Sublethal effects of copper on some biological traits of the amphipod Gammarus aequicauda reared under laboratory conditions

Chemosphere 93 (2013) 1015–1022

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Sublethal effects of copper on some biological traits of the amphipod Gammarus aequicauda reared under laboratory conditions E. Prato ⇑, I. Parlapiano 1, F. Biandolino 1 Institute for Coastal Marine Environment, CNR – Via Roma, 3 74100 Taranto, Italy

h i g h l i g h t s  Copper (Cu) is common and widespread in coastal marine environment.  We investigated the sublethal concentrations of Cu to a marine amphipod.  Survival, growth and reproduction were chosen as endpoints. 1

 Adverse effects of Cu were detected at concentrations of 0.05 mgCu L

and 0.1 mgCu L1.

 Hence it is crucial to consider its sublethal toxic effect in the marine environment.

a r t i c l e

i n f o

Article history: Received 7 January 2013 Received in revised form 13 May 2013 Accepted 25 May 2013 Available online 22 June 2013 Keywords: Chronic test Copper Gammarus aequicauda Growth Reproduction

a b s t r a c t The common and widespread copper contamination in marine coastal environments make toxicity data necessary to assess the aquatic hazard and risk of this metal. In the present study, the sublethal effects of copper on survival, growth and reproduction of Gammarus aequicauda were investigated. Amphipods were exposed for 77 d to 2 nominal copper concentrations (50, 100 lg L1). Survival was the most sensitive measure of effect and was significantly reduced, especially during early life stage (juveniles). Growth of amphipods was also negatively affected by copper and the growth impairment in G. aequicauda increases with increasing metal concentration. The reproductive traits were impaired by each of the copper concentrations, even if there were not any significant differences between control and copper treatments. The size at maturity increased with increasing copper, so the smallest ovigerous females in the control and copper treatments were 0.83 mm and 1.35 mm head length, respectively. There was a positive correlation between the brood size and the body size of the female in all treatments, whilst the fecundity (n°juveniles/female) decreased in the order control, 50 and 100 lgCu L1. Copper demonstrates chronic toxicity to G. aequicauda at realistic environmental concentrations. The results of this study entail that the understanding of chronic toxicity of a substance, especially on population level effects, is crucial to assess the long-term effect of the substance in the ecosystem. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Urban and industrial activities in marine and estuarine coastal areas introduce significant amounts of pollutants. Heavy metals are of primary concern as they persist in the environment, moving up the food chain (Suedel et al., 1994), and could cause several disorders to wild life and humans (Tam and Wong, 2000). Among heavy metals, copper (Cu) is common and widespread in coastal marine environments (Hall and Anderson, 1999). The Cubased antifouling coating on ship hulls has been one of the major sources of Cu contamination (Valkirs et al., 2003). The mode of ⇑ Corresponding author. Tel.: +39 4542210; fax: +39 4542215. 1

E-mail address: [email protected] (E. Prato). Tel.: +39 4542210; fax: +39 4542215.

0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.05.071

toxic action of copper upon aquatic invertebrates is well known, though the ecotoxicological mode of action of Cu to aquatic organisms, and their Cu detoxification pathways vary among different species (Eisler, 1998; Stenersen, 2004; Kwok et al., 2008; Bao et al., 2013). This metal is essential for the maintenance of living organism metabolism, but beyond certain threshold levels can be extremely toxic and readily available to aquatic organisms. Indeed, copper, after mercury and silver, is the most toxic metal for marine phytoplankton (Sunda and Guillard, 1976), bivalve (His et al., 1999), echinoderms (Kobayashi, 1981) crustacean larvae (Young et al., 1979). Therefore, biomonitoring of this contaminant, by the use of toxicity tests, is considered an important tool for assessing the degree of contamination in coastal waters (Morillo et al., 2002).

