Ecotoxicology and Environmental Safety 117 (2015) 81–88
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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Adverse effects of bisphenol A on water louse (Asellus aquaticus) Maja Plahuta a,n, Tatjana Tišler a, Albin Pintar a, Mihael Jožef Toman b a b
Laboratory for Environmental Sciences and Engineering, National Institute of Chemistry, Hajdrihova 19, SI-1001 Ljubljana, Slovenia University of Ljubljana, Biotechnical Faculty, Department of Biology, Jamnikarjeva 101, SI-1000 Ljubljana, Slovenia
art ic l e i nf o
a b s t r a c t
Article history: Received 22 October 2014 Received in revised form 21 March 2015 Accepted 25 March 2015
Experiments were performed to study the effects of short and long-term exposure to bisphenol A (BPA) on a freshwater crustacean isopod Asellus aquaticus (L.). Two life stages of isopods were exposed to a range of BPA concentrations, from aqueous and two dietary sources, in the form of with BPA spiked conditioned alder leaf (Alnus glutinosa) discs, or spiked formulated sediment, to determine the relative importance of each source of exposure on the uptake of this contaminant. Several lethal and sublethal endpoints were evaluated in this study to measure the potential effects of BPA on A. aquaticus, including mortality, growth and feeding rate inhibition, mobility inhibition, de-pigmentation and molting disturbances. They signify a correlation to BPA levels and a difference in BPA uptake efficiency from different uptake sources. Results of acute exposure to BPA show a greater sensitivity of test systems using juvenile specimens with a 96 h LC50 of 8.6 mg L 1 BPA in water medium and a 96 h LC50 of 13.5 mg L 1 BPA in sediment. In comparison, adult isopods show a 96 h LC50 of 25.1 mg L 1 BPA in water medium and a 96 h LC50 of 65.1 mg L 1 BPA in sediment. Observed endpoints of chronic exposures suggest the alder leave discs to be the most efficient uptake source of BPA, in contrast to uptake from water or heterogeneous sediment. Significant (po0.05) growth inhibition, with a 21 d NOEC of 0.5/2.5 mg L 1 (for juvenile/adult organisms), and feeding rate inhibition, with a 21 d NOEC of 0.5/1.0 mg L 1 (for juvenile/adult organisms), were proven to be the most sensitive toxicity endpoints. An even more sensitive effect turned out to be molting frequency, which was significantly reduced; a 21d NOEC was 1.0 mg L 1 of BPA for adult organisms and an even lower 21d NOEC of 0.05 mg L 1 of BPA for juveniles. The observed endpoints are recorded at very low, non-toxic exposure concentrations, indicating that BPA acts as an endocrine disrupting compound, as well as a toxic substance. We also determined the importance of the direct dietary uptake of the pollutants, significant for juveniles as well as adult animals. & 2015 Elsevier Inc. All rights reserved.
Keywords: Asellus aquaticus Bisphenol A Aqueous uptake Dietary uptake Endocrine disrupting effects
1. Introduction Aquatic ecosystems are daily polluted with man-made toxic organic pollutants, which are biologically active and can impose adverse health effects on wildlife and humans. In the past two decades, there has been an increasing concern over some classes of environmental contaminants known as endocrine-disrupting chemicals (EDCs), which have the ability to interfere with the endocrine system of exposed organisms. Often they mimic or block endogenous hormones (Petrović et al., 2004; Schug et al., 2011). Due to a high production and environmental release of various synthetic industrial chemicals and their by-products, these compounds have resulted in developmental deficits and reproductive impairments in a wide range of aquatic species (Crain n
Corresponding author. E-mail addresses: [email protected] (M. Plahuta), [email protected] (T. Tišler), [email protected] (A. Pintar), [email protected] (M.J. Toman). http://dx.doi.org/10.1016/j.ecoenv.2015.03.031 0147-6513/& 2015 Elsevier Inc. All rights reserved.
et al., 2007; Zou, 2010; Rubin, 2011). In this study, a well-known organic contaminant, used worldwide in the production of polycarbonate plastic and epoxy resins, bisphenol A (BPA, 2,2-bis-(4hydroxyphenyl)-propane; CAS Registry No. 80-05-7), was investigated (Staples et al., 1998). It is generating concerns due to its extensive production, widespread use and consequently frequent presence in various environmental matrices (Vandenberg et al., 2007; Flint et al., 2012). Wildlife (as well as humans) is exposed to BPA from a variety of environmental releases like waste landfills leachate and wastewater treatment plants (Staples et al., 1998). The concentrations of BPA vary greatly depending on location and time of sampling. European surface water analyzes report BPA concentrations ranging from 0.5 up to 21 mg L 1 (Fromme et al., 2002; Belfroid et al., 2002). Levels of BPA in waste landfill leachate can be even higher, 5.4 mg L 1 (Yamada et al., 1999) and even up to 17.2 mg L 1 of BPA (Yamamoto et al., 2001). From the obtained values we can see that such concentrations could cause adverse effects on Asellus aquaticus in long-term exposures. BPA has been
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the subject of considerable aquatic toxicity testing in recent years. The majority of available effect data on aquatic organisms has been obtained from studies using fish (Lahnsteiner et al., 2005; Staples et al., 2011), some aquatic invertebrates (Hill et al., 2002; Pascoe et al., 2002; Watts et al., 2003) and macrophyte plants (Mihaich et al., 2009). De Kermoysan et al. (2013) also studied the long-term effects of BPA on macrophytes, macro-invertebrates and fish in an aquatic mesocosm. BPA also acts as an estrogenically active endocrine disrupting compound (Okada et al., 2008), and as such impacts wildlife populations by interfering with some aspects of endocrine-mediated processes, causing impairment of developmental, reproductive and other hormonally mediated processes (Rodriguez et al., 2007). In the aquatic environment organisms can uptake lipophilic toxicants from contaminated water via gills or the skin. Dietary sources also play an important role in contaminants uptake through ingestion and the gastrointestinal tract, especially in sediment dwelling organisms. They are exposed directly to contaminants by ingesting organic matter from the sediment and uptaking the sediment-associated contaminants from interstitial and overlying water. As uptake from dietary sources can explain 100% of total residue, it is possible that the source of BPA, to which an organism is exposed, plays an important role in effects occurrence in animals (Landrum and Robbins, 1990; Peeters et al., 2000). The water louse, Asellus aquaticus (Linnaeus, 1758), was chosen as a model species in this study for several reasons. Firstly, there are no previous reports of the effects of BPA to A. aquaticus. It is native in Slovenia and abundantly present in surface waters all throughout western Eurasia (Graca et al., 1994). It is in continuous contact with water and the sediment (McCahon et al., 1990), easy to culture in the laboratory, tolerant to varying physicochemical characteristics of sediment, has negligible biotransformation, intermediate sensitivity to contaminants and shows specific effects of toxicity and endocrine disruption (van Hattum et al., 1989). It is a shredder that feeds on detritus associated with fungi, bacteria and periphyton (Graca et al., 1993). The dietary preference of A. aquaticus toward naturally fungally colonized leaf material was presented by Bloor (2011). A. aquaticus also presents an important food source for predatory invertebrates, fish, and waterfowl (Van Hattum et al., 1989) and thus plays an important role in the benthic-pelagic coupling of the food chain transfer of pollutants (Peeters et al., 2000). Their small size and robust nature make them ideally suited for application in toxicity tests. Furthermore, early life stages of the species offer all the key attributes of a complete in vivo system. The objective of this study was to investigate the effects of BPA uptake from aqueous or dietary sources (inoculated alder (Alnus glutinosa) leaf discs or formulated sediment) on survival, growth, feeding rate, mobility, pigmentation and molting of two different life stages of the sediment dwelling freshwater isopod Asellus aquaticus (L.) in a laboratory experiment, which will enable us to better characterize the toxic and endocrine disrupting effects of BPA.
2. Materials and methods Laboratory bred aquatic isopods A. aquaticus were exposed to BPA from aqueous and dietary sources (in the form of spiked alder leaf discs and spiked formulated sediment) to determine the importance of each source of exposure on the uptake of this pollutant. For the purposes of determining the effects of BPA on different life stages of A. aquaticus, both adult and juvenile life stages were exposed to BPA for 4 d (acute exposure) as well as 3 weeks (chronic exposure). Control groups consisted of 12 specimens per test, exposed to unpolluted water, food or sediment.
2.1. Test organisms and culture procedures Specimens of A. aquaticus were collected from an unpolluted stream in Črna vas, Ljubljana (Slovenia, 45° 59′ N, 14° 28′ E) and laboratory cultured as a population giving rise to new generations used in tests. As such, animals used in testing, had no prior history of contamination. Isopods were kept in four liter glass aquaria tanks containing reconstructed water (M4 medium, ISO 10706, 2000) with pH 7.57 0.5, dissolved oxygen (4 60% of air saturation value), at 20 72 °C, with a 16:8 h light/dark photoperiod, illumination around 500 lx, and continuously purged with purified air at a flow rate of 50 mL min 1. Organisms were weekly fed on conditioned alder leaves (Alnus glutinosa), colonized by micro-organisms to maintain a healthy population. 2.2. Chemicals and preparation for test conduction The chemical used in the tests was Bisphenol A (BPA) (SigmaAldrich, Germany, purity 99%) dissolved in ultrapure water (18.2 MΩ cm 1, ELGA LabWater, UK) as a stock solution of 100 mg L 1. For testing purposes salts were added to the solution, forming reconstructed water (Elendt M4 medium, ISO 10706, 2000) or dilution water (ISO 6341, 2013). Initial BPA stock solution was diluted for purposes of acute and chronic exposure to BPA from water, leaves and sediment. 24 h prior to each test, randomly chosen juvenile and adult laboratory bred males and females were placed in 250 mL lidded glass containers, containing deionized water in oxygen depleting conditions without nutritional support. 12 adults or juveniles in intermolt stage were then placed in 6 well test plates (Ø 33.78 22 mm) (Techno Plastic Products AG, Trasadingen, Switzerland), arranged one animal per test chamber. 2.3. Experimental design Each experiment was performed in parallels, and replicated three times. Exposure was conducted in 6 well test plates with Ø 35 mm, 10 mL wells. In tests, juvenile (size range 1.0–1.4 mm) and adult (size range 4–5 mm) specimens of A. aquaticus were exposed at 20 71 °C. Acute exposure tests were performed to derive 96 h LC50 values, which were further used to set up the chronic exposures. Organisms were exposed to test media for 96 h in static exposure system without aeration. Juvenile isopods were acutely exposed to aqueous and sediment BPA concentration range from 2.5 to 20 mg L 1. Adult organisms were acutely exposed to the range from 5 to 70 mg L 1 BPA. Chronic exposure tests were performed for 21 d in a semi-static exposure system with medium renewal every three days. Organisms were exposed to aqueous, food and sediment bound BPA in concentrations ranging from 0.05 to 5.0 mg L 1 BPA for juveniles and from 0.5 to 10.0 mg L 1 BPA for adults. The 21 d NOEC and LOEC for each effect were calculated (P o0.05, ANOVA). Test solutions were taken when water renewed, for measuring the concentrations of test chemical by means of HPLC. At the end of the experiment pH value and dissolved oxygen concentration were measured, to assure the effects only be attributed to the toxic chemicals. Survival of the control group was above 80% after acute and chronic exposure to unpolluted media. In tests determining effects of aqueous BPA on exposed isopods, each test chamber was filled with 10 mL of test medium. In long term exposure tests, one Ø 10 mm unpolluted alder leaf disk was added to each chamber, acting as food as well as shelter. As no specific sediment preferences were recognized in the literature, for purposes of short and long term sediment exposure tests with Asellus aquaticus, formulated sediment was prepared according to the OECD 218 (2004) for Chironomid sediment toxicity test, with a few modifications, as follows. The sediment
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consists of 94% of natural quartz sand (Termit d.d., Slovenia) with granulation r1 mm, 2% of kaolin clay ( Z30% kaolinite, Merck, Darmstadt, Germany), CaCO3 (Merck, Darmstadt, Germany), and 4% organic content in the form of grounded unpolluted conditioned Alder leafs (Alnus glutinosa). Prior to grinding, leafs were dried for 24 h at 207 1 °C. Deionized water was added to the sediment mixture, in the ratio 1:4 of mixture/water. Calcium carbonate was added to adjust the pH of the final sediment mixture to 7.0 70.5. The prepared mixture was filtrated using a 1.2 mm CN membrane filter (Sartorius AG, Goettingen, Germany), and the solid residue was re-suspended in BPA solutions of different concentrations, in the ratio 1:4 of sediment/solution. Sediment testing included a control (unpolluted) sediment. Exposure test, conducted in test plates, contained 2 g of dry weight of the sediment and 5 mL of dilution water per well. Overlying water was renewed every three days, the sediment was not renewed. For feeding and testing purposes autumn-shed alder leaves (Alnus glutinosa) were collected at an Alder forest and used as a food source for isopods. The leaves were dried at 40 °C for 24 h, and stored in a dry place, until used. The dry leaves were soaked in 500 mL of river water (collected from the Asellus aquaticus sampling place) for at least 10 d in order to be colonized by microorganisms (Gollady et al., 1983). For testing purposes the conditioned leaf material was cut into Ø 10 mm discs. The leaf midrib was avoided when cutting out discs. Spiking was performed by soaking the discs in BPA solutions for 96 h, with daily solution renewal, while constantly stirring on a magnetic stirrer. Before solution renewal, samples were taken to be analyzed using HPLC in order to determine the amount of BPA spiked on leaf discs. Prepared discs were air dried for 24 h, and weighed before placed each into one testing chambers containing one A. aquaticus and 10 mL of reconstructed water. Leaf consumption was daily visually monitored and after 21 d, the discs were removed, air dried for 24 h, and reweighted. The weight loss of the leaf discs, not caused by A. aquaticus activity, was determined by exposing the discs to test medium for 21 d in the absence of isopods.
isopods that assumed a lateral position, with extended legs, with only pleopod and/or antennal movement were considered immobile. The animals did not move even after gentile mechanic agitation. To determine the de-pigmentation, the exposed animals were divided in two groups, either a specimen was considered pigmented or de-pigmented, in which case less than approximately 20% of body surface, determined by visual inspection, contained black pigment. The pigmentation results of tested BPA concentrations were compared to a control group, which was not exposed to BPA.
