Toxicity and bioaccumulation of copper and lead in five marine microalgae

ARTICLE IN PRESS Ecotoxicology and Environmental Safety 72 (2009) 1503–1513

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Toxicity and bioaccumulation of copper and lead in five marine microalgae Bibiana Debelius a,, Jesu´s M. Forja a, A´ngel DelValls a, Luis M. Lubia´n b a b

´diz, Spain ´ blica Saharaui, s/n, 11510 Puerto Real, Ca Departamento de Quı´mica-Fı´sica, Universidad de Ciencias del Mar y Ambientales, Avda Repu ´diz, Spain ´ blica Saharaui, 2, 11510 Puerto Real, Ca Instituto de Ciencias Marinas de Andalucı´a, Avda Repu

a r t i c l e in fo

abstract

Article history: Received 11 September 2008 Received in revised form 2 April 2009 Accepted 4 April 2009 Available online 8 May 2009

On five marine microalgae with the same biovolume quantity (Tetraselmis chuii, Rhodomonas salina, Chaetoceros sp., Isochrysis galbana (T-iso) and Nannochloropsis gaditana) 72-h exposure toxicity tests with copper and lead were performed. For both metals, 72-h EC50s showed T. chuii as the most tolerant and R. salina as one of the most sensitive. Besides copper and lead EC50 concentrations, metal concentrations in solution and accumulated on/in the cell where also analysed. T. chuii, the most tolerant species accumulated high copper concentrations (EC50(Cu) ¼ 330 mg L1; EC50(Pb) ¼ 2600 mg L1), and R. salina the most sensitive to copper, accumulated the highest amount of this metal (EC50(Cu) ¼ 50 mg L1). Results of this study show that there is no specific relationship between cell tolerance and accumulated metal on/in the cell. On the other hand, due to an established evidence of the influence of cellular density in microalgae toxicity tests, this effect was also studied. Results showed reduced EC50 values when initial cellular densities decreased. In this study, the term ‘‘toxic cellular quota’’ was used to express all data. This allowed, in a single expression, the combination of two parameters that clearly influence growth, cellular density and toxic concentration. & 2009 Elsevier Inc. All rights reserved.

Keywords: Copper Lead Marine microalgae Exposition toxicity tests Toxic cellular quota Accumulation Growth inhibition

1. Introduction The microbial community, including protozoa, algae, bacteria and fungi, comprises the bulk of the biomass in many systems and fulfils crucial roles in energy flow pathways and nutrient regeneration (Cairns et al., 1992). Microalgae are the primary producers at the base of the aquatic food chain. They are one of the first groups to be affected by metal contamination, and therefore they provide important information for predicting the environmental impact of pollution. Besides, some authors, e.g. Stauber and Davies (2000), established that microbial tests, using bacteria and single-celled algae, in particular, are very sensitive to toxic compounds and that they do not suffer from the animalethics constraints associated with using higher organisms such as invertebrates and fish. Several protocols for algal toxicity testing are already established for freshwater microalgae, but there are few that can be described as ‘‘standard’’ for marine microalgae, although guidelines have been published and some marine species have been recommended as test strains (Walsh, 1988, 1994). Due to their frequent involvement and persistence in marine environmental pollution, metals are one group of contaminants

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E-mail address: [email protected] (B. Debelius). 0147-6513/$ - see front matter & 2009 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2009.04.006

most commonly assessed in the literature (e.g., Sunda and Lewis, 1978; Canterford and Canterford, 1980; Levy et al., 2008). Several metals are essential for living organisms at very low concentrations, but at high concentrations most are toxic and have a direct and adverse influence on various physiological and biochemical processes. Copper belongs to the category of ‘‘essential metals’’ and participates in growth, metabolism and enzyme activities. The effect of copper is of interest, as this element has become a widespread contaminant due to its use as an algaecide and a fungicide in agriculture (Baro´n et al., 1995). It is one of the most toxic metals to microalgae, and can be toxic at concentrations as low as 1 mg L1. In contrast, there is limited data on lead toxicity in marine microalgae (Vasconcelos and Leal, 2001; Yap et al., 2004; Satoh et al., 2005). One of the most studied toxic effects of metals on microorganisms is growth inhibition. Attempts to standardize growth inhibition tests with microalgae for regulatory purposes have revealed a number of methodological problems. Inter-laboratory comparisons carried out demonstrate that, to date, most of the test procedures in common use do not give comparable results (Nyholm and Ka¨llqvist, 1989). Most authors conclude that the metal concentration that affects growth in microalgae is largely variable and depends on the species used, cell density, composition of medium or physical culture conditions (Genter, 1996; Moreno-Garrido et al., 2000; Stauber and Davies, 2000; Franklin et al., 2002). This causes large differences in toxicity results and

