In Vitro Zn-Nitzschia closterium Interactions: Partition ... · closterium were evaluated on exponential growth rates, µ, and also regarding 96-h exposure time for estimating the

In: Ecotoxicology Research Developments ISBN 978-1-60692-167-8 Editor: Eduardo B. Santos © 2009 Nova Science Publishers, Inc.

Chapter 5

In Vitro Zn-Nitzschia closterium Interactions: Partition, Toxicity, Bioaccumulation and Biological

Concentration Factors

P. Pedrosa11, C. S. Karez2 and W. C. Pfeiffer3

1. Laboratório de Ciências Ambientais (LCA), Centro de Biociências e Biotecnologia (CBB), Universidade Estadual do Norte Fluminense (UENF).

2.Oficina Regional de Ciencia para América Latina y el Caribe de la Organización de las Naciones Unidas para la Educación, la Ciência y la Cultura

(UNESCO), UNESCO Montevideo 3. Laboratório de Radioisótopos Eduardo Penna Franca, Instituto de Biofísica Carlos

Chagas Filho (IBCCF), Centro de Ciências e da Saúde (CCS), Universidade Federal do Rio de Janeiro (UFRJ)

Abstract

In this study we carried out batch cultures for investigating Zn-Nitzschia closterium (Ehrenberg) W. Smith interactions: partition, toxicity, bioaccumulation and biological concentration factors (BCFs). Individuals of the marine diatom were isolated from Sepetiba Bay (Rio de Janeiro, Brazil) — a semi enclosed body-water mostly contaminated for heavy metals like Zn, Cd, Cr and Pb. In the laboratory N. closterium was grown and exposed to a series of zinc concentrations (from 7 (control) to 1,1120 µg L-1) spiked with a fixed amount of radioactive 65Zn (110 Bq mL-1). Toxicity effects of Zn-water concentrations in N. closterium were evaluated on exponential growth rates, µ, and also regarding 96-h exposure time for estimating the effective concentration to reduce, relative to the control, 50% N. closterium standing crop. No inhibition was observed on µ with Zn-water increase, but we found an estimated 96h-EC50 of 1,019 µg Zn L-1. The other interactions, partition,

1 Corresponding Author: UENF, CBB, LCA. Av. Alberto Lamego, 2000.28013-602. Campos dos Goytacazes, RJ,

Brazil.Phone: +55 22 2726 1470. Fax: +55 22 27261472. E-mail: [email protected].

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P. Pedrosa, C. S. Karez and W. C. Pfeiffer 148

bioaccumulation and BCFs, varied a great deal with momentum. In all cultures zinc was transferred from water (dissolved phase) to algal populations (particulate phase), reaching in some cases up to ~90%. This trend was partially reversed with ageing of the cultures. During the exposure time bioaccumulation ranged from ~1 to >13,000 fg Zn cell-1 and BCFs from 102 to 104. Such great variations seemed to be influenced by, or related with Zn-water concentrations, Zn toxicity and growth phase to N. closterium, and water pH. Our results were compared with reported results on zinc concentrations in natural phytoplankton from Sepetiba Bay, and suggested some compatibility between in vitro experiments and in situ data. Objectively, this crossed information quite aggregate worth and reliability to our ecotoxicological interpretations and predictions at a site specific scale.

Introduction Contaminations for heavy metals in fresh and marine waters have been reported for

decades. In such environments phytoplankters are typically known for being important primary producers. They supply part of the particulate and dissolved organic matter and oxygen, which are fundamental to the structure and metabolism of aquatic ecosystems (Roberts 2006). Metal-phytoplankton interactions are, thus, an important issue to be addressed in ecotoxicology. This issue is very complex because hydrological, physicochemical, chemical and biological factors can significantly alter the concentration, speciation and bioavailability of metals in aquatic ecosystems (Luoma 1983). The sensitivity of phytoplankton species to a metal toxicant or contaminant is thus not easy to predict. In addition, these organisms can also accumulate heavy metals, presenting high biological concentration factors (BCFs) (Lowman et al. 1971; Fisher 1986). Consequently, living and dead phytoplankton can affect the fate and distribution of metals in marine ecosystems through sorption-desorption processes coupled with physical transport — within and across water masses and currents they are suspended in — and through biological transference intermediated by specific heterotrophs.

