg IB l−1 treatments, but remained elevated at 80 mg IB l−1 where algae had accumulated (Table 1). To quantify IB, 1.5 ml was sampled from each replicate on days 1, 2, 4, 8 and 14, and stored at −20 ◦ C. Subsequently, IB was quantified at 217 nm by UV-spectrophotometry using a Shimadzu UV-1201 (Shimadzu, Europa GmbH, Duisburg, Germany) following the method of Pascoe et al. (2003). All the measured IB concentrations were within ±20% of the nominal concentration during the experiment, except on day 14 where the measured concentration of IB was >25% less than the nominal concentration in one of the treatments (Table 1). This was due to a slight reduction (average 16%) in the measured concentration of the highest treatment on day 14 compared with day 1 (Table 1). 2.5. Data analysis and statistical methods All data, except PGR and pH, were log10 -transformed. Two-way ANOVAs were performed in Minitab® Release 14.1 (Minitab Inc., State College, PA, USA). Results on all data verified that no block effect was present. Differences between treatments for water chemical parameters and biological endpoints were analysed in SPSS 12.0.1 for Windows (SPSS, Chicago, IL, USA) using regression analysis, and one-way ANOVA together with Tukey’s honestly significant difference for post hoc comparisons. Equality of variance was tested

Table 1 Water chemistry and quantification of ibuprofen Day 1

Day 2

Day 4

Day 8

Day 14

Dissolved oxygen (mg l−1 ) Control 20 mg IB l−1 40 mg IB l−1 80 mg IB l−1

– – – –

8.15 ± 0.03a 8.18 ± 0.03a 8.15 ± 0.03a 8.10 ± 0.00a

8.45 ± 0.03a 8.48 ± 0.03a 8.48 ± 0.03a 8.50 ± 0.00a

9.03 ± 0.09a 9.23 ± 0.09ab 9.08 ± 0.02ab 9.43 ± 0.10b

7.83 ± 0.13a 8.18 ± 0.24a 8.08 ± 0.28a 9.48 ± 0.15b

Conductivity (␮S cm−1 ) Control 20 mg IB l−1 40 mg IB l−1 80 mg IB l−1

– – – –

421 ± 0.9a 427 ± 0.8b 431 ± 0.3c 442 ± 0.3d

426 ± 1.2a 431 ± 0.6b 437 ± 0.5c 447 ± 0.9d

423 ± 1.3a 428 ± 0.5b 432 ± 0.8c 444 ± 0.9d

417 ± 1.6a 422 ± 0.5b 427 ± 0.8c 435 ± 0.9d

pH Control 20 mg IB l−1 40 mg IB l−1 80 mg IB l−1

– – – –

7.72 ± 0.01a 7.73 ± 0.00a 7.73 ± 0.01a 7.72 ± 0.01a

7.99 ± 0.02a 8.03 ± 0.02a 8.03 ± 0.02a 8.06 ± 0.01a

8.32 ± 0.06a 8.52 ± 0.05b 8.45 ± 0.01a 8.66 ± 0.02b

7.74 ± 0.06a 8.00 ± 0.16a 7.96 ± 0.18a 8.83 ± 0.04b

Nominal concentration (mg IB l−1 ) Control 20 40 80

Measured concentration (mg IB l−1 ) n.d. n.d. 20.8 ± 0.04 18.5 ± 0.40 41.8 ± 0.29 37.4 ± 0.23 80.5 ± 1.11 72.4 ± 0.73

n.d. 18.0 ± 0.85 37.3 ± 0.98 76.2 ± 1.87

n.d. 18.6 ± 0.84 38.5 ± 0.75 78.0 ± 2.80

n.d. 18.1 ± 0.95 (18.2 ± 0.41) 36.3 ± 1.15 (37.3 ± 0.57) 67.3 ± 4.01 (72.7 ± 1.93)

Values are mean ± S.E. (n = 4); n.d. signifies “none detected”. Values in parantheses represent the time weighted average (TWA) concentrations following 14 days of exposure. Different letters within the same day signify a significant difference (P < 0.05).

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using Levene’s test. For all tests, a significant level of 5% was applied. ECx values for reproduction were estimated from the relationship between number of juveniles and the IB concentration by fitting an exponential decay model: RE = RC e−aIB

(1)

where RE is the reproduction in the exposed treatment, RC the control reproduction (total number of juveniles), a the decay constant and IB represents the tested concentration of IB. PROC NLMIXED (SAS Institute Inc., 2004) was used to fit the data to the model and provide estimates and 95% confidence limits for the ECx values. PGR was based on manually counting of population numbers and estimated as: PGR = loge

(Nt /N0 ) t

(2)

where N0 is the initial number of individuals at time 0 and Nt is the final number of individuals t days later. Positive values of PGR indicate a growing population, PGR = 0 indicates a stable population and negative PGR values indicate a population in decline and headed toward probable extinction (Sibly and Hone, 2002).