2.4. Test endpoints
3.1. Acute exposure effects
The primary short exposure endpoint was survival. In addition mobility, growth rate, feeding rate, de-pigmentation, and ecdysis were observed in long term studies. Survival, mobility and molting of test organisms were quantitatively determined through visual observations daily, in all tests. Death was determined by absence of respiratory movement. The dead animals were regularly removed from test chambers. Total body length was measure from the tip of the head to the end of the pleotelson. The animals were measured individually prior the test, under a stereo microscope (Nikon SMZ1000, resolution 80) and NIS-Element D Microscope Imagining Software (Nikon, Tokyo, Japan). After the experiments were stopped the live animals were re-measured, and growth rate was calculated by subtracting the body length at the beginning of the test from the body length at the end of the test. Inhibition of growth rate was determined as a decrease in growth rate per day in test groups, in comparison to control group. The feeding rate was calculated by subtracting the end disc weight from the weight at the beginning of the test. The consumed weight was then divided by the number of days the disc was exposed, to acquire the average leaf consumption per day. Inhibition of feeding rate was calculated as a difference in relation to the control group, which was not exposed to BPA. The molting disturbances, including effects on molting frequency, and the time span until the first ecdysis, were determined by daily visual inspection. The exuviae (molted exoskeletons) were removed from test chambers on daily basis. Immobility and de-pigmentation were also visually determined at the end of the test. According to Green et al. (1988),
Acute exposure of A. aquaticus to aqueous and sediment bound selected organic pollutant and endocrine disruptor BPA, demonstrated significant effects on the mortality rate (Po 0.05, ANOVA) of exposed isopods. Median lethal concentrations (96 h LC50) for aqueous BPA were obtained at 9.5 mg L 1 for juveniles and 25.1 mg L 1 for adult organisms (Fig. 1). Acute exposure to BPA bound to formulated sediment, showed a less significant effect on mortality in juvenile as well as adult organisms. The median lethal concentration (96 h EC50) for juvenile organisms was 4.7 mg L 1 and 65.1 mg L 1 for adult, regarding sediment bound BPA. Mortality is presented by means of dose-mortality curve, together with standard deviation and LC50 values (Fig. 1). The juvenile and adult LC50 differed significantly (po 0.05), where juveniles were more sensitive to BPA in comparison to adults. Comparison of effects of the test medium also demonstrated a significantly different LC50 between uptake from water and from sediment. Comparing the above results to data obtained from the literature for other species, or previously reported (Plahuta et al., 2014), it is obvious that the sensitivity of the test system using juvenile A. aquaticus is higher than the one using Daphnia magna (24 h EC50 was 12.5 mg L 1), but still lower than the test system using Gammarus pulex (48 h LC50 was 5.6 mg L 1 BPA) (Watts et al., 2001).
2.5. HPLC analysis The concentrations of BPA in water samples were analyzed by means of high performance liquid chromatography (HPLC; Agilent Technologies, USA) equipped with a quaternary solvent pump, autosampler and UV detection (λ ¼ 210 nm) at a constant temperature of 30 °C and a limit of detection (LOD) at 0.2 mg L 1. 2.6. Statistical analysis LC50 values for tests with Asellus aquaticus were determined using the probit analysis based on results from acute exposure tests and responses of organisms under various concentrations of BPA. The normality of our results was analyzed with the Shapiro– Wilk test. The test deducted that the results are normally distributed. Differences between groups were tested using one-way ANOVA, on results from chronic exposure tests, with BPA concentrations as the explanatory variable. The critical level of significance was po 0.05. T-test was used as a post hoc test, to test between-concentration differences (po 0.05). Statistical analyzes were performed using Origin 8.1 SR3 data analysis and graphing software (OriginLab Corporation, MA, USA).
3. Results and discussion
3.2. Chronic sublethal exposure effects Chronic exposure to the selected organic pollutant BPA
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100
80 Immobilization (%)
Mortality after 96 h (%)
80
100
Uptake from water / A Uptake from sediment / A Uptake from sediment / J Uptake frome water /J
60
40
60
Uptake from water / A Uptake from food / A Uptake from sediment / A Uptake from water / J Uptake from food / J Uptake from sediment / J
40
20 20
0 0.1
0 0.1
1
10
1
10
BPA concentration in media (mg L )
100
BPA concentration in media (mg L )
demonstrated visible effects, which were related to BPA concentration levels, and the types of uptake routes (water, leaf or sediment). An experiment was conducted where desorption of BPA from the sediment, and spiked alder leaf discs (Alnus glutinosa), into the overlaying water was observed. It was concluded that desorption from the sediment was minimal and below detection limits of our chemical analysis, whereas the maximum desorption of BPA from spiked alder leaf discs was 12%.
100
80 Pigmentation (%)
Fig. 1. The percentage of adult and juvenile A. aquaticus mortalities in acute exposure tests to aqueous BPA, or from sediment bound BPA, is presented. The results are the means of three repeated tests with the standard deviation between tests indicated by the error bars on the graph. The 96 h LC50 values are indicated by the dotted lines. A and J in the legends refer to adult or juvenile specimens, exposed in the experiment.
60
40
20
0 0.1
Uptake from water / A Uptake from food / A Uptake from sediment / A Uptake from water / J Uptake from food / J Uptake from sediment / J
1
10
BPA concentration in media (mg L ) 3.2.1. Immobilization The impact of chronic exposure to BPA from water, leaf discs and formulated sediment, on locomotion of juvenile and adult A. aquaticus, are presented as concentration dependent response curves in Fig. 2. Exposed animals showed a range of sensitivity to BPA depending on the age of animals and the uptake route of BPA. Highest statistically significant sensitivity in comparison to control group, was seen with juvenile isopods exposed to BPA spiked on alder leaf discs, with a 21 d NOEC for immobilization of 1.5 mg L 1 of BPA, a 21 d LOEC for immobilization of 2.0 mg L 1 BPA, and 21 d EC50 of 2.15 ( 70.16) mg L 1 BPA (ANOVA, F ¼183.1, df ¼5, p o0.05; t-Test, p o0.05) (Table 1). Similar results were drawn from adult exposure to BPA with a 21 d LOEC for immobilization of 5.0 mg L 1, and a 21 d NOEC of 2.0 mg L 1, and a 21 d EC50 of 5.33 (70.33) mg L 1 BPA (ANOVA, F¼38.6, df ¼4, p o0.05; t-Test, p o0.05) after dietary exposure to BPA from alder leaf discs. Aqueous exposure to BPA caused a 21 d NOEC/LOEC of 2.0/2.5 mg L 1 of BPA, and a 21 d EC50 of 2.35(70.11) mg L 1 BPA (ANOVA, F¼110.7, df ¼ 5, p o0.05; t-Test, po 0.05) for juveniles, and 5/10 mg L 1 of BPA and a 21 d EC50 of 4.42 ( 70.58) mg L 1 BPA (ANOVA, F¼ 114.4, df ¼4, p o0.05; t-Test, p o0.05) for adults. Least effects were seen when the animals were exposed to artificial sediment spiked with BPA, with a 21 d NOEC/LOEC of 5/10 mg L 1 BPA for both life stages and a 21 d EC50 of 5.10 (70.44) mg L 1 BPA (ANOVA, F¼92.6, df ¼6, p o0.05; t-Test, p o0.05) for juveniles and a 21 d EC50 of 5.64 ( 70.82) mg L 1 BPA (ANOVA, F ¼20.9, df ¼4, p o0.05; t-Test, p o0.05) for adults (Table 1). Immobilization symptoms are characteristic for A. aquaticus exposure to phenol (McCahon et al., 1990; Green et al., 1988). After longer exposures the animals ceased to move, but the pleopods
Fig. 2. (A) presents the percentage of immobilized animals after a chronic exposure to BPA. (B) presents the percentage of pigmented specimens after chronic exposure to BPA. The results are presented in the concentration response curves as the means of three separate tests with the standard deviation between tests indicated by the error bars on the graph. Hill model was used to perform the curve fitting. A and J in the legends refer to adult or juvenile specimens, exposed in the experiment.
still beat, where the animals were paralyzed, and subsequently died (Green et al., 1988). 3.2.2. Pigmentation Statistically significant reduction of body pigmentation was observed after chronic exposure to BPA. The results are presented in concentration dependent response curves in Fig. 2. The 21 d NOEC/LOEC values are presented in Table 1. The observed effect was most pronounced in juvenile organisms exposed to alder leaf discs spiked with BPA, with 21 d NOEC/LOEC of 1.0/2.0 mg L 1 BPA and 21 d EC50 of 1.72 ( 70.64) mg L 1 BPA (ANOVA, F¼ 180.2, df ¼6, po 0.05; t-Test, p o0.05). The 21 d NOEC/LOEC for adults exposed to BPA from alder leaves was 2.0/5.0 mg L 1, while 21 d EC50 of 1.83 (7 0.51) mg L 1 BPA (ANOVA, F¼262.2, df ¼ 5, po 0.05; t-Test, p o0.05). Exposure to aqueous and sediment bound BPA caused less significant effects on pigmentation with juveniles, with 21 d NOEC/LOEC of 2.0/2.5 mg L 1 BPA and 2.5/5.0 mg L 1 BPA (ANOVA, F¼95.8, df ¼5, p o0.05; t-Test, po 0.05) respectively. 21 d EC50 for aqueous uptake of juveniles was 2.24 ( 70.52) mg L 1 BPA, and 21 d EC50 for sediment uptake of juveniles was 2.21 ( 70.81) mg L 1 BPA. Less effect was observed also with adults exposed to aqueous as well as for those exposed to sediment bound BPA, with 21 d NOEC/LOEC of
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Table 1 Overview of the 21 d NOEC/LOEC values of the chronic exposure of juvenile and adult A. aquaticus to BPA from water, food or sediment. Age
Immobilization Pigmentation Molting Growth Feeding
Water
Juvenile Adult Juvenile Adult Juvenile Adult Juvenile Adult Juvenile Adult
Food
21 d NOEC (mg L 1)
21 d LOEC (mg L 1)
21 d NOEC (mg L 1)
21 d LOEC (mg L 1)
21 d NOEC (mg L 1)
21 d LOEC (mg L 1)
2.0 5.0 2.0 5.0 0.1 2.0 2.5 5.0 2.0 2.0
2.5 10.0 2.5 10.0 0.5 5.0 5.0 10.0 2.5 5.0
1.5 2.0 1.0 2.0 0.05 1.0 0.5 2.0 0.5 1.0
2.0 5.0 2.0 5.0 0.1 2.0 1.0 5.0 1.0 2.0
5.0 5.0 2.5 5.0 / / 2.5 5.0 / /
10.0 10.0 5.0 10.0 / / 5.0 10.0 / /
5.0/10.0 mg L 1 BPA (ANOVA, F¼104.9, df ¼ 4, po 0.05; t-Test, p o0.05), and 21 d EC50 of 3.32 ( 74.03) mg L 1 of aqueous BPA, and 21 d EC50 of 3.63 (74.23) mg L 1 of sediment bound BPA. The results of body de-pigmentation intensity shows that dietary exposure to BPA from alder leaf discs presents the most efficient BPA uptake route among tested. The color changes in crabs are controlled hormonally
*
80
(Fingerman et al., 1981; Liu et al., 2012), and the dispersion of the black pigment is controlled by BPDH (black pigment dispersing hormone), produced principally in the eyestalk ganglia. It was reported that EDC as PCBs caused sublethal effects, like inhibited black pigment dispersion in the crustaceans. Reportably it inhibited the release of BPDH (Staub and Fingerman, 1984a). Naphthalene also inhibits the black pigment dispersion by inhibiting
Week3 Week2 Week1
100
Molting (%)
Sediment
Week3 Week2 Week1
100
80
60
*
*
40
20
*
*
*
*
Molting (%)
Effect
*
60
* 40
*
2.5
5
20
0
0 0
0.05
0.1
0.5
1
2
2.5
5
0
0.05
BPA concentration in media (mg L )
*
80
0.1
0.5
1
2
BPA concentration in media (mg L )
100
*
Week3 Week2 Week1
Week3 Week2 Week1
100
80
*
Molting (%)
* Molting (%)
*
60
40
*
60
* 40
* 20
20
0
0 0
0.5
1
2
5
BPA concentration in media (mg L )
10
15
0
0.5
1
2
5
10
15
BPA concentration in media (mg L )
Fig. 3. Inhibition in molting frequency and distribution of molts per week after a chronic exposure to BPA. (A) Results of juvenile A. aquaticus after chronic exposure to aquatic BPA. (B) Results of juvenile A. aquaticus after chronic exposure to BPA bound on alder leaf discs. (C) Results of adult A. aquaticus after chronic exposure to aquatic BPA. (D) Results of adult A. aquaticus after chronic exposure to BPA bound on alder leaf discs. The results are expressed as mean 7S.D. Asterisks indicate significant differences (po 0.05) in relation to the control group.