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prevents reliable comparison of the data obtained by different authors. Luoma and Rainbow (2005) established that the toxicity effect is not only a function of exposure to contaminants, but also of internal biological sequestration. Besides, relatively little is known about the accumulation of the different contaminants by marine microalgae, which may be an important step in the understanding of inter-species tolerance differences (Levy et al., 2007, 2008). From an ecotoxicological point of view, standard guidelines of algal growth inhibition also use unrealistically high cell densities, mostly 104–106 cells mL1. This cell density parameter is very important because: on the one hand, there is established evidence that the sensitivity of toxicity tests increases with a decrease in initial cell density (Vasseur et al., 1988; Stratton and Giles, 1990; Moreno-Garrido et al., 2000; Franklin et al., 2002) and, on the other hand, metal bioavailability studies show that the cells themselves can alter the speciation of metals in solution and consequently this can lead to over- or underestimation of metal bioavailability in natural waters. Steemann and Wium-Andersen (1970) referred to the existence of a phenomenon that they called the biomass effect, and concluded that copper toxicity decreased as the concentration of cells in suspension increased. Therefore, toxicity tests should use lower initial cell densities to develop more environmentally relevant toxicity tests using densities that more closely reflect those found in aquatic systems (102–103 cells mL1). To date, the use of such low cell densities has not been specified in standard tests, due to the difficulty of reliably counting cells at such low densities. Franklin et al. (2002) and other authors suggest that standard static laboratory bioassays using 104–105 algal cells mL1 may seriously underestimate metal toxicity in natural waters. Flow cytometry, which enables the measurement of very low cell densities and rapid multi-parameter analysis of individual cells, makes an excellent tool to solve this problem. Flow cytometry has been applied in the study of the aquatic environment since the 1980s (Franqueira et al., 2000). As an ecotoxicological tool, it has been used to develop more environmentally relevant bioassays for the rapid assessment of the bioavailability and toxicity of contaminants in aquatic systems (Franqueira et al., 2000, 2001; MorenoGarrido et al., 2001; Adams and Stauber, 2004; Binet and Stauber, 2006). This study has several objectives. The first is to assess the toxicity of copper and lead, comparing sensitivities and metal bioaccumulation, at the same biovolume, for five marine microalgae Tetraselmis chuii, Rhodomonas salina, Chaetoceros sp., Isochrysis galbana (T-iso) and Nannochloropsis gaditana. These five marine microalgae were chosen because they represent different classes of marine microalgae and are all easy to culture. The second objective is to consider the influence of initial cell density on the EC50 values; different cell densities were used to demonstrate the importance of this parameter in toxicity tests. The third objective is to analyse the complete study using the term ‘‘toxic cellular quota’’; this term combines the two main factors studied in marine toxicity tests, contaminant concentration and cell density.

2. Materials and methods 2.1. Microalgal cultures The species of marine microalgae used in this experiment are shown in Table 1. All species, corresponding to different classes, Tetraselmis chuii Butcher (Prasinophyceae), Rhodomonas salina (Wislouch) Hill and Wetherbee (Cryptophyceae), Chaetoceros sp. (Bacillariophyceae), Isochrysis aff. galbana Haines T-iso (Prymnesiophyceae) and Nannochloropsis gaditana Lubia´n (Eustigmatophyceae) were obtained from the Marine Microalgal Culture Collection of the Instituto de Ciencias Marinas de Andalucia (Lubia´n and Yu´fera, 1989). These algal strains were maintained with natural seawater from the Bay of Cadiz, previously filtered (GF/F Whatman). Seawater was sterilized by autoclaving. Background copper and lead concentrations resulted under 1 mg L1 (limit detection value of voltammetry). The medium was enriched with a final 3 (4 mM) and, in the case of concentration composed of NO 3 (124 mM), PO4 Chaetoceros sp., SiO2 (50 mM). This medium, rather than another standard medium, such as Guillard’s f/2 formulation, was chosen for two reasons: it has been demonstrated that EDTA greatly decreases the toxicity of metals due to the chelating properties of the molecules (Sunda and Guillard, 1976; Moreno-Garrido, 1997); and it is particularly interesting to perform a bioassay in conditions most similar to those found naturally. Cultures were maintained at 2071 1C, under continuous white light (30 mmolm2 s2) in a culture chamber and aseptically transferred to fresh media weekly to maintain cells in logarithmic growth phase.

2.2. Algal growth inhibition toxicity tests All five microalgae were inoculated from cultures in a log phase growth to obtain the initial cell densities shown in Table 1. Based on the volume of each strain, measured with a Coulter Counter, strains were brought to the same biovolume quantity, 325  104 mm3 mL1; T. chuii was chosen as the reference. This allows to compare the results of metal sensitivity obtained for the different strains. All toxicity tests were kept at the same conditions and fresh media established for the stock of five microalgae, at 2071 1C, under continuous white light (30 mmolm2 s2) in a culture chamber. Continuous light was used instead of light-dark period, following the same culture conditions established in our laboratory for the Marine Microalgal Culture Collection of the Instituto de Ciencias Marinas de Andalucia (Lubia´n and Yu´fera, 1989) In prior tests, a wide range of copper and lead concentrations was assayed in order to find the appropriate range of concentrations for metal toxicity to each species of microalgae. For the 72-h toxicity tests, glass Pyrex tubes were used to hold 10 mL of test medium, placed on an orbital shaker to homogenize the medium during the bioassay. Exponentially growing populations of each microalgal species were exposed to ten progressively increasing nominal concentrations of each metal, including a control (Cu: 0, 5, 10, 20, 40, 80, 120, 200, 300 and 600 mg L1; Pb: 0, 50, 100, 250, 500, 800, 1000, 1600, 3000 and 6000 mg L1). Considering metal adsorption to the glass Pyrex tubes, samples with copper and lead without microalgae were also included in the assay. All toxicity tests were performed in triplicate. All glassware was cleaned with dilute nitric acid (10%) and rinsed several times with Milli-Q water before experiments. Copper and lead were added as Merck standard. The cell density in the control cultures increased by a factor of 8 within 3 days. All toxicity tests were kept at 2071 1C, under continuous white light (30 mmolm2 2 s ) in a culture chamber. The pH was recorded initially and after 72 h, there was no variations and the pH value was 8.0. EC50 values based on growth inhibition for each strain were calculated by flow cytometry analysis. From each 10 mL glass Pyrex tube, a 2 mL subsample was separated to determinate cell density by flow cytometry.