Interactions between metals and phytoplankton (and particles in suspension as a whole, also referred to as seston or suspended particulate matter (SPM)) are, however, usually too complex for being accurately related with, and interpreted to, a single or specific environmental factor. Biological (e.g., physiological state, species composition, biomass), physical (e.g. temperature, light, particle size and specific surface area), and chemical variables (e.g. pH, Eh, dissolved and particulate organic matter, metal-metal interactions, major elements), which in great extent are not determined or controlled in situ represent factors that potentially influences metal-SPM interactions (Salomons and Förstner, 1984, Widdows and Donkin, 1991; Aidar and Sigaud, 1992, Turner and Millward 2002, Pedrosa et al. 2007). Despite the complexity of in situ metal-SPM interactions such kind of information represents a valuable reference for ecotoxicological research.

In contrast with in situ studies, physical, chemical and biological variables are usually known, controlled or semi-controlled in in vitro assays. Such systems of study — albeit in certain cases excessively artificial — also represent a useful method to studying ecotoxicology. Many studies on metal-phytoplankton toxicity are, however, restricted to physiological or lethality functions of metal(s) (or other substances) to a specific species

Zn-Nitzschia closterium Interactions 149

(Stirn et al. 1973; Nyholm and Källqvist, 1989; Matulová, 1991; McCormick and Cairns, 1994; Wong et al. 1995).

In the present study, beyond lethality of Zn to N. closterium — a ubiquitous marine diatom, which was isolated from Sepetiba Bay (Rio de Janeiro, Brazil) — we also assessed simultaneous information on partition, bioaccumulation and BCFs from batch cultures. We additionally compare and relate our in vitro results with Zn concentrations measured in natural phytoplankton from the same site (Sepetiba Bay), which is mostly contaminated for zinc.

Methods

Site Description Sepetiba Bay, a 519 Km2 semi enclosed body-water distant about 60 Km southern of Rio

de Janeiro city (RJ, Brazil), is of a great importance, serving for many economic activities, such as: commercial fishing, tourism, industrialization and agriculture. This Bay has also been used as industrial and urban waste disposal from a variety of human activities (IFIAS,1988). It is situated at parallels 22º 54' - 23º 04' S and 43º 33' - 44º 02' W (Figure 1).

Figure 1. Localization of Sepetiba Bay (Baía de Sepetiba) in Brazil (top map). The Bay as well as the sampling point are also shown in a more detailed view (below map).


Decadal anthropogenically-mediated activities like mining and industrial plants have quite contributed to the contamination of bottom sediments, suspended particulate matter, rain and biota for heavy metals (Zn, Cd, Cr and Pb) (Lacerda, 1983; Rezende, 1988; Pedlowski, 1990; Barcellos, 1991). Among the metal contaminants in Sepetiba Bay, zinc (as well as Cd) is one of the most commonly reported in the literature (Lacerda, 1983, Lacerda and Molissani, 2006). The Bay has been historically contaminated with these metals from a large Zn smelting plant. Although this industry was closed in 1996, environmental contamination for zinc still appears to occur from the remaining waste pool. Other potential sources of Zn contamination (beside other organic and inorganic substances) derive from approximately 400 industries, mostly metallurgical plants (Lacerda et al., 1987). In fact, recent estimations of annual Zn emissions (ton y-1) to soils and rivers from anthropogenic sources to Sepetiba Bay indicate persistent and even enhanced values for a ~30-years period (Lacerda and Molissani, 2006).

Laboratory Assays Nitzschia closterium was isolated in December 1990 from Sepetiba Bay through a serial

dilution method (Hudson and Hay, 1989). Artificial seawater was prepared according to the International Standard Organization - ISO/TC (1988), being enriched with f/2 medium (Guillard, 1975). Stock solutions of Zn (as ZnCl2.2H2O; Merck, P.A.) were prepared with deionized water and their concentrations checked with a flame atomic absorption spectrophotometer (model Varian 1475). Fixed amounts of Zn were then added into each batch culture in order to generate nominal concentrations that ranged from 7 (control) to 1,120 µg L-1. Zinc concentrations in water and N. closterium were radiolabeled by spiking 110 Bq 65Zn mL-1. At the beginning of the experiments borosilicate Erlenmeyer flasks (ca. 1 liter) were filled with 450 mL medium plus 45 mL inoculums, yielding 495 mL. The cultures were grown at 20 ± 2 ºC and under 40 Watts power fluorescent lighting provided at an intensity of 33 µmol PAR photons m-2 s-1 on 12:12 h light/dark cycle. Over the experiments the flasks were shaken once a day to avoid cell deposition and promoting gas-exchange.