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the release of BPDH (Staub and Fingerman, 1984b). Therefore it is possible that BPA as an endocrine disrupting compound also disrupts the dispersion of the black pigment, which is expressed as body de-pigmentation of A. aquaticus. 3.2.3. Molting Chronic exposure to the selected endocrine disrupting compound caused concentration dependent molting disturbances. Observed were significant inhibitions in molting frequency as well as an increase of the intermolt period (the time span between two ecdysis) (Fig. 3). Significantly reduced molting frequency (21 d LOEC) was observed at 5.0 mg L 1 (ANOVA, F ¼6.3, df ¼7, p o0.05; t-Test, po 0.05) of aqueous BPA for adult organisms and 0.5 mg L 1 (ANOVA, F¼8.2, df ¼7, p o0.05; t-Test, p o0.05) of aqueous BPA for juveniles. 21 d NOEC was observed at 2.0 mg L 1 of aqueous BPA for adults, and 0.1 mg L 1 of aqueous BPA for juveniles (Table 1). Dietary exposure to spiked alder leaf discs caused significantly delayed molting, and reduced molting frequency (21 d LOEC) at a concentration as low as 2.0 mg L 1 of BPA (ANOVA, F¼ 4.6, df ¼7, p o0.05; t-Test, p o0.05) for adult organisms and as low as 0.1 mg L 1 of BPA (ANOVA, F¼7.8, df ¼ 7, p o0.05; tTest, p o0.05) for juveniles. No observed effects (NOEC) were
documented at 1.0 mg L 1 (ANOVA, F¼9.3, df ¼7, p o0.05; t-Test, po 0.05) of aqueous BPA for adults, and 0.05 mg L 1 (ANOVA, F¼5.6, df ¼7, p o0.05; t-Test, po 0.05) of aqueous BPA for juveniles (Table 1). Increasing BPA concentration also seemed to drastically delay the molting. When juveniles were exposed to 2.5 mg L 1 of BPA spiked on alder leaf discs and 5.0 mg L 1 of aqueous BPA, only 18.2 and 27.3% of moltings were observed in the first week, respectively. The highest tested concentration, 10.0 mg L 1 of BPA spiked on alder leaf discs, caused a significant molting delay in adults with only 16.7% of molts being observed in the first week and another 16% of moltings in the second week. No further ecdyses were observed in the next two weeks. At these highest tested BPA concentrations other toxic effects were observed. We can attribute the lack of molting in highest tested BPA concentrations to toxicity in addition to endocrine disruption. Molting in crustaceans is under immediate control of steroid molt promoting hormones called ecdysteroids (Chang, 1993; DeFur et al., Tattersfield; Lachaise et al., 1993). Observed effects of BPA on A. aquaticus molting in low, sublethal concentrations, may indicate that BPA acts as an ecdysteroid antagonist. It is possible that a certain interaction exists between the steroid molting hormones and estrogen-mimicking BPA that would allow it to
0.05
0.16
Control Uptake from water Uptake from food
Control Uptake from water Uptake from food
0.14
0.12
0.03
*
0.02
*
*
* *
0.01
*
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Fig. 4. Box plot graphs show the effects of chronic exposures on exposed specimens. (A) Presents the feeding rate inhibition of juvenile A. aquaticus in relation to the BPA concentration in water or food; n¼ 18. (B) Presents the feeding rate inhibition of adult A. aquaticus in relation to the BPA concentration in water or food; n¼ 18. (C) Presents the body growth rate inhibition of juvenile A. aquaticus in relation to the BPA concentration in water, food or sediment; n¼ 18. (D) Presents the body growth rate inhibition of adult A. aquaticus in relation to the BPA concentration in water, food or sediment; n¼ 18. Asterisks indicate significant differences (ANOVA, p o 0.05) in relation to the control group. The ends of the box are the upper and lower quartiles, the median is marked by a vertical line inside the box. The square inside the box represents the mean of the data, and the dotted line presents the average growth rate of the control group. The whiskers present the highest and lowest observations.
M. Plahuta et al. / Ecotoxicology and Environmental Safety 117 (2015) 81–88
compete with ecdysone at the ecdysteroid receptor, by influencing its release, or retaining its transformation into active 20-hydroxyecdosone. Because of structural similarities Molting inhibition in crustaceans has also been observed in Daphnia magna exposed to endocrine disrupting compounds (Zou and Fingerman, 1997a; Zou and Fingerman, 1997b). A delay in molting of Chironomus riparius after exposure to BPA was also observed by Watts et al. (2003), as a result of estrogenic activity of BPA. 3.2.4. Growth rate inhibition The present study showed a concentration dependent reduction in body length increase of A. aquaticus when exposed to BPA (Fig. 4). The test system using juvenile specimens showed a higher sensitivity to BPA regardless the uptake route in comparison to adults, although both used test systems showed a decrease in growth rate when exposed to with BPA spiked alder leaf discs. A statistically significant 42.7% reduction was observed in growth of juveniles with 21 d NOEC/LOEC of 2.5/5.0 mg L 1 of aqueous BPA (ANOVA, F¼3.8, df ¼ 5, po 0.05; t-Test, p o0.05), and a 43.8% growth reduction in adults with 21 d NOEC/LOEC of 5.0/10.0 mg L 1 of aqueous BPA (ANOVA, F¼ 4.2, df ¼4, p o0.05; tTest, p o0.05). Exposure to contaminated alder leaf discs resulted in a further reduction in growth with a significant 38.8% reduction in growth of juveniles and a 21 d NOEC/LOEC of 0.5/1.0 mg L 1 BPA and 21 d EC50 of 2.56 (7 0.29) mg L 1 BPA (ANOVA, F ¼4.3, df ¼5, p o0.05; t-Test, po 0.05), and 66.8% growth reduction in adults at a 21 d NOEC/LOEC of 2.5/5.0 mg L 1 BPA (2.5 mg L 1 BPA cause a 28.7% non significant growth rate reduction in adults) (ANOVA, F¼20.5, df ¼4, po 0.05; t-Test, p o0.05). Exposure to sediment bound BPA caused growth reduction effects comparable to that of aqueous BPA exposure. These results confirm that the dietary route of BPA uptake, from direct feeding with alder leaf discs spiked with BPA, presents the most efficient uptake route of BPA. However, both utilized test systems showed a prominent decrease in growth rate when exposed to BPA adsorbed to leaves. It encompasses an up to 70% growth reduction with juvenile A. aquaticus exposed to 5 mg L 1 BPA in leaves and an up to 81% growth reduction in adult organisms exposed to 10 mg L 1 BPA in leaves. BPA dissolved in water and adsorbed on the sediment caused a lesser amount of effect, which was found to be statistically significant only at highest tested concentrations. The growth results suggest that feeding with contaminated organic matter presents a dominant uptake rout of BPA for A. aquaticus. As in crustacean significant growth can only occur through periodic molting, disruption of molting, as well as inhibited feeding, will result in reduced growth (DeFur et al., Tattersfield; Toda et al., 1987). Another reason for decreased body growth after exposure to contaminants presents a reduction in available energy for growth (Van Brummelen and Stuijfzand, 1993). The organisms exposed to toxicants have to allocate energy to resist the toxicant by avoidance, exclusion, removal, or biochemical complexation (Donker, 1992). 3.2.5. Feeding rate inhibition The feeding results confirm what growth results showed, that the dietary uptake route presents an important route of contaminants uptake into aquatic organisms (Fig. 4). Feeding was significantly reduced in comparison to the control group, with a 21 d NOEC/LOEC of 2.0/2.5 mg L 1 of BPA and 21 d EC50 of 1.14 (7 0.83) mg L 1 BPA (ANOVA, F ¼3.7, df ¼5, p o0.05; t-Test, p o0.05) for juveniles and 21 d NOEC/LOEC of 2.0/5.0 mg L 1 of BPA and (ANOVA, F¼9.5, df ¼ 4, p o0.05; t-Test, p o0.05) for adults, exposed to aquatic BPA. Even more sensitive was the feeding rate with a dietary uptake of 21 d NOEC/LOEC equal to 0.5/1.0 mg L 1 of BPA and 21 d EC50 of 0.83 ( 70.97) mg L 1 BPA (ANOVA, F¼ 5.4, df ¼5, p o0.05; t-Test, p o0.05) for juveniles, and
87
a 21 d NOEC/LOEC of 1.0/2.0 mg L 1 of BPA and 21 d EC50 of 1.3 (71.03) mg L 1 BPA (ANOVA, F¼25.7, df ¼4, po 0.05; t-Test, po 0.05) for adults. The obtained results suggest that the dietary uptake of BPA directly from contaminated food presents a more efficient uptake route than respiratory uptake of aquatic BPA through gills, as BPA is a liphophillic compound, with a log Kow of 3.40, and is expected to adsorb substantially to organic compounds of the sediment (Staples et al., 1998). In contrast to these findings, Felten et al. (2008) concluded the uptake through the gills of being the primary exposure route of aquatic species to contaminants. They found out that dietary uptake of hydrophobic xenobiotics increases in importance, directly with log Kow. It is also said that the transition between aqueous and dietary uptake is the main contributor to lipophillic chemical uptake at log Kow between 3.0 and 6.0, with predominant uptake route of dietary uptake at higher log Kow. Qiao et al. (2000) observed an aqueous uptake of 98% for a chemical with log Kow value of 5.05, and only 15% for a chemical with a log Kow value of 7.55. The present study opposes the findings of Opperhuizen (1991), who predicted that the gills and gastrointestinal tract were of equal importance for the uptake of hydrophobic chemicals. However, studies into the different uptake routes of contaminants show that water (Van Hattum et al., 1989), sediment and food (Peeters et al., 2000) are important.
4. Conclusions The present study has shown that bisphenol A causes lethal as well as sublethal toxic and endocrine disrupting effects in adult and juvenile Asellus aquaticus at acute and chronic exposures. We examined the relative importance of aqueous and dietary uptake of BPA, by comparing the intensity of observed exposure effects. The intensity of observed effects was considerably higher in animals exposed to with BPA spiked alder leaf discs, than when exposed to BPA dissolved in water or spiked on formulated sediment, which can be explained as food being a primary uptake route of BPA. Although expected, the sediment did not present a very efficient uptake source of BPA, causing subsequently an effect comparable to aquatic exposure. Of all observed effects, growth rate inhibition, feeding rate inhibition and molting inhibition were found to be the most sensitive exposure endpoints, at which statistically significant effects are expressed at lowest BPA concentrations, thus providing a realistic representation of possible endocrine-mediated effects in addition to toxic effects. The disruptions in molting were the most sensitive of all endpoints, recorded at the lowest tested concentrations, with a 21 d LOEC of 0.1 mg L 1 for juvenile A. aquaticus (21 d NOEC of 0.05 mg L 1). Other observed effects, including mortality, immobility and de-pigmentation are observed at higher concentrations, as a result of both endocrine disruption as well as general toxicity to the whole organism. The test system, using A. aquaticus, proved to be a rather useful indicating system for determining the presence and effects of the toxic as well as endocrine disrupting contaminant BPA, considering different uptake routes at low, environmentally relevant concentrations.
