Archives of Environmental Contamination and Toxicology

, Volume 62, Issue 1, pp 13–21

A Multidisciplinary Approach for Assessing the Toxicity of Marine Sediments: Analysis of Metal Content and Elutriate Bioassays Using Metal Bioavailability and Genotoxicity Biomarkers

Authors

  • Stefania Frassinetti
    • Institute of Agricultural Biology and BiotechnologyCNR
  • Emanuela Pitzalis
    • Institute of Chemistry of Organometallic CompoundsCNR
  • Marco Carlo Mascherpa
    • Institute of Chemistry of Organometallic CompoundsCNR
  • Leonardo Caltavuturo
    • Institute of Agricultural Biology and BiotechnologyCNR
    • Institute of BiophysicsCNR
Article

DOI: 10.1007/s00244-011-9667-x

Cite this article as:
Frassinetti, S., Pitzalis, E., Mascherpa, M.C. et al. Arch Environ Contam Toxicol (2012) 62: 13. doi:10.1007/s00244-011-9667-x

Abstract

The goal of this article is to verify the applicability of two different biological assays for studying a coastal area that is subject to anthropogenic inputs. Phytochelatins in the marine diatom Thalassiosira weissflogii were used as a biomarker of metal bioavailability. The frequency of genetic damage in the sensitive D7 strain of the yeast Saccharomyces cerevisiae was used to estimate the mutagenic potential. Biological assays were carried out using sediment elutriates. Sediments were collected at three selected sites located in the Gulf of Follonica (Tuscany, Italy), during a 2-year sampling period: Cala Violina (reference site) and the mouths of the rivers Pecora and Cornia, named sites V, P and C, respectively. The chemical characterization of each site was determined in terms of metal concentrations (As, Cd, Cr, Cu, Ni, Pb), measured in 11 sediment samples for each site. The results showed that metal concentrations in sediments from sites C and P were 2–10 times higher than the reference values (site V, year 2004). In addition, we found generally higher metal concentrations in the 2007 sediments than in the 2008 ones, including those of site V, due to the occurrence of an unexpected pollution event. This enabled us to obtain a pollution gradient to validate the proposed bioassays. In fact, the bioassays showed a potential biological hazard in the 2007 elutriates. Significant mutagenic effects were found in samples exhibiting higher concentrations of Cd and Cr. The induction of phytochelatins in T.weissflogii correlated positively with the Cd concentration in the elutriates.

Heavy metals in marine sediments have a natural and anthropogenic origin. Their distribution and accumulation are influenced by several factors, such as their redox state, physical transport, and adsorption processes (Buccolieri et al. 2006). The risk of metal pollution is related to the remobilization of the metals from sediment to a water column and, consequently, through the food chain (Pacifico et al. 2007). It is now accepted that marine pollution events should not be documented only in terms of the chemical concentration of pollutants (Depledge and Hopkin 1995; US EPA 2001). To assess environmental risks, chemical and biological methodologies need to be integrated together.

Integrated ecotoxicological approaches have been applied to describe the quality of marine sediment (Beiras et al. 2003; Cesar et al. 2007; Lee et al. 2006; Macken et al. 2009; Usero et al. 2008). Biological assays, involving a number of species at different trophic levels, have been widely used to assess the potential toxicity of different types of pollutants in marine sediments (Nendza 2002). In addition, toxicity tests performed on sediment elutriates can assess the hazards of sediment-bound pollutants, which can be released in the water column when sediments are resuspended (US EPA 1991).

