Ecotoxicology

, Volume 21, Issue 3, pp 688–697 | Cite as

Toxic effects of pentachlorophenol, azinphos-methyl and chlorpyrifos on the development of Paracentrotus lividus embryos

  • Silvia Buono
  • Sonia Manzo
  • Giovanna Maria
  • Giovanni Sansone
Article

Abstract

The application of many current-use pesticides has increased after the disuse of persistent, bioaccumulative or toxic ones as DDT or chlordane. Many of the used pesticides are considered less dangerous towards the environment for their physico-chemical properties. This study investigated the toxic effects of three current-use pesticides, pentachlorophenol (PCP), azinphos-methyl (AZM), and chlorpyrifos, on Mediterranean sea urchin Paracentrotuslividus early development and offspring quality. The experimental results showed that the most toxic pesticides were PCP and AZM at EC50 level. Nevertheless at low concentration PCP resulted the less toxic compound and showed EC1 value more protective than NOEC. PCP at high concentration seemed to modify cytoskeleton assembly, while at low concentrations, it could alter the deposition of the larval skeleton. OPs at low concentrations until 300 μg/l showed a similar toxicological behaviour with a trend corresponding to the pesticide concentrations. At high concentration (500 μg/l) the effect mainly observed was the embryos pre-larval arrest. This investigation highlighted the relevance to evaluate, in coastal seawaters, the levels of the used pesticides to understand the real impact on benthic populations mainly in sites characterized by intensive agriculture or floriculture activities, such as the coastal areas of the Mediterranean Sea.

Keywords

Paracentrotus lividus Pentachlorophenol Azinphos-methyl Chlorpyrifos 

Introduction

Current-use pesticide applications have increased after the disuse of persistent, bioaccumulative or toxic biocides such as the organochlorine insecticides DDT (1,1,1-trichloro-2,2-bis (4-chlorophenyl) ethane) or chlordane (octachloro-4,7-methanohydroindane). Actually the used pesticides are considered less dangerous towards environment because of their physico-chemical properties such as short half-life or decreased potential for bioaccumulation due to lower octanol/water partitioning coefficients (Kow).

Pesticides can be introduced into the aquatic environment by drift, surface run-off, leaching from soil, accidental spills, and atmospheric deposition. Moreover some pesticides, similar to the legacy pesticides, have the potential for long range transport (Muir et al. 2004).

Despite the concentrations of pesticides found in the environment are much lower than those which cause direct lethality in non-target aquatic species, their sub-lethal effects are only partly known (Shelley et al. 2009).

Chronic exposure to xenobiotics for some key species could cause a significant impact at the ecosystem since they may be affected in reproductive ability, behaviour, growth, homeostasis, or susceptibility to diseases (LeBlanc and Bain 1997; Spromberg and Meador 2006). Therefore these substances, introduced into the marine environment, represent a threat to non-target marine species (Bellas et al. 2005).

We investigated the toxic effect on Paracentrotuslividus embryos of three pesticides: two Organo Phosphorus compounds (OPs: AZM and CPF) and one organochlorurate (PCP).

Pentachlorophenol (PCP), an uncoupler of mitochondrial oxidative phosphorylation, acts by destroying the electrochemical potential across the inner membrane of mitochondria (Martello et al. 1998; Mitchell and Moyte 1967; Zha et al. 2006). Formerly, in Europe PCP was extensively used as a biocide in the protection of timber and textiles, consequently it presence was often detected in the air, water and soil (Muir and Eduljee 1999).

Actually, since 1980s, the concern about the toxicity of PCP and the potential adverse effects on human being and the environment led to a regulatory action to limit its use.

In the EU the use of PCP, its salts and esters is currently limited to two industrial applications: wood preservation (91 173 EEC, which includes sapstain control) and the impregnation of heavy duty textiles. The reduction of the organochlorurate pesticides use induced an increase in the consumption of organophosphate and carbamates; these compounds are more dangerous to the environment respect to the organochlorurate, but they are faster degradable (Nimmo and McEwen 1994).

Azinphos-methyl (AZM) is a broad spectrum organophosphate insecticide that, like all OP insecticides, exerts its insecticidal properties acting as acetylcholinesterase (AChE) inhibitor. In the US, it is registered for use on select nut trees, vegetable crops, and fruit trees. However the U.S. Environmental Protection Agency (EPA) considered a denial of reregistration, citing, “concern to farm workers, pesticide applicators, and aquatic ecosystems” (EPA 2009). In EU, AZM has been banned since 2006 (Scott 2008). The New Zealand Environmental Risk Management Authority established to phase out AZM over a 5 year period starting from 2009 (ERMA 2009).

Chlorpyrifos (CPF) is manufactured by reacting 3,5,6-trichloro-2-pyridinol with diethylthiophosphoryl chloride (Muller 2000). In the US, CPF is registered only for agricultural use, where it is “one of the most widely used organophosphate insecticides,” as reported by EPA (EPA 2002).

These compounds showed a half life ranging from 15 days to several weeks, according to the pH, oxygen availability and other parameters, including the nature of the substrate (Aluigi and Falugi 2010). Actually no data are available about the presence of OPs and PCP in sea water due to their short persistence into the environment.

Despite many compounds, like pesticides, are often present in the environment at concentrations far below their individual median toxic Effect Concentration 50% (EC50), and also below their individual no observed effect concentration (NOEC), they could give rise to substantial consequences. NOEC values are usually derived from experimental data by applying statistical hypothesis testing procedures such as Dunnett’s (1964), and, therefore, they only denote the highest test concentration at, and below which, the response of exposed organisms does not depart significantly (in a statistical sense) from untreated controls (Skalski 1981).

In addition, NOEC values derived from standard toxicity tests have been typically shown to correspond to effects ≥10% (Moore and Caux 1997).

The use of regression-base statistical estimates of low-effect concentrations, Concentration causing an Effect of x% (ECx) estimations (Van der Hoeven et al. 1997), which are discussed to replace the NOEC in risk assessment procedures (Moore and Caux 1997; Van der Hoeven et al. 1997), could overcome this problem.

On the other hand, Shief et al. (2001) showed that the endpoint response and the nature of the toxicity test could be an important consideration for the selection of NOEC or ECx.

In this study the EC1 value was also considered in order to determine the pesticide concentration that induce the minimum effect (1%) statistically detectable and therefore to identify the more protective parameter for sea urchin embryos.

