Ecotoxicology, 10, 263–271, 2001 © 2001 Kluwer Academic Publishers. Manufactured in The Netherlands.
Combined Toxicity of Dissolved Mercury with Copper, Lead and Cadmium on Embryogenesis and Early Larval Growth of the Paracentrotus lividus Sea-Urchin ´ N. FERNANDEZ AND R. BEIRAS∗ Departamento de Ecolox´ıa e Biolox´ıa Animal, Universidade de Vigo, E-36200, Vigo, Galicia Received 14 September 2000; Accepted 28 October 2000; Revised 15 December 2000
Abstract. The individual and combined toxicity of dissolved mercury, copper, lead and cadmium has been investigated by using the Paracentrotus lividus sea-urchin embryo-larval bioassay. Embryogenesis success and early larval growth have been recorded after incubation of fertilised eggs in seawater, both with single metals and binary combinations of Hg with every other metal. For individual metals the ranking of toxicity was Hg > Cu > Pb > Cd, with EC50 values of 21.9, 66.8, 509 and 9240 µg/l, respectively. Lowest observed effect concentrations (LOEC) for early larval growth were approximately three times lower than the EC50 values for Hg, Cu and Pb, and more than two orders of magnitude lower for Cd, emphasizing the danger of underestimating toxicity when only lethal effects are recorded. Marking & Dawson’s additive indices ranged from 0.10 to 0.19, indicating additive effects with a slight trend to synergism, which was statistically significant for the Hg–Pb combination only. Hayes’ additive indices were within the margins considered acceptable to describe additive interactions. Keywords: toxicity test; sea-urchin; mercury; trace metals; interaction; embryogenesis; larval growth
Introduction Human activities strongly increase the background levels of toxic trace metals such as mercury, copper, lead and cadmium in natural waters. Chemical analyses allow a determination of the degree and nature of pollution, but they do not provide evidence as to the biological consequences (Chapman et al., 1987). Bioassays allow the detection of these effects by measuring the biological responses of marine organisms, particularly in their highly sensitive early life stages (His et al., 1999a). Sea-urchins have been widely used to provide biological material for embryo-larval ∗ To whom correspondence should be addressed: Tel.: (34)986 812648; Fax: (34)986 812556; E-mail:
[email protected]
bioassays (e.g. Bougis et al., 1979; Kobayashi, 1981; Pagano et al., 1986; Dinnel and Stober, 1987; Zhadan et al., 1992), due to the ease of obtaining gametes and in vitro fertilisation. Many studies have demonstrated the sensitivity of sea-urchin embryos to single metals at concentrations in the 0.01–0.1 mg/l range for Hg and Cu, and 0.1–10 mg/l for Cd and Pb (Waterman, 1937; Kobayashi, 1981; Carr, 1996; Warnau et al., 1996). In contrast, few studies on sea urchins have addressed the interactions of metals, despite the fact that effluents running into the marine environment normally contain more than one trace metal in significant amounts. The toxicity of a chemical can be enhanced (positive interaction or synergism), reduced (negative interaction or antagonism), or unaffected (no interaction) by
264 Fern´andez and Beiras the presence of another toxicant. If two chemicals in a mixture do not affect the toxicity of each other, simple joint action is present. The combined effects of two chemicals with simple joint action are additive, i.e. they may be predicted by the summation of the doses of the individual chemicals after adjustment for the differences in toxic potencies (toxic units). For two chemicals A and B, dose-additivity can be experimentally tested by the calculation of the interaction factor (IF) IF = dA /DA + dB /DB where di and Di are the doses of each chemical producing the same effect in a mixture (d) and individually (D) (Cassee et al., 1999). Additive effects are demonstrated when IF = 1. Usually the dose referred to is the median effective concentration (EC50 ), because this is the parameter that can be calculated with greatest precision from a regression of concentration on effect. Investigating the interaction between two toxicants may be complicated by the fact that this may be dose-dependent; i.e. it may vary with the proportions of the chemicals used in the mixture. Therefore, it is advisable to test mixtures of equal proportions of the expected individual EC50 values. The usual experimental approach to study interactions consists of testing the toxicity of the chemicals in combination and comparing the results with those from tests with single chemicals. Additive effects are considered to exist when no significant deviations from the expected combined effect are found. Materials and methods Biological material Adult sea-urchins (Paracentrotus lividus) were collected from a subtidal population at Canido