Acknowledgements This article is a result of doctoral research, in part financed by the European Union, European Social Fund and the Republic of Slovenia, Ministry for Education, Science and Sport in the framework of the Operational program for human resources development for the period 2007–2013. The authors gratefully
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acknowledge the financial support of the Ministry of Education, Science and Sport of the Republic of Slovenia through Research program No. P2-0150.
References Belfroid, A., van Velzen, M., van der Horst, B., Vethaak, D., 2002. Occurrence of bisphenol A in surface water and uptake in fish: evaluation of field measurements. Chemosphere 49, 97–103. Bloor, M.C., 2011. Dietary preference of Gammarus pulex and Asellus aquaticus during a laboratory breeding programme for ecotoxicological studies. Int. J. Zool. . http://dx.doi.org/10.1155/2011/294394, Article ID 294394, 5 pages Chang, E.S., 1993. Comparative endocrinology of molting and reproduction: insects and crustaceans. Annu. Rev. Entomol. 38, 161–180. Crain, D.A., Eriksen, M., Iguchi, T., Jobling, S., Laufer, H., LeBlanc, G.A., Guillette, L.J., 2007. An ecological assessment of bisphenol A: evidence from comparative biology. Reprod. Toxicol. 24, 225–239. DeFur P.L., Crane M., Ingersoll C., Tattersfield L. Endocrine disruption in invertebrates: Endocrinology, testing and assessment. Pensacola FL: SETAC Press. De Kermoysan, G., Joachim, S., Patrick, B., Lonjaret, M., Tebby, C., Lesaulnier, F., Lestremau, F., Chatellier, C., Akrour, Z., Pheron, E., Porcher, J.M., Pery, A.R.R., Beaudouin, R., 2013. Effects of bisphenol A on different trophic levels in a lotic experimental ecosystem. Aquat. Toxicol. 144–145, 186–198. Donker, M.H., 1992. Energy reserves and distribution of metals in populations of the isopod Porcellio scaber from metal contaminated sites. Funct. Ecol. 6, 445–454. Felten, V., Charmantier, G., Mons, R., Geffars, A., Rousselle, P., Coquery, M., Garric, J., Geffard, O., 2008. Physiological and behavioral responses of Gammarus pulex (Crustacea: Amphipoda) exposed to cadmium. Aquat. Toxicol. 86, 413–425. Fingerman, M., Hanumante, M.M., Fingerman, S.W., Reinschmidt, D.C., 1981. Effects of norepinephrine and norepinephrine agonists and antagonists on the melanophores of the fiddler crab Ulca pugilator. J. Crustac. Biol. 1, 16–27. Flint, S., Markle, T., Thompson, S., Wallace, E., 2012. Bisphenol A exposure, effects, and policy: a wildlife perspective. J. Environ. Manag. 104, 19–34. Fromme, H., Kuchler, T., Otto, T., Pilz, K., Muller, J., Wenzel, A., 2002. Occurrence of phthalates and bisphenol A and F in the environment. Water Res. 36, 1429–1438. Gollady, S.W., Webster, J.R., Benfield, E.F., 1983. Factors affecting food utilization by a leaf shredding aquatic insect: leaf species and conditioning time. Ecography 6 (2), 157–162. Graca, M.A.S., Maltby, L., Calow, P., 1993. Importance of fungi in the diet of Gammarus pulex and Asellus aquaticus. Oecologia 96, 304–309. Graca, M.A.S., Maltby, L., Calow, P., 1994. Comparative ecology of Gammarus pulex (L.) and Asellus aquaticus (L.) I: population dynamics and micro distribution. Hidrobiology 281, 155–162. Green, D.W.J., Williams, K.A., Huges, D.R.L., Shaik, G.A.R., Pascoe, D., 1988. Toxicity of phenol to Asellus aquaticus (L.)–effects of temperature and episodic exposure. Water Res. 22, 225–231. Hill, M., Stabile, C., Steffen, L.K., Hill, A., 2002. Toxic effects of endocrine disrupters on freshwater sponges: common developmental abnormalities. Environ. Pollut. 117, 295–300. ISO 10706. 2000. Water quality–Determination of long term toxicity of substances to Daphnia magna Straus (Cladocera, Crustacea). ISO 6341. 2013. Water quallity–Determination of the inhibition of the mobility of Daphnia magna Straus (Cladocera, Crustacea). Lachaise, F., Lerous, A., Hubert, M., Lafont, R., 1993. The molting gland of crustaceans – Localization, activity, and endocrine control (a review). J. Crustac. Biol. 13, 198–234. Lahnsteiner, F., Berger, B., Kletz, M., Weismann, T., 2005. Effects of Bisphenol A on maturation and quality of semen and egg in the brown trout, Salmo trutta f. fario. Aquat. Toxicol. 75, 213–224. Landrum, P.F., Robbins, J.A., 1990. Bioavailability of sediment-associated contaminants to benthic invertebrates In: Baudo, R., Giesy, J., Muntau, H. (Eds.), Sediments: Chemistry and Toxicity of in-place Pollutants. Lewis Publishers, Florida, pp. 237–263. Liu, C., Yan, W., Zhou, B., Guo, Y., Liu, H., Yu, H., et al., 2012. Characterization of a bystander effect induced by the endocrine-disrupting chemical 6-propyl-2thiouracil in zebrafish embryos. Aquat. Toxicol. 118–119, 108–115. McCahon, C.P., Barton, S.F., Pascoe, D., 1990. The toxicity of phenol to the freshwater crustacean Asellus aquaticus (L.) during episodic exposure–relationship between sub–lethal responses and body phenol concentrations. Arch. Environ. Contam. Toxicol. 19 986–929. Mihaich, E.M., Friderich, U., Caspers, N., Hall, A.T., Klečka, G.M., Dimond, S.S., Staples, C.A., Ortego, L.S., Hentges, S.G., 2009. Acute and chronic toxicity testing of bisphenol A with aquatic invertebrates and plants. Ecotoxicol. Environ. Saf. 72,
1392–1399. OECD 2004. Test guideline 218 Sediment-water Chironomid toxicity test using spiked sediment. Okada, H., Tokunaga, T., Liu, X.H., Takayanagi, S., Matsushima, A., Shimohigashi, Y., 2008. Direct evidence revealing structural elements essential for the high binding ability of bisphenol A to human estrogen-related receptor-gamma. Environ. Health Perspect. 116, 32–38. Opperhuizen, A., 1991. Bioconcentration and biomagnifications: is a distinction necessary In: Nagel, R., Loskill, R. (Eds.), Bioaccumulation in Aquatic Systems: Contributions to the Assessment. VCH, New York, pp. 43–66. Pascoe, D., Carroll, K., Karntanut, W., Watts, M.M., 2002. Toxicity of 17 alpha-ethinylestradiol and bisphenol A to the freshwater Cnidarian Hydra vulgaris. Arch. Environ. Contam. Toxicol. 43, 56–63. Peeters, E.T.H.M., De Jager, T.J., Beijer, J.A.J., Koelmans, A.A., 2000. Effects of benzo (a) pyrene and size of organic matter particles on bioaccumulation and growth of Asellus aquaticus. Arch. Environ. Contam. Toxicol. 39, 307–314. Petrović, M., Eljarrat, E., Lopez de Alda, M.J., Barcelo', D., 2004. Endocrine disrupting compounds and others emerging contaminants in the environment: a survey on new monitoring strategies and occurrence data. Anal. Bioanal. Chem. 378, 549–562. Plahuta, M., Tišler, T., Toman, M.J., Pintar, A., 2014. Efficiency of advanced oxidation processes in lowering bisphenol A toxicity and oestrogenic activity in aqueous samples. Arch. Ind. Hyg. Toxicol. 65, 77–87. Rodriguez, E.M., Medesani, D.A., Fingerman, M., 2007. Endocrine disruption in crustaceans due to pollutants: a review. Comp. Biochem. Physiol. 146, 661–671. Rubin, B.S., 2011. Bisphenol A: an endocrine disruptor with widespread exposure and multiple effects. J. Steroid Biochem. Mol. Biol. 127, 27–34. Qiao, P., Gobas, F.A.P.C., Farrell, A.P., 2000. Relative contributions of aqueous and dietary uptake of hydrophobic chemicals to the body burden in juvenile rainbow trout. Arch. Environ. Contam. Toxicol. 39, 369–377. Schug, T.T., Janesick, A., Blumberg, B., Heindel, J.J., 2011. Endocrine disrupting chemicals and disease susceptibility. J. Steroid Biochem. Mol. Biol. 127, 204–215. Staples, C.A., Dorn, P.B., Klecka, G.M., O'Block, S.T., Harris, L.R., 1998. A review of the environmental fate, effects, and exposures of bisphenol A. Chemosphere 36, 2149–2173. Staples, C.A., Tilghman, H.A., Friederich, U., Caspers, N., Klecka, G.M., 2011. Early lifestage and multigeneration toxicity study with bisphenol A and fathead minnows (Pimephales promelas). Ecotoxicol. Environ. Saf. 74 (6), 1548–1557. Staub, G.C., Fingerman, M.A., 1984a. . Mechanism of action for the inhibition of black pigment dispersion in the fiddler crab, Uca pugilator, by naphthalene. Comp. Biochem. Physiol. Part C: Toxicol. Pharmacol. 79, 447––453. Staub, G.C., Fingerman, M., 1984b. Effect of naphthalene on color changes of the sand fiddler crab, Uca pugilator. Comp. Biochem. Physiol. Part C: Toxicol. Pharmacol. 77 (1), 7–12. Toda, H., Arima, T., Takahashi, M., Inhimura, S., 1987. Physiological evaluation of temperature effect on the growth processes of the mysid, Neomysis intermedia Czerniawsky. J. Plankton Res. 9, 51–63. Van Brummelen, T.C., Stuijfzand, S.C., 1993. Effects of benzo (a) pyrene on survival, growth and energy reserves in the terrestrial isopods Oniscus asellus and Porcellio scaber. Sci. Total Environ. 134 (2), 921–930. Van Hattum, B., De Voogt, P., Van den Bosch, L., Van Straalen, N.M., Joosse, E.N., Govers, H., 1989. Bioaccumulation of cadmium by the freshwater isopod Asellus aquaticus (L.) from aqueous and dietary sources. Environ. Pollut. 62 (2–3)), 129–151. Vandenberg, L.N., Hauser, R., Marcus, M., Olea, N., Welshons, W.W., 2007. Human exposure to bisphenol A (BPA). Reprod. Toxicol. 24 (2), 137–177. Watts, M.M., Pascoe, D., Carroll, K., 2001. Survival and precopulatory behavior of Gammarus pulec (L.) exposed to two xenoestrogens. Water Res. 35 (10), 2347–2352. Watts, M.M., Pascoe, D., Carroll, K., 2003. Exposure to 17 alpha-ethinylestradion and bisphenol A-effects on larval moulting and mouthpart structure of Chironomus riparius. Ecotoxicol. Environ. Saf. 54, 207–215. Yamada, K., Urase, T., Matsuo, T., Suzuki, N., 1999. Constituents of organic pollutants in leachate from different types of landfill sites and their fate in the treatment process. J. Jpn. Soc. Water Environ. 22, 40–45. Yamamoto, T., Yasuhara, A., Shiraishi, H., Nakasugi, O., 2001. Bisphenol A in hazardous waste landfill leachate. Chemosphere 42, 415–418. Zou, E., 2010. Aquatic invertebrate endocrine disruption In: Breed, M.D., Moore, J. (Eds.), Encyclopedia of Animal Behavior. Academic Press, London, pp. 112–123. Zou, E., Fingerman, M., 1997a. Effects of estrogenic xenobiotics on molting of the water flea Daphnia magna. Ecotoxicol. Environ. Saf. 38, 281–285.. Zou, E., Fingerman, M., 1997b. Synthetic estrogenic agents do not interfere with sex differentiation but do inhibit molting of the cladoceran Daphnia magna. Bull. Environ. Contam. Toxicol. 58, 596–602.