Taking into account that coastal sediments can release hazardous metal concentrations into seawater, it is important to choose appropriate biomarkers to assess the metal bioavailability and toxicity (Martin-Diaz et al. 2004). For this purpose, biomarkers of effect on organisms, enabling molecular and physiological alterations to be evaluated, might be considered a suitable tool. Several studies have focused on physiological detoxification, such as the synthesis of metallothioneins in many animal species (Amiard et al. 2006; Geffard et al. 2007; Martin-Diaz et al. 2007) or of phytochelatins (PCs) in plants and algae (Ahner et al. 1997; Kawakami et al. 2006; Le Faucheur et al. 2005; Morelli and Fantozzi 2008; Pawlik-Skowronska 2000) as biomarkers of metal exposure. In a recent study (Morelli et al. 2009), we developed a new bioassay based on the presence of PCs in marine microalgae as a response to metal bioavailability in the elutriates of marine sediments. PCs are intracellular metal-binding peptides (Grill et al. 1985), enzymatically synthesized in response to metal ions, thereby constituting a specific cellular signal of metal exposure.

Because environmental samples consist of a complex mixture of contaminants, in addition to using metal-specific bioassays, alternative biological approaches can help in assessing ecotoxicological risks. Of these, an evaluation of the mutagenic risk is significant. Eukaryotic microorganisms, such as protozoa and fungi, are located in one of the first trophic levels in sediments and can be used as an indicator for toxicity in aquatic systems (Weber et al. 2006). Yeast species are excellent models for basic biological research and are an important tool in understanding the environmental impact on cellular systems (Forsburg 2005). Yeasts have been widely used as model systems in the environmental screening of genotoxicity (Bronzetti et al. 1997; OECD 1986a, b). The yeast Saccharomyces cerevisiae has been used for the genotoxicity evaluation of surface water, drinking water (Guzzella et al. 2004; Pellacani et al. 2006), as well as of freshwater sediments (Frassinetti et al. 2006; Weber et al. 2006).

The goal of this work is to verify the applicability of biological assays carried out with two different species in order to study a coastal area subjected to anthropogenic inputs. The sediments were collected from the coast of the Gulf of Follonica, southern Tuscany, Italy, during a 2-year sampling period (2007 and 2008). This area, although not highly polluted, has received inputs from mining, industrial, tourist, and agricultural activities. Metal distribution in sediments and soils from this part of the Tyrrhenian coast and from the surrounding area has been surveyed by several studies (Baroni et al. 2004; Focardi and Tiezzi 2009; Frassinetti et al. 2006; Leoni and Sartori 1997): Some positive anomalies, although of low absolute intensities, have been observed for As, Pb, and Zn.

By using the pollution gradient occurring in the 2 years of the sampling period, we measured (1) the contamination levels of As, Cd, Cr, Cu, Ni, and Pb in marine sediments sampled from selected sites, (2) the metal bioavailability by determining the concentration of PCs in the marine diatom Thalassiosira weissflogii grown in the elutriates of these sediments, and (3) the genotoxic potential of the sediment elutriates using short-term in vitro bioassays carried out with the yeast S. cerevisiae.

Materials and Methods

Chemical Analysis of Sediments

The sediments used in this study were collected from three coastal sites located in the Gulf of Follonica (Tuscany, Italy), as shown in Fig. 1a and b). For each site, the samples were taken from 11 different points, at 50 and 100 m from the central point, along each of five transects. The sampling strategy is illustrated in Fig. 1c.
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Fig. 1

Map of the sampling sites (a) and its location in Italy (b). Strategy of sampling adopted in each site (c). Sites C, P, and V correspond to the Cornia River mouth, Pecora River mouth and Cala Violina, respectively

The sediments were sampled using a hand-held shovel from the seabed and transferred into prewashed polyethylene flasks. After collection, the sediments were frozen at −20°C and stored until analysis. A subsample of sediments (~5 g) was weighed after slow defrosting, dried at 120°C until weight stabilization, and weighed dry.