Sea urchin embryos and gametes are often utilized to assess the toxicity of chemical compounds in the marine ecosystem (Pagano et al. 1986; Kobayashi and Okamura 2002; Manzo et al. 2008) due to their availability and sensitivity in the short and medium time. The sea urchins toxicity test has been utilized for several decades to evaluate the toxicity of some xenobiotics and their future in the marine ecosystem (Bay et al. 1993; His et al. 1999; Kobayashi and Okamura 2002; Manzo 2004; Marin et al. 2000; Pagano et al. 1996a, 1996b). P. lividus is one of the most commonly used organisms in biomonitoring studies, which require simple, rapid, and inexpensive but sensitive methods (Kobayashi 1991; Manzo and Torricelli 2000; Pagano et al. 1989). In particular P. lividus early life stages are very sensitive to many pollutants (His et al. 1999; Ringwood Huffman 1992).

Many studies have demonstrated the sensitivity of sea urchin embryos to single pesticides as pure substances (Aluigi and Falugi 2010; Aluigi et al. 2008; Bellas et al. 2005; Buznikov et al. 2007; Pesando et al. 2003). However effects on sea urchin early development and offspring quality are quite unknown respect to freshwater toxicology based on the use of Daphnia (Hutchinson et al. 1998; Leung et al. 2001). More toxicological data are needed for a proper assess of environmental risk posed by pesticides, and to implement sea water quality standards protective for marine organisms. The aim of this study was to determine the toxic effects of three current-use pesticides—one organochlorurate (PCP) and two organophosphates (AZM and CPF)—on Mediterranean sea urchin P.lividus early development and offspring quality. For each pesticide EC50, EC1, NOEC, and lowest observed effect concentration (LOEC) were defined to provide biological criteria for the implementation of water quality standards to protect marine organisms. In the present investigation the EC50 was evaluated with two different methods (ICp-methods and best fit procedure) in order to compare them and to identify the more protective one for sea urchin embryos.

Materials and methods

Test organisms

Adult P. lividus (Lamark) were collected from the Tyrrhenian Sea (Bay of Naples) by the staff of the Zoological Station of Naples (Italy). Sea urchins were then acclimatized for 24 h in natural Filtered Sea Water (FSW 0.45 μm Ø) at 18 ± 1°C (salinity 38%, pH 8 ± 0.2). In fact, the use of the animals immediately after their collection produced a decrease of normal plutei in the control, probably due to stress induced by the collection activity itself. An abrupt increase in temperature or salinity might not only induce spawning, but more seriously harm the gametes (ASTM 2004). Besides, we observed (data not shown) that the permanence in aquarium after collection, could provoke remarkable sea urchin mortality and a substantial reduction of gamete quality.

Toxicity test

Gametes were harvested and embryos were reared according to Pagano et al. (1986). Spawning was induced in sea urchins by injection of 1 ml of 0.5 M KCl through the perioral membrane. Eggs were collected by separately placing each spawning female in a different 250 ml beaker with FSW, while ‘‘dry’’ sperm from each male was collected with an automatic pipette and stored in a sterile tube placed on ice. For each experiment, six female individuals were selected for their appropriate egg quality (no immature forms, no debris, and no fertilized eggs) and high amount. Males were selected for sperm motility (checked under the microscope) and amount. Then the gametes of the best three males and three females were pooled and filtered through nylon cheesecloth (Ø = 200 μm for eggs and 50 μm for sperm). The egg suspension was diluted in order to obtain the final density of 250–300 eggs/ml.

In the embryotoxicity protocols, fertilization was carried out by adding 1 ml of pooled-sperm, diluted 1:1,000 in FSW, to the egg suspension and by incubating if at 18°C for 20 min. Fertilization success in the stock solution was verified by the presence of the fertilization membrane in a random sample of 100 eggs.

Zygotes were then employed in embryotoxicity test (T = 18 ± 1°C, exposure time = 48–50 h) (Pagano et al. 1996a, 1996b modified). The experiments were carried out at least six times, where the control and each pesticide treatment were carried out at least in triplicate.

The utilized embryotoxicity test procedure has been previously described in Manzo et al. (2006).

Developmental abnormalities were determined after an exposition time of 48–50 h in each replicate by direct observation. For each treatment schedule, 100 plutei were scored for the frequencies of: (1) normal (N) larvae, according to their symmetry, shape, and size; (2) retarded (R) larvae with shape and symmetry the same as normal, but with reduced size (<1/2 with respect to N); (3) malformed larvae (P1), affected in skeletal and/or gut differentiation and/or pigmentation; and (4) pre-larval arrest (P2), embryos unable to go to larval differentiation, as abnormal blastula or gastrulae (Pagano et al. 2001). The viability of embryos (P2, P1, R, and N) was evaluated at microscope observation. Mean percentage abnormalities and 95% confidence limits were calculated for all samples and compared to the results obtained from the controls. If abnormalities in the controls were 20% or more, the test was considered invalid and repeated. However, to evaluate the test’s reproducibility a positive control was carried out with a reference toxicant (Cu) (Arizzi Novelli et al. 2002; Volpi Ghirardini and Arizzi Novelli 2001).

Test Solutions

Pentachlorophenol (C6Cl5OH, purity > 99.50%), CPF (C9H11Cl3NO3PS, purity > 99.00%), and AZM (C10H12N3O3PS2, purity > 99.00%) were purchased from ALDRICH company (US).

Pentachlorophenol, environmentally persistent fungicide, has a log Kow of 5.05 (Kaiser and Valdmanis 1981) and a water solubility of 10–20 mg/l (Arcand et al. 1995). AZM has a log Kow of 2.96 (Tomlin 1994) and a water solubility of 28 mg/l (20°C—Tomlin 1994; Worthing 1991). The solubility of CPF in water is 1.4 mg/l at 25°C, and it has a log Kow value of 4.7 (Tomlin 2006).

The pesticides stock solutions were prepared by dissolving reagent grade in a non toxic concentration of organic dissolvent, DMSO (Sciarrino and Matranga 1995).

Test solutions were obtained by diluting the stock solution in FSW. In this study the exposure concentrations in this study ranged between 0.1 a 500 μg/l for each pesticide. Experimental concentrations were chosen on the basis of preliminary trials and on literature data (Foster et al. 1998; Hartgers et al. 1999; His and Seaman 1993; Key and Fulton 1993; Mansueto et al. 1993; Whiting et al. 1996). In the preparation of test solutions the final dilution factor of FSW was always maintained at 10%, approximately corresponding to a final salinity of 34‰ that is near to the value selected by His et al. (1999) for their tests on fertilized eggs of P. lividus and it is well within the range of salinity “tolerance” (33–38‰) for sea urchin embryos and larvae (Bressan et al. 1995). In previous experiments it was also verified that this procedure did not affect the results of the tests.