According to the total zinc concentrations, 7 (control), 70, 140, 280, 800 and 1,120 µg L-

1, cultures were identified as I (control), II, III, IV, V and VI, respectively. Three similar (not identical) experiments were carried out simultaneously, which are referred to as A, B and C (see Table 1) (A-B exponetial and C-senescent innoculum). The pH was measured potentiometrically, while cell concentrations were determined through a Fuchs-Rosenthal chamber using a light microscope. The specific growth rates (µ) of N. closterium were determined on exponential phase. Toxic effects of zinc were investigated on µ and also on cell densities over 96-h exposition only to experiment B, which could be referenced against the pertinent literature. For cell densities, the effective concentration capable to produce, in relation to the control, a 50% reduction in N. closterium standing crop (EC50) was estimated by interpolating the measured data.

Depending on cell densities, 2-10 mL culture were filtered onto 0.45 µm Millipore membrane filter (HAWP 0025) in a low pressure (<0.25 atm) followed by an additional


filtration with radionuclide-free synthetic seawater (ISO/TC, 1988). This procedure aimed to remove eventual 65Zn adsorption on both cells and filtering membrane.

Table 1. Characteristics of the experiments A, B, and C.

Experiment A B C Inoculum growth phase Exponential Exponential Senescent Length (days) 7 31 35 Initial Zn concentrations (μg L-1) 7 (I) 7 (I) 7 (I) 70 (II) 280 (IV) 280 (IV)

140 (III) 800 (V) 800 (V)

280 (IV) 1,120 (VI) 1,120 (VI) Filter contamination for gamma (γ)-radiation from Zn-water was also checked by using

two filters positioned one below (control-filter) another as described by Davies (1973). 65Zn γ-radiation was measured with an automated gamma counter (Compugamma LKB-1282, Wallac gama spectrometer) and the results were recorded as counts per minute (cpm). The resulting counts in the filter samples, hereafter P (particulate phase), were corrected for the average background value (control-filter) and normalized to mL. The distribution of zinc between water (dissolved phase) and P was monitored over periods of time that ranged from hours to a few days. At the end of every experiment each flask was rinsed with HCl 0.1 N to check loss of 65Zn onto the wall flasks interior. Average results indicated that 65Zn adsorption was in general negligible; less than 3% of the total activity added.

To determine average cellular Zn concentrations in N. closterium (Zn cell quota), γ-counting from P were divided by the number of cells mL-1 as well as by the γ-counting from water plus filter (S), and then multiplied by the total concentration of stable zinc [Zn] in the culture (Eq. 1). The results were expressed as fg Zn cell-1 (femtograma = 10-15 g):

[Zn] cell-1 = [(cpm ml-1 P)/(number of cells mL-1)] . [Zn] stable mL-1of culture [Equation 1]

(cpm.mL-1 S)

where cpm mL-1 is gamma counts per minute normalized to milliliter, P is the particulate (algae) retained on filter and S is the sum (water + P).

Average N. closterium dry weight cell was estimated gravimetrically in an analytical balance with ±0.1 mg precision by weighing eight dried-oven filter samples containing cells at exponential phase. The resulting value, 1.0 ± 0.11 ng, was used to calculate their BCFs for zinc as follows. 65Zn γ-radiation was normalized to 1 mL and cpm adjusted per cell (65Zn ng-

1). Converted zinc concentrations in a cell were multiplied by 109 in order to obtain [Zn] g-1

(dry weight cell) and then divided by either the Zn content present in 1 mL water at the momentum of filtration (Eq. 2) or divided by the nominal concentration of Zn (mL-1) in each assay medium (Eq. 3). These approaches allowed us to compare two alternative methods for estimating BCFs, which are more probable related to in situ (Eq. 2) and in vitro (Eq. 3) measurements.


BCF = [Zn] g-1 dry weight cell [Equation 2] [Zn] mL-1 water(momentum of sampling)

BCF = [Zn] g-1dry weight cell [Equation 3] [Zn] mL-1 (nominal concentration in culture) Note: for such Equations (2 and 3), resulting values are dimensionless, assuming that 1

mL water = 1 g water. As a whole, cell dimensions were similar between Zn-water concentrations (or cultures).

At the end of experiments (B and C), however, the aged cultures tended to present somewhat smaller and thinner cells. On this respect, since the average dry weight cell of N. closterium was obtained from cultures on exponential phase of growth, BCFs associated to the end of the experiments (B and C) or presenting a senescent growth phase were likely underestimated.