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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Adverse effects of bisphenol A on water louse (Asellus aquaticus) Maja Plahuta a,n, Tatjana Tišler a, Albin Pintar a, Mihael Jožef Toman b a b
Laboratory for Environmental Sciences and Engineering, National Institute of Chemistry, Hajdrihova 19, SI-1001 Ljubljana, Slovenia University of Ljubljana, Biotechnical Faculty, Department of Biology, Jamnikarjeva 101, SI-1000 Ljubljana, Slovenia
art ic l e i nf o
a b s t r a c t
Article history: Received 22 October 2014 Received in revised form 21 March 2015 Accepted 25 March 2015
Experiments were performed to study the effects of short and long-term exposure to bisphenol A (BPA) on a freshwater crustacean isopod Asellus aquaticus (L.). Two life stages of isopods were exposed to a range of BPA concentrations, from aqueous and two dietary sources, in the form of with BPA spiked conditioned alder leaf (Alnus glutinosa) discs, or spiked formulated sediment, to determine the relative importance of each source of exposure on the uptake of this contaminant. Several lethal and sublethal endpoints were evaluated in this study to measure the potential effects of BPA on A. aquaticus, including mortality, growth and feeding rate inhibition, mobility inhibition, de-pigmentation and molting disturbances. They signify a correlation to BPA levels and a difference in BPA uptake efficiency from different uptake sources. Results of acute exposure to BPA show a greater sensitivity of test systems using juvenile specimens with a 96 h LC50 of 8.6 mg L 1 BPA in water medium and a 96 h LC50 of 13.5 mg L 1 BPA in sediment. In comparison, adult isopods show a 96 h LC50 of 25.1 mg L 1 BPA in water medium and a 96 h LC50 of 65.1 mg L 1 BPA in sediment. Observed endpoints of chronic exposures suggest the alder leave discs to be the most efficient uptake source of BPA, in contrast to uptake from water or heterogeneous sediment. Significant (po0.05) growth inhibition, with a 21 d NOEC of 0.5/2.5 mg L 1 (for juvenile/adult organisms), and feeding rate inhibition, with a 21 d NOEC of 0.5/1.0 mg L 1 (for juvenile/adult organisms), were proven to be the most sensitive toxicity endpoints. An even more sensitive effect turned out to be molting frequency, which was significantly reduced; a 21d NOEC was 1.0 mg L 1 of BPA for adult organisms and an even lower 21d NOEC of 0.05 mg L 1 of BPA for juveniles. The observed endpoints are recorded at very low, non-toxic exposure concentrations, indicating that BPA acts as an endocrine disrupting compound, as well as a toxic substance. We also determined the importance of the direct dietary uptake of the pollutants, significant for juveniles as well as adult animals. & 2015 Elsevier Inc. All rights reserved.
Keywords: Asellus aquaticus Bisphenol A Aqueous uptake Dietary uptake Endocrine disrupting effects
1. Introduction Aquatic ecosystems are daily polluted with man-made toxic organic pollutants, which are biologically active and can impose adverse health effects on wildlife and humans. In the past two decades, there has been an increasing concern over some classes of environmental contaminants known as endocrine-disrupting chemicals (EDCs), which have the ability to interfere with the endocrine system of exposed organisms. Often they mimic or block endogenous hormones (Petrović et al., 2004; Schug et al., 2011). Due to a high production and environmental release of various synthetic industrial chemicals and their by-products, these compounds have resulted in developmental deficits and reproductive impairments in a wide range of aquatic species (Crain n
Corresponding author. E-mail addresses: [email protected] (M. Plahuta), [email protected] (T. Tišler), [email protected] (A. Pintar), [email protected] (M.J. Toman). http://dx.doi.org/10.1016/j.ecoenv.2015.03.031 0147-6513/& 2015 Elsevier Inc. All rights reserved.
et al., 2007; Zou, 2010; Rubin, 2011). In this study, a well-known organic contaminant, used worldwide in the production of polycarbonate plastic and epoxy resins, bisphenol A (BPA, 2,2-bis-(4hydroxyphenyl)-propane; CAS Registry No. 80-05-7), was investigated (Staples et al., 1998). It is generating concerns due to its extensive production, widespread use and consequently frequent presence in various environmental matrices (Vandenberg et al., 2007; Flint et al., 2012). Wildlife (as well as humans) is exposed to BPA from a variety of environmental releases like waste landfills leachate and wastewater treatment plants (Staples et al., 1998). The concentrations of BPA vary greatly depending on location and time of sampling. European surface water analyzes report BPA concentrations ranging from 0.5 up to 21 mg L 1 (Fromme et al., 2002; Belfroid et al., 2002). Levels of BPA in waste landfill leachate can be even higher, 5.4 mg L 1 (Yamada et al., 1999) and even up to 17.2 mg L 1 of BPA (Yamamoto et al., 2001). From the obtained values we can see that such concentrations could cause adverse effects on Asellus aquaticus in long-term exposures. BPA has been
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the subject of considerable aquatic toxicity testing in recent years. The majority of available effect data on aquatic organisms has been obtained from studies using fish (Lahnsteiner et al., 2005; Staples et al., 2011), some aquatic invertebrates (Hill et al., 2002; Pascoe et al., 2002; Watts et al., 2003) and macrophyte plants (Mihaich et al., 2009). De Kermoysan et al. (2013) also studied the long-term effects of BPA on macrophytes, macro-invertebrates and fish in an aquatic mesocosm. BPA also acts as an estrogenically active endocrine disrupting compound (Okada et al., 2008), and as such impacts wildlife populations by interfering with some aspects of endocrine-mediated processes, causing impairment of developmental, reproductive and other hormonally mediated processes (Rodriguez et al., 2007). In the aquatic environment organisms can uptake lipophilic toxicants from contaminated water via gills or the skin. Dietary sources also play an important role in contaminants uptake through ingestion and the gastrointestinal tract, especially in sediment dwelling organisms. They are exposed directly to contaminants by ingesting organic matter from the sediment and uptaking the sediment-associated contaminants from interstitial and overlying water. As uptake from dietary sources can explain 100% of total residue, it is possible that the source of BPA, to which an organism is exposed, plays an important role in effects occurrence in animals (Landrum and Robbins, 1990; Peeters et al., 2000). The water louse, Asellus aquaticus (Linnaeus, 1758), was chosen as a model species in this study for several reasons. Firstly, there are no previous reports of the effects of BPA to A. aquaticus. It is native in Slovenia and abundantly present in surface waters all throughout western Eurasia (Graca et al., 1994). It is in continuous contact with water and the sediment (McCahon et al., 1990), easy to culture in the laboratory, tolerant to varying physicochemical characteristics of sediment, has negligible biotransformation, intermediate sensitivity to contaminants and shows specific effects of toxicity and endocrine disruption (van Hattum et al., 1989). It is a shredder that feeds on detritus associated with fungi, bacteria and periphyton (Graca et al., 1993). The dietary preference of A. aquaticus toward naturally fungally colonized leaf material was presented by Bloor (2011). A. aquaticus also presents an important food source for predatory invertebrates, fish, and waterfowl (Van Hattum et al., 1989) and thus plays an important role in the benthic-pelagic coupling of the food chain transfer of pollutants (Peeters et al., 2000). Their small size and robust nature make them ideally suited for application in toxicity tests. Furthermore, early life stages of the species offer all the key attributes of a complete in vivo system. The objective of this study was to investigate the effects of BPA uptake from aqueous or dietary sources (inoculated alder (Alnus glutinosa) leaf discs or formulated sediment) on survival, growth, feeding rate, mobility, pigmentation and molting of two different life stages of the sediment dwelling freshwater isopod Asellus aquaticus (L.) in a laboratory experiment, which will enable us to better characterize the toxic and endocrine disrupting effects of BPA.
2. Materials and methods Laboratory bred aquatic isopods A. aquaticus were exposed to BPA from aqueous and dietary sources (in the form of spiked alder leaf discs and spiked formulated sediment) to determine the importance of each source of exposure on the uptake of this pollutant. For the purposes of determining the effects of BPA on different life stages of A. aquaticus, both adult and juvenile life stages were exposed to BPA for 4 d (acute exposure) as well as 3 weeks (chronic exposure). Control groups consisted of 12 specimens per test, exposed to unpolluted water, food or sediment.
2.1. Test organisms and culture procedures Specimens of A. aquaticus were collected from an unpolluted stream in Črna vas, Ljubljana (Slovenia, 45° 59′ N, 14° 28′ E) and laboratory cultured as a population giving rise to new generations used in tests. As such, animals used in testing, had no prior history of contamination. Isopods were kept in four liter glass aquaria tanks containing reconstructed water (M4 medium, ISO 10706, 2000) with pH 7.57 0.5, dissolved oxygen (4 60% of air saturation value), at 20 72 °C, with a 16:8 h light/dark photoperiod, illumination around 500 lx, and continuously purged with purified air at a flow rate of 50 mL min 1. Organisms were weekly fed on conditioned alder leaves (Alnus glutinosa), colonized by micro-organisms to maintain a healthy population. 2.2. Chemicals and preparation for test conduction The chemical used in the tests was Bisphenol A (BPA) (SigmaAldrich, Germany, purity 99%) dissolved in ultrapure water (18.2 MΩ cm 1, ELGA LabWater, UK) as a stock solution of 100 mg L 1. For testing purposes salts were added to the solution, forming reconstructed water (Elendt M4 medium, ISO 10706, 2000) or dilution water (ISO 6341, 2013). Initial BPA stock solution was diluted for purposes of acute and chronic exposure to BPA from water, leaves and sediment. 24 h prior to each test, randomly chosen juvenile and adult laboratory bred males and females were placed in 250 mL lidded glass containers, containing deionized water in oxygen depleting conditions without nutritional support. 12 adults or juveniles in intermolt stage were then placed in 6 well test plates (Ø 33.78 22 mm) (Techno Plastic Products AG, Trasadingen, Switzerland), arranged one animal per test chamber. 2.3. Experimental design Each experiment was performed in parallels, and replicated three times. Exposure was conducted in 6 well test plates with Ø 35 mm, 10 mL wells. In tests, juvenile (size range 1.0–1.4 mm) and adult (size range 4–5 mm) specimens of A. aquaticus were exposed at 20 71 °C. Acute exposure tests were performed to derive 96 h LC50 values, which were further used to set up the chronic exposures. Organisms were exposed to test media for 96 h in static exposure system without aeration. Juvenile isopods were acutely exposed to aqueous and sediment BPA concentration range from 2.5 to 20 mg L 1. Adult organisms were acutely exposed to the range from 5 to 70 mg L 1 BPA. Chronic exposure tests were performed for 21 d in a semi-static exposure system with medium renewal every three days. Organisms were exposed to aqueous, food and sediment bound BPA in concentrations ranging from 0.05 to 5.0 mg L 1 BPA for juveniles and from 0.5 to 10.0 mg L 1 BPA for adults. The 21 d NOEC and LOEC for each effect were calculated (P o0.05, ANOVA). Test solutions were taken when water renewed, for measuring the concentrations of test chemical by means of HPLC. At the end of the experiment pH value and dissolved oxygen concentration were measured, to assure the effects only be attributed to the toxic chemicals. Survival of the control group was above 80% after acute and chronic exposure to unpolluted media. In tests determining effects of aqueous BPA on exposed isopods, each test chamber was filled with 10 mL of test medium. In long term exposure tests, one Ø 10 mm unpolluted alder leaf disk was added to each chamber, acting as food as well as shelter. As no specific sediment preferences were recognized in the literature, for purposes of short and long term sediment exposure tests with Asellus aquaticus, formulated sediment was prepared according to the OECD 218 (2004) for Chironomid sediment toxicity test, with a few modifications, as follows. The sediment
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consists of 94% of natural quartz sand (Termit d.d., Slovenia) with granulation r1 mm, 2% of kaolin clay ( Z30% kaolinite, Merck, Darmstadt, Germany), CaCO3 (Merck, Darmstadt, Germany), and 4% organic content in the form of grounded unpolluted conditioned Alder leafs (Alnus glutinosa). Prior to grinding, leafs were dried for 24 h at 207 1 °C. Deionized water was added to the sediment mixture, in the ratio 1:4 of mixture/water. Calcium carbonate was added to adjust the pH of the final sediment mixture to 7.0 70.5. The prepared mixture was filtrated using a 1.2 mm CN membrane filter (Sartorius AG, Goettingen, Germany), and the solid residue was re-suspended in BPA solutions of different concentrations, in the ratio 1:4 of sediment/solution. Sediment testing included a control (unpolluted) sediment. Exposure test, conducted in test plates, contained 2 g of dry weight of the sediment and 5 mL of dilution water per well. Overlying water was renewed every three days, the sediment was not renewed. For feeding and testing purposes autumn-shed alder leaves (Alnus glutinosa) were collected at an Alder forest and used as a food source for isopods. The leaves were dried at 40 °C for 24 h, and stored in a dry place, until used. The dry leaves were soaked in 500 mL of river water (collected from the Asellus aquaticus sampling place) for at least 10 d in order to be colonized by microorganisms (Gollady et al., 1983). For testing purposes the conditioned leaf material was cut into Ø 10 mm discs. The leaf midrib was avoided when cutting out discs. Spiking was performed by soaking the discs in BPA solutions for 96 h, with daily solution renewal, while constantly stirring on a magnetic stirrer. Before solution renewal, samples were taken to be analyzed using HPLC in order to determine the amount of BPA spiked on leaf discs. Prepared discs were air dried for 24 h, and weighed before placed each into one testing chambers containing one A. aquaticus and 10 mL of reconstructed water. Leaf consumption was daily visually monitored and after 21 d, the discs were removed, air dried for 24 h, and reweighted. The weight loss of the leaf discs, not caused by A. aquaticus activity, was determined by exposing the discs to test medium for 21 d in the absence of isopods.
isopods that assumed a lateral position, with extended legs, with only pleopod and/or antennal movement were considered immobile. The animals did not move even after gentile mechanic agitation. To determine the de-pigmentation, the exposed animals were divided in two groups, either a specimen was considered pigmented or de-pigmented, in which case less than approximately 20% of body surface, determined by visual inspection, contained black pigment. The pigmentation results of tested BPA concentrations were compared to a control group, which was not exposed to BPA.