2.3. Determination of copper and lead concentrations in solution and on/in microalgae Copper and lead concentrations in solution and on/in each microalga were determined for 5 of the ten nominal concentrations established for toxicity tests (Cu: 0, 5, 20, 80, 200 and 600 mg L1; Pb: 0, 50, 250, 800, 1600 and 6000 mg L1).

Table 1 Main characteristics of the five microalgae used in the study. Microalgae species

Classes

Volume (mm3)

Surface (mm2)

Initial density (  104 cell/mL)

Nannochloropsis gaditana Isochrysis galbana (T-iso) Chaetoceros sp. Rhodomonas salina Tetraselmis chuii

Eustigmatophyceae Prymnesiophyceae Bacillariophyceae Cryptophyceae Prasinophyceae

7.7 49 77 212 325

19 65 88 171 228

42 6.6 4.2 1.5 1

ARTICLE IN PRESS B. Debelius et al. / Ecotoxicology and Environmental Safety 72 (2009) 1503–1513 The total dissolved copper and lead concentrations in solution were determined after removing algae by filtration—with a low vacuum o100 mmHg, to eliminate any possible cell losses and contamination from the dissolved metal in solution. Filtration was made with Millipore 0.45 mm pore size cellulose nitrate filters, and immediately acidified to pH 1–2 for later analysis. Multiple tubes were necessary to get enough biomass for cellular metal analysis. To each 0.45 mm filter for each strain and metal concentration, 1.5 mL of suprapur concentrated nitric acid was added for determination of total copper and lead concentrations in and on the algae. Control filters (without microalgae) were carried out through the same procedure, to measure possible copper and lead interference. Also copper and lead concentrations in enriched seawater were analysed.

2.4. Influence of cellular density on toxicity tests In order to study the influence of cell density on toxicity tests, a concentration for lead and copper was set at the approximate EC50 values obtained in the toxicity tests for each strain. Tests with five different initial cell concentrations were then analysed. The established initial cell concentrations for each microalgal strain were the result of multiplying the initial cellular concentrations used for the toxicity tests, by the factors 0.1, 0.5, 1, 5 and 10. All experiments were performed in triplicate.

2.5. Flow cytometry analysis Samples for analysis by flow cytometry were collected from the toxicity tests after 72 h of treatment. These were analysed using a FACScalibur flow cytometer equipped with a 488 nm excitation argon laser and the data were computed with CellQuest software (Beckton-Dickinson). Each culture was immediately analysed for 30–60 s (6000–10,000 events per measurement) from samples previously fixed with 3–4% formaldehyde. Counts, signals of side-angle light scatter (SSC), and autofluorescence (FL3, 4630 nm) were recorded and used as indicators of the cellular size and chlorophyll fluorescence, respectively (Sobrino et al., 2004).

2.6. Copper and lead analysis All copper and lead concentrations in solution and on/in cells were analysed at the end of the toxicity tests using a Metrohm 663 VA stand equipped with an Hg electrode, with 1 mg L1 limit detection (DP-VRA) (Van den Berg, 1989). All samples, when measured, were at a pH E4.6 by addition of 2 mL of sodium acetate buffer. In the case of cellular metal samples, since samples were strongly acidified, they had to be slightly neutralized with a sodium hydroxide ultra pure solution (1 M). Solutions were purged with high-purity nitrogen for 8 min before measurements. A deposition potential of 1.1 V was applied for 300 s, with stirring, followed by a non-stirring period of 10 s. Potential was scanned at 13.3 mV s1 at a pulse amplitude of 30 mV. Besides, metal standard solutions were used to calibrate the ASV response.

2.7. Statistical analysis The 72-h EC50 values were obtained from the adjustment of growth inhibition data (respect to a growth control) to a sigmoid equation corresponding to each contaminant and microalga. Study of significant differences in sensitivity between species was followed by a T-test analysis with Instat3 Software.

3. Results 3.1. Toxicity of copper and lead The degree of growth inhibition was affected in different ways, depending on the strain of microalga and metal concentration added in solution (Fig. 1). All samples after 72 h of metal exposure were analysed by flow cytometry and EC50 values, shown in Table 2, were calculated from fitted sigmoid equations with the form f(x) ¼ a/(1+e((xx0)/b)) with r2 values ranging from 0.88 to 0.99. When ranked in the order of increasing sensitivity to copper we have T. chuii, N. gaditana, Chaetoceros sp., I. galbana (T-iso) and R. salina. However, in the case of lead, the order of increasing sensitivity is T. chuii, I. galbana (T-iso), R. salina, N. gaditana, and Chaetoceros sp. There are large differences in species sensitivities to these metals: for all the five strains, copper EC50 values are all