(42◦ 11 36 N, 8◦ 49 30 W). Gametes were obtained by dissection from a pair of adults and observed under the microscope to check their maturity (spherical eggs and mobile sperm). The eggs were transferred into a sterile measuring cylinder containing artificial seawater (ASW) (see below). A few µl of dry sperm were collected directly from the gonad with a Pasteur pipette, added to the suspension of eggs and carefully stirred with a plunger to allow fertilisation. Four samples of 20 µl were taken to record fertilisation success (assessed by the percentage of eggs showing a fertilisation membrane) and egg density. Six hundred eggs were placed in 20 ml polypropylene vials for each test, and five replicates per treatment were assayed. Metal solutions Metal solutions were obtained by dissolving HgCl2 , CuCl2 · 2H2 O, CdCl2 · H2 O and Pb(NO3 )2 in chemically defined ASW made up from double distilled water and analytical grade reagents following the formulation of His et al. (1997). Preliminary trials showed no significant differences in embryogenesis success when using either 0.22 µm filtered natural seawater or ASW. ASW was chosen as an experimental medium because the background trace metal levels are known and are below those causing toxic effects to P. lividus, and especially because it is free of organic matter, which may otherwise influence the results of toxicity tests. In order to test the toxicity of single chemicals (individual series), 5 concentrations of each metal in geometric progression (2 higher and 2 lower than the expected EC50 ) were prepared. The toxicity of Hg was also tested in binary combination with every other metal. For each combination, five treatments were prepared by addition of equal fractions of the metal concentrations tested in the individual series. Table 1
Table 1. Nominal concentrations (µg/l) of metals used in the individual and mixed series Number of tests
Concentrations (µg/l)
Metal Hg Cu Pb Cd Hg + Cu Hg + Pb Hg + Cd
41 2 2 2 1 1 1
4 16 250 2000 2+8 1 + 64 2.8 + 1024
8 32 500 4000 4 + 16 2 + 128 5.6 + 2048
1 Two additional tests were performed in teflon vials.
16 64 1000 8000 8 + 32 4 + 256 11.3 + 4096
32 128 2000 16,000 16 + 64 8 + 512 22.6 + 8192
64 256 4000 32,000 32 + 128 16 + 1024 45.2 + 16,384
Embryotoxicity of metal combinations to sea-urchin summarises the concentrations of metals tested in the individual and mixed series. Incubation and endpoints The vials containing the fertilised eggs were incubated at 20 ◦ C for 48 h. These conditions allowed complete development of embryos into pluteus larvae, but minimised background mortality and the duration of the tests (Fern´andez, 1999). Five replicates per concentration and five controls with ASW only, were prepared. After the incubation period the larvae were fixed with a few drops of 40% formalin. The endpoints recorded in each vial were the percentage of fully developed 4-arm pluteus larvae (n = 100) and the mean larval length (n = 25).
265
The most sensitive endpoint recorded, the Lowest Observed Effect Concentration (LOEC) was determined from the larval length data by using ANOVA and Tukey’s test. The LOEC was defined as the lowest test concentration showing significant differences in the median length of larvae, compared to the control. To identify the kind of interaction in binary mixtures of Hg with other metals, the additive toxicity index (S) of Marking and Dawson (1975) and its 95% confidence intervals (CI) were calculated. Additive, more-than-additive (synergism) and less-than-additive (antagonism) effects are indicated by zero, positive and negative values of this index respectively. The inclusion of zero within the 95% CI suggests a lack of significant deviation from additivity, whilst 95% CI lying above or below zero indicate significant synergism or antagonism, respectively.
Chemical analyses Actual metal concentrations were recorded in samples taken at the beginning and the end of the incubation (48 h). Verification of Cu, Pb and Cd concentrations utilised electrochemical methods. Differential pulse anodic stripping voltammetry was used for Cu and Pb, and square wave anodic stripping voltammetry for Cd (Fern´andez et al., 1992; Bersier et al., 1994). Mercury was analysed by automated cold vapour atomic absorption spectrometry (Weltz and Schubert-Jacobs, 1991). Statistical methods Replication and sample sizes were designed to provide a significance level of 95% for both endpoints; an error lower than 5% in larval percentage data; and a power of 90% in the detection of differences of 15 µm in larval length (Burstein, 1971; Sokal and Rohlf, 1981; Zar, 1984). The median effective concentrations (EC50 ) for each metal, i.e. the metal concentrations reducing embryogenesis success to 50% of the control values, were obtained by linear regression of the larval percentage data (normalised by using angular transformation) against the logarithm of concentration (µg of metal ion per litre). The 95% confidence intervals were obtained by the Litchfield–Wilcoxon method (Newman, 1995). When data were not suitable for this method (because not enough partial responses were obtained), the binomial method was used (Newman, 1995).