2.4. Test endpoints
3.1. Acute exposure effects
The primary short exposure endpoint was survival. In addition mobility, growth rate, feeding rate, de-pigmentation, and ecdysis were observed in long term studies. Survival, mobility and molting of test organisms were quantitatively determined through visual observations daily, in all tests. Death was determined by absence of respiratory movement. The dead animals were regularly removed from test chambers. Total body length was measure from the tip of the head to the end of the pleotelson. The animals were measured individually prior the test, under a stereo microscope (Nikon SMZ1000, resolution 80) and NIS-Element D Microscope Imagining Software (Nikon, Tokyo, Japan). After the experiments were stopped the live animals were re-measured, and growth rate was calculated by subtracting the body length at the beginning of the test from the body length at the end of the test. Inhibition of growth rate was determined as a decrease in growth rate per day in test groups, in comparison to control group. The feeding rate was calculated by subtracting the end disc weight from the weight at the beginning of the test. The consumed weight was then divided by the number of days the disc was exposed, to acquire the average leaf consumption per day. Inhibition of feeding rate was calculated as a difference in relation to the control group, which was not exposed to BPA. The molting disturbances, including effects on molting frequency, and the time span until the first ecdysis, were determined by daily visual inspection. The exuviae (molted exoskeletons) were removed from test chambers on daily basis. Immobility and de-pigmentation were also visually determined at the end of the test. According to Green et al. (1988),
Acute exposure of A. aquaticus to aqueous and sediment bound selected organic pollutant and endocrine disruptor BPA, demonstrated significant effects on the mortality rate (Po 0.05, ANOVA) of exposed isopods. Median lethal concentrations (96 h LC50) for aqueous BPA were obtained at 9.5 mg L 1 for juveniles and 25.1 mg L 1 for adult organisms (Fig. 1). Acute exposure to BPA bound to formulated sediment, showed a less significant effect on mortality in juvenile as well as adult organisms. The median lethal concentration (96 h EC50) for juvenile organisms was 4.7 mg L 1 and 65.1 mg L 1 for adult, regarding sediment bound BPA. Mortality is presented by means of dose-mortality curve, together with standard deviation and LC50 values (Fig. 1). The juvenile and adult LC50 differed significantly (po 0.05), where juveniles were more sensitive to BPA in comparison to adults. Comparison of effects of the test medium also demonstrated a significantly different LC50 between uptake from water and from sediment. Comparing the above results to data obtained from the literature for other species, or previously reported (Plahuta et al., 2014), it is obvious that the sensitivity of the test system using juvenile A. aquaticus is higher than the one using Daphnia magna (24 h EC50 was 12.5 mg L 1), but still lower than the test system using Gammarus pulex (48 h LC50 was 5.6 mg L 1 BPA) (Watts et al., 2001).
2.5. HPLC analysis The concentrations of BPA in water samples were analyzed by means of high performance liquid chromatography (HPLC; Agilent Technologies, USA) equipped with a quaternary solvent pump, autosampler and UV detection (λ ¼ 210 nm) at a constant temperature of 30 °C and a limit of detection (LOD) at 0.2 mg L 1. 2.6. Statistical analysis LC50 values for tests with Asellus aquaticus were determined using the probit analysis based on results from acute exposure tests and responses of organisms under various concentrations of BPA. The normality of our results was analyzed with the Shapiro– Wilk test. The test deducted that the results are normally distributed. Differences between groups were tested using one-way ANOVA, on results from chronic exposure tests, with BPA concentrations as the explanatory variable. The critical level of significance was po 0.05. T-test was used as a post hoc test, to test between-concentration differences (po 0.05). Statistical analyzes were performed using Origin 8.1 SR3 data analysis and graphing software (OriginLab Corporation, MA, USA).
3. Results and discussion
3.2. Chronic sublethal exposure effects Chronic exposure to the selected organic pollutant BPA
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100
80 Immobilization (%)
Mortality after 96 h (%)
80
100
Uptake from water / A Uptake from sediment / A Uptake from sediment / J Uptake frome water /J
60
40
60
Uptake from water / A Uptake from food / A Uptake from sediment / A Uptake from water / J Uptake from food / J Uptake from sediment / J
40
20 20
0 0.1
0 0.1
1
10
1
10
BPA concentration in media (mg L )
100
BPA concentration in media (mg L )
demonstrated visible effects, which were related to BPA concentration levels, and the types of uptake routes (water, leaf or sediment). An experiment was conducted where desorption of BPA from the sediment, and spiked alder leaf discs (Alnus glutinosa), into the overlaying water was observed. It was concluded that desorption from the sediment was minimal and below detection limits of our chemical analysis, whereas the maximum desorption of BPA from spiked alder leaf discs was 12%.
100
80 Pigmentation (%)
Fig. 1. The percentage of adult and juvenile A. aquaticus mortalities in acute exposure tests to aqueous BPA, or from sediment bound BPA, is presented. The results are the means of three repeated tests with the standard deviation between tests indicated by the error bars on the graph. The 96 h LC50 values are indicated by the dotted lines. A and J in the legends refer to adult or juvenile specimens, exposed in the experiment.
60
40
20
0 0.1
Uptake from water / A Uptake from food / A Uptake from sediment / A Uptake from water / J Uptake from food / J Uptake from sediment / J
1
10
BPA concentration in media (mg L ) 3.2.1. Immobilization The impact of chronic exposure to BPA from water, leaf discs and formulated sediment, on locomotion of juvenile and adult A. aquaticus, are presented as concentration dependent response curves in Fig. 2. Exposed animals showed a range of sensitivity to BPA depending on the age of animals and the uptake route of BPA. Highest statistically significant sensitivity in comparison to control group, was seen with juvenile isopods exposed to BPA spiked on alder leaf discs, with a 21 d NOEC for immobilization of 1.5 mg L 1 of BPA, a 21 d LOEC for immobilization of 2.0 mg L 1 BPA, and 21 d EC50 of 2.15 ( 70.16) mg L 1 BPA (ANOVA, F ¼183.1, df ¼5, p o0.05; t-Test, p o0.05) (Table 1). Similar results were drawn from adult exposure to BPA with a 21 d LOEC for immobilization of 5.0 mg L 1, and a 21 d NOEC of 2.0 mg L 1, and a 21 d EC50 of 5.33 (70.33) mg L 1 BPA (ANOVA, F¼38.6, df ¼4, p o0.05; t-Test, p o0.05) after dietary exposure to BPA from alder leaf discs. Aqueous exposure to BPA caused a 21 d NOEC/LOEC of 2.0/2.5 mg L 1 of BPA, and a 21 d EC50 of 2.35(70.11) mg L 1 BPA (ANOVA, F¼110.7, df ¼ 5, p o0.05; t-Test, po 0.05) for juveniles, and 5/10 mg L 1 of BPA and a 21 d EC50 of 4.42 ( 70.58) mg L 1 BPA (ANOVA, F¼ 114.4, df ¼4, p o0.05; t-Test, p o0.05) for adults. Least effects were seen when the animals were exposed to artificial sediment spiked with BPA, with a 21 d NOEC/LOEC of 5/10 mg L 1 BPA for both life stages and a 21 d EC50 of 5.10 (70.44) mg L 1 BPA (ANOVA, F¼92.6, df ¼6, p o0.05; t-Test, p o0.05) for juveniles and a 21 d EC50 of 5.64 ( 70.82) mg L 1 BPA (ANOVA, F ¼20.9, df ¼4, p o0.05; t-Test, p o0.05) for adults (Table 1). Immobilization symptoms are characteristic for A. aquaticus exposure to phenol (McCahon et al., 1990; Green et al., 1988). After longer exposures the animals ceased to move, but the pleopods
Fig. 2. (A) presents the percentage of immobilized animals after a chronic exposure to BPA. (B) presents the percentage of pigmented specimens after chronic exposure to BPA. The results are presented in the concentration response curves as the means of three separate tests with the standard deviation between tests indicated by the error bars on the graph. Hill model was used to perform the curve fitting. A and J in the legends refer to adult or juvenile specimens, exposed in the experiment.
still beat, where the animals were paralyzed, and subsequently died (Green et al., 1988). 3.2.2. Pigmentation Statistically significant reduction of body pigmentation was observed after chronic exposure to BPA. The results are presented in concentration dependent response curves in Fig. 2. The 21 d NOEC/LOEC values are presented in Table 1. The observed effect was most pronounced in juvenile organisms exposed to alder leaf discs spiked with BPA, with 21 d NOEC/LOEC of 1.0/2.0 mg L 1 BPA and 21 d EC50 of 1.72 ( 70.64) mg L 1 BPA (ANOVA, F¼ 180.2, df ¼6, po 0.05; t-Test, p o0.05). The 21 d NOEC/LOEC for adults exposed to BPA from alder leaves was 2.0/5.0 mg L 1, while 21 d EC50 of 1.83 (7 0.51) mg L 1 BPA (ANOVA, F¼262.2, df ¼ 5, po 0.05; t-Test, p o0.05). Exposure to aqueous and sediment bound BPA caused less significant effects on pigmentation with juveniles, with 21 d NOEC/LOEC of 2.0/2.5 mg L 1 BPA and 2.5/5.0 mg L 1 BPA (ANOVA, F¼95.8, df ¼5, p o0.05; t-Test, po 0.05) respectively. 21 d EC50 for aqueous uptake of juveniles was 2.24 ( 70.52) mg L 1 BPA, and 21 d EC50 for sediment uptake of juveniles was 2.21 ( 70.81) mg L 1 BPA. Less effect was observed also with adults exposed to aqueous as well as for those exposed to sediment bound BPA, with 21 d NOEC/LOEC of
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Table 1 Overview of the 21 d NOEC/LOEC values of the chronic exposure of juvenile and adult A. aquaticus to BPA from water, food or sediment. Age
Immobilization Pigmentation Molting Growth Feeding
Water
Juvenile Adult Juvenile Adult Juvenile Adult Juvenile Adult Juvenile Adult
Food
21 d NOEC (mg L 1)
21 d LOEC (mg L 1)
21 d NOEC (mg L 1)
21 d LOEC (mg L 1)
21 d NOEC (mg L 1)
21 d LOEC (mg L 1)
2.0 5.0 2.0 5.0 0.1 2.0 2.5 5.0 2.0 2.0
2.5 10.0 2.5 10.0 0.5 5.0 5.0 10.0 2.5 5.0
1.5 2.0 1.0 2.0 0.05 1.0 0.5 2.0 0.5 1.0
2.0 5.0 2.0 5.0 0.1 2.0 1.0 5.0 1.0 2.0
5.0 5.0 2.5 5.0 / / 2.5 5.0 / /
10.0 10.0 5.0 10.0 / / 5.0 10.0 / /
5.0/10.0 mg L 1 BPA (ANOVA, F¼104.9, df ¼ 4, po 0.05; t-Test, p o0.05), and 21 d EC50 of 3.32 ( 74.03) mg L 1 of aqueous BPA, and 21 d EC50 of 3.63 (74.23) mg L 1 of sediment bound BPA. The results of body de-pigmentation intensity shows that dietary exposure to BPA from alder leaf discs presents the most efficient BPA uptake route among tested. The color changes in crabs are controlled hormonally
*
80
(Fingerman et al., 1981; Liu et al., 2012), and the dispersion of the black pigment is controlled by BPDH (black pigment dispersing hormone), produced principally in the eyestalk ganglia. It was reported that EDC as PCBs caused sublethal effects, like inhibited black pigment dispersion in the crustaceans. Reportably it inhibited the release of BPDH (Staub and Fingerman, 1984a). Naphthalene also inhibits the black pigment dispersion by inhibiting
Week3 Week2 Week1
100
Molting (%)
Sediment
Week3 Week2 Week1
100
80
60
*
*
40
20
*
*
*
*
Molting (%)
Effect
*
60
* 40
*
2.5
5
20
0
0 0
0.05
0.1
0.5
1
2
2.5
5
0
0.05
BPA concentration in media (mg L )
*
80
0.1
0.5
1
2
BPA concentration in media (mg L )
100
*
Week3 Week2 Week1
Week3 Week2 Week1
100
80
*
Molting (%)
* Molting (%)
*
60
40
*
60
* 40
* 20
20
0
0 0
0.5
1
2
5
BPA concentration in media (mg L )
10
15
0
0.5
1
2
5
10
15
BPA concentration in media (mg L )
Fig. 3. Inhibition in molting frequency and distribution of molts per week after a chronic exposure to BPA. (A) Results of juvenile A. aquaticus after chronic exposure to aquatic BPA. (B) Results of juvenile A. aquaticus after chronic exposure to BPA bound on alder leaf discs. (C) Results of adult A. aquaticus after chronic exposure to aquatic BPA. (D) Results of adult A. aquaticus after chronic exposure to BPA bound on alder leaf discs. The results are expressed as mean 7S.D. Asterisks indicate significant differences (po 0.05) in relation to the control group.