1505

much lower than lead, confirming that copper is more toxic than lead to marine microalgae. Since all data were obtained by flow cytometry, light scattering and fluorescent properties of algal cells were also determined; these provide additional information regarding the toxic mode of action of copper and lead, which differs between species. All of the strains studied showed an increase in the SSC signal (side-scatter), corresponding to an increase in cell size or change in shape, when exposed to the highest copper and lead concentrations of this study. As is shown in Fig. 2, this increase of cell size with increase of metal concentration also showed a larger variability of SSC signals at higher metal concentrations. This high variability of SSC signal at the higher exposure concentrations of metals may be due to two possible effects. Firstly, at higher metal concentrations there are lower cell densities, this may provoke a major deviation in the statistical results. Secondly, SSC signal does not only correspond to cell size, but can also reflect cell complexity; at higher metal concentrations the existing damaged cells may increase their size as a function of their individual complexity. Besides side-scatter, data with respect to FL3 values, related to chlorophyll a, were also studied. This factor is considered of interest because the effect of metal is related both to growth inhibition and a possible change in cellular chlorophyll content. The result of the average signal values multiplied by the respective cell density in each case, termed ‘‘total FL3’’, were represented as degrees of inhibition (% growth inhibition), and they are functions of toxic concentration. The curves obtained for the relationship between total FL3 and metal concentration (data not shown) showed a strong similarity to those obtained in Fig. 1. No significant differences in the EC50 calculations were found compared to those obtained from the growth inhibition effect of copper and lead.

3.2. Copper and lead accumulation on/in cells Besides calculations of EC50 based on % of growth inhibition, the study of copper and lead distribution in solution ([Cu]S, [Pb]S), celladsorbed metal (XCu, XPb) and metal adsorbed by the glass tubes were also measured. Data of adsorbed metal to glass and to algae were fitted using the typical Freundlich adsorption isotherm (e.g. Fig. 3): X M ¼ K F ½M nS where XM is the accumulated metal per cell in pg cell1, [M]S is the metal concentration in solution in mg L1 and KF and n are constant values. Adsorption of metal to glass was higher for copper than for lead. In the case of copper, adsorbed concentration varied from 6% at the lowest exposure concentration, up to 29% at the highest copper concentration 600 mg L1. In contrast, lead showed a much lower KF value with maximum adsorption at the highest Pb concentration (6000 mg L1), 9%. For copper and lead distribution in solution and on/in cells, final measured copper and lead concentrations varied as a function of microalgae and exposure to metal, as shown in Fig. 4 and Table 3. For the maximum concentration of copper, dissolved copper concentrations in solution varied from 37% for T. chui, to 77% for R. salina when measured at the beginning of the bioassay. In the case of lead, concentrations varied from 55% for T. chuii and N. gaditana, to 100% for R. salina and Chaetoceros sp. Cellular copper and lead (i.e. metal on and in cells) in each strain increased upon exposure to both metals. In the case of R. salina, for copper, and Chaetoceros sp., in the case of lead, showed maximum Kd (solution-cell metal concentration partition coefficient measured at 72 h, shown in Table 5), conferring these strains a large

ARTICLE IN PRESS B. Debelius et al. / Ecotoxicology and Environmental Safety 72 (2009) 1503–1513

% Growth Inhibition

1506

100

100

80

80

60

60

40

40 20

20

N. gaditana

N. gaditana 0

0

% Growth Inhibition

0

200

400

100

80

80

60

60

40

40

% Growth Inhibition

2000

4000

6000

20

20 I. galbana 0

200

400

I. galbana

0

600

0

100

100

80

80

60

60

40

40

1000

2000

3000

4000

20

20 Chatoceros sp.

0 0

% Growth Inhibition

0

100

0

200

400

Chaetoceros sp. 0 0

600

100

100

80

80

60

60

40

40

2000

4000

6000

20

20 R. salina

0 0

% Growth Inhibition

600

200

400

R. salina

0

600

0

100

100

80

80

60

60

40

40

2000

4000

6000

20

20 T. chuii 0

200

400 600 Cu (µg L-1)

T. chuii

0

0 800

0

2000 4000 Pb (µg L-1)

6000

Fig. 1. Seventy-two-hour growth inhibition curves for the five marine microalgae exposed to different measured concentrations of copper and lead.

accumulation of metal, differing significantly from the rest of the strains studied. In this case, this high accumulation corresponds to the most sensitive strains for each of these two metals.

There was no clear relationship between cellular copper or lead and the sensitivity of each algae to copper or lead (based on comparison with EC50 values). Binary representations of cellular

ARTICLE IN PRESS B. Debelius et al. / Ecotoxicology and Environmental Safety 72 (2009) 1503–1513

copper/lead expressed in the different ways (XM, KF/cell volume and Kd) were made vs. EC50 values, with the intention of finding a graphical relationship. These combination did not show any clear tendency for the five strains studied.

4. Discussion From the EC50 values calculated for each microalga, it is found that the contaminant effect of copper results in lower EC50 concentrations than that of lead (Table 2). Of the five strains studied, R. salina and I. galbana showed the lowest values, and are therefore the most sensitive to copper exposure. R. salina, which has already been used in other bioassays, shows a very high sensitivity to many metals (Moreno-Garrido et al., 1999, 2001). The high toxicity observed for copper has already been described by other authors (e.g., Cid et al., 1995; Nalewajko and Olaveson, 1995; Danilov and Ekelund, 2001, etc.). Most of these authors reported on freshwater species of algae commonly used in standard methods. Franklin et al. (2001) obtained EC50 values of 10 mg L1 for the freshwater algae Chlorella sp. and Selenastrum capricornutum. Even in the case of dinoflagellates, Anderson and Morel (1978) stated that 1 mg L1 of copper was enough to interfere with some of the cellular processes of Gonyaulax tamariensis, causing inhibition in their motility. From the previous test performed to find the appropriate range of toxicity for each species of microalga, lead already showed a considerable difference in toxicity range from copper. Lead is shown to be less toxic; therefore its toxicity range is wider. R. salina and I. galbana are the most sensitive species to copper but not to lead, whereas Chaetoceros sp. is more sensitive to lead. T. chuii resulted with the highest EC50 values for both metals. Therefore, this microalga was the most tolerant. These results are in accordance with those obtained by other authors, e.g. Maeda and Sakaguchi (1990), Moreno-Garrido et al. (2001), Hampel et al. (2001) and Ismail et al. (2002).