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Although numerous studies have determined the acute lethal toxicity of copper in various invertebrate taxa, little is known about the effects of chronic exposure to this metal. Most of them consist of short-term exposures (acute toxicity tests) considering only survival as the endpoint (Traunspurger and Drews, 1996; Chapman and Wang, 2001; USEPA, 1999). These tests provide useful information for hazard assessments, however, they are not representative of typical environmental exposures (i.e., low concentrations over extended periods) nor do they account for the potential subtle sublethal responses (e.g., growth and reproduction) elicited by such exposures. Consequently, chronic toxicity tests, that more closely approximate low-level, long-term environmental exposure and that measure sublethal responses are being developed are important to characterize in assessing risk to aquatic ecosystems. However, the amount of chronic toxicity tests is limited and not routinely used to evaluate sediment toxicity in marine and estuarine sediments (USEPA-USACE, 2001). Both lethal and sublethal toxic effects can manifest themselves over the course of chronic exposure to toxicants, resulting in deleterious consequences not only for individual organisms, but also for a population as a whole. Development of chronic bioassays involves much greater scientific and technological effort than acute bioassays, which may explain why they are not used so often. Selection of test organisms is one of the critical topics. Few test organisms have appropriate characteristics, and not all species that are adequate for acute toxicity tests can be used in chronic testing. The present paper would provide information about the effects of sub-lethal concentration of copper on some biological traits of Gammarus aequicauda, which will have implications for future developments in ecotoxicological testing. This amphipod species was chosen as the test species for many reasons. Firstly, it is widely distributed along the Mediterranean and Black Sea. It is particularly abundant in the Mar Piccolo estuary (Ionian Sea, Southern Italy) and available in large numbers, supports the food chains of many predators and hence its susceptibility to a pollutant reflects on the whole ecosystem. Previous studies on its biology and life cycle in the field and laboratory have been investigated (Prato and Biandolino, 2003; Biandolino and Prato, 2006; Prato et al., 2006c, 2008a). Secondly, it is ecologically relevant, sensitive to contaminants, and ease of culture in the laboratory. In previous studies G. aequicauda was evaluated as test organism and fulfills most of the criteria required for suitable sediment toxicity tests (Prato and Biandolino, 2005), so has been used in a variety of ecotoxicological investigations (Prato et al., 2006a,b; Annicchiarico et al., 2007; Narracci et al., 2009; Prato and Biandolino, 2009a,b; Prato et al., 2009a). The objective of this study was to evaluate potential injury to G. aequicauda populations exposed to sub-lethal copper concentrations by evaluating some biological traits. For this purpose we addressed the following questions: (1) Does the copper affect the lifehistory traits? (2) Do the stress responses appear different at different copper concentrations?

2. Materials and methods 2.1. Collection and maintenance of amphipod Specimens of G. aequicauda were collected in April 2012 from the Mar Piccolo estuary (Ionian Sea, Italy; 40°290 1700 N; 17°140 2300 E), with a hand net (mesh size 750 lm), among the green macroalga Chaetomorpha linum and red algae Gracilaria sp., and were transported to the laboratory (15-min trip) in plastic containers filled with native seawater. The collection site is used as a