Results and Discussion

Overall Data The data on cell densities, Zn concentrations in water and zinc-algae (P), and water pH

are plotted against day of experiment (Figure 2), and their descriptive statistics are summarized in Table 2. As a whole, cell densities varied from 2 x 104 to 2 x106 cells mL-1, zinc-water concentrations from 0.3 to ~1,000 µg L-1, zinc-algae (P) concentrations from 0.1 to ~900 µg L-1, and pH from 7.8 to 9.5. These values integrate and reflect variations over and between the experiments.

Partition The biphasic, dissolved-particulate, distribution of Zn showed a clear decrease in water with

increasing cell density, achieving threshold values in the mid-experiment (B and C), ~15 days. After that, a shifting in Zn distribution was evidenced, especially in cultures I and IV. This behavior seemed to be related with nutrient impoverishment as the cultures rapidly entered in a senescent growth phase. In cultures V and VI this trend was somewhat delayed or less apparent in both experiments B and C, which may be related with lower average cell densities with increasing Zn concentrations in mediums. This was particularly evident when cultures I and IV are compared with cultures V and VI (Table 2).


Table 2. Descriptive statistics to pH, cell densities, Zn-water concentrations, Zn-filter (algae, P) concentrations, and %Zn-water concentrations for experiments A, B, and C.

Table 2 (Continued)

Descriptive statistics

Experiment pH Cells mL-1 μg Zn L-1 (water)

μg Zn L-1 (P)

% Zn-water

median 8.6 935313 5 2 65.2 mean 8.7 924664 4 3 56.1 sd B-I 0.6 678974 3 3 37.1 minimum 7.8 67031 0.3 0.1 4.8 maximum 9.5 2120000 6.9 6.7 98.8 median

8.6 951250 156 124 55.7

mean 8.7 933821 152 128 54.3 sd B-IV 0.6 714540 99 99 35.5 minimum 7.8 54537 11 18 4.1 maximum 9.5 2350000 262 269 93.4 median

8.8 767813 410 391 51.3

mean 8.7 804038 396 415 49.5 sd B-V 0.6 621867 157 166 19.6 minimum 7.8 55938 131 150 16.3 maximum 9.5 1880000 650 670 81.3 median 8.7 512813 526 594 47.0


Table 2 (Continued)

mean 8.6 706680 554 566 49.5 sd B-VI 0.5 660904 188 188 16.8 minimum 7.8 32500 186 239 16.6 maximum 9.4 1865000 881 934 78.7 median

8.8 396250 5.5 1.5 77.9

mean 8.6 601512 4.5 2.5 64.3 sd C-I 0.6 626268 2.1 2.1 29.9 minimum 7.8 25000 0.8 0.5 11.0 maximum 9.2 2000000 6.5 6.2 93.5 median 8.7 387500 217 63 77.5 mean 8.6 497183 186 94 66.6 sd C-IV 0.6 470407 67 67 24.1 minimum 7.8 33906 71 25 25.4 maximum 9.5 1487500 255 209 91.1 median

8.5 233750 452 348 56.4

mean 8.5 361879 468 332 58.5 sd C-V 0.4 348344 105 105 13.2 minimum 7.8 32632 347 86 43.4 maximum 9.1 1150000 714 453 89.2

median 8.3 140000 647 473 57.8 mean 8.3 240045 644 476 57.5 sd C-VI 0.3 240831 155 155 13.9 minimum 7.8 32656 391 143 34.9 maximum 8.8 717500 977 729 87.2

Analyzing Zn partition on the first three initial samples, when algae biomasses were similar

between cultures, we observe a trend of increasing Zn-P (cpm) with increasing Zn-water concentration (Figure 2). For example, instantaneous (time zero) cpm measurements showed values that jumped from ~3 to ~35 and from ~18 to ~33 in experiments B and C, respectively. Studying zinc uptake in the marine diatom Phaeodactylum tricornutum, Davies (1973, 1978) reported a rate proportional to the concentration gradient between extracellular zinc and zinc inside the membrane. Comparatively, also at the first day of experiment (samples collected at 0, 0.2 and 1 day of experiment), Zn migration from dissolved to particulate phase was also higher in experiment C than in experiment A and B (Figure 2). This result indicates that inoculums on a senescent phase of growth can be more reactive to sorption Zn from water than inoculums in an exponential growth phase. But, it was in the experiment B that Zn-water uptake by N. closterium populations achieved the higher, yet temporarily, values: >90% in cultures I/IV and ∼80% in cultures V/VI (see Table 2).