3.3. Influence of cellular density test For the range of metal concentrations studied, growth inhibition for all five microalgae and the two metals decreased with an increase in the initial cell density of the bioassay. This is shown in Fig. 5, where for the same metal concentration, at lower values of initial cell density, higher values of inhibition percentage are obtained. For example, in the case of N. gaditana and lead , the same lead concentration was more than eight times more toxic at an initial cell density of 4.2  104 cell mL1 (82% growth inhibition) than at a cell density of 2.1 106 cell mL1 (10% growth inhibition). Table 2 Results of EC50 (7SD) values for each metal and microalga tested, ordered in increase of size. Microalgae/EC50 (7SD)

Copper (ppb)

Lead (ppb)

Copper (mM)

Lead (mM)

Nannochloropsis gaditana Isochrysis galbana (T-iso) Chaetoceros sp. Rhodomonas salina Tetraselmis chuii

137720 58730*4 88710* 487104 3307100

740720 1340720 105760 900790 26407200

2.2 0.9 1.4 0.7 5.2

3.6 6.5 0.5 4.3 13

*4

Values resulted with no significant t-tests differences (P40.05).

SSC (a.u.)

1507

600

600

500

500

400

400

300

300

200

200

100

100 0

200

400 Cu (µg

600

0

800

2000

L-1)

4000

Pb (µg

6000

L-1)

Fig. 2. Increase of SSC signal (volume/size) for T. chuii, exposed to increasing concentrations of copper and lead.

1 n = 1.437 KF = 9.170x10-3

2

LOG XPb (pg cell-1)

LOG XCu (pg cell-1)

4

0 -2 -4

n= 0.718 KF = 2.022x10-3

0 -1 -2 -3

-1

0

1

LOG [Cu]S (µg

2 L-1)

3

0

1

2

LOG [Pb]S (µg

3 L-1)

Fig. 3. Linearization of Freundlich adsorption isotherms for copper and lead sorption by I. galbana.

4

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[Cu]s (µg L-1)

600

N. gaditana I. galbana Chaetoceros sp. R. salina T. chuii

400

6000 [Pb]s (µg L-1)

1508

200

4000 2000 0

0 200 400 Cu (µg L-1)

600

6000

60

60

8

cell-1)

40

XPb (pg

XCu (pg cell-1)

0

20

0

2000 4000 Pb (µg L-1)

6000

0

2000 4000 Pb (µg L-1)

6000

6 4 2

0

0 0

200 400 Cu (µg L-1)

600

Fig. 4. Copper and lead concentrations in solution ([M]S) and cellular associated (XM) by each of the five microalgae studied when exposed to different copper and lead concentration.

Table 3 Copper and lead distribution in solution, and adsorbed on/in cells by each microalgal strain for the different metal exposure concentrations studied. Initial measured [Cu] mg L1

[Cu]s mg L1

XCu pg cell1

Initial measured [Pb] mg L1

[Pb]s mg L1

XPb pg cell1

N. gaditana

5.2 16. 7 87.2 186 590

2.7 14 etc 53 107.2 325.1

0.004 0.02 0.13 0.30 6.8

51.4 275.1 824.1 1704 6348

9 130 438.7 1501 3246.6

0.003 0.04 0.1 0.2 1.5

I. galbana

5.2 16. 7 87.2 186 590

0.8 5.1 30.5 98.3 372

0.004 0.01 1.9 9.4 59.9

51.4 275.1 824.1 1704 6348

14 83.2 421 1833.2 6289

0.01 0.1 0.4 0.3 0.6

Chaetoceros sp.

5.2 16.7 87.2 186 590

2.6 3.6 5.6 8.9 303

0.005 0.01 0.02 0.8 8.9

51.4 275.1 824.1 1704 6348

0.2 126.5 712.3 1553.3 6340

0.04 1.8 4.9 7.9 53.7

R. salina

5.2 16.7 87.2 186 590

8.8 15.6 33.7 129.4 453.2

0.1 3.3 53.4 5624

51.4 275.1 824.1 1704 6348

12.6 99.7 472.4 1495.3 6348

0.02 0.8 2.2 3.6

T. chuii

5.2 16.7 87.2 186 590

1.9 10.4 27.0 41 218.3

0.2 1.4 3.4 3.5 12.1

51. 275.1 824.1 1704 6348

13 80.3 301.5 685.8 3718.4

0.04 0.3 1.3 3.1 5.4

Generally, direct comparisons of EC50 values are difficult because of the use of different species, initial cell densities and laboratory set-ups in respect of light illumination, temperature, composition of culture media and exposure time (Table 4). Some authors, e.g. Blanck et al. (1984) have reported variation of more than three orders of magnitude for 13 freshwater species of algae. Kasai et al. (1993) have shown, for 56 strains of algae representing seven taxonomic groups, values between 6.5 and 1500 mg L1 for

the pesticide triazine. The EC50 values for copper and lead in the case of the marine Haptophyte, I. galbana, have been reported by several authors: Cu 110–1000 mg L1 (Wilson and Freeburg, 1980); Cu 30–410 mg L1 (Ismail et al., 2002); Cu 910 mg L1 and Pb 1400 mg L1 (Yap et al., 2004); Cu 4200 mg L1 and Pb 2500 mg L1 (Satoh et al., 2005). Comparing the values of Table 2 with those cited, this study shows lower EC50 values for copper, although they agree in part with those of Ismail et al. (2002). However, as it