Results Chemical analyses The nominal and actual metal concentrations at the beginning (t = 0 h) and (except for Hg) at the end (t = 48 h) of tests are shown in Table 2. The measured concentrations followed the intended nominal values closely, ranging from 83% to 133% at the beginning of the assays and from 61% to 127% at the end of the tests. Metal loss from solution during the assays was negligible, in general. Embryogenesis success As illustrated in the concentration-response curves (Figs. 1–4), trace metals inhibited embryogenesis success (defined as the percentage development of 4-arm pluteus larvae), following typical sigmoidal patterns. In addition to this quantitative response, increasing metal concentrations arrested embryo development at earlier stages, causing a progressive shift from pre-plutei to blastula in the developmental stage of the individuals obtained after the incubation period. For Cd, 100% pre-plutei were found at 16,000 µg/l, while differentiation was arrested at the gastrula and blastula stages at 32,000 and 64,000 µg/l respectively. At 64 µg/l of Cu, embryo development was arrested at the gastrula stage, whilst at 128 µg/l only morulae were found. For Hg, pre-plutei and morulae were found at 16 and 32 µg/l respectively, whilst at 128 µg/l development
266 Fern´andez and Beiras Table 2. Nominal and actual concentrations of metals at the beginning (t = 0 h) and the end (t = 48 h) of the tests, and the percentage of the nominal concentration as measured at each time Concentration (µg/l) Metal
Nominal
Actual 0 h
%
Actual 48 h
%
Hg
4 8 16 32 64 16 32 64 128 256 128 256 512 1024 2048 2 4 8 16 32
5.32 10.13 16.31 35.28 52.80 17.91 ± 0.0 32.6 ± 1.1 62.3 ± 5.5 154.1 ± 4.9 272.7 ± 4.1 137.9 ± 5.3 295 ± 9.5 553.3 ± 17.2 1017.6 ± 2.5 1915.1 ± 162.1 2.00 ± 0.04 3.97 ± 0.07 8.08 ± 0.19 15.19 ± 0.38 31.56 ± 0.75
133 126.6 101.9 110.2 82.5 111.9 101.9 97.3 120.4 106.5 107.7 115.2 108.1 99.4 93.5 100.0 99.25 101.0 94.9 98.6
n.m.1 n.m. n.m. n.m. n.m. 17.9 ± 1.9 38.6 ± 1.6 72.1 ± 2.7 117.8 ± 3.3 277.1 ± 8.7 113.2 ± 0.0 238.8 ± 7.0 652.1 ± 18.9 862.7 ± 48.9 1247.6 ± 162.1 2.01 ± 0.04 3.94 ± 0.07 8.11 ± 0.19 15.19 ± 0.38 29.20 ± 0.75
111.9 120.6 112.6 92.0 108.2 88.4 93.3 127.4 84.2 60.9 100.5 98.5 101.4 94.9 91.3
Cu
Pb
Cd2
1 n.m: not measured. 2 mg/l.
Figure 1. Concentration-response curves for mercury tests with Paracentrotus lividus sea-urchin embryos incubated at 20 ◦ C for 48 h.
was arrested after the first cellular division. Similarly, for Pb at 500 µg/l embryos developed to the preplutei stage, while at 1000 µg/l only gastrulae were found. Table 3 shows that the lowest EC50 was obtained for Hg (21.9 µg/l), followed by Cu (66.8 µg/l), Pb (509 µg/l) and Cd (9240 µg/l). The four elements
Figure 2. Concentration-response curves for copper tests with Paracentrotus lividus sea-urchin embryos incubated at 20 ◦ C for 48 h.
tested were ranked consistently in the following order from highest to lowest toxicity: Hg > Cu > Pb > Cd. Larval growth Concentration dependence in growth inhibition was also found for all the metals tested, as illustrated in
Embryotoxicity of metal combinations to sea-urchin
267
Figure 3. Concentration-response curves for lead tests with Paracentrotus lividus sea-urchin embryos incubated at 20 ◦ C for 48 h.