Figure 2. Y1-axis: 65Zn Gamma radiation counts per minute (cpm) in water and filter (algae, P) normalized to mililiter, and cell densities also normalized to mL. Y2-axis: pH. All data ploted on day of experiment and referred to Nitszchia closterium batch cultures for experiments A, B, and C with 7, 70, 140, 280, 560, 800, and 1,120 µg Zn L-1 (medium), respectively, I, II, III, IV, V, VI, and VII.

As a whole, these results suggest that Zn can be dynamically mobilized from dissolved phase to particulate phase and vice versa, which is relevant to monitoring metal contamination in natural waters based on phytoplankton assessment. This observation can be particularly important or critical in diagnosing the environmental status of natural waters and aquatic ecosystems when a discrete or pulse contamination occurs.

Zn Toxicity (Lethal Effect): 96h-50% Effective Concentrations (96h-EC50)

Although we have counted cell densities in all the three experiments, only experiments A and B were considered suitable for comparisons with the specific literature we reference herein, which also worked with N. closterium inoculums in exponential phase and in batch cultures. In the experiment C, Zn toxicity on senescent inoculums seemed to reduce the number of viable cells, delaying the beginning of the exponential and senescent growth phases in cultures with higher zinc concentrations (in relation to the control). As a result,


internal comparisons referenced against the control would be partially out of phase and, therefore, such kind of comparison was avoided as well. From these considerations, we shall focus Zn toxicity to Nitzschia closterium on experiment B; otherwise the experiment should be properly referred.

Regarding that Zn concentrations typically range from ~5 to10 µg L-1 in marine waters (Goldberg et al. 1971), then it is expected that no toxicity effect should occur to N. closterium within this range of concentration. We noted however that even in assay mediums containing 280, 800 and 1,120 µg Zn L-1 the exponential growth rates (µ) did not change consistently in relation to the control. That is, it was not possible evidencing a clear pattern on µ of increasing or decreasing with increasing Zn-water concentrations. These results as well as other growth related population parameters like doubling time (td) and number of divisions (k2), all in a day unit basis, are shown for the three experiments (Table 3). Taking into account these parameters as a whole, we conclude that individual cells of N. closterium, which escaped lethal effect of Zn-water concentrations tested here, were capable to maintain its physiological status similar to the control.

In contrast with this study, Stauber and Florence (1990) and Zhang and Florence (1987) found, respectively, IC50 (50% inhibition concentrations) of 65 µg Zn L-1 and 60 µg Zn L-1 to produce a 50% reduction on µ of N. closterium. But Stauber and Florence (1989) also reported an IC50 of 850 µg Zn L-1 on µ to the same species when assayed in nutrient medium f, containing chelators.

Table 3. Some growth population related parameters: growth rate (µ), duplication time (td), and generation time (k2), also showing the coefficients of linear regressions, which

referred growth rate calculations of N. closterium

These results show how dramatically Zn IC50 on µ of N. closterium can vary. Part of our

contrasting results might in addition be related with intrinsic differences associated to N. closterium ‘strains’, which could emerge from different pressures of Zn contamination.

Although no consistent effect was observed on µ for the tested Zn concentrations, our experiments showed clear depressions on N. closterium cell densities with 800 and 1,120 µg Zn L-1. For instance, over 96-h exposure cell densities in cultures V and VI were,


respectively, ~27 and 62% lower than observed in the control (experiment B). Absolute values ranged from 114,063 to 301,250 cells mL-1 against 562,500 cells mL-1 in the control. The effective Zn-water concentration capable of produce a 50% reduction on N. closterium cell densities (EC50) over 96-h exposition was 1,019 µg L-1 (Figure 3). This concentration is not commonly reported in natural waters and, therefore, should likely be related to extreme, severe or acute environmental contamination. Comparatively, Rosko and Rachlin (1975) found a 96h-EC50 of 271 µg Zn L-1 in N. closterium isolated from New York Bight. When chelating agents like citric acid and NCTC-135 were added to culture mediums these authors found a 96h-EC50 of 360 µg Zn L-1. Rachlin et al. (1983), working with the same species, but isolated from the culture centre of algae and protozoa in Cambridge, found a 96h-EC50 of 176.5 µg Zn L-1 (without addition of chelant agent). Analyzing these results Rachlin et al. pointed out that more resistant specimens of N. closterium to Zn toxicity effects were likely selected in the New York Bight waters in which large amounts of pollutants, anthropogenically-mediated, should occur. Therefore, in comparison to the findings reported by Rosko and Rachlin and Rachlin et al., our results suggest that a relatively high zinc contamination in Sepetiba Bay could result in more resistant specimens. Consistently, as early mentioned, Zn contamination has been reported in Sepetiba Bay for decades.