ARTICLE IN PRESS B. Debelius et al. / Ecotoxicology and Environmental Safety 72 (2009) 1503–1513

100

100

I. galbana

N. gaditana 80

% Growth Inhibition

% Growth Inhibition

1509

60 40 20

80 60 40 20 0

0 0

1

2 3 4 x106 cell mL-1

0

5

4 8 12 x 105 cell mL-1

16

% Growth Inhibition

100 Chaetoceros sp.

80 60 40 20 0 0

1

2 3 4 x105 cell mL-1 100

R. salina

80 60 40 20

% Growth Inhibition

100 % Growth Inhibition

5

T. chuii

80 60 40 20 0

0 0

4 8 12 x 104 cell mL-1

16

0

2

4 6 8 10 12 x 104 cell mL-1

Fig. 5. Growth inhibition curves for different initial cell densities in five marine microalgae, exposed to a constant concentration of copper (K) and lead (.).

Table 4 EC50 copper values for different microalgae, obtained by different authors. Microalgae species

EC50 (ppb)

References

Aphanizomenon gracile Chlorella sp. D. tertiolecta D. tertiolecta Gonyaulax tamariensis Phaeodactylum tricornutum Phaeodactylum tricornutum Phaeodactylum tricornutum Rhodomonas salina Selenastrum capricornutum Tetraselmis sp.

64 7–16 530 1000 1000 100,000 8 20 30 7–17 47

Lu¨deritz and Nicklish (1989) Franklin et al. (2001, 2002) Levy et al. (2008) Franklin et al. (2001) Anderson and Morel (1978) Cid et al. (1995) Levy et al. (2008) Franklin et al. (2001) Moreno-Garrido et al. (1999) Franklin et al. (2001, 2002) Levy et al. (2008)

can be seen, there are notable differences, even if the organism is the same (as already reported, for copper, by Thomas and Seibert, 1977). Inter-laboratory comparisons have revealed that the results of algal toxicity tests carried out using current or past testing practices are extremely variable. Consequently, the data on algal sensitivity reported in the literature are only of value or use if the method employed is described in detail.

As is shown in Table 2 comparing the toxicity concentrations of copper and lead in mM yields in the following conclusion. The lowest metal concentration inhibits growth by 50% of that for lead in the five microalgae strains studied (with a value of 0.50 mM for Chaetoceros sp.). This is unlike the results in terms of ppb values. Hence, for T. chuii more than twice the mM concentration of lead is needed to produce the same effect as copper. Almost all toxicity bioassays are expressed in terms of ppb/ppm, but it is considered of interest to mention the use of mM in toxicity comparisons between different compounds. From flow cytometry data, SSC signal increases in most of the marine microalgae studied, corresponding to an increase in cell size or change in shape, when exposed to the copper and lead concentrations used (e.g., Fig. 2). This increase/change depends on the strain, concentration and exposure (Franklin et al., 2002). Several authors, using optical and/or electron microscopy, and flow cytometry, have previously observed an increase in cell volume of several species of microalgae in response to toxic levels ˜ os of metals (Fisher et al., 1981; Stauber and Florence, 1987, Bolan et al., 1992; Abalde et al., 1995; Cid et al., 1996; Franqueira et al., 2000; Franklin et al., 2001, 2002). Metals may alter the cell membrane permeability to small cations (Overnell, 1975), and the increase in cell volume is probably due to the increased permeability of the cell membrane to Na+. Alternatively, cells

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continue to fix carbon but cannot divide, leading to enlarged cells as photosynthetic products accumulate. Since all the bioassays were performed with an equal biomass, it is of interest to analyse if there is a relationship between cell volume, size or weight and the EC50 values obtained. Quigg et al. (2006) established that small cells have larger surface area-tovolume ratios and they have been reported to be more sensitive to copper than larger species. As it is observed in Fig. 6 there appears to be a higher tolerance for larger strains. It is also observed that a larger cell surface does not always correspond to a higher EC50 value, and therefore to a lower sensitivity to contaminant. It is important to consider that we are studying the effect on only five different strains. To be able to establish a tendency a larger range of size cells should be followed. In the current study, T. chuii is the largest cell and is also the most tolerant of all five strains for both metals studied, but for the other strains studied results are not clear enough, probably due to other structural and physiological factors that have more influence. For example, N. gaditana is the smallest of the five microalgae studied, but presents a greater tolerance than others; this can probably be due to the very low permeability of its cellular membrane, which gives it a great metal adsorption capacity. It is able to remove 100% of the copper after 24 h in accumulation experiments when exposed to concentrations of 500 mg L1 (Moreno-Garrido et al., 1998), and this could explain why it presents higher EC50 values than other species with a larger cell size, such as I. galbana, Chaetoceros sp., or even R. salina. With reference to this last species, R. salina, a special sensitivity to copper has been detected in spite of its relative cell size, as already pointed out; for this reason other aspects of its physiology need to be studied to find what determines its sensitivity. It is considered that a much wider bioassay in respect of cell size should be performed to establish that tendency of toxicity tolerance to increase with cell size. Some authors, e.g. Levy et al. (2007), establish that there is no clear evidence that small microalgae are more sensitive to copper than larger species.