3. Results 3.1. Somatic growth Somatic growth, measured as body surface area, was significantly and strongly positively correlated (r131 = 0.855; P < 0.001) with age in the control populations (data not shown). Interestingly, somatic growth of the founder members of the populations increased significantly with increasing IB concentrations from day 8 and onwards (Fig. 1). Initially, the increase was only significant at 20 mg IB l−1 , but on days 12 and 14, a significantly larger body surface area was also evident at 40 and 80 mg IB l−1 , respectively (Fig. 1). On day 14, daphnids from the treated populations were on average 22%–29% larger than control daphnids.

Fig. 1. Somatic growth of founding Daphnia magna measured as body surface area during a 14-day exposure to ibuprofen (n = 23–40, mean ± S.E.). Letters a, b and c signify a significant difference (P < 0.05) between control and 20 mg IB l−1 ; control and 20–40 mg IB l−1 and control and 20–80 mg IB l−1 , respectively.

3.2. Reproduction The most significant effects observed were on reproduction, realised through delayed onset and reduced fecundity (the latter supported by observations of egg abortions). The day of first reproduction was delayed significantly at 40 mg IB l−1 (Table 2), and total reproduction and number of offspring per female was significantly reduced in all the IB treatments with cessation at 80 mg IB l−1 (Table 2). There was a significant and strongly negative concentration-dependent relationship (r15 = −0.985, P < 0.001) between total reproduction and IB (data not shown) with an estimated 14-day EC10 and EC50 of 2.04 mg IB l−1 (95% C.L. 1.62; 2.46) and 13.4 mg IB l−1 (95% C.L. 10.7; 16.2), respectively. However, extrapolations based on these ECx values should be treated with caution, especially EC10 , as both estimates are below the tested concentrations.

Table 2 Reproduction and survival of Daphnia magna following a 14-day exposure to ibuprofen

Time of first reproduction (d) Total reproduction Offspring/female Survival (%)

Control

20 mg IB l−1

40 mg IB l−1

80 mg IB l−1

10.8 ± 0.25a 290 ± 10.9a 31.3 ± 0.74a 92.5 ± 2.50a

10.8 ± 0.25a 127 ± 24.4b 12.7 ± 2.44b 100 ± 0a

12.3 ± 0.25b 28.5 ± 4.91c 28.5 ± 0.49c 100 ± 0a

n.a. 0 ± 0d 0 ± 0d 75.0 ± 6.45b

Values are mean ± S.E. (n = 4). Different letters signify a significant difference (P < 0.05); n.a. signifies “not applicable”.

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3.3. Survival There was no effect of IB on survival at 20 and 40 mg IB l−1 , but significant mortality was apparent at 80 mg IB l−1 (Table 2). Previous results on D. magna indicated that total mortality occurs at and above 80–100 mg IB l−1 after 12 days exposure (Heckmann et al., 2005) demonstrating a threshold-like relationship between IB and survival. 3.4. Population dynamics PGR was significantly reduced in all the IB treatments compared with the control, and followed a significant and strongly negative linear relationship (Fig. 2). PGR was positive at 0, 20 and 40 mg IB l−1 , but at 80 mg IB l−1 populations were in decline due to failure to reproduce and reduced survivorship. Individuals were classified into three size classes, corresponding to neonates, juveniles and adults, to evaluate how population size structure was affected by IB (Fig. 3). Ibuprofen significantly affected all the treated populations, and at 80 mg IB l−1 the populations consisted entirely of adults. The relative percentage of adults in the population increased more than 2-fold at 20 mg IB l−1 and almost 5-fold at 40 mg IB l−1 ; and the percentage of juveniles was more than 14-fold lower at 40 mg IB l−1 compared with control populations (Fig. 3). These results show that IB had a targeted effect on

Fig. 2. Population growth rate (PGR) of Daphnia magna following a 14-day exposure to ibuprofen (n = 4, mean ± S.E.). The line represents a significant concentration-dependent linear relationship (r15 = −0.945, P < 0.001). Different letters signify a significant difference (P < 0.05).