Beside Zn-water concentrations, it is also important to assess bioaccumulation of the contaminant in the test-organism. In microalgae in general, metal toxicity has been described to be primarily associated with binding to sulphydryl groups in proteins or disrupting protein structure or displacing essential elements (Viarengo and Nicotera 1991, Arunakumara and Zhang 2008). On the other hand, it has been shown that many marine microalgae (including bacillariophyceae) are able to produce phytochelatin (Ahner et al. 1995). These biomolecules chelate metals through coordination with the sulfhydryl group in cysteine and are assumed to acting either as a metal detoxifying system or as a buffer (Ahner et al. 1995). Ahner and Morel (1995) demonstrated that among eight metals Cd, beyond Cu and Zn were the most effective inducers of phytochelatin in the marine diatom Thalassiosira weissflogii. However, whether these processes of detoxification operates in N. closterium and how they could influence Zn bioaccumulation is uncertain.

In experiment B, Zn bioaccumulation ranged from 1.2 to 10,419 fg cell-1. In experiment A, concentrations were within the ranges verified to experiment B in cultures I-IV. In experiment C, maximal bioaccumulations were even greater than those observed in experiment B, with values ranging from 22.2 (I) to 13,408 (VI) fg Zn cell-1 (Table 4). These enhanced Zn bioaccumulations in N. closterium inoculums on senescent growth phase suggest the exposition of biogenic sites, which are likely reactive to Zn adsorption. Losses in cell integrity and/or viability — which might be caused by senescence and Zn toxicity — could support enhanced Zn adsorption on the particulate phase.

Although speculative, the importance of adsorption process to total Zn bioaccumulation to N. closterium is found in the literature. For example, using an EDTA treatment followed by seawater washes for removing external Zn on the same species, Stauber and Florence (1990) estimated that only 3-4% of the total Zn associated to N. closterium was internal. This result, however, does not represent a perfect reference for extrapolating such data to our study because, differently of us, in their study N. closterium was grown in a seawater medium


without chelant agents, which may affect estimations of external/internal Zn bioaccumulation to N. closterium.

Figure 3. Relationship between Zn-water concentrations and relative reduction of cell densities, referenced to culture I (control) in experiment B. The polynomial equation inside the plot represent the best fitness for the measured data.

Table 4. Descriptive statistics to bioaccumulation in N. closterium (fg Zn cell-1) (=µg Zn g-1 dry weight cell) for experiments A, B, and C


Normalized to N. closterium cell dry weight, Zn bioaccumulation varied from 1.9 to 3,731 µg g-1 — which equals to fg Zn cell-1 in this study — over 96-h exposition in the experiment B (Figure 4).

Figure 4. Relationship between Zn bioaccumulation (µg Zn g-1 dry weight cell) and relative reduction of cell densities of N. closterium, referenced to culture I (control) in experiment B. The polynomial equation inside the plot represent the best fitness for the measured data.

The estimated 96h-EC50 on cell densities normalized to Zn cell quota was ~2,848 fg, which, according with our study, should equal to a pulse concentration between 800 and 1,120 µg Zn L-1 water. It is interesting to note that maximum values of bioaccumulation in N. closterium from cultures V and VI were even higher than that estimated to EC50 (for Zn cell quota). But it is also important to note that practically all such (likely toxic) concentrations were found before 96-h exposition. For example, in the experiment B, 4,004 and 10,419 fg Zn cell-1 were found over 48-h exposition. After that, with the beginning of the exponential phase, Zn concentrations tended to decrease in N. closterium cells, and — as early mentioned — growth rates as well other related population parameters did not change consistently with increasing Zn concentrations tested here. In this sense, Zn toxicity was rather likely related to Zn bioaccumulation in the beginning of the experiments. From an ecotoxicological point of view, therefore, the actual concentration of Zn to produce 50% reduction on N. closterium cell densities could exceed ∼2,848 fg cell-1 (for this study).