Figure 4. Concentration-response curves for cadmium tests with Paracentrotus lividus sea-urchin embryos incubated at 20 ◦ C for 48 h. Table 3. Median effective concentrations (EC50 ; µM and µg/l) and their coefficients of variation (C.V.) and the lowest observed effect concentrations (LOEC; µM and µg/l) for Hg, Cu, Cd and Pb Metal
EC50 (µM)
EC50 (µg/l)
C.V.
LOEC (µM)
LOEC (µg/l)
Hg Cu Pb Cd
0.11 1.1 2.5 82
21.95 66.76 509.5 9240
29.8 11.2 3.2 1.7
0.04 0.25 1.21 0.44
8 16 250 50
Fig. 5. Concentrations of 250 and 500 µg/l of Pb caused a larval length reduction of 5.3% and 13%, respectively. For Cu, concentrations of 16, 32 and 64 µg/l caused reductions of 2.5%, 13.7% and 45.1% in mean larval lengths. An inhibition of 6% in larval growth was
Figure 5. Mean larval length (n = 25) at different concentrations of Cd (A), Pb (B), Cu (C) and Hg (D), after incubation for 48 h at 20 ◦ C. Error bars show the standard deviation.
observed at 8 µg/l of Hg and this increased to 27% at 16 µg/l. The LOEC inhibiting early larval growth were markedly lower than those causing 50% embryogenesis failure (EC50 ) for all the metals tested, and especially for Cd (Table 3). Early larval growth is therefore a more sensitive response than embryogenesis success in P. lividus. For example, 50 µg/l of Cd reduced mean larval length by 33.7% compared to control values, whilst the embryogenesis inhibition EC50 for this metal was 9240 µg/l.
268 Fern´andez and Beiras Table 4. The additive toxicity index (S) and 95% confidence intervals (±CI 95%) for binary combinations of mercury with the other metals studied
Hg + Cu Hg + Cd Hg + Pb
S
−CI 95%
+CI 95%
0.10 0.11 0.19
−0.01 n.c.1 0.07
0.22 n.c. 0.33
1 n.c: not calculable.
(with no partial effects) prevented calculation of the 95% CI of the index by the method of Marking and Dawson (1975). Discussion
Figure 6. Percentage of successful embryogenesis for Hg and Cu (A), Hg and Pb (B), and Hg and Cd (C), tested individually and in combination. The X axis shows the toxic units (TU = concentration/EC50 ).
Interactions Figure 6 shows the concentration-response curves for the pairs of metals, tested individually and in combination. Additive index (S) and 95% CI for Hg + Cu and Hg + Pb combinations were 0.10 (−0.01, 0.22) and 0.19 (0.07, 0.33), respectively (Table 4). The zero value is contained within the 95% CI for the index for Hg + Cu, and this implies a simple additive effect. The 95% confidence interval for Hg + Pb are just over zero, so a weak but significant positive interaction (synergism) existed between these two metals. For the Hg + Cd combination, S was 0.11, which is very similar to the additive index for the Hg + Cu combination that was considered to have additive effects. However, in this case, the nature of the response
The results on the embryotoxicity of single metals presented in this study (Table 3) demonstrate that the ranking of toxicity of trace metals to P. lividus embryos decreases as follows: Hg > Cu > Pb > Cd. This agrees with data obtained for embryos of other marine invertebrate species. EC50 values for mercury ranged from 15 to 30 µg/l in the current work, compared to 4–8 µg/l (His et al., 1999b) and 20–40 µg/l (Warnau et al., 1996) respectively. His et al. (1999b) used a more restrictive criterion for larval normality, that considered certain morphological features of the larval rods disregarded here. Concerning Cu, we found an EC50 of 67 µg/l, while His et al. (1999b) reported 50–100 µg/l, and Bougis et al. (1979) cited an EC50 of 45 µg/l. For Pb, we found an EC50 of 510 µg/l; Warnau and Pagano (1994) reported an EC50 for this species of 414 µg/l, and His et al. (1999b) cited a value of 482 µg/l for Pb acetate. Although the results found for Cd are more variable, both previous results and this study demonstrate that Cd is the least toxic metal among these four for sea-urchin embryos. The EC50 values reported for Cd and P. lividus are 3800 µg/l (Heyvang, 1994); 3372–11,241 µg/l (Warnau et al., 1996), and 9240 µg/l in the present study. Similarly, for Arbacia punctulata the EC50 value for Cd was 7380 µg/l (Carr, 1996), and for Dendraster excentricus, 5200–10,800 µg/l (Dinnel et al., 1989). The present study shows that the sensitivity of the sea-urchin embryo bioassay to metals can be markedly improved if the 48 h larval length is used as an endpoint, in addition to the percentage of developed pluteus larvae. By using image analysis, the former method could easily be automated and rendered suitable for routine