reference when conducting toxicity tests (Prato and Biandolino, 2009a,b; Prato et al., 2009a). In the laboratory, the amphipods, both males and females, were transferred to 30 L plastic aquaria with their native seawater and sediment. The culture aquaria were maintained at a temperature of 18 ± 2 °C and a salinity of 36‰ under a 12: 12-h light: dark regime in a climate-controlled room for 10 d to acclimate the individuals to experimental conditions and continuously aerated. These conditions were selected according to the annual ranges of the water variables registered in the Mar Piccolo. The culturing system was semi-static, with 100% water renewal twice a week. Food consisted of the macroalga C. linum fed ad libitum and Tetra-Min fish food. Small stones (about 10 cm2 in surface area) were used to mimic the natural environment and provide shelter. 2.2. Test substance Stock solution of copper was prepared by dissolving copper(II) sulfate pentahydrate (Cu-SO4 5H2O), in Milli-QTM water to obtain metals concentration of 1 g L1. The stock solution was further diluted in filtered seawater (0.2 lm filter fiber) in volumetric flasks to obtain working solutions at designated nominal concentrations before dosing. The test solutions were aerated for at least 24 h prior to the addition of animals to allow both the equilibration of the system. Sublethal concentrations assayed were chosen based on value of <96-h LC50 values estimated for G. aequicauda in previous studies (Prato et al., 2006b, 2007a, 2007b; Annicchiarico et al., 2007). 2.3. Chronic water-only toxicity tests After 10 d acclimation, ovigerous females with eggs at the same stage of maturation, were transferred into separate cultures, until the release of new-borns, after which the females were removed to avoid cannibalism. Two week later, the seawater was sieved through 275 lm mesh screen size to collect the newborns that will be used for chronic toxicity exposures. The experiments were started with 610 d old amphipods of a size class of 0.4 ± 0.03 mm head length. They were randomly stocked in groups of 80 into 3000-mL test aquaria filled with 2000-mL of test solution. Long-term responses were obtained by evaluating survival, sexual maturation, fecundity and growth in individuals treated with 50 and 100 lg L1 of copper, during 77 d. The experiments were conducted twice with three replicates per concentration per experiment. Each test was accompanied by a negative control (filtered seawater only), which measures the response of the amphipods in the absence of contaminant stress. The negative control furnishes an acceptability measure of the experiment, by providing evidence of the test organism’s health, test conditions and handling procedures. It is also used for statistical comparison with the test solutions. Culture conditions that are optimal for the continuous culture of G. aequicauda in the laboratory have been investigated and described previously (Prato et al., 2006c, 2008a). In order to ensure adequate oxygen levels for the duration of the toxicity tests, aquaria were gently aerated by attaching an slight air line to the end of a Pasteur pipette which produced no or minimal disturbance to organisms. The aquaria were covered with polypropylene lids with small holes punched for insertion of glass rods for aeration. Water quality parameters in the chronic toxicity test cultures were kept constant throughout the experiment, since they can affect the toxicity of some trace metals (temperature 18 ± 2 °C, salinity of 36‰, pH 8 ± 0.4, dissolved oxygen 8.3 ± 0.3 mg L1) (Prato et al., 2007c, 2008b, 2009a).

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Additional food was added during the week as required while a reduction of food was necessary if uneaten food accumulates and starts to rot. The test solution was completely renewed twice per week. Surviving animals from each replicate were removed sieving through a 325 lm mesh, and carefully transferred into fresh chamber test with renewed copper solution (prepared as described previously). These transfers minimized algal growth and accumulation of waste metabolic products such as ammonia. Water quality measurements (dissolved oxygen and pH) were determined three times per week, both before and after test solution renewal, whilst the temperature was recorded at 24-h intervals. At 7 d intervals, pools of 10 animals from each replicate were removed, and carefully manipulated for the following measures collection: survival, head length (mm), number of antennal segments, sexual maturation time, precopula duration, percentage of gravid females, fecundity and embryonic development time. 2.4. Growth Head length (HL) was used instead of the total length (TL), due the difficulty of measuring an amphipod’s recurved body. Head length was measured, to the nearest 0.02 mm, from the anterior end of the rostrum to the posterior margin of the cephalon. The relation between TL and HL was previously determined in a sample of 200 animals, and it can be described by the following linear regression equation:

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cannibalism by the female. After newborns releasing, the female head length was measured to the nearest 0.02 mm. Maturity was calculated as the number of ovigerous females in the population and fecundity as the number of newborns released by a female in each replicate. The partial fecundity index (PF), defined as mean brood size/ mean breeding female size and relative size when reaching maturity index (Mind), defined as minimal/mean breeding female size, were determined. 2.6. Statistical analysis All data were tested for normality (Kolmogorov–Smirnov’s test) and homogeneity (Bartlett’s test). Two-way analysis of variance (two-way ANOVA) was employed to test the effect of experimental treatments and time on G. aequicauda survival, growth and fecundity. Treatments were regarded as statistically significant at P < 0.05. Tukey test was used for a posteriori comparisons of means with control. All statistical analyses were performed using SPSS software (Version 12.0, Chicago, IL). Consequently, a 1-way ANOVA was used to confirm the significant difference between the treatments. A posteriori Tukey test showed that the difference between each treatment was significant at the P = 0.05 level. In order to test the relationship between the number of juveniles and the length of the female, regression analysis was performed (StatgraphicsÒ, version 2.1).

TL ¼ 0:2248 þ 8:4211HL; R ¼ 0:92: The segments of the 1st pair of antennae were counted on the primary flagellae with a binocular microscope. After measurement, living individuals were returned to the respective aquarium. The von Bertalanffy growth model was used to estimate growth rates for the whole life cycle using FISAT II program package (Gayanilo et al., 2005):

HLt ¼ HL1 ð1  ekðtt0 Þ Þ where HL1 is the asymptotic length parameter (mm), k is the intrinsic growth rate, and t0 is a parameter representing the theoretical age (in years) at which Lt = 0. 2.5. Reproductive traits Sexual dimorphism was determined by the size and shape of the gnathopods. Male sexual maturation was determined by the appearance of an enlarged propodus on the 2nd gnathopod; female sexual maturation was determined by the appearance, presence, and condition of the oostegites. Oostegites of immature females lacked setae; mature females had fully developed, setose oostegites, and ovigerous females had eggs in the brood pouch. Animals without these characteristics were considered to be juveniles. The pair of amphipods in precopula, in which the male holds and carries the female, were counted, removed from the aquaria and placed individually inside a 0.50 L glass beaker with 0.25 L of the correspondent copper solution at the same conditions. They were observed daily until the female separated from the male and was determined to be ovigerous. When separation occurred, the male was measured to the nearest 0.02 mm and returned in the 1L beaker, whilst the ovigerous female was monitored daily in order to determine how many days are required for embryonic development and how many newborns each female could produce under these experimental conditions. The embryonic development time was recorded as the period from oviposition to release of juveniles from the brood pouch. Newborns were removed to avoid

3. Results 3.1. Survival Results of the experiments examining the impact of Cu on survival of G. aequicauda are shown in Fig. 1. Control individuals displayed a survival below 50% after 49th d of exposure that continued to decrease, reaching 10% at the end of the test period. The survival of the organisms exposed to 50 lgCu L1 was dropped by 50% to the 42th d of exposure. Animal exposed to 100 lgCu L1 responded similarly to control up to 14th d after which survival was significantly reduced below 50% after approximately 25th d of exposure. From the 28th to 42th d survival was constant, then started to decline until it reached a survival of 12% at the end of the chronic exposure. The concentration of 50 lgCu L1 showed the highest survival percentage (16%) at the end of the experiment. Two-way ANOVA showed that there were significant concentration (F = 24.81, P < 0.001, df = 2) and time effects (F = 192.33, P < 0.001, df = 10), and significant interaction between these variables (F = 4.86, P < 0.001, df = 20). A one-way ANOVA confirmed the significant difference between the treatments and an a posteriori Tukey test showed that the difference between each treatment was significant at the P = 0.05 level (Fig. 1). 3.2. Growth The experiment permitted the observation of G. aequicauda growth after exposure to negative control and two concentrations of copper for 77 d. Initial average head length of the specimens were 0.4 ± 0.03 mm (Fig. 2). In general mean head length in individuals exposed to 50 and 100 lgCu L1 followed a similar pattern to the control group. However, the marked increase in head length in the control treatment between 28 and 35 d was not observed in the both copper treatments. Instead, a marked and rapid increase in mean head length