Seawater used for the test solutions (and for acclimatization) was sampled in an uncontaminated area far from the coast and it was routinely used in the laboratory for ecotoxicological tests and optimization of analytical methods. As a consequence, seawater samples from this area were analyzed several times for trace elements and organic micropollutants using wide-spectrum-screening analytical methods.

It was also carried out a control with DMSO at a maximum concentration utilized in the test solutions and exhibited no observable effects on the studied organisms.

Statistical analysis

Differences in development success (comparisons between the control group and each of the experimental groups) were tested for significance using the multiple comparisons Dunnett’s test.

The EC50 and EC1 were calculated using the linear interpolation method (inhibition concentration procedure or ICp) (Cesar et al. 2004; US EPA 1993). The bootstrap method is used to obtain the 95% confidence interval, because standard statistical methods for confidence intervals calculations are not applicable. Analysis of variance (ANOVA) was applied, using raw data, in order to test for significant differences in effects among treatments (significance level was always set at p = 0.05); then NOEC and LOEC were determined by Dunnett’s procedure. When data did not meet the assumptions of normality and homocedasticity, non parametric Kruskall-Wallis test was employed to compare individual treatments.

Concentration–response analyzes

Concentration–response analysis was performed in the same way for each toxicant. Concentration–response functions were statistically determined by applying a best fit procedure. With this approach, different regression models (Boltzmann, Logistic, Exponential), provided by Origin® 8 SR2 (Northampton, MA) statistical software, were applied to each data set in order to determine, on the basis of statistical criteria, the regression model that best described the trend of the observed toxicity data.

Regression curves were obtained and analyzed with Origin® 8 SR2 software, using the least-square method and the analysis of residuals. Models that have passed the residual analysis as reliable candidates are then subjected to a second selection step.

By this procedure, we calculated the sum of absolute residuals (SAE) and the sum of absolute deviations (SAD), the model that showed the minimum SAE and SAD values was selected as the best fitting one. At this stage, the most appropriate model was chosen by applying a goodness of fit criterion. However, the first results obtained from a simulation study carried out by Scholze et al. (2001) indicated that the SAE is much more sensitive than the SAD measure.

Results

EC50 and EC1 values obtained with ICp EPA method (US EPA 1993), NOEC and LOEC values calculated with Dunnett’s procedure, for each tested substances, are given in Table 1 both μg/l and molar units, for comparison.
Table 1

Pesticides toxic effects on P. lividus: no observed effect concentration (NOEC μg/l and nM), lowest observed effect concentration (LOEC μg/l and nM) were determinate with Dunnett’s procedure

 

Pentachlophenol

Azinphos-methyl

Chlorpyrifos

EC1(μg/l)

(nM)

0.31 ± 0.09

1.16 ± 0.34

0.24 ± 0.11

0.76 ± 0.35

0.29 ± 0.11

0.83 ± 0.31

NOEC (μg/l)

(nM)

30

112.64

0.7

2.205

0.8

2.28

LOEC (μg/l)

(nM)

50

187.72

1.6

5.04

1

2.85

EC50 (μg/l)

(nM)

147.84 ± 12.30

555.08 ± 46.18

141.23 ± 28.18

445.07 ± 88.80

194.6 ± 16.11

555.05 ± 45.95

EC50 (50% Effect concentration μg/l and nM) and EC1 (1% effect concentration μg/l and nM) ± SE (ICp approach: U.S. EPA 1993)

AZM (EC50 141.23 ± 28.18 μg/l) revealed as the most toxic pesticide. The NOEC value exceeded the corresponding EC1 (ICp) by a factor of 97 for PCP, 2.9 for AZM and 2.8 for CPF.

Pentachlorophenol

Results from each treatment and the corresponding regression fit curve, are shown in Fig. 1a. The PCP effects were not significant at concentrations below 30 μg/l (p < 0.05) and the LOEC was 50 μg/l, but increased at higher values reaching 100% of abnormalities at 300 μg/l, and yielding an EC50 of 147.84 (±12.30) μg/l (Table 1).
Fig. 1

Pentachlorophenol embryotoxicity on P. lividus. a Percentage of malformed out of 100 individuals normalized with respect to control, as a function of tested concentrations. Horizontal dotted lines indicate the 95% confidence limits of the control mean (n = 6). 50% effect level is also represented. Solid circles represent PCP treatments. Thick solid line indicates the regression fit of the observations. b Number of individuals with different developmental anomalies obtained after 48 h exposure. N normal plutei; R retarded larvae; P1 malformed larvae; P2 blastulae or gastrulae (developmental arrest); see also ‘‘Materials and methods’’ (mean of at least six experiments, each in triplicate ± SE)

The effect percentage up to 100 μg/l concentration was always lower than 20% and the developmental defects in treated P. lividus larvae (Fig. 1b) were mainly P1 type (larvae affected in skeletal or gut differentiation). P2 (total arrest at prelarval stadium) and R (retarded larvae) developmental alterations were always present, but at very low percentage. At concentration of 500 μg/l the developmental defects in treated P. lividus larvae were only P2 type (Fig. 1b).

The best fit function of PCP embriotoxicity data was a sigmoid growth function (Logistic) and showed an EC50 value of 126.03 [123.57–128.53 (limits 95% confidence bands)] μg/l.

Chlorpyrifos

The CPF toxic effects on P. lividus embryos, together with the corresponding regression fit curve, are reported in Fig. 2a.
Fig. 2

Chlorpyrifos embryotoxicity on P. lividus. a Percentage of malformed out of 100 individuals normalized with respect to control, as a function of tested concentrations. Horizontal dotted lines indicate the 95% confidence limits of the control mean (n = 6). 50% effect level is also represented. Solid circles represent CPF treatments. Thick solid line indicates the regression fit of the observations. b Number of individuals with different developmental anomalies obtained after 48 h exposure. N normal plutei; R retarded larvae; P1 malformed larvae; P2 blastulae or gastrulae (developmental arrest); see also ‘‘Materials and methods’’ (mean of at least six experiments, each in triplicate ± SE)

The CPF effects were not significant at concentrations below 0.8 μg/l (p < 0.05) and the LOEC was 1 μg/l, but increased at higher values reaching 100% of abnormalities at 500 μg/l, and yielding an EC50 of 194.6 (±16.11) μg/l (Table 1).

The effect percentage up to 90 μg/l concentration was always lower than 20% and the developmental defects in treated P. lividus larvae (Fig. 2b) were mainly P1 type. From 300 μg/l concentration the P1 type developmental defect decreased (30%) and P2 type appeared, representing the mainly developmental defects at 500 μg/l (Fig. 2b).