Embryotoxicity of metal combinations to sea-urchin use. Larval length provides a more gradual response to metal concentration, and is also a normally distributed variable. Thus ANOVA and a posteriori tests can be applied to obtain the lowest concentration that significantly inhibits larval growth; i.e. the LOEC. Whereas for Hg, Cu and Pb this procedure increased the bioassay sensitivity approximately 3-fold, the LOEC value for Cd was more than two orders of magnitude lower than the EC50 for embryogenesis inhibition. In Galician coastal waters, local maximum concentrations of 190 µg/l for Cu and 44 µg/l for Pb have been recorded (unpublished data). The present study shows that 16 µg/l of Cu decreases the larval growth of P. lividus, and 67 µg/l arrests the embryo development of 50% of the test population. The Cu concentrations found in natural waters are thus occasionally above the effective levels, indicating the presence of ecological risk. For Pb, the concentrations found in Galician waters are below the toxic values reported here, but its contribution in toxicity units (0.1) to overall metal toxicity is not negligible. By contrast the concentrations measured in coastal waters for Hg and Cd are orders of magnitude below those causing toxic effects to P. lividus. The toxicity of Hg in combination with Cu, Pb and Cd was also investigated in the current study. The resulting additive indices for these combinations were 0.10, 0.19 and 0.11 respectively, where a value of 0 indicates strict additivity. Thus, the combined toxicity in each case was slightly higher (10% to 19%) than predicted by the additive model. However, these differences were only significant for the Hg + Pb combination, which would be considered as synergistic following the method of Marking & Dawson (1975). This method relies on measurement of the confidence intervals of the EC50 values recorded for the individual toxicants and their combined effects. These parameters were obtained using the Litchfield-Wilcoxon method, which is based on the slope of the toxicity curve, disregarding the goodness of fit of the experimental data to that curve. Thus, when the slope is high, the confidence intervals are narrow, no matter how closely the points fit the regression model. Other authors propose alternative views. Cassee et al. (1999) warned about the use of simple additive indices, and state in their recent review that the statistical basis for testing whether a specific index deviates from additivity is very complex. Hayes (1991) proposed that combinations yielding additive indices between 0.5 and 2 (corresponding to the range from −1 to 1 for the index of Marking and Dawson, 1975)
269
should not be considered as significantly different from additivity. Following this criterion, the interactions studied in the present study would be all considered as additive. We have observed additive or very slightly synergistic effects for combinations of Hg with Cu, Pb and Cd. Additive effects have been observed in bivalves for mixtures of Cu–Ag (Coglianese and Martin, 1981) and Mn–Mo (Morgan et al., 1986). MacInnes (1980) found combined effects within 25% of those predicted by strict additivity for Hg–Cu and Hg–Zn combinations, and a synergistic interaction for a Cu–Zn mixture. A Zn–Cd mixture has been reported to be synergistic for white shrimp, Penaeus setiferus (Vanegas et al., 1997), but antagonistic for bivalve embryos (Pavicic, 1980). Interactions described for mixtures of Hg with Cu or Zn for C. gigas larvae were generally antagonistic (Pereira et al., 1998), but again they did not deviate more than 25% from the predictions of the additive model. Further, the interaction changed depending on the concentrations employed in combination. MacInnes (1980) also found that the degree of synergism