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100

a a

90

b

80

a

CTR

b

Survival %

0.1 mg/L

a

70

b

60

0.05 mg/L

a a,b

50

a,b a b

b

a

40

a

a

a b aa a aa a

30 20

aaa a

10

b

a,b

0 7

14

21

28

35

42

49

56

63

70

77

Days Fig. 1. Survival (%) of G. aequicauda in control and copper treatments during the exposure time. a, b letters denote significant differences (p < 0.05) between groups.

was recorded between 35 and 42 d in all control and treated animals (more than double compared to the previous week). After this time it was observed that the growth was not significantly different among treatment at the same times. The maximum head length attained by individuals under experimental conditions was 2.35 mm in the control (corresponding to body size of 20 mm), 2.13 mm at 50 lgCu L1 (corresponding to body size of 18.16 mm) and 2.12 mm at 100 lgCu L1 (corresponding to body size of 18.07 mm). Two-way ANOVA showed a significant difference between the treatments (F = 26.366, P < 0.001, df = 2) and time effects (F = 712.97, P < 0.001, df = 10) and significant interaction between these variables (F = 2.34, P < 0.001, df = 20). However, an a posteriori Tukey test indicated that significant differences occurred only between the control and both copper treatments at 35 d. The von Bertalanffy growth model, calibrated with data obtained from this study, showed that the growth was continuous throughout the study period in the control and copper treatments (Fig. 3). The highest growth rate was observed in the control and the lowest in organisms exposed to 100 lgCu L1. Indeed, as can be observed, the growth parameters were different for each of the studied treatments. The average size (HL1) and the growth coefficient (k) were significantly greater in control animals than to those of animals exposed to 50 and 100 lg L1 of copper (Table 1). Furthermore, there was a strong correlation between mean head length (mm) and the mean number of antennular flagellum segments in the control animals (R2 = 0.68, P < 0.01) and in 50 and 100 lgCu L1 (R2 = 0.72 and 0.80 respectively, P < 0.01).

3.3. Fecundity The sex of males and females could be confirmed at about 25 d for control individuals and at 30 d and 31 d for 50 and 100 lgCu L1, respectively (Table 2). For all experimental groups (CTR and copper treatments) first precopulatory pairing was observed  a week later. Size at maturity was smaller in control individuals than those exposed to copper (Table 2). The exposure to sublethal concentrations of copper did not significantly affect the average time spent by G. aequicauda male and female in precopula that was 36.63 h in control and 32.84 h in both copper treatments. During observations it was revealed that mating was based on a preference to size with a positive correlation occurring between male and female size: in all cases larger males preferred larger females (Fig. 4). There was no significant difference in the embryonic development time between the control and copper treatments as shown by one-way ANOVA (F = 1.58, P = 0.207, df = 2). In particular hatching was observed after 7.31 ± 1.25 d in the control and after 7.48 ± 1.27 d and 7.18 ± 1.25 d in 50 and 100 lgCu L1 groups, respectively. In relation to the number of precopula pairs the statistical analysis showed no significant difference between the treatments (ANOVA, F = 2.376, P > 0.05, df = 2) and time effects (F = 0.669, P > 0.05, df = 10) and no significant interaction between these variables (F = 1.521, P > 0.05, df = 20). However, the highest number of precopula pairs in the control experiments was observed at 35–42 d, while in the organisms

Fig. 2. Head lenght (mean ± s.d.) of G. aequicauda in control and copper treatments, during the exposure time.

Fig. 3. Von Bertalanffy growth models calibrated with laboratory data to estimate growth of G. aequicauda in the control and copper treatments during the study period.