The regression model that best described the observed trend toxicity data was a sigmoid growth function (Logistic) and showed an EC50 value of 231.31 [213.52–250.59 (limits 95% confidence bands)] μg/l.

Azinphos-methyl

The toxicity of this pesticide on P. lividus embryos, together with the corresponding regression fit curve, is shown in Fig. 3a. The AZM effects were not significant at concentrations below 0.7 μg/l (p < 0.05) and the LOEC was 1.6 μg/l, but increased at higher values reaching 100% of abnormalities at 500 μg/l, and yielding an EC50 of 141.23 (±28.18) μg/l (Table 1).
Fig. 3

Azinphos-methyl embryotoxicity on P. lividus. A: % of malformed out of 100 individuals normalized with respect to control, as a function of tested concentrations. Horizontal dotted lines indicate the 95% confidence limits of the control mean (n = 6). 50% effect level is also represented. Solid circles represent AZM treatments. Thick solid line indicates the regression fit of the observations. B: Number of individuals with different developmental anomalies obtained after 48 h exposure. N normal plutei, R retarded larvae, P1 malformed larvae, P2 blastulae or gastrulae (developmental arrest); see also ‘‘Materials and methods’’ (mean of at least six experiments, each in triplicate ± SE)

The developmental defects (Fig. 3b) were mainly P1 type, showing a dose-dependent trend up to 300 μg/l (45%): P2 and R developmental alterations are always present but at very low percentage (<20%).

The best fit function of AZM embriotoxicity data was an exponential function (ExpDec1) and showed an EC50 value of 118.23 [113.68–122.97 (limits 95% confidence bands)] μg/l.

Discussion

In the present study both PCP and AZM revealed as the most toxic compounds to the early developmental stages of P. lividus (Table 1, EC50 values) with EC50 (ICp) values not statistically different. At the same time the three pesticides showed toxicity values statistically not different at EC1 level (Table 1). However, OP always resulted more embryotoxic than PCP at low concentrations (EC1, NOEC, and LOEC) in fact considering NOEC values, the rank of pesticide toxicity to P. lividus embryos decreases as follows AZM > CPF > PCP, with NOEC values exceeding the corresponding EC1 (ICp-method) of a factor of 97 for PCP, 2.9 for AZM, and 2.8 for CPF. Besides the EC1 values of AZM and CPF resulted in the same order of magnitude compared to NOEC, underlining the good level of NOEC accuracy. For PCP, instead, the EC1 values resulted more protective compared to NOEC. The rank of pesticide toxicity, considering LOECs decreases as follows CPF > AZM > PCP. Comparison of toxicity data in molar units (Table 1) showed that the rank of pesticide toxicity did not change for EC1, NOEC, and LOEC levels but at EC50 the three pesticides resulted not statistically different.

The logistic regression model better describe the observed toxicity data for PCP and AZM, where a fast increase of toxicity is observable, whereas a sigmoid growth function was more appropriate for CPF toxicity. Besides, CPF, and AZM EC50s obtained from the best fit curves resulted not statistically different from the EC50 values determined with ICp-method, instead for PCP they resulted of the same order of magnitude but statistically different. The use of different methods of data interpolations could produce different results; however, in our study, the two applied methods of EC50 calculation produced equally protective values for P. lividus embryos.

With the exception of Daphnia magna and others freshwater organisms, toxic effects (EC50, EC1, NOEC, and LOEC) of PCP, CPF, and AZM on P. lividus are scarcely reported in literature (Hutchinson et al. 1998; Leung et al. 2001). Organochlorine and organophosphorus compound toxic effects (LOEC) reported in literature, upon different sea urchin embryos and larvae, spanned concentrations between 1000 and 3,000 μg/l (Kobayashi 1995). In the present study, the AZM, CPF, PCP LOEC range was from 1 to 50 μg/l, showing the highest sensitivity of P. lividus embryos to these pesticides.

Pentachlorophenol exerts its mechanism of toxic action as uncoupler of mitochondrial oxidative phosphorylation, therefore destroying the electrochemical potential across the inner membrane of mitochondria (Martello et al. 1998; Mitchell 1961; Mitchell and Moyte 1967; Zha et al. 2006). The low toxic effects observed on P. lividus embryos at low concentrations (up to 100 μg/l) could be linked to PCP interaction with larval skeleton deposition that give rise to malformed larvae (P1) (acting on calcium homeostasis). At a PCP concentration of 500 μg/l, the maximum effect was observed and the developmental defects in treated P. lividus larvae were mainly of P2 type (Fig. 2a). PCP at this concentration, could completely block the development of P. lividus embryos at blastula or gastrula stage (modifying cytoskeleton assembly during blastomere division) producing P2 type alterations.

Ozretic and Krajnovic-Ozretic (1985) reported PCP EC50 for P. lividus (0.40–0.60 mg/l) higher than value found in the present study. This could be linked to the different test procedure utilized: cleaving eggs and early embryonic stages were exposed to various PCP concentrations for only 30 min. Sea urchin PCP spermiotoxicity (EC50) of 4.23 mg/l for Arbacia spatuligera (Zuñiga et al. 1995) and of 0.9 mg/l for Arbacia punctulata (Burgess et al. 1993) were also reported in literature. Therefore PCP seemed to exert lower effect on sea urchin sperms compared to embryos: this could be due to the higher resistance of sperms—differentiated cells—to environmental pollutants (Manzo et al. 2006).

Pentachlorophenol acute toxicity on estuarine/marine invertebrate evaluated by Scow et al. (1980) exhibited a good sensitivity of these organisms:Crassostrea gigas (shell deposition) showed an EC50 of 0.048 mg/l and Palaemon elegans LC50 of 0.084 mg/l. Regarding to freshwater organisms, Repetto et al. (2001) reported for Daphnia magna an EC50 of 0.40 mg/l while Johnson and Finley (1980) found a LC50 of 0.05 mg/l for Oncorhynchus mykiss.

Although PCP employ has been significantly limited in recent years, due to its persistence and use in some areas of the world, environmental concentrations still remain very high. Water concentration of 2.94 μg/l recently reported (Santiago and Kwan 2007) was in the toxicity range identified in this study for P. lividus embryos which is also the environmental risk to which these organisms were exposed.

Organo phosphorus pesticides cause the inhibition of AChE, damaging the transmission of the nervous impulse from the central nervous system (Timchalk 2001).Then, OPs are not expected to be highly toxic to sea urchins, in comparison to organisms with a more complex nervous system (Bellas et al. 2005). However, using P. lividus early development stages as a model, Aluigi et al. (2008) found that OP insecticides could affect development and cause embryonic anomalies. Sea urchin model revealed to be very sensitive to organophosphate exposure (Morale et al. 1998) because AChE and AChE activity were present in sea urchin blastomeres as early as the unfertilized egg stage (Falugi et al. 1992; Falugi and Prestipino 1987). At first cell cycles, Aluigi et al. (2008) also found an acetylcholine (ACh) muscarinic receptor that is capable of determining an increase in Ca2+ concentration when activated by ACh or its antagonist substances (Harrison et al. 2002).