increased as the concentrations in the mixtures increased, and Moreau et al. (1999) stated that additive indices relying on EC50 values can be applied to concentrations near the EC50 values only. From the environmental standpoint, the main goal is to predict the toxicity of mixtures of pollutants, rather than to classify the combinations of toxicants by their types of interaction. From the experimental data reviewed above and the present study it can be concluded that using an additive model, generally allows the prediction of the combined toxicity of pairs of metals within a 25% error margin. To predict the impact of metal-polluted effluents in the marine environment it would be necessary to test not only binary mixtures, but also combinations of three or more metals, or even combinations of metals with organic pollutants. This approach becomes increasingly difficult because of the exponential multiplication of the number of test groups with the increasing numbers of chemicals in a mixture (Cassee et al., 1999). Furthermore, metals not only interact among themselves, but their combined toxicity also depends on environmental factors which affect metal speciation (such as pH, salinity, and especially the levels and types of dissolved organic matter). Estuarine waters are often rich in organic matter such as humic acids, with the ability to complex free metal ions and reduce their bioavailability. Estuaries are also
270 Fern´andez and Beiras exposed to periods of reduced pH and salinity, and these conditions are known to commonly increase metal toxicity compared to oceanic seawater. All these factors should be taken into account in order to extrapolate results from laboratory toxicity tests to the marine environment. References Bersier, P.M., Howell, J. and Brunlett, C. (1994). Advanced electroanalytical techniques versus atomic absortion spectrometry inductively coupled plasma atomic-emission spectrometry and inductively coupled plasma mass spectrometry in environmental analyses. Analyst, February 119–219. ´ Bougis, P., Corre, M.C. and Etienne, M. (1979). Sea-urchin larvae as a tool for assessment of the quality of sea water. Ann. Inst. Oc´eanograph. Paris 55(1), 21–6. Burstein, H. (1971). Attribute Sampling: Tables and Explanations. New York: Mcgraw-Hill Company. Carr, R.S. (1996). Sediment quality assessment studies of Tampa Bay, Florida. Environ. Toxicol. Chem. 15(7), 1218–31. Cassee, F.R., S¨uhnel, J., Groten, J.P. and Feron, V. (1999). Toxicology of chemical mixtures. In B. Ballantyne, T. Marrs and T. Syversen (eds) General and Applied Toxicology, Vol. 1, pp. 303–19. London: MacMillan. Chapman, P.M., Dexter, R.N. and Long, E.R. (1987). Synoptic measures of sediments contamination, toxicity and infaunal community composition (the Sediment Quality Triad) in San Francisco Bay. Marine Ecol. Progr. Ser. 37, 75–93. Coglianese, M.P. and Martin, M. (1981). Individual and interactive effects of environmental stress on the embryonic development of the Pacific Oyster, Crassostrea gigas. I. The toxicity of copper and silver. Marine Environ. Res. 5, 13–27. Dinnel, P.A. and Stober, Q.J. (1987). Application of the sea urchin sperm bioassay to sewage treatment efficiency and toxicity in marine waters. Marine Environ. Res. 21, 121–33. Dinnel, P.A., Link, J.M., Stober, Q.J., Letorneau, M.W. and Roberts, W.E. (1989). Comparative sensitivity of sea urchin sperm bioassays to metals and pesticide toxicity tests. Arch. Environ. Contam. Toxicol. 18, 748–55. Fern´andez, J.M., Hayes, P.J. and Smyth, M.R. (1992). Analytical voltammetry in environmental science. I. Inorganic species. In M.R. Smyth and J.G. Vos (eds) Analytical Voltametry (Wilson & Wilson’s Comprehensive Analytical Chemistry) Vol 27. Amsterdam: Elsevier. Fern´andez, N. (1999). Toxicidad del mercurio, cobre, cadmio y plomo sobre la embriog´enesis del erizo de mar (Paracentrotus lividus). MSc Thesis, Universidade de Vigo. Hayes, W.J. (1991). Dosage and other factors influencing toxicity. In Handbook of Pesticide Toxicology, Vol. 1. General Principles. London: Academic Press. Heyvang, I. (1994). Toxicit´e des micropollutants en milieu marin. Mise au point d’un test simplifi´e bas´e sur l’utilisation d’oeufs d’embryons et de pluteus de Paracentrotus lividus. Exemples d’applications. IFREMER. His, E., Budzinski, H., Geffard, O. and Beiras, R. (1997). Effects of a hydrocarbon-polluted sediment on Crassostrea gigas (Thunberg)