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M. Plahuta et al. / Ecotoxicology and Environmental Safety 117 (2015) 81–88
the release of BPDH (Staub and Fingerman, 1984b). Therefore it is possible that BPA as an endocrine disrupting compound also disrupts the dispersion of the black pigment, which is expressed as body de-pigmentation of A. aquaticus. 3.2.3. Molting Chronic exposure to the selected endocrine disrupting compound caused concentration dependent molting disturbances. Observed were significant inhibitions in molting frequency as well as an increase of the intermolt period (the time span between two ecdysis) (Fig. 3). Significantly reduced molting frequency (21 d LOEC) was observed at 5.0 mg L 1 (ANOVA, F ¼6.3, df ¼7, p o0.05; t-Test, po 0.05) of aqueous BPA for adult organisms and 0.5 mg L 1 (ANOVA, F¼8.2, df ¼7, p o0.05; t-Test, p o0.05) of aqueous BPA for juveniles. 21 d NOEC was observed at 2.0 mg L 1 of aqueous BPA for adults, and 0.1 mg L 1 of aqueous BPA for juveniles (Table 1). Dietary exposure to spiked alder leaf discs caused significantly delayed molting, and reduced molting frequency (21 d LOEC) at a concentration as low as 2.0 mg L 1 of BPA (ANOVA, F¼ 4.6, df ¼7, p o0.05; t-Test, p o0.05) for adult organisms and as low as 0.1 mg L 1 of BPA (ANOVA, F¼7.8, df ¼ 7, p o0.05; tTest, p o0.05) for juveniles. No observed effects (NOEC) were
documented at 1.0 mg L 1 (ANOVA, F¼9.3, df ¼7, p o0.05; t-Test, po 0.05) of aqueous BPA for adults, and 0.05 mg L 1 (ANOVA, F¼5.6, df ¼7, p o0.05; t-Test, po 0.05) of aqueous BPA for juveniles (Table 1). Increasing BPA concentration also seemed to drastically delay the molting. When juveniles were exposed to 2.5 mg L 1 of BPA spiked on alder leaf discs and 5.0 mg L 1 of aqueous BPA, only 18.2 and 27.3% of moltings were observed in the first week, respectively. The highest tested concentration, 10.0 mg L 1 of BPA spiked on alder leaf discs, caused a significant molting delay in adults with only 16.7% of molts being observed in the first week and another 16% of moltings in the second week. No further ecdyses were observed in the next two weeks. At these highest tested BPA concentrations other toxic effects were observed. We can attribute the lack of molting in highest tested BPA concentrations to toxicity in addition to endocrine disruption. Molting in crustaceans is under immediate control of steroid molt promoting hormones called ecdysteroids (Chang, 1993; DeFur et al., Tattersfield; Lachaise et al., 1993). Observed effects of BPA on A. aquaticus molting in low, sublethal concentrations, may indicate that BPA acts as an ecdysteroid antagonist. It is possible that a certain interaction exists between the steroid molting hormones and estrogen-mimicking BPA that would allow it to
0.05
0.16
Control Uptake from water Uptake from food
Control Uptake from water Uptake from food
0.14
0.12
0.03
*
0.02
*
*
* *
0.01
*
Feeding rate (mg day )
Feeding rate (mg day )
0.04
* 0.10
0.08
* *
0.06
*
0.04
* 0.02
0.00
0.00 10
2 0.5 3
1
4
5
2
6
8 2.5 9
7
10 5
00
11
1
1
1
BPA concentration (mg L )
2
5
5
5
10
10
10
BPA concentration (mg L )
0.020
0.030
Control Uptake from food Uptake from water Uptake from sediment
*
*
*
* *
0.010
*
0.005
Control Uptake from water Uptake from food Uptake from sediment
0.025
Growth rate (mm day )
0.015
Growth rate (mm day )
2
2
0.020
*
* 0.015
*
0.010
*
0.005
0.000 0.000 00
0.5
0.5 0.5
0.5
1
11
1
2
22
2
2.5
BPA concentration (mg L )
2.5 2.5
2.5
5
5 5
5
00
1
11
1
2
22
2
5
55
5
10
10 10
10
BPA concentration (mg L )
Fig. 4. Box plot graphs show the effects of chronic exposures on exposed specimens. (A) Presents the feeding rate inhibition of juvenile A. aquaticus in relation to the BPA concentration in water or food; n¼ 18. (B) Presents the feeding rate inhibition of adult A. aquaticus in relation to the BPA concentration in water or food; n¼ 18. (C) Presents the body growth rate inhibition of juvenile A. aquaticus in relation to the BPA concentration in water, food or sediment; n¼ 18. (D) Presents the body growth rate inhibition of adult A. aquaticus in relation to the BPA concentration in water, food or sediment; n¼ 18. Asterisks indicate significant differences (ANOVA, p o 0.05) in relation to the control group. The ends of the box are the upper and lower quartiles, the median is marked by a vertical line inside the box. The square inside the box represents the mean of the data, and the dotted line presents the average growth rate of the control group. The whiskers present the highest and lowest observations.
M. Plahuta et al. / Ecotoxicology and Environmental Safety 117 (2015) 81–88
compete with ecdysone at the ecdysteroid receptor, by influencing its release, or retaining its transformation into active 20-hydroxyecdosone. Because of structural similarities Molting inhibition in crustaceans has also been observed in Daphnia magna exposed to endocrine disrupting compounds (Zou and Fingerman, 1997a; Zou and Fingerman, 1997b). A delay in molting of Chironomus riparius after exposure to BPA was also observed by Watts et al. (2003), as a result of estrogenic activity of BPA. 3.2.4. Growth rate inhibition The present study showed a concentration dependent reduction in body length increase of A. aquaticus when exposed to BPA (Fig. 4). The test system using juvenile specimens showed a higher sensitivity to BPA regardless the uptake route in comparison to adults, although both used test systems showed a decrease in growth rate when exposed to with BPA spiked alder leaf discs. A statistically significant 42.7% reduction was observed in growth of juveniles with 21 d NOEC/LOEC of 2.5/5.0 mg L 1 of aqueous BPA (ANOVA, F¼3.8, df ¼ 5, po 0.05; t-Test, p o0.05), and a 43.8% growth reduction in adults with 21 d NOEC/LOEC of 5.0/10.0 mg L 1 of aqueous BPA (ANOVA, F¼ 4.2, df ¼4, p o0.05; tTest, p o0.05). Exposure to contaminated alder leaf discs resulted in a further reduction in growth with a significant 38.8% reduction in growth of juveniles and a 21 d NOEC/LOEC of 0.5/1.0 mg L 1 BPA and 21 d EC50 of 2.56 (7 0.29) mg L 1 BPA (ANOVA, F ¼4.3, df ¼5, p o0.05; t-Test, po 0.05), and 66.8% growth reduction in adults at a 21 d NOEC/LOEC of 2.5/5.0 mg L 1 BPA (2.5 mg L 1 BPA cause a 28.7% non significant growth rate reduction in adults) (ANOVA, F¼20.5, df ¼4, po 0.05; t-Test, p o0.05). Exposure to sediment bound BPA caused growth reduction effects comparable to that of aqueous BPA exposure. These results confirm that the dietary route of BPA uptake, from direct feeding with alder leaf discs spiked with BPA, presents the most efficient uptake route of BPA. However, both utilized test systems showed a prominent decrease in growth rate when exposed to BPA adsorbed to leaves. It encompasses an up to 70% growth reduction with juvenile A. aquaticus exposed to 5 mg L 1 BPA in leaves and an up to 81% growth reduction in adult organisms exposed to 10 mg L 1 BPA in leaves. BPA dissolved in water and adsorbed on the sediment caused a lesser amount of effect, which was found to be statistically significant only at highest tested concentrations. The growth results suggest that feeding with contaminated organic matter presents a dominant uptake rout of BPA for A. aquaticus. As in crustacean significant growth can only occur through periodic molting, disruption of molting, as well as inhibited feeding, will result in reduced growth (DeFur et al., Tattersfield; Toda et al., 1987). Another reason for decreased body growth after exposure to contaminants presents a reduction in available energy for growth (Van Brummelen and Stuijfzand, 1993). The organisms exposed to toxicants have to allocate energy to resist the toxicant by avoidance, exclusion, removal, or biochemical complexation (Donker, 1992). 3.2.5. Feeding rate inhibition The feeding results confirm what growth results showed, that the dietary uptake route presents an important route of contaminants uptake into aquatic organisms (Fig. 4). Feeding was significantly reduced in comparison to the control group, with a 21 d NOEC/LOEC of 2.0/2.5 mg L 1 of BPA and 21 d EC50 of 1.14 (7 0.83) mg L 1 BPA (ANOVA, F ¼3.7, df ¼5, p o0.05; t-Test, p o0.05) for juveniles and 21 d NOEC/LOEC of 2.0/5.0 mg L 1 of BPA and (ANOVA, F¼9.5, df ¼ 4, p o0.05; t-Test, p o0.05) for adults, exposed to aquatic BPA. Even more sensitive was the feeding rate with a dietary uptake of 21 d NOEC/LOEC equal to 0.5/1.0 mg L 1 of BPA and 21 d EC50 of 0.83 ( 70.97) mg L 1 BPA (ANOVA, F¼ 5.4, df ¼5, p o0.05; t-Test, p o0.05) for juveniles, and
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a 21 d NOEC/LOEC of 1.0/2.0 mg L 1 of BPA and 21 d EC50 of 1.3 (71.03) mg L 1 BPA (ANOVA, F¼25.7, df ¼4, po 0.05; t-Test, po 0.05) for adults. The obtained results suggest that the dietary uptake of BPA directly from contaminated food presents a more efficient uptake route than respiratory uptake of aquatic BPA through gills, as BPA is a liphophillic compound, with a log Kow of 3.40, and is expected to adsorb substantially to organic compounds of the sediment (Staples et al., 1998). In contrast to these findings, Felten et al. (2008) concluded the uptake through the gills of being the primary exposure route of aquatic species to contaminants. They found out that dietary uptake of hydrophobic xenobiotics increases in importance, directly with log Kow. It is also said that the transition between aqueous and dietary uptake is the main contributor to lipophillic chemical uptake at log Kow between 3.0 and 6.0, with predominant uptake route of dietary uptake at higher log Kow. Qiao et al. (2000) observed an aqueous uptake of 98% for a chemical with log Kow value of 5.05, and only 15% for a chemical with a log Kow value of 7.55. The present study opposes the findings of Opperhuizen (1991), who predicted that the gills and gastrointestinal tract were of equal importance for the uptake of hydrophobic chemicals. However, studies into the different uptake routes of contaminants show that water (Van Hattum et al., 1989), sediment and food (Peeters et al., 2000) are important.