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per brood and female head length (HL) was statistically significant in all treatments (P < 0.01) (Table 3). For each experimental treatment, the number of new-borns released by females belonging at different HL size classes, is shown in Fig. 7. The largest number of newborns was 51 and released by a female with a head length of 1.87 mm (corresponding to 16.01 mm in total length) exposed to 50 lgCu L1. The number of new-borns per female did not differ significantly between control and the copper treatments, however, showed a concentration-dependent reduction. The brood size (mean, maximum and minimum) and partial fecundity were reported in Table 4. The mean number of juveniles released by females ranging between 1.6 and 1.8 mm head length decreased from 14.22 eggs/female in control to 11.44 eggs/female in animals treated with 50 lgCu L1, to 10.28 ggs/female in the group exposed to 100 lgCu L1. The index of fecundity, calculated to account for possible size differences in breeding females, as in Ladewig et al. (2006) showed the highest values in the control experimental samples (7.92).

4. Discussion Chronic exposure to metals may be frequent in natural environments, therefore when considering any invertebrate species as a bioindicator or metal biomonitor, it is necessary to have a good knowledge of its taxonomy distribution and life history characteristics (Rainbow, 1995). This study provides a first assessment of the performance of G. aequicauda under sub lethal concentration of copper in the laboratory. The effects reported here have been observed at environmentally realistic concentrations registered in the Mar Piccolo basin, since the copper concentrations generally range from 6 lg g1 dw to 150 lg g1 dw, with the highest concentrations a result of significant anthropogenic inputs. The sampling area showed a copper concentrations in a range of 30–35 lg g1 dw in the sediment (Annicchiarico et al., 2007; Prato, unpublished data), while literature data did not exist regarding the water copper concentrations. The acute toxicity of Cu for G. aequicauda has been calculated to be 0.70 lgCu L1 (Prato, unpublished data); however, sublethal long-term copper exposure (concentrations 7–14 lower than the LC50) reduced the survivorship and reproductive performance of this amphipod. Exposure to levels as low as 50–100 lgCu L1 significantly reduced survival of G. aequicauda, especially during early life stage (juveniles), which are shown to be the most sensitive (Prato and

Fig. 4. Relationship between HL of males and females engaged in precopula in all experiments (CTR and copper treatments).

exposed to 50 and 100 lgCu L1 at 43–50 and 59–62, respectively (Fig. 5). The maturity index was lower in the control (0.52), than to copper treatments with 0.83 and 0.79 in 50 and 100 lgCu L1, respectively. The highest number of ovigerous females was found in the range size of 1.6–1.75 mm HL in the control and 50 lgCu L1, while 100 lgCu L1 showed a more wide range between 1.6 and 1.9 mm HL (Fig. 6). The relationship between the number of juveniles (NJs)

Total number precopula pairs

25

CTR 0.05mgCu/L 0.1mgCu/L

20 15 10 5

75-78

71-74

67-70

63-66

59-62

55-58

51-54

47-50

43-46

39-42

35-38

31-34

0

Days Fig. 5. Total number of precopula pairs in the control and copper treatments during the study period.

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Ovigerous females (%)

100% 80% 0.1mgCu/L

60%

0.05 mgCu/L

40%

CTR

20%

2.05-2.2

1.90-2.050

1.75-1.9

1.6-1.75

1.45-1.6

1.3-1.45

1.15-1.3

1.1-1.150

0.95-1.1

0.8-0.95

0%

HL size classes (mm) Fig. 6. Percentage (%) of ovigerous females in control and copper treatments at different HL size classes.