From our results, we could hypothesize that CPF applies its toxic action, mainly on calcium homeostasis up to 100 μg/l, in fact an increasing P1 developmental defects was observable (Fig. 2b). At higher doses CPF exerted an early effect on sea urchin embryological development, therefore abnormal blastula or gastrulae (P2) appeared and became the developmental defects most represented at 500 μg/l.

Buznikov et al. (2001) described CPF toxic effects (LOEC) on sea urchins Strogylocentrotus droebachiensis and S. purpuratus embryos at 5,400 μg/l, highlighting their poor sensitivity for these compounds. Bellas et al. (2005) reported an EC50 (300 μg/l) for P. lividus embryos that is in the same order of magnitude of the value measured in the present study.

Chlorpyrifos toxicity data on freshwater organisms are showed by Van Wijngaarden et al. (2005) in a comprehensive review. D. magna revealed to be very sensitive to CPF with an EC50 of 1.3 μg/l.

Azinphos-methyl caused embryotoxic effects on P. lividus at concentrations as low as 30 μg/l, but toxicity still remain under 25% up to 90 μg/l. This pesticide, for all tested doses, altered sea urchin embryos regular development especially after gastrulation and indeed the registered abnormalities were in particular P1 type, with a dose-dependent trend up to 300 μg/l. P2 and R developmental alterations are always present but at very low percentage. As reported by Pesando et al. (2003) OPs could affect nuclear and cytoskeletal status as well as DNA synthesis during embryonic development, but from the gastrulation stage onwards, the main effects were exerted on the rate of primary mesenchyme cells migration, larval size, perioral arm length, and AChE activity distribution.

Azinphos-methyl toxicity has been shown on both freshwater and marine species, however only limited data are available and no acute-to-chronic ratio for any marine species has been determined (Morton et al. 1997).

The most sensitive saltwater organism to AZM, appeared to be the common shrimp (Crangon crangon), with a 48-h LC50 of 0.67 μg/l (Portmann and Wilson 1971). AZM 48-h EC50 s of 2.4 μg/l (Mayer 1987), and 4.4 μg/l (Butler 1963) for the brown shrimp (Penaeus aztecus) have also been reported. Davis and Hidu (1969) showed developmental effects to 50% on Eastern oysters (Crassostrea virginica) after 48-h exposure to 620 μg/l of AZM.

Data of chronic AZM exposure on marine invertebrates are extremely limited and not very up to date. Davis and Hidu (1969) reported a 12-day LC50 of 860 μg/l for northern quahog (Mercenaria mercenaria), larvae. Cripe et al. (1984) determined that 7 day exposure to 1.9 μg/l of AZM produced significant mortality in sheepshead minnow (Cyprinodon variegatus), while exposure to 0.06 μg/l for 107 days significantly inhibited brain AChE activity in the same species. Actually, sea urchin embryotoxicity test is a sub-chronic test, in fact the exposure to toxicant is performed for all developmental stage from zygote to pluteus. However the endpoint tested (correct embryological development) showed variable sensitivity to AZM compared to data about marine invertebrate mortality.

Conclusion

The known sensitivity of sea urchins to several environmental pollutants revealed to be noticeable also for pesticides (two organophosphate AZM and CPF and one organochlorurate PCP) tested in this study. The most toxic pesticides to the early developmental stages of P. lividus were PCP and AZM at EC50 level. Nevertheless OP toxicity measured was always higher than PCP one at low concentration (EC1, NOEC, and LOEC). The good level of NOEC accuracy compared to EC1 was measured. In fact the EC1 values of AZM and CPF resulted of the same order of magnitude of NOEC, whereas for PCP the EC1 values resulted more protective.

Pentachlorophenol at higher concentration could modify cytoskeleton assembly during blastomere division, producing P2 type alterations, while at lower concentrations, acting on calcium homeostasis, it could alter the deposition of the larval skeleton, giving rise to malformed larvae P1. OPs showed similar toxicity behaviour: at low concentrations it was hypothized that they can interfere with calcium metabolism; in fact skeletal malformation were mainly produced, with a trend corresponding to the pesticide concentrations (until 300 μg/l). At high concentration (500 μg/l) the main effects were early exerted on the embryos development that was blocked at gastrula and blastula stage.

Organo phosphorus EC50s obtained with both methods (best fit procedure and ICp) resulted not statistically different, unlike for PCP they resulted of the same order of magnitude but statistically different. The use of different interpolation methods of raw data could produce different results; however, in our investigation, the two applied methods of EC50 calculation produced equally protective values for P. lividus embryos.

Further experiments are in progress in our laboratories to assess the toxicity effects on P. lividus of PCP, AZM, and CPF mixtures at levels corresponding to and below not significant effect concentration in order to investigate the potential risk linked to their joint action. In addition it would be remarkable to evaluate the pesticide toxic action on target organisms of different trophic levels.

This research highlighted the importance of evaluating, in coastal seawaters, levels of used pesticides in order to understand the real impact on benthic populations mainly in sites characterized by intensive agriculture or floriculture activities, such as the coastal areas of the Mediterranean Sea.