4.1. Copper and lead distribution

4.2. Influence of cell density Similar effects, to those obtained in this work, of reduced toxicity with increase in algal inoculum, have been reported for a number of contaminants, including Cu, Cd and Zn (Vasseur et al., 1988; Stratton and Giles, 1990; Moreno-Garrido et al., 2000; Franklin et al., 2001, 2002; Stauber et al., 2002). Vasseur et al. (1988) suggested that decreased toxicity with increasing initial cell density was due, in part, to adsorption onto the algal biomass. Measured concentrations of intra-cellular and extra-cellular copper at different cell densities have been studied, e.g. by Franklin et al. (2002), who state that measurements of extra- and intra-cellular copper for Chlorella sp. and S. capricornutum confirmed that less copper was bound to the cells, resulting in less copper uptake and lower toxicity at higher initial cell densities. Toxicity was proportional to the amount of copper associated with the cell at each cell density. As cell density increases, less copper is bound at the cell surface, leading to less copper uptake into the cell and consequently less disruption of cell division. It is possible that adsorbed metal would be available

500

3000

400

2500 EC50Pb (ppb)

EC50Cu (ppb)

Differences in species sensitivity is difficult to determine even when studying toxicant distribution during toxicity tests. Data obtained for cellular-associated copper and lead by microalgae showed that the strains that were the most sensitive to these two metals also had the maximum Kd values, and therefore they adsorbed the most metal per cell. For example R. salina for copper, and Chaetoceros sp. for lead. This agrees with results obtained by other authors, e.g. Koshmanesh et al. (1997) who established that Closterium lunula is most tolerant to copper but at the same time it adsorbed the least amount of copper of all the microalgae studied. This relationship between more-tolerance-less metal adsorbed

did not hold true for T. chuii, which was the most tolerant strain but had higher cellular copper and lead compared to other more sensitive strains, e.g., Chaetoceros sp. and N. gaditana in the case of copper, and N. gaditana and I. galbana in the case of lead. Levy et al. (2007, 2008) established that Tetraselmis sp. had a much higher Kd value for copper than other marine microalgae, suggesting that it adsorbed relatively more copper to cell surfaces compared with other species. Nassiri et al. (1995) established the existence of a correlation of Cu and Cd concentrations with the metal-binding polypeptides phytochelatin in Tetraselmis suecica that could explain its higher resistance, given its ability to accumulate more metal. In an approach of trying to relate cellular metal with cell metal tolerance, the different values corresponding to the XM at the EC50, Freundlich isotherm constants and solution-cell metal partition coefficient measured at the end of the assay (72 h) Kd, are shown in Table 5. In this work T. chuii shows a copper KF value of 0.262 which is much higher than for the rest of the strains’ Freundlich constant values. R. salina and I. galbana with similar sensitivity (similar EC50 values) to this metal showed huge differences in Kd and XCu. In the case of lead, except for Chaetoceros sp. (the most sensitive, besides showing a maximum KF), the rest of the strains appear to accumulate more metal with an increase in cell size. Hence, as results in this study show, it is very difficult to find any existing relationship of toxic metal effect with inter-species sensitivity by cell size, cell wall type and accumulation on/in the cell. Authors such as Quigg et al. (2006) hypothesized that differences in the cell surface reactivity for Cu could contribute to differences in algal sensitivity to this metal.

300 200 100

2000 1500 1000 500

0 0

50

100

150

200

Cell Surface (µm2)

250

0 0

50

100

150

200

Cell Surface (µm2)

Fig. 6. Data plot of cell surface vs. EC50 copper and lead values for the five strains studied.

250

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for cells, and this would explain the increase in toxicity when initial cellular density decreases. As the number of cells increases, less contaminant is bioavailable to each cell, leading to less binding to the cells. This is why Stratton and Giles (1990) state that toxicity values which consider the contaminant that corresponds to each organism in terms of weight per organism or total weight of organisms are more interesting than calculated values of EC50. A possible approach can be the expression already introduced by Moreno-Garrido et al. (2000) and Debelius et al. (2008), the ‘‘toxic cellular quota’’ (Q; mass of metal per cell). Although it is true that metal toxicity in microalgae depends on the strain, its metal uptake into the cell and/or detoxification, the idea of introducing this term, pretends to be an approach to a better unification of literature data by eliminating the effect of cellular density. This term allows cellular density and contaminant concentration values to be combined in one single expression, and therefore fixes two parameters that clearly influence inhibition growth in one. Fig. 7 shows Q values vs. % of growth

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inhibition. Considering the mean cellular volumes for each strain (established in Table 1), the cellular volumes of these five microalgae have a clear influence on the Q value: more metal per cell is necessary for a larger volume cell, previously established by Moreno-Garrido (1997). The Q calculated in the second part of the bioassay, on the influence of cellular densities, showed very similar curves to those obtained in the first part of toxicity tests. This can be seen, by way of an example, in Fig. 8, where all data for T. chuii are shown in terms of Q. From what has been concluded, in toxicity tests the quantity of metal taken up per cell has a bigger influence on toxicity than the number of cells in the medium. The expression in terms of Q provides an easier way to standardize protocols, through the use of a single expression containing the values for metal concentration on/in the cells and number of cells. In addition, the physiology itself may have an overall effect on the way in which the metal is accumulated in the cell. It is demonstrated that there are two phases in metal adsorption by

Table 5 Cellular concentrations of copper and lead at 50% inhibition of growth corresponding to each metal and microalgae studied, and values corresponding to the Freundlich isotherm constant (KF) and solution-cell metal partition coefficient at the end of the assay (Kd).

