Toxic effects of pentachlorophenol, azinphos-methyl and chlorpyrifos on the development of Paracentrotus lividus embryos
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.
KeywordsParacentrotus lividus Pentachlorophenol Azinphos-methyl Chlorpyrifos
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
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.
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).
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.
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 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.
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
0.31 ± 0.09
1.16 ± 0.34
0.24 ± 0.11
0.76 ± 0.35
0.29 ± 0.11
0.83 ± 0.31
147.84 ± 12.30
555.08 ± 46.18
141.23 ± 28.18
445.07 ± 88.80
194.6 ± 16.11
555.05 ± 45.95
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.
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.
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.
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.
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.
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.
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