References

  1. Aluigi MG, Falugi C (2010) Dose-dependent effects of chlorpyriphos, an organophosphate pesticide, on metamorphosis of the sea urchin, Paracentrotus lividus. Ecotoxicology 19:520–529CrossRefGoogle Scholar
  2. Aluigi MG, Angelini C, Corte G, Falugi C (2008) The sea urchin, Paracentrotus lividus, embryo as a “bioethical” model for neurodevelopmental toxicity testing. Cell Biol Toxicol 24:587–601CrossRefGoogle Scholar
  3. Arcand Y, Hawari J, Guiot SR (1995) Solubility of pentachlorophenol in aqueous solutions: the pH effect. Water Res 29:131–136CrossRefGoogle Scholar
  4. Arizzi Novelli A, Argese E, Tagliapietra D, Bettiol C, Volpi Ghirardini A (2002) Toxicity of tributyltin and triphenyltin towards early life stages of Paracentrotus lividus (Echinodermata: Echinoidea). Environ Toxicol Chem 21:859–864Google Scholar
  5. ASTM (American Society for Testing, Materials) (2004) Standard guide for conducting static acute toxicity tests with echinoid embryos. ASTM Standard Guide E 1563–98. In: Annual book of ASTM standards, Section 11: biological effects and environmental fate; biothecnology; pesticides, vol 115. ASTM, West ConshohochenGoogle Scholar
  6. Bay S, Burgess R, Nacci D (1993) Status and applications of echinoid (Phylum: Echinodermata) toxicity test methods. In: Landis WG, Hughes JS, Lewis MA (eds) Environmental toxicology risk assessment. ASTM STP 1179. American Society for Testing and Materials, Philadelphia, pp 281–302CrossRefGoogle Scholar
  7. Bellas J, Beiras R, Marino-Balsa JC, Fernandez N (2005) Toxicity of organic compounds to marine invertebrate embryos and larvae: a comparison between the sea urchin embryogenesis bioassay and alternative test species. Ecotoxicology 14:337–353CrossRefGoogle Scholar
  8. Bressan M, Marin M, Brunetti R (1995) Influence of temperature and salinity on embryonic-development of Paracentrotus lividus (lmk, 1816). Hydrobiologia 304:175–184CrossRefGoogle Scholar
  9. Burgess MR, Schweitzer KA, McKinney RA, Phelps DK (1993) Contaminated marine sediment: water column and interstitial toxic effects. Environ Toxicol Chem ETOCDK 12:127–138CrossRefGoogle Scholar
  10. Butler PA (1963) A review of fish, wildlife service investigations during 1961 and 1962. Circular 167. US Fish Wildl Serv, Washington DCGoogle Scholar
  11. Buznikov GA, Nikitina LA, Bezuglov VV, Lauder JM, Padilla S, Slotkin TA (2001) An invertebrate model of the developmental neurotoxicity of insecticides: effects of chlorpyrifos and dieldrin in sea urchin embryos and larvae. Environ Health Perspect 109(7):651–661Google Scholar
  12. Buznikov GA, Nikitina LA, Rakic LM, Milosevic I, Bezuglov VV, Lauder JM, Slotkin TA (2007) The sea urchin embryo, an invertebrate model for mammalian developmental neurotoxicity, reveals multiple neurotransmitter mechanisms for effects of CPF: therapeutic interventions and a comparison with the monoamine deplete, reserpine. Brain Res Bull 74:221–231CrossRefGoogle Scholar
  13. Cesar A, Marin A, Marin-Guirao L, Vita R (2004) Amphipod and sea urchin tests to assess the toxicity of Mediterranean sediments: the case of Portman Bay. Sci Mar 68:205–213CrossRefGoogle Scholar
  14. Cripe GM, Goodman LR, Hansen DJ (1984) Effect of chronic exposure to EPN and to guthion on the critical swimming speed and brain acetylcholinesterase activity of Cyprinodon variegatus. Aquatic Toxicol. 5:255–266CrossRefGoogle Scholar
  15. Davis HC, Hidu H (1969) Effects of pesticides on embryonic development of clams and oysters and on survival and growth of the larvae. Fish Bull 67(2):393–404Google Scholar
  16. Dunnett CW (1964) New tables for multiple comparisons with a control. Biometrics 20:482–491CrossRefGoogle Scholar
  17. ERMA (2009) Environmental Risk Management Authority Decision ERMA New Zealand Decision: Application HRC07002. http://www.ermanz.govt.nz/news-events/archives/media-releases/2009/mr-20091106.html
  18. Falugi C, Prestipino G (1987) Effects of some inhibitors on the cholinergic system active during the sea urchin Paracentrotus lividus development. In: Boudouresque CF (ed) Colloque International sur Paracentrotus lividus. Gis Posidonie publ, Marseilles, pp 147–155Google Scholar
  19. Falugi C, Pieroni M, Drews U, Stengel P, Lammerding-Koppel M (1992) Possible functions of the “embryonic” cholinergic system present in gametes at fertilization. In: Scalera Liaci L, Canicattì C (eds) Echinoderm research. Balkema press, Rotterdam, pp 161–164Google Scholar
  20. Foster S, Thomas M, Korth W (1998) Laboratory derived acute toxicity of selected pesticides to Ceriodaphnia dubia Aust. J Ecotoxicol 4:53–59Google Scholar
  21. Harrison PK, Falugi C, Angelini C, Whitaker MJ (2002) Muscarinic signalling affects intracellular calcium concentration during the first cell cycle of sea urchin embryos. Cell Calcium 31(6):289–297CrossRefGoogle Scholar
  22. Hartgers EM, Heugens EHW, Deneer JW (1999) Effect of lindane on the clearance rate of Daphnia magna. Arch Environ Contam Toxicol 36:399–404CrossRefGoogle Scholar
  23. His E, Seaman MNL (1993) Effects of twelve pesticides on larvae of oysters (Crassosstrea gigas) and on two species of unicellular marine algae (Isochrysis galbana and Chaetoceros calcitrans). Comm Meet Int Counc Explor Sea CM-ICES/E 22:1–8Google Scholar
  24. His E, Heyvang I, Geffard O, De Mountadouin X (1999) A comparison between oyster (Crassostrea gigas) and sea urchin (Paracentrotus lividus) larval bioassay for toxicological studies. Wat Res 7:1706–1718CrossRefGoogle Scholar
  25. Hutchinson TT, Scholz N, Guhl W (1998) Analysis of the ECETOC aquatic toxicity (EAT) database IV–Comparative toxicity of chemical substances to freshwater versus saltwater organisms. Chemosphere 36(1):143–153CrossRefGoogle Scholar
  26. Johnson WW, Finley MT (1980) Handbook of acute toxicity of chemicals to fish and aquatic invertebrates. Resource publication 137. US Department of Interior, Fish and Wildlife Service, Washington DC, p 65Google Scholar