Biandolino, 2005). Therefore in long term exposure, changes in survival would represent a reasonable estimate of the copper chronic effect on G. aequicauda. The determination of the survival endpoint requires less time and fewer costs than the determination of most other endpoints, and it appears adequate for hazard assessments in the initial stage of estimating chronic toxicity. Growth rate is frequently used as a sublethal endpoint in ecotoxicological studies as it is usually reduced by environmental stressors (Maund et al., 1992; Moore and Dillon, 1993; Conradi and Depledge, 1998). To better understand the impact of Cu on the G. aequicauda’s population, chronic exposure with a longer experimental duration (77 d) has been conducted, as a longer period can allow us to assess different biological traits and make use of such data to obtain information on the fitness of the population. In this study results revealed that the exposure to sublethal copper concentrations determined on specimens a lower growth than that observed in the control, but only in a significant way around 35th d. In addition growth impairment in G. aequicauda increases with increasing copper concentration. This impairment in growth could be due to a reduction in the amount of metabolic energy available in the animal and could be explained by the depletion of energy reserves in one or more tissues. Copper is a superoxide generator, and can induce oxidative stress in tissues of aquatic organisms. It has ability to induce the formation of hydrogen peroxide (H2O2), which might be transformed to hydroxyl radical OH (Stenersen, 2004), that is extremely reactive and modifies all biomolecules in its vicinity, including lipids, proteins and DNA (Eisler, 1998; Stenersen, 2004).

Number of juveniles released

30 CTR

25

0.05 mgCu/L

0.1mgCu/L

20 15 10 5

2.05-2.2

1.9-2.05

1.75-1.9

1.6-1.75

1.45-1.6

1.3-1.45

1.15-1.3

1.1-1.15

0.95-1.1

0.8-0.95

0

HL female classes (mm) Fig. 7. Mean number (±s.d.) of juveniles per female at different HL size classes, in control and copper treatments.

The severe toxic effect on survival was not apparent in growth of surviving animals. However, caution must be taken when interpreting growth data from chronic tests when there were significant reductions in survival (Sibley et al., 1997; Ingersoll et al., 1998; Green et al., 1999). One main concern is that growth may have been affected by the reduction of organism density on the course of the test. Another concern, particularly relevant in tests with gammarids, is cannibalism. Borgmann et al. (1990) reported no detectable copper effects on the growth of amphipoda Hyalella azteca, where no significant reduction in growth occurred at any concentration that did not cause significant chronic mortality. A most relevant aspect of the environmental impact of heavy metal may be related to its effects on the reproductive traits. It has been observed that growth responses in G. aequicauda (and as well in other amphipods/invertebrates), have a direct reflection in reproductive performance, since female maturation and brood size are a function of growth (Correia et al., 2001; Neuparth et al., 2002). A variety of contaminants have been shown to affect sexual maturity, fecundity, brood size and embryonic development in amphipods, including metals (Sundelin, 1989; Conradi and Depledge, 1998; Lawrence and Poulter, 2001; Zulkosky et al., 2002). Based on the results of this study the reproductive traits of G. aequicauda were impaired by each of the copper concentrations, even if were not found significant differences between control and copper treatments. The copper sub-lethal exposure showed a lengthened the time required to reach maturity. Gentile et al. (1983) reported for Mysidopsis bahia that exposure to high levels of heavy metal such as mercury, significantly lengthened the time required for mysids to reach maturity, which was expressed as delay in the appearance of eggs in the brood pouch and the release of young. However, Khan et al. (1992) asserted that sexual maturity in which they described the presence of gonads or a brood pouch, a feasible endpoint for reproductive tests, because gonad maturation is essentially the first step toward reproductive output. Most contaminant effects are likely to lengthen the time to release of the first brood. Lussier et al. (1985) found that several metals (mercury, zinc, nickel) significantly increased the time to first brood release, whereas others (e.g., cadmium, copper, silver) did not. In this study the first brood in both copper treatments showed a delay of about a week than to control group. The disruption of precopula pairs is frequently used as a toxicological endpoint. In this study the precopula ranged from 37 h in the control to 33 h in both copper treatment groups. Pascoe et al. (1994) and Poulton and Pascoe (1990) described a behavioral bioassay, with Gammarus pulex, using the disruption of precopulatory pairing as a signal of the presence of copper and cadmium, respectively. In both studies, the mean induced separation time decreased with increasing of metals concentrations. It can be suggested that precopula disruption is a very sensitive and rapid method of detecting environmental stress for a wide range of pollutants, both in the laboratory and in the field. The embryonic developmental time of G. aequicauda, in this study, was also not influenced, according to Gentile et al. (1983) for M. bahia exoposed to mercury. However, Kwok et al. (2008) reported a significantly lengthen of the embryonic developmental time at 100 lgCu L1 for copepod Tigriopus japonicus. The most likely effect of contaminants is a reduction in fecundity (Lussier et al., 1985; Gale et al., 2006; Ringenary et al., 2007) which in some cases is the only response to contaminant exposure (Lussier et al., 1999). Although no apparent differences were found in the mean brood size between the two copper concentrations and control, the mean brood size separated in a decreasing order from control > 50 > 100 suggesting that fecundity may be affected by