N.gaditana I. galbana Chaetoceros sp. R. salina T. chuii

XCu at 50% I pg cell1

XPb at 50% I pg cell1

KF [Cu]  103

KF [Pb]  103

Kd [Cu]  108 cell L1

Kd [Pb]  108 cell L1

0.17 1 0.02 17 2.2

0.06 0.2 0.5 0.7 5.7

1.1 9.2 1.3 0.05 262

1.4 2 83 4.3 5.5

2.2 16.4 3 4922 5.1

0.04 0.007 0.7 0.2 0.1

100

80 60 N. gaditana R. salina T. chuii I. galbana Chaetoceros sp.

40 20

% Growth Inhibition

% Growth Inhibition

100

80 60 40 20 0

0 0

20

40 60 80 Q (10-12 g Cu cell-1)

0

100

200 400 600 Q (10-12 g Pb cell-1)

Fig. 7. Growth inhibition in five marine microalgae for increasing values of toxic cellular quota (mass of Cu or Pb per cell).

100 % Growth Inhibition

% Growth Inhibition

100 80 60 40 20

80 60 40 20 0

0 0

100 200 300 Q (10-12gCu cell-1)

400

0

1000 2000 3000 Q (10-12gPb cell-1)

4000

Fig. 8. All growth inhibition data for T. chuii expressed in terms of Q of copper and lead. (K) represents data fixing cellular densities; (m) represents data fixing the metal concentration. Dotted corresponds to the fitted sigmoid curve of all data.

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microalgae: a first phase, not dependent on cellular metabolism, where metal binds to the cellular surface, and a second, slower phase, dependent on metabolism, where metal is accumulated in the interior of the cell (Moreno-Garrido et al., 2000). In the field, additional complications can arise from the effects in the environment that can affect the organisms at the same or different trophic levels. This may affect their tolerances and accumulations of trace metals.

5. Conclusions There are notable differences between toxicity tests, even if the organism is the same. One important fact is cellular density, due to the established variations of metal toxicity effect as a function of cellular density exposure. The use of the term ‘‘toxic cellular quota’’, included in this study, provides an easier way to standardize protocols, through the use of a single expression containing the values for metal concentration and number of cells. Also, the use of flow cytometry is of great advantage in these kind of studies. It enables more environmentally relevant toxicity tests to be performed, by permitting a much better simulation of certain inherent variables of phytoplankton, like low cellular densities or the presence of different species at the same time in the same bioassay. Of all the strains studied T. chuii was the most tolerant to both metals (more sensitive to copper than to lead). It also represents the largest size (cell volume) of all five strains, and in the case of copper, high KF and Kd values. On the other hand, R. salina was the most sensitive in the case of copper (along with I. galbana) and also resulted as the strain with the largest cellular copper concentrations at a 50% of growth inhibition. Results obtained in this study conclude that there appears to be no clear relationship of copper and lead microalgal tolerance with cellular size, taxonomic class, or sorption capacity. This agrees with other studies followed with copper that demonstrate that sensitivity of algal species to copper is not related to external copper binding, intra-cellular copper concentrations nor uptake rates (Levy et al., 2008). Understanding the interaction of metals with surfaces can reveal the mechanisms of metal internalization and toxicity. Therefore, future studies should research on the different copper tolerance mechanisms observed on each of the different marine microalgae.

Acknowledgments This work has been supported by the Spanish CICYT (Spanish Commission for Research and Development) in the contract of the project PETRI 95-0971. The authors thank Dr. I. Moreno-Garrido for his assistance in marine microalgae toxicity tests and Dr. J.L. Stauber for the intensive review work. References Abalde, J., Cid, A., Reiriz, S., Torres, E., Herrero, C., 1995. Response of the marine microalgae Dunaliella tertiolecta (Chlorophyceae) to copper toxicity in short time experiments. Bull. Environ. Contam. Toxicol. 54, 317–324. Adams, M.S., Stauber, J.L., 2004. Development of a whole-sediment toxicity test using a benthic marine microalga. Environ. Toxicol. Chem. 23, 1957–1968. Anderson, D.M., Morel, F.M.M., 1978. Copper sensitivity of Gonyaulax tamarensis. Limnol. Oceanogr. 23, 283–294. Baro´n, M., Arellano, J.B., Lo´pez-Gorge´, J., 1995. Copper and photosystem II: a controversial relationship. Physiol. Plant. 94, 174–180. Binet, M.T., Stauber, J.L., 2006. Rapid flow cytometric method for the assessment of toxic dinoflagellate cyst viability. Mar. Environ. Res. 62, 247–260. Blanck, H., Wallin, G., Wa¨nberg, D-A˚., 1984. Species dependent variation in algal sensitivity to chemical compounds. Ecotoxicol. Environ. Saf. 8, 339–351.

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