  27. Kaiser KLE, Valdmanis I (1981) Apparent octanol/water partition coefficients of pentachlorophenol as a function of pH. Can J Chem 60:2104–2106CrossRefGoogle Scholar
  28. Key PB, Fulton MH (1993) Lethal and sublethal effects of chlorpyrifos exposure on adult and larval stages of the grass shrimp Palaemonetes pugio. J Environ Sci Health B 28(5):621–640CrossRefGoogle Scholar
  29. Kobayashi N (1991) Marine pollution bioassay by using sea urchin eggs in the Tanabe Bay, Wakayama Prefecture, Japan, 1970–1987. Marine Pollut Bull 23:709–713CrossRefGoogle Scholar
  30. Kobayashi N (1995) Bioassay data for marine pollution using echinoderms. In: Cheremisinoff PN (ed) Encyclopedia of environmental control technology. Gulf Publ Co, Houston, pp 539–609Google Scholar
  31. Kobayashi N, Okamura H (2002) Effects of new antifouling compounds on the development of sea urchin. Mar Pollut Bull 44:748–751CrossRefGoogle Scholar
  32. LeBlanc GA, Bain LJ (1997) Chronic toxicity of environmental contaminants: sentinels and biomarkers. Environ Health Perspect 105:65–80Google Scholar
  33. Leung KMY, Morritt D, Wheeler JR, Whitehouse P, Sorokin N, Toy R, Holt M, Crane M (2001) Can saltwater toxicity be predicted from freshwater data? Mar Pollut Bull 42(11):1007–1013CrossRefGoogle Scholar
  34. Mansueto C, Gianguzza M, Dolcemascolo G, Pellerito L (1993) Effects of tributyltin (IV) chloride exposure on early embryonic stages of Ciona intestinalis in vivo and ultrastructural investigations. Appl Organomet Chem 7:391–399CrossRefGoogle Scholar
  35. Manzo S (2004) Sea urchin embryotoxicity test: proposal for a simplified bioassay. Ecotox Environ Safety 57(2):123–128CrossRefGoogle Scholar
  36. Manzo S, Torricelli L (2000) Preliminary findings about Regione Campania (South Italy) coastal water ecotoxicological data. Abstract book of II, National Conference of Sea Science, Geneva, 21/25 November, 223Google Scholar
  37. Manzo S, Buono S, Cremisini C (2006) Toxic effects of Irgarol and Diuron on sea urchin Paracentrotus lividus early development, fertilization, and offspring quality. Arch Environ Contam Toxicol 51:61–68CrossRefGoogle Scholar
  38. Manzo S, Buono S, Cremisini C (2008) Predictability of copper, irgarol, and diuron combined effects on sea urchin Paracentrotus lividus. Arch Environ Contam Toxicol 54:57–68CrossRefGoogle Scholar
  39. Marin MG, Moschino V, Cima F, Celli C (2000) Embryotoxicity of butyltin compounds to the sea urchin Paracentrotus lividus. Mar Environ Res 50:231–235CrossRefGoogle Scholar
  40. Martello LB, Tjeerdema RS, Smith WS, Kauten RJ, Crosby DG (1998) Influence of salinity on the actions of pentachlorophenol in Haliotis as measured by in vivo 31P-NMR spectroscopy. Aquat Toxicol 41:229–250CrossRefGoogle Scholar
  41. Mayer FL Jr (1987) Acute toxicity handbook of chemicals to estuarine organisms. EPA/600/8–87/017. Environmental Research Laboratory, Gulf Breeze, p 274Google Scholar
  42. Mitchell P (1961) Coupling of phosphorylation to electron and hydrogen transfer by a chemiosmotic type of mechanism. Nature 191:144–148CrossRefGoogle Scholar
  43. Mitchell P, Moyte J (1967) Acid-base titration across the membrane system of rat liver mitochondria. Catalysis by uncouplers. Biochem J 104:588–600Google Scholar
  44. Moore DRJ, Caux PY (1997) Estimating low toxic effects. Environ Toxicol Chem 16:794–801CrossRefGoogle Scholar
  45. Morale A, Coniglio L, Angelini C, Cimoli G, Bolla A, Alleteo D, Russo P, Falugi C (1998) Biological effects of a neurotoxic pesticide at low concentrations on sea urchin early development. A teratogenic assay. Chemosphere 37(14–15):3001–3010CrossRefGoogle Scholar
  46. Morton MG, Mayer FL Jr, Dickson KL, Waller WT, Moore JC (1997) Acute and chronic toxicity of azinphos-methyl to two estuarine species, Mysidopsis bahia and Cyprinodon variegatus. Arch Environ Contam Toxicol 32:436–441CrossRefGoogle Scholar
  47. Muir J, Eduljee G (1999) PCP in the freshwater and marine environment of the European Union. Sci Total Environ 236:41–56CrossRefGoogle Scholar
  48. Muir DC, Teixeira C, Wania F (2004) Empirical and modeling evidence of regional atmospheric transport of current-use pesticides. Environ Toxicol Chem 23:2421–2432CrossRefGoogle Scholar
  49. Muller F (2000) Agrochemicals: composition, production, toxicology, applications. Wiley, Toronto, p 541Google Scholar
  50. Nimmo DR, McEwen LC (1994) Pesticides. In: Calow P (ed) Handbook of Ecotoxicology. Blackwell Scientific Publications, Oxford, pp 155–203Google Scholar
  51. Ozretic B, Krajnovic-Ozretic (1985) Morphological and biochemical evidence of the toxic effect of pentachlorophenol on the developing embryos of the sea urchin. Aquat Toxicol 7:255–263CrossRefGoogle Scholar
  52. Pagano G, Cipollaro M, Corsale G, Esposito A, Ragucci E, Giordano GG, Trieff NM (1986) The sea urchin: bioassay for the assessment of damage from environmental contaminants. In: Cairns J (ed) Community toxicity testing. ASTM STP920. American Society for Testing and Materials, Philadelphia, pp 66–92CrossRefGoogle Scholar
  53. Pagano G, Corsale G, Esposito A, Dinnel PA, Romana LA (1989) Use of sea urchin sperm and embryo bioassay in testing the sublethal toxicity of realistic pollutant level. Adv Appl Biotech Ser 5:153–163Google Scholar
  54. Pagano G, Iaccarino M, Guida M, Manzo S, Oral R, Romanelli R, Rossi M (1996a) Cadmium toxicity in spiked sediment to sea urchin embryos and sperm. Mar Environ Res 42:54–55CrossRefGoogle Scholar
  55. Pagano G, His E, Beiras R, De Biase A, Korkina LG, Iaccarino M, Oral R, Qiuniou Warnau M, Trieff NM (1996b) Cytogenetic, developmental, and biochemical effects of aluminium, iron, and their mixture in sea urchins and mussels. Arch Environ Contam Toxicol 31:466–474CrossRefGoogle Scholar
  56. Pagano G, Korkina LG, Iaccarino M, De Biase A, Deva IB, Doroniu YK, Guida M, Melluso G, Oral R, Trieff NM, Warnau M (2001) Developmental, cytogenetic and biochemical effects in sea-urchinbioassays. In: Garrignos P, Walker CH, Barth H, Narbonne JF (eds) Biomarkers in marine ecosystems: a pratical approach. Elsevier, Amsterdam, pp 85–129Google Scholar