Fig. 3. Distribution of Daphnia magna populations in three size classes following a 14-day exposure to ibuprofen (n = 4, mean ± S.E.). Different letters within size classes signify a significant difference (P < 0.05).

reproduction. The changes to population size structure appeared to be induced by delayed reproduction and reduced fecundity (Table 2), rather than by increased juvenile mortality. 4. Discussion The results from this study indicate that the longterm population consequences of a chronic IB exposure at environmentally realistic concentrations (ng l−1 to ␮g l−1 ) would most likely be of minor importance (see also overview of invertebrate chronic-effect data in Table 3). However, no definite conclusion can be made on the risks of environmentally relevant IB concentrations before potential effects following multi-generational exposure have been assessed. In support of our results, induced egg abortion and reduced PGR has also been reported previously in D. magna exposed chronically to a metabolite (o-hydroxyhippuric at 10 mg l−1 ) of the NSAID acetylsalicylic acid (Marques et al., 2004b); although the parent compound had no impact at the same concentration (Marques et al., 2004a). A recent study on killifish (Oryzias latipes) revealed that, although reproduction was delayed following a 6-week chronic exposure to ␮g l−1 levels of IB, total reproduction of killifish did not differ between treatments (Flippin et al., 2007). Whereas, an in vitro study revealed endocrine disruption at sub-molar IB concentrations decreasing cortisol production by approximately 40% in interrenal cells of rainbow trout Oncorhynchus mykiss; which potentially may impair overall stress response (Gravel and Vijayan, 2006). Thus, clearly more chronic studies need to be conducted on aquatic organisms to assess the ecological risks of IB and NSAIDs in general. In mammals, IB and related NSAIDs are known to interrupt the production of various eicosanoids, mainly by inhibiting the cyclooxygenase (COX) pathway, which

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Table 3 Overview of invertebrate chronic-effect data on ibuprofen Test species

Endpoint

Exposure range (mg IB l−1 )

Effect concentrationa (mg IB l−1 )

Reference

Hydra vulgaris (Cnidarian)

0.01–10

7d NOEC = 10

Pascoe et al. (2003)

Hydra vulgaris (Cnidarian) Planorbis carinatus (Mollusc)

Reproduction (bud formation) Survival Growth

0.01–10 0.32–5.36

Pascoe et al. (2003) Pounds et al. (submitted for publication)

Planorbis carinatus (Mollusc)

Reproduction

0.32–5.36

Planorbis carinatus (Mollusc)

Survival

0.32–5.36

Daphnia magna (Crustacean) Daphnia magna (Crustacean)

Reproduction Reproduction

0.1–80 20–80

Daphnia magna (Crustacean)

Survival

20–80

Daphnia magna (Crustacean)

Population growth rate (PGR)

20–80

7d NOEC = 10 21d NOEC = 1.02 21d LOEC = 2.43 21d NOEC = 2.43 21d LOEC = 5.36 21d NOEC = 5.36 21d LOEC > 5.36 21d NOEC = 20 14d EC10 = 2.04 (1.62; 2.46) 14d EC50 = 13.4 (10.7; 16.2) 14d NOEC = 20 14d LOEC = 80 14d NOEC < 20 14d LOEC = 20

Pounds et al. (submitted for publication) Pounds et al. (submitted for publication) Han et al. (2006) This study

This study This study

a

Values are measured concentrations (with the exception of Han et al., 2006). Numbers in parantheses signify the lower and upper 95% confidence limits.

is one of the three major pathways involved in eicosanoid biosynthesis. All eicosanoids derive from a common precursor, arachidonic acid (an omega-6 fatty acid), which is converted into different eicosanoids through either the COX pathway (e.g. prostaglandins); the lipoxygenase (LOX) pathway (e.g. leukotrienes and lipoxins); or the cytochrome P450 epoxygenase pathway (e.g. epoxyeicosatrienoic acids) (Fig. 4). Eicosanoids act as autocrine or paracrine signallers (also referred to as local hormones) mediating proximal cell responses to stimuli. They differ from hormones in that the compound is usually broken down quickly, and, thus, their effects are normally local (i.e. within the same tissue from which they were synthesised). In mammals, eicosanoids operate as important regulators of inflammation, ion flux and neural and reproductive function, and several studies reveal that they also play vital roles in reproduction, immune response and temperature regulation of insects (reviewed in Rowley et al., 2005; Stanley-Samuelson, 1994). Recently, a mechanism corresponding to the mammalian eicosanoid biosynthesis was proposed in the coral Plexaura homomalla (Valmsen et al., 2001). This indicates that the NSAID mode of action in invertebrates could be similar to that of mammals, which is further supported by evidence of inhibition of eicosanoid generation by NSAIDs in a wide range of invertebrate species (e.g. Rowley et al., 2005). We have recently provided gene expression data, which suggests an IB concentration-

dependent induction of the expression of the D. magna ortholog of Leukotriene B4 12-hydroxydehydrogenase (Heckmann et al., 2006). The translated enzyme is known to metabolise Leukotriene B4 in eicosanoid biosynthesis (Stanley-Samuelson, 1994). Interestingly, Leukotriene B4 has been shown to have an important role in regulating yolk uptake during oogenesis in insects

Fig. 4. Simplified overview of eicosanoid biosynthesis based on the present knowledge from mammalian models. Arachidonic acid originates from different phospholipids (e.g. diacylglycerol), and is then further metabolised through one of the three shown pathways: cyclooxygenase, lipooxygenase and cytochrome P450 epoxygenase.