Zn Bioaccumulation in N. closterium (Beyond Zn Toxicity)

After the initial ~2-5 days of the experiments, beyond Zn toxicity — as well as Zn-water concentrations, and growth phase of inoculums — other factors like exponential growth phase and water pH also seemed to influence bioaccumulation in N. closterium. The


variations in Zn bioaccumulation to N. closterium are illustrated only to experiment B (Figure. 5). One observation refers to the fact that the quantity of Zn per cell decreased with beginning exponential phase in all cultures, despite increased Zn partition onto algal populations (P) (Figure 2). This behavior is consistent with findings of Knauer and Martin (1973) who related lower heavy metal concentrations with higher primary productivity in natural phytoplankton from Monterey Bay, California. Then as the slope of the growth curve became virtually lower (cell densities tended to grow linearly), Zn cell quota back to increase in cultures I and IV and, in cultures V and VI, a slightly but consistent decrease was found. In general, a similar but less prominent pattern was also observed to experiment C (data not show). Focusing on experiment B, after ~14 days there was a conspicuous decreasing in Zn bioaccumulation despite cell densities were in a quite steady state (stationary phase) in culture I. We speculate that this observation could be a function of nutrient impoverishment in the assay medium, leading to a decreasing cell quota protein with ageing of the culture. As known, proteins are one of the most important macromolecules able to bind metals, and their biosynthesis, comparatively to exponential phase, tend to reduce on stationary and senescent phases.

Water pH is another potential factor that could be related with bioaccumulation. Zinc speciation is greatly affected by pH, and the most bioactive form, Zn+2, tend to be converted to species like ZnOH2 and ZnCO3

= when pH>8.5 (Bernhard and Zattera, 1975). As a result, it is probable that Zn toxicity had diminished with increasing pH. Indeed inverse relationships between metal toxicity and pH have been shown (Rai et al., 1993). However, the contrary has also been reported in many studies with microalgae-metal interactions (Franklin et al., 2000; De Schamphelaere et al., 2003, Wilde et al. 2006). In such cases, it is suggested that pH increasing would decrease competition between the metal ion and H+ at the cell surface.

In our experiments pH increased with increasing cell densities, varying from 7.6 to ∼9.5. This variation may also influence adsorption (Förstner and Wittman, 1981; Salomons and Förstner, 1984), and was partially suggested here.

Although bioaccumulation and water pH correlated positive and significantly (r = 0.65, p<0.05) in culture I, an opposite trend was observed to the remaining cultures, ranging from -0.38 (culture IV) to -0.78 (culture V) and -0.73 (culture VI).

These results suggest that viable cells of N. closterium can excrete Zn by eliminating excessive Zn cell quota, which might explain the observed dichotomy between cultures I and IV-V-VI. Some mechanisms of detoxification like production of binding factors and proteins, exclusion of metals from cells by ion-selective transporters and excretion or compartmentalization have been suggested with regard to reducing heavy metal toxicity (Arunakumara and Zhang, 2008). Among these mechanisms of detoxification, Zn cell quota could be diminished by exclusion and excretion processes. It is valid to note however that even assuming that N. closterium promoted some Zn detoxification, water pH still should influence bioaccumulation in the species because physicochemical speciation and sorption processes must occur independently of biological processes of detoxification.


Figure 5. Zinc bioaccumulation and cell densities (growth curve) of N. closterium. Note that Zn bioaccumulation fall with beginning exponential phase (2-6 days approximately) in all cultures, back to increase in cultures I and IV but not in cultures V and VI. At the end, however, all cultures tend to show lower values of Zn bioaccumulation. (See comments on such apparent dichotomy between cultures I-IV and V-VI in the main text).

Zn Biological Concentration Factors (BCFs) Biological concentration factors (BCFs) allow estimating or comparing the capacity or

ability of one organism concentrate a substance from its external medium (e.g. water). Conceptually, the establishment of BCFs for a species should be rather obtained at an equilibrium state. From a practical purpose, however, this condition is not easily satisfied, especially in situ conditions where metal concentrations are likely not in steady state. As a result there have been researchers who consider BCFs irrespective to meeting an equilibrium status (Canterford et al., 1978; Radwan et al., 1990).