4. Conclusions The present study has shown that bisphenol A causes lethal as well as sublethal toxic and endocrine disrupting effects in adult and juvenile Asellus aquaticus at acute and chronic exposures. We examined the relative importance of aqueous and dietary uptake of BPA, by comparing the intensity of observed exposure effects. The intensity of observed effects was considerably higher in animals exposed to with BPA spiked alder leaf discs, than when exposed to BPA dissolved in water or spiked on formulated sediment, which can be explained as food being a primary uptake route of BPA. Although expected, the sediment did not present a very efficient uptake source of BPA, causing subsequently an effect comparable to aquatic exposure. Of all observed effects, growth rate inhibition, feeding rate inhibition and molting inhibition were found to be the most sensitive exposure endpoints, at which statistically significant effects are expressed at lowest BPA concentrations, thus providing a realistic representation of possible endocrine-mediated effects in addition to toxic effects. The disruptions in molting were the most sensitive of all endpoints, recorded at the lowest tested concentrations, with a 21 d LOEC of 0.1 mg L 1 for juvenile A. aquaticus (21 d NOEC of 0.05 mg L 1). Other observed effects, including mortality, immobility and de-pigmentation are observed at higher concentrations, as a result of both endocrine disruption as well as general toxicity to the whole organism. The test system, using A. aquaticus, proved to be a rather useful indicating system for determining the presence and effects of the toxic as well as endocrine disrupting contaminant BPA, considering different uptake routes at low, environmentally relevant concentrations.
Acknowledgements This article is a result of doctoral research, in part financed by the European Union, European Social Fund and the Republic of Slovenia, Ministry for Education, Science and Sport in the framework of the Operational program for human resources development for the period 2007–2013. The authors gratefully
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acknowledge the financial support of the Ministry of Education, Science and Sport of the Republic of Slovenia through Research program No. P2-0150.
References Belfroid, A., van Velzen, M., van der Horst, B., Vethaak, D., 2002. Occurrence of bisphenol A in surface water and uptake in fish: evaluation of field measurements. Chemosphere 49, 97–103. Bloor, M.C., 2011. Dietary preference of Gammarus pulex and Asellus aquaticus during a laboratory breeding programme for ecotoxicological studies. Int. J. Zool. . http://dx.doi.org/10.1155/2011/294394, Article ID 294394, 5 pages Chang, E.S., 1993. Comparative endocrinology of molting and reproduction: insects and crustaceans. Annu. Rev. Entomol. 38, 161–180. Crain, D.A., Eriksen, M., Iguchi, T., Jobling, S., Laufer, H., LeBlanc, G.A., Guillette, L.J., 2007. An ecological assessment of bisphenol A: evidence from comparative biology. Reprod. Toxicol. 24, 225–239. DeFur P.L., Crane M., Ingersoll C., Tattersfield L. Endocrine disruption in invertebrates: Endocrinology, testing and assessment. Pensacola FL: SETAC Press. De Kermoysan, G., Joachim, S., Patrick, B., Lonjaret, M., Tebby, C., Lesaulnier, F., Lestremau, F., Chatellier, C., Akrour, Z., Pheron, E., Porcher, J.M., Pery, A.R.R., Beaudouin, R., 2013. Effects of bisphenol A on different trophic levels in a lotic experimental ecosystem. Aquat. Toxicol. 144–145, 186–198. Donker, M.H., 1992. Energy reserves and distribution of metals in populations of the isopod Porcellio scaber from metal contaminated sites. Funct. Ecol. 6, 445–454. Felten, V., Charmantier, G., Mons, R., Geffars, A., Rousselle, P., Coquery, M., Garric, J., Geffard, O., 2008. Physiological and behavioral responses of Gammarus pulex (Crustacea: Amphipoda) exposed to cadmium. Aquat. Toxicol. 86, 413–425. Fingerman, M., Hanumante, M.M., Fingerman, S.W., Reinschmidt, D.C., 1981. Effects of norepinephrine and norepinephrine agonists and antagonists on the melanophores of the fiddler crab Ulca pugilator. J. Crustac. Biol. 1, 16–27. Flint, S., Markle, T., Thompson, S., Wallace, E., 2012. Bisphenol A exposure, effects, and policy: a wildlife perspective. J. Environ. Manag. 104, 19–34. Fromme, H., Kuchler, T., Otto, T., Pilz, K., Muller, J., Wenzel, A., 2002. Occurrence of phthalates and bisphenol A and F in the environment. Water Res. 36, 1429–1438. Gollady, S.W., Webster, J.R., Benfield, E.F., 1983. Factors affecting food utilization by a leaf shredding aquatic insect: leaf species and conditioning time. Ecography 6 (2), 157–162. Graca, M.A.S., Maltby, L., Calow, P., 1993. Importance of fungi in the diet of Gammarus pulex and Asellus aquaticus. Oecologia 96, 304–309. Graca, M.A.S., Maltby, L., Calow, P., 1994. Comparative ecology of Gammarus pulex (L.) and Asellus aquaticus (L.) I: population dynamics and micro distribution. Hidrobiology 281, 155–162. Green, D.W.J., Williams, K.A., Huges, D.R.L., Shaik, G.A.R., Pascoe, D., 1988. Toxicity of phenol to Asellus aquaticus (L.)–effects of temperature and episodic exposure. Water Res. 22, 225–231. Hill, M., Stabile, C., Steffen, L.K., Hill, A., 2002. Toxic effects of endocrine disrupters on freshwater sponges: common developmental abnormalities. Environ. Pollut. 117, 295–300. ISO 10706. 2000. Water quality–Determination of long term toxicity of substances to Daphnia magna Straus (Cladocera, Crustacea). ISO 6341. 2013. Water quallity–Determination of the inhibition of the mobility of Daphnia magna Straus (Cladocera, Crustacea). Lachaise, F., Lerous, A., Hubert, M., Lafont, R., 1993. The molting gland of crustaceans – Localization, activity, and endocrine control (a review). J. Crustac. Biol. 13, 198–234. Lahnsteiner, F., Berger, B., Kletz, M., Weismann, T., 2005. Effects of Bisphenol A on maturation and quality of semen and egg in the brown trout, Salmo trutta f. fario. Aquat. Toxicol. 75, 213–224. Landrum, P.F., Robbins, J.A., 1990. Bioavailability of sediment-associated contaminants to benthic invertebrates In: Baudo, R., Giesy, J., Muntau, H. (Eds.), Sediments: Chemistry and Toxicity of in-place Pollutants. Lewis Publishers, Florida, pp. 237–263. Liu, C., Yan, W., Zhou, B., Guo, Y., Liu, H., Yu, H., et al., 2012. Characterization of a bystander effect induced by the endocrine-disrupting chemical 6-propyl-2thiouracil in zebrafish embryos. Aquat. Toxicol. 118–119, 108–115. McCahon, C.P., Barton, S.F., Pascoe, D., 1990. The toxicity of phenol to the freshwater crustacean Asellus aquaticus (L.) during episodic exposure–relationship between sub–lethal responses and body phenol concentrations. Arch. Environ. Contam. Toxicol. 19 986–929. Mihaich, E.M., Friderich, U., Caspers, N., Hall, A.T., Klečka, G.M., Dimond, S.S., Staples, C.A., Ortego, L.S., Hentges, S.G., 2009. Acute and chronic toxicity testing of bisphenol A with aquatic invertebrates and plants. Ecotoxicol. Environ. Saf. 72,
1392–1399. OECD 2004. Test guideline 218 Sediment-water Chironomid toxicity test using spiked sediment. Okada, H., Tokunaga, T., Liu, X.H., Takayanagi, S., Matsushima, A., Shimohigashi, Y., 2008. Direct evidence revealing structural elements essential for the high binding ability of bisphenol A to human estrogen-related receptor-gamma. Environ. Health Perspect. 116, 32–38. Opperhuizen, A., 1991. Bioconcentration and biomagnifications: is a distinction necessary In: Nagel, R., Loskill, R. (Eds.), Bioaccumulation in Aquatic Systems: Contributions to the Assessment. VCH, New York, pp. 43–66. Pascoe, D., Carroll, K., Karntanut, W., Watts, M.M., 2002. Toxicity of 17 alpha-ethinylestradiol and bisphenol A to the freshwater Cnidarian Hydra vulgaris. Arch. Environ. Contam. Toxicol. 43, 56–63. Peeters, E.T.H.M., De Jager, T.J., Beijer, J.A.J., Koelmans, A.A., 2000. Effects of benzo (a) pyrene and size of organic matter particles on bioaccumulation and growth of Asellus aquaticus. Arch. Environ. Contam. Toxicol. 39, 307–314. Petrović, M., Eljarrat, E., Lopez de Alda, M.J., Barcelo', D., 2004. Endocrine disrupting compounds and others emerging contaminants in the environment: a survey on new monitoring strategies and occurrence data. Anal. Bioanal. Chem. 378, 549–562. Plahuta, M., Tišler, T., Toman, M.J., Pintar, A., 2014. Efficiency of advanced oxidation processes in lowering bisphenol A toxicity and oestrogenic activity in aqueous samples. Arch. Ind. Hyg. Toxicol. 65, 77–87. Rodriguez, E.M., Medesani, D.A., Fingerman, M., 2007. Endocrine disruption in crustaceans due to pollutants: a review. Comp. Biochem. Physiol. 146, 661–671. Rubin, B.S., 2011. Bisphenol A: an endocrine disruptor with widespread exposure and multiple effects. J. Steroid Biochem. Mol. Biol. 127, 27–34. Qiao, P., Gobas, F.A.P.C., Farrell, A.P., 2000. Relative contributions of aqueous and dietary uptake of hydrophobic chemicals to the body burden in juvenile rainbow trout. Arch. Environ. Contam. Toxicol. 39, 369–377. Schug, T.T., Janesick, A., Blumberg, B., Heindel, J.J., 2011. Endocrine disrupting chemicals and disease susceptibility. J. Steroid Biochem. Mol. Biol. 127, 204–215. Staples, C.A., Dorn, P.B., Klecka, G.M., O'Block, S.T., Harris, L.R., 1998. A review of the environmental fate, effects, and exposures of bisphenol A. Chemosphere 36, 2149–2173. Staples, C.A., Tilghman, H.A., Friederich, U., Caspers, N., Klecka, G.M., 2011. Early lifestage and multigeneration toxicity study with bisphenol A and fathead minnows (Pimephales promelas). Ecotoxicol. Environ. Saf. 74 (6), 1548–1557. Staub, G.C., Fingerman, M.A., 1984a. . Mechanism of action for the inhibition of black pigment dispersion in the fiddler crab, Uca pugilator, by naphthalene. Comp. Biochem. Physiol. Part C: Toxicol. Pharmacol. 79, 447––453. Staub, G.C., Fingerman, M., 1984b. Effect of naphthalene on color changes of the sand fiddler crab, Uca pugilator. Comp. Biochem. Physiol. Part C: Toxicol. Pharmacol. 77 (1), 7–12. Toda, H., Arima, T., Takahashi, M., Inhimura, S., 1987. Physiological evaluation of temperature effect on the growth processes of the mysid, Neomysis intermedia Czerniawsky. J. Plankton Res. 9, 51–63. Van Brummelen, T.C., Stuijfzand, S.C., 1993. Effects of benzo (a) pyrene on survival, growth and energy reserves in the terrestrial isopods Oniscus asellus and Porcellio scaber. Sci. Total Environ. 134 (2), 921–930. Van Hattum, B., De Voogt, P., Van den Bosch, L., Van Straalen, N.M., Joosse, E.N., Govers, H., 1989. Bioaccumulation of cadmium by the freshwater isopod Asellus aquaticus (L.) from aqueous and dietary sources. Environ. Pollut. 62 (2–3)), 129–151. Vandenberg, L.N., Hauser, R., Marcus, M., Olea, N., Welshons, W.W., 2007. Human exposure to bisphenol A (BPA). Reprod. Toxicol. 24 (2), 137–177. Watts, M.M., Pascoe, D., Carroll, K., 2001. Survival and precopulatory behavior of Gammarus pulec (L.) exposed to two xenoestrogens. Water Res. 35 (10), 2347–2352. Watts, M.M., Pascoe, D., Carroll, K., 2003. Exposure to 17 alpha-ethinylestradion and bisphenol A-effects on larval moulting and mouthpart structure of Chironomus riparius. Ecotoxicol. Environ. Saf. 54, 207–215. Yamada, K., Urase, T., Matsuo, T., Suzuki, N., 1999. Constituents of organic pollutants in leachate from different types of landfill sites and their fate in the treatment process. J. Jpn. Soc. Water Environ. 22, 40–45. Yamamoto, T., Yasuhara, A., Shiraishi, H., Nakasugi, O., 2001. Bisphenol A in hazardous waste landfill leachate. Chemosphere 42, 415–418. Zou, E., 2010. Aquatic invertebrate endocrine disruption In: Breed, M.D., Moore, J. (Eds.), Encyclopedia of Animal Behavior. Academic Press, London, pp. 112–123. Zou, E., Fingerman, M., 1997a. Effects of estrogenic xenobiotics on molting of the water flea Daphnia magna. Ecotoxicol. Environ. Saf. 38, 281–285.. Zou, E., Fingerman, M., 1997b. Synthetic estrogenic agents do not interfere with sex differentiation but do inhibit molting of the cladoceran Daphnia magna. Bull. Environ. Contam. Toxicol. 58, 596–602.