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(Medeiros et al., 2004). This apparent link between eicosanoids and reproduction may help to explain the effects of IB on reproduction in D. magna observed in this study. But further work is needed to confirm the genomic observations preferably including transcriptomics, proteomics and enzyme activity assays. Moreover, linking the impact of NSAIDs on mammals and invertebrates should be approached with caution till more evidence exists. The IB concentrations tested here (≤80 mg l−1 ) had little effect on survival, demonstrating the low acute toxicity of IB. This is similar to the majority of pharmaceuticals, with the exception of, e.g. anti-depressive drugs (Fent et al., 2006; K¨ummerer, 2004). Throughout the duration of the experiment, Daphnia populations were fed an equal daily ration of algae, and based on the results and observations on somatic growth, there did not seem to be any major differences in energy intake. However, one could argue that the observed increase in somatic growth was an artefact of increased availability of algae due to fewer or no offspring being present amongst IB treated populations. Compared to the other treatments on day 14, the steep increase in somatic growth at 80 mg IB l−1 (Fig. 1) could well be influenced by a greater per capita availability and intake of algae. But individuals exposed to 20 mg IB l−1 had an increased somatic growth compared with the controls on day 8. At this time, there was a similar survival within the two treatments, and no offspring had yet been produced. Consequently, increased somatic growth was probably not caused by an increasing amount of available food/energy, but may be interpreted as a direct effect of IB. The effects on somatic growth and survival, together with a strong negative impact of IB on reproduction support the “Principle of Allocation”, where the amount of energy that an individual can invest in maintenance, growth and reproduction is limited and highly dependent on energy input (Sibly and Calow, 1986). Thus, the quenching of reproduction would liberate energy to be invested in somatic growth given that maintenance and energy input remain constant. In previous acute immobility studies on D. magna, Cleuvers (2003, 2004) has stated that the IB mode of action is acting non-specific by narcosis or baseline toxicity (sensu Verhaar et al., 1992). We have clearly shown that there is a very specific chronic effect of IB on reproduction. But why is such a high IB concentration necessary to cause any effect in Daphnia; and is the mode of action truly non-specific especially in the light of our recent molecular evidence? In mammals, at the molecular level, IB reversibly inhibits the enzymatic activity of COX through competing with its

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substrate, arachidonic acid, for the catalytic sites of COX (McGettigan and Henry, 2000). Flippin et al. (2007) suggest that high effect concentrations in fish may be due to inefficient IB inhibition of teleost COX. Based on evidence from invertebrates, one could hypothesize that the eicosanoid precursor arachidonic acid is very abundant, which would require high IB concentrations to produce any effect. For instance, Weinheimer and Spraggins (reviewed in Rowley et al., 2005) report that up to 8% of the dry mass of the coral P. homomalla are eicosanoids. Such high content of eicosanoids may be extraordinary, but if these concentrations are directly in proportion to their precursor arachidonic acid, and if a similar level is found in Daphnia, it would explain why such high IB concentrations are needed to cause an effect in Daphnia. A high body content of eicosanoids could also imply a high availability of the major enzymes (e.g. COX or LOX) involved in eicosanoid biosynthesis. This is the situation in humans where COX-1 (the only COX isoform present in invertebrates) is constitutively expressed in almost all tissues (e.g. McGettigan and Henry, 2000). Further work is currently underway to assess the global expression profile of Daphnia and attempt to identify the key genes involved in stress responses to IB. We hope this work will shed further light on the effects of IB on eicosanoid biosynthesis and reproduction in Daphnia. As a concluding remark, and as encouraged by other investigators (Iguchi et al., 2006), further experimentation should focus on invertebrate genomic and other molecular toxicology studies, taking care to consider quality control (Pritchard et al., 2001 and references herein). This could reveal similarities with the genomic effects of pharmaceuticals in mammalian systems, which could have major consequences for the future testing of pharmaceuticals by, e.g. reducing the need for vertebrate models such as the use of fish in environmental safety assessments. Additionally, further chronic studies on pharmaceuticals are recommended at the population or community level to increase our understanding of the long-term impact of pharmaceuticals in aquatic environments. Conflict of interest The authors state that they have no competing financial or other interests that could inappropriately influence the current study. Acknowledgements We gratefully acknowledge the financial support of AstraZeneca, Syngenta, NERC (project NER/

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D/S/2002/00413 “The population and molecular stress responses of an ecotoxicology indicator species”) and The Research Endowment Trust Fund of the University of Reading. We would also like to thank Paul Henning Krogh for statistical assistance on ECx modelling, and two anonymous reviewers for their valuable comments on the manuscript.

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