It is important to note that BCFs like estimated in Eq. 2 can be apparent because final results can fluctuate irrespective to bioaccumulation in cells. In coastal waters where the quantity of biogenic and abiogenic particles is typically great and their nature is quite different, competitive sorption for metals likely occurs. This would lead to a misinterpretation (overestimation) of BCFs to a specific species. Consistently, the values obtained from Eq. 2 were in general higher than those from Eq. 3 and, particularly in culture I, more variable as well. Eq. 2 is however instrumental to demonstrate the fragility in accurately estimating BCFs when based on an in situ sampling. Hence, Eq. 3 seems to be more suitable for


assessing the capacity of a species to concentrate a substance. In this study, equations 2 and 3 (see Material and methods) aid us to illustrate how BCFs can vary as a function of the method of calculation (Figure 6). The relationship between equations 2 and 3 is additionally illustrated in Figure 7 but only to contrasting Zn-water concentrations (cultures I and VI, experiments B and C). Note that BCFs calculated through Eq. 2 presented higher values than those from Eq. 3, especially in the control cultures.

This pattern was somewhat confirmed to cultures IV (similar to the control cultures) and V (similar to culture VI) (data not shown). Hence, comparisons between BCFs from literature require qualification and care.

In spite of such differences in estimating BCFs, a common trend was found here. For example, fixing the second day of experiment when cultures V and VI showed the highest BCFs (for both equations), BCFs increased with increasing Zn-water concentrations (Figure 6). This kind of relationship has also been reported for several freshwater algae (Coleman et al., 1971 apud Canterford et al., 1978). But the contrary has been documented as well; Canterford et al. (1978) studying Ditylum brightwellii observed a decrease of BCFs as Zn concentrations in water were enhanced.

With the continuity of the experiments, all cultures tended to present BCFs near 7 x 102 and, at the end of the experiment, 1-3 x 102. The patterns found in experiment B were in general also found in experiment C, but — similarly to found to bioaccumulation — BCFs were quite greater in this experiment. Sakagushi et al. (1979) also related higher BCFs in dead cells than in living ones, but Fisher et al. (1983, 1984) apud Fisher (1986) pointed out that, in general, both living and dead phytoplankton have similar ability to concentrate metals from water. As a whole, however, both experiments (B and C) ranged from 102 to 104, independently of what sort of estimation is applied.

Coupling Zn in Vitro Results with Natural Phytoplankton in Sepetiba Bay

Whether collectively BCFs (Zn) of phytoplankton communities in marine waters are or not comparable to our findings is uncertain. But assuming that similar BCFs for an especific metal are applicable to different algae species (Fisher, 1986) and considering 7 x 102 a conservative value for N. closterium (isolated from Sepetiba Bay), we can roughly relate in vitro to in situ conditions. Zinc concentrations in phytoplankton samples collected in Sepetiba Bay (at Coroa Grande inlet) in 1989 ranged from 2.9 x 102 and 8.2 x 10 2 µg g-1 dry wt (Pedrosa et al. 1995). These concentrations are similar to that encountered in cultures IV and V (perhaps also in VI), which suggests maximum or pulse concentrations between 280 and 800 µg Zn L-1 in that site. Basing on this crossed information a critical contamination of zinc should occur in that Bay. As early mentioned, this notion has been consistently converged through independent methods and by different authors (Lacerda 1983, Rezende 1988, Pedlowski 1990, Barcellos 1991, Lacerda and Molissani 2006)

Suspended particulate matter (Lacerda, 1983) and more specifically the phytoplankton fraction (Pedrosa et al. 1995) have been suggested as one of the most important components that potential affect the fate of heavy metals in Sepetiba Bay. In this context, through transport and trophic pathways phytoplankton can exert an important role in redistributing Zn in the contaminated site. A simple extrapolation of our results to Sepetiba Bay suggests the system would present or experience relatively high Zn concentrations (280-800 µg Zn L-1),


Figure 6. Biological concentration factors (BCFs) provided by Equations 2 and 3. Note that BCFs provided by Equation 2 tend to be more variable and higher than those obtained from Equation 3 (see Methods for details on Equations 2 and 3).


Figure 7. Relationships between BCFs provided by Equations 2 and 3. Note that coefficients of correlation are more strong at high Zn-water concentrations (cultures b-VI and c-VI, right panel) than in the control cultures (b-I and c-I, left panel), and that Equation 2 tend to overestimate BCFs in relation to BCFs provided by Equation 3. Dashed lines represent a 1:1 proportion between the equations.

which lead us to think — also based on our results on Zn-N.closterium interactions — that the trophic structure and productivity of the ecosystem might be negatively affected for decades. In conclusion, although limitations and care are needed to compare in vitro with to in situ environmental conditions, we believe this methodology constitutes a desirable and valuable effort in studying ecotoxicology.

Acknowledgements The authors would like to thank Dr. Sandra M. F. de Oliveira e Azevedo who allowed us

to use her laboratory facilities and space for cultivating N. closterium.


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