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copper. The lowest mean brood size at the highest copper concentration may be a result of the copper effect on embryonic survival at sensitive life stages through determining passive embryo loss. In addition it should be highlighted that brood size increases with body size among female amphipods (Steele and Steele, 1991). A linear correlation between mean brood and female size is a common feature in amphipods (Beare and Moore, 1996; Costa and Costa, 1999; Persson, 1999; Cunha et al., 2000; Maranhão and Marques, 2003), as observed in this study for all treatments. Larger females generally carry more eggs than smaller ones because of the greater body length, as the marsupium capacity is proportional to body size (Sheader, 1977). Therefore, reduced fecundity in response to exposure to toxicant needs to be carefully evaluated comparing with control organisms, to distinguish direct interruption of reproductive processes from a simple reduction in growth. In stress conditions amphipods may divert more energy to growth than reproduction and may abort broods (Maltby and Naylor, 1990). In Corophium volutator from Sweden, there was reduced egg-production in females exposed to dissolved copper concentrations as low as 50 lgCu L1 (Eriksson and Weeks, 1994). Conradi and Deplege (1998, 1999) using Cu and Zinc-doses sediments, have shown that this metals reduced the maximal body size and decreased amphipod fertility. Understanding the potential impacts of this contaminant on invertebrate such as G. aequicauda is crucial because of the ecological implications of such effects. The metal studied produced a reduction of survival, growth and fecundity, lengthened of time reaching sexual maturity, shortened of time spent in precopula. Many estuarine invertebrates, including amphipods, provide an important food source for juvenile commercial fish and birds. Costa and Elliott (1991) have highlighted the central role of small epibenthic crustaceans, including amphipods and shrimp, as the main link between detritus and fishes. Indeed, in the specific case of the Mar Piccolo basin this gammarids is an important prey within the foodweb, occurring in the diet of most fish species (Biandolino and Prato, 2006; Prato and Biandolino, 2009a,b). Consequently, any impact on the survival, reproduction and fecundity of this species may have significant effects on estuarine food chains and secondary production.

5. Conclusion The effective protection of aquatic populations from the potential impact of contaminants requires consideration of effects on individual species, survival, growth and reproduction, since these are fundamental in determining the fitness of populations in natural ecosystems. The results of this study provide further evidence to support the use of G. aequicauda in ecotoxicology studies particularly in estuarine areas where the choice of appropriate species is limited. The main limitation with the G. aequicauda chronic test is the length of time required for the test species to reach maturity and reproduce i.e. 35 d in this study compared with 28 d for both the Leptocheirus plumulosus (USEPA, 2001) and the Gammarus locusta (Neuparth et al., 2005) chronic tests. It may also prove useful to conduct partial lifecycle tests by initiating the test using juvenile of 15 d; thus providing a shorter test (28 d) but retaining reproduction traits. Laboratory tests are important in ecological assessment, but care must be taken when we extrapolate conclusions from laboratory to field conditions (Sasson-Brickson and Burton, 1991). In the field, invertebrates are exposed to many sources of stress, which could act separately or synergistically, while in the laboratory experiments are carried out at controlled conditions.

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Acknowledgement The authors are grateful to Dr. Jonathan Poole for his technical support in the English revision.

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