metamorphosis. Comptes Rendus de l’Acad´emie des Sciences Paris, S´erie Sciences de la Vie 320, 797–803 His, E., Beiras R. and Seaman, M.N.L. (1999a). The assessment of marine pollution. Bioassays with bivalve embryos and larvae. In A.I. Southward, P.A. Tyler and C.M. Young (eds) Advances in Marine Biology, Vol. 37. London: Academic Press. His, E., Heyvang, I., Geffard, O. and De Montaudouin, X. (1999b). A comparison between oyster (Crassostrea gigas) and sea urchin (Paracentrotus lividus) larval bioassay for toxicological studies. Water Res. 33(7), 1706–18. Kobayashi, N. (1981). Comparative toxicity of various chemicals, oil extracts and oil dispersant extracts to Canadian and Japanese sea urchin eggs. Publications of Seto Marine Biology Laboratory XXVI(1/3), 123–33. MacInnes, J.R. (1980). Response of embryos of the American Oyster, Crassostrea virginica, to heavy metals mixtures. Marine Environ. Res. 4, 217–27. Marking, L.L. and Dawson, V.K. (1975). Method of assessment of toxicity or efficacy of mixtures of chemicals. US Fish Wildlife Service Invest. Control 67, 1–8 Moreau, C.J., Klerks, P.L. and Haas, C.N. (1999). Interaction between phenanthrene and zinc in the toxicity to the sheepshead minnow (Cyprinodon variegatus). Archiv. Environ. Contam. Toxicol. 37, 251–7. Morgan, J.D., Mitchell, D.G. and Chapman, P.M. (1986). Individual and combined toxicity of manganese and molybdenum to mussel, Mytilus edulis, larvae. Bull. Environ. Contam. Toxicol. 37, 303–7. Newman, M.C. (1995). Quantitative Methods in Aquatic Toxicology. Boca Raton: CRC Press. Pavicic, J. (1980). Interaction of Cd and Zn in relation to oxygen consumption in early stages of marine bivalve molluscs. V Journ´ees ´ d’Etudes des Pollutions 627–34, Cagliari C.I.E.S.M. Pagano, M., Cipollaro, M., Corsale, G., Esposito, A., Ragucci, E., Giordano, G.G. and Trieff, N.M. (1986). The sea urchin: bioassay for the assessment of damage from environmental contaminants. Community Toxicity Testing, ASTM STP 920, John Cairns, Jr. ed. American Society for Testing and Materials, Philadelphia. Pereira, S.A., Nascimento, A., Smith, H., Leite, N.L., de Ara´ujo, M.S. and Silva, A. (1998). The combined effects of temperature and metals copper, zinc and mercury on the embryological development of the mangrove oyster, Crassostrea rhizophorae. Ecotoxicol. Environ. Restoration 1(1), 21–32. Sokal, R. and Rohlf, F. (1981). Biometry, 2nd edn. New York: W.H. Freeman and Company. Vanegas, C., Espina, S., Botello, A.V. and Villanueva, S. (1997). Acute toxicity and synergism of cadmium and zinc in white shrimp, Penaeus setiferus, juveniles. Bull. Environ. Contam. Toxicol. 58, 87–92. Warnau, M., Iaccarino, M., De Biase, A., Temara, A., Jangoux, M., Dubois, P. and Pagano, G. (1996). Spermiotoxicity and embryotoxicity of heavy metals in the Echinoid Paracentrotus lividus. Environ. Toxicol. Chem. 15(11), 1931–6. Warnau, M. and Pagano, G. (1994). Developmental toxicity of PbCl2 in the Echinoid Paracentrotus lividus (Echinodermata) Bull. Environ. Contam. Toxicol. 53, 434–41. Waterman, A.J. (1937). Effects of salts of heavy metals on development of sea urchin, Arbacia punctulata. Marine Biol. Lab. LXXIII(3), 401–20.
Embryotoxicity of metal combinations to sea-urchin Weltz, B. and Schubert-Jacobs, M. (1991). Evaluation of the flow injection system and optimization of parameters for hydride generation atomic absorption spectrometry. Atom. Spectrosc. 12(4), 91–103. Zar, J.H. (1984) Biostatistical Analysis, 2nd edn. London: PrenticeHall International.
271
Zhadan, P.M., Vashchenco, M.A., Malakhov, V.V., Medvedeva, L.A. and Gareyeva, R.V. (1992). The effect of environmental pollution, hydrocarbons and heavy metals on reproduction of sea urchins and bivalves. In V.I. Ilyichev and V.V. Anikiev (eds) Oceanic and Anthropogenic Controls of Life in the Pacific Ocean, pp. 267–86. Dordrecht: Kluwer Academic Publishers.