  57. Pesando D, Huitorelb P, Dolcinia V, Angelinic C, Guidettid P, Falugic C (2003) Biological targets of neurotoxic pesticides analysed by alteration of developmental events in the Mediterranean sea urchin, Paracentrotus lividus. Mar Environ Res 55:39–57CrossRefGoogle Scholar
  58. Portmann JE, Wilson KW (1971) The toxicity of 140 substances to the brown shrimp and other marine animals. MAFF Shellfish information leaflet No 22. Essex, England, p 12Google Scholar
  59. Repetto G, Jos A, Hazen MJ, Molero ML, del Peso A, Salguero M, Castillo PD, Rodriguez-Vicente MC, Repetto M (2001) A test battery for the ecotoxicological evaluation of pentachlorophenol. Toxicol In Vitro 15(4–5):503–509CrossRefGoogle Scholar
  60. Ringwood Huffman A (1992) Comparative sensitivity of gametes and early developmental stages of a sea urchin species (Echinometra mathei) and a bivalve species (Isognomon californicum) during metal exposures. Arch Environ Contain Toxicol 22:265–288Google Scholar
  61. Santiago EC, Kwan CS (2007) Endocrine-disrupting phenols in selected rivers and bays in the Philippines. Mar Pollut Bull 54:1036–1046CrossRefGoogle Scholar
  62. Scholze M, Boedeker W, Faust M, Backhaus T, Altenburger R, Grimme H (2001) A general best-fit method for concentration response curves and the estimation of low effect concentrations. Environ Toxicol Chem 20:448–457Google Scholar
  63. Sciarrino S, Matranga V (1995) Effects of retinoic acid and dimethylsulfoxide on the morphogenesis of the sea urchin embryo. Cell Biol Int 19(8):675–680CrossRefGoogle Scholar
  64. Scott A (2008) “Europe Rejects Appeal for Use of Azinphos-methyl Pesticide”. Chemical Week. http://www.chemweek.com/envirotech/regulatory/13435.html. Accessed Aug 11 2008
  65. Scow K, Goyer M, Payne E, Perwak J, Thomas R, Wallace D, Walker P, Wood M, Delpire L (1980) An exposure and risk assessment for pehtachlorophenol. Final report. U.S. Environmental Protection Agency (EPA-440/4–81/021, Washington DCGoogle Scholar
  66. Shelley LK, Balfry KS, Ross PS, Kennedy CJ (2009) Immunotoxicological effects of a sub-chronic exposure to selected current-use pesticides in rainbow trout (Oncorhynchus mykiss). Aquatic Toxicol 92:95–103CrossRefGoogle Scholar
  67. Shief JN, Choa ML, Chen CY (2001) Statistical comparisons of the no-observed-effect concentration and the effective concentration at 10% inhibition (EC10) in algal toxicity test. Water Sci Technol 43:141–146Google Scholar
  68. Skalski JR (1981) Statistical inconsistencies in the use of no-observed-effect levels in toxicity testing. In: Branson DR, Dickson KL (eds) Aquatic Toxicology and Hazard Assessment: Fourth Conference. ASTM STP 737. American Society for Testing and Materials, Philadelphia, PA, pp 377–387Google Scholar
  69. Spromberg JA, Meador JP (2006) Relating chronic toxicity responses to population level effects: a comparison of population-level parameters for three salmon species as a function of low-level toxicity. Ecol Model 199:240–252CrossRefGoogle Scholar
  70. Timchalk C (2001) Organophosphate pharmacokinetics. In: Krieger R (ed) Handbook of pesticide toxicology, vol 2. Academic Press, San Diego, pp 929–951CrossRefGoogle Scholar
  71. Tomlin C (ed) (1994) The pesticide manual. Tenth edition. British Crop Protection Council, p 1341Google Scholar
  72. Tomlin C (ed) (2006) The pesticide manual. 14th edition British Crop Protection Council (BCPC), p 1350Google Scholar
  73. US EPA (2002) Interim Reregistration Eligibility Decision for Chlorpyrifos. http://www.epa.gov/oppsrrd1/REDs/chlorpyrifos_ired.pdf
  74. US EPA (2009) Reregistration Eligibility Decision for Azinphos-Methyl. www.epa.gov/oppsrrd1/reregistration/azm/phaseout_fs.htm
  75. US EPA (1993) A linear interpolation method for sublethal toxicity: the inhibition concentration (ICp) approach. National Effluent Toxicity Assessment Center Technical Report 03–93. Environmental Research Laboratory, DuluthGoogle Scholar
  76. Van Der Hoeven N, Noppert F, Annegaaike L (1997) How to measure no effect. Part I: towards a new measure of chronic toxicity in ecotoxicology. Introduction and workshop results. Environmetrics 8:241–248CrossRefGoogle Scholar
  77. Van Wijngaarden RPA, Brock TCM, Douglas MT (2005) Effects of chlorpyrifos in freshwater model ecosystems: do experimental conditions change ecotoxicological threshold levels? Pest Manag Sci 61:923–935CrossRefGoogle Scholar
  78. Volpi Ghirardini A, Arizzi Novelli A (2001) A sperm cell toxicity test procedure for the Mediterranean species Paracentrotus lividus (Echinodermata: Echinoidea). Environ Technol 22:439–445CrossRefGoogle Scholar
  79. Whiting VK, Gripe GM, Lepo JE (1996) Effects of the anionic surfactant, sodium dodecyl sulfate, on newly-hatched Blue crabs, Callinectes sapidus, and other routinely tested estuarine crustaceans Arch. Environ Contam Toxicol 31:293–295CrossRefGoogle Scholar
  80. Worthing CR (ed) (1991) The pesticide manual: a world compendium. Ninth Edition. The British Crop Protection CouncilGoogle Scholar
  81. Zha J, Wang Z, Schlenk D (2006) Effects of pentachlorophenol on the reproduction of Japanese medaka (Oryzias latipes). Chemico-Biological Interactions 161:26–36CrossRefGoogle Scholar
  82. Zuñiga M, Roa R, Larrain A (1995) Sperm cell bioassay with the sea urchin Arbacia spatuligera on samples from two polluted chilean coastal sites. Mar Pollut Bull 30:313–319CrossRefGoogle Scholar

Copyright information

© Springer Science+Business Media, LLC 2011

Authors and Affiliations

  • Silvia Buono
    • 1
  • Sonia Manzo
    • 2
  • Giovanna Maria
    • 3
  • Giovanni Sansone
    • 3
  1. 1.CRIAcq Università degli Studi di Napoli—Federico IIPortici (NA)Italy
  2. 2.ENEA C. R. PorticiPortici (NA)Italy
  3. 3.Dipartimento Delle Scienze BiologicheUniversità degli Studi di Napoli—Federico IINaplesItaly

Personalised recommendations