The sediments were decomposed under pressure in a Milestone Start D microwave system equipped with MPR 600/12S rotor (12 Teflon vessels; vessel volume = 100 ml; operating pressure = 35 bar; operating temperature = 260°C) and temperature sensor ATC 400CE. One reference material portion and one blank were added to each set of 10 subsamples; they were then processed like the other samples and analyzed randomly. The digestion procedure was as follows: Each portion of wet sample (~1 g) was weighed in the Teflon vessel and 3 ml of 69% HNO3 (Aristar, BDH) and 2 ml of 30% H2O2 (Suprapur, Merck) were added; after 60 min, 3 ml HNO3 and 1 ml H2O2 were added and the mixture left to react overnight at room temperature. The vessels were then closed and put in the microwave oven, heated to 200°C for 10 min and held at 200°C for 20 min. The vessels were allowed to cool slowly to room temperature; then the contents were put in preconditioned 25-ml polypropylene flasks and diluted to the volume with ultrapure water. The final HNO3 concentration in the samples was 15%. If not analyzed immediately, the samples were stored at −20°C.

The As, Cd, Cr, Cu, Ni, and Pb contents were determined by electrothermal atomic absorption spectrometry (AAS). The measurements were carried out with a Perkin–Elmer model 4100 ZL atomic absorption spectrometer equipped with a transverse heated graphite atomizer (THGA), a longitudinal Zeeman effect background corrector, and an AS-71 autosampler. THGA graphite tubes with integrated platforms (part No. B300-0653) were either used as purchased or pretreated with a carbide forming element (referred to as an Ir–Zr-treated platform), as described elsewhere (Tsalev et al. 1995; De Giglio et al. 2000). Instrumental parameters for AAS measurements and the temperature program for each element analyzed were set following the manual. The measurements were recorded in peak area mode (integrated absorbance, Aint). Perkin–Elmer electrodeless discharge lamps, system 2, were used for As and Cd; Perkin–Elmer hollow-cathode lamps were used for the other elements.

Stock standard solutions for metal measurements were prepared by dilution from 1 mg/ml commercial standard solutions for AAS (Fluka). Ultrapure grade water was produced with USF ELGA Purelab Classic. Certified reference materials MESS-1 and BCSS-1 from the Institute of National Measurement Standards of the National Research Council of Canada were used as references.

Elutriates

Elutriates were prepared according to US EPA methods (2001). The sediment was mixed with natural seawater in a 1:4 (w/v) ratio (dry weight) and shaken at 500 rpm for 1 h, using a vertical stirrer (Velp Scientific). After settling (30 min), the aqueous fraction was centrifuged at 7000 rpm for 15 min and filtered (0.45-μm membrane filters). The elutriates were used for bioassays within a few hours or, alternatively, stored at −20°C. The procedure of freezing samples was adopted to allow the simultaneous testing of samples coming from different sites, according to Volpi Ghirardini et al. (2005). Natural seawater was collected in an uncontaminated area, 3 miles offshore from the Island of Capraia (Tyrrhenian Sea, Italy), filtered through 0.45-μm membrane filters and stored in the dark at 4°C.

The total dissolved metal concentration in the elutriates was determined by anodic stripping voltammetry (ASV) after dilution with natural seawater and acidification with HCl Suprapur (Carlo Erba) at pH 2. Voltammetric measurements were carried out by a Metrohm model 646 VA processor in conjunction with a 647 VA stand, equipped with a rotating Hg-plated glassy carbon disk as a working electrode (MFE). The preplated MFE was obtained by plating at a potential of −1 V for 600 s from a solution 5 × 10−5 M of Hg(NO3)2. Voltammetric conditions for the analytical measurements were as follows: rotating speed = 1500 rpm, deposition potential (Edep) = −1 V, deposition time = 300 s. During the rest time (15 s) before the anodic scan, the electrode potential was set at −1.4 V in order to minimize the interferences produced by the adsorption of organic matter on the electrode surface (Scarano and Bramanti 1993).

PC-Based Bioassay

The bioassay was carried out using the marine diatom T. weissflogii (strain CCAP 1085/1) obtained from the Culture Collection of Algae and Protozoa, Dunstaffnage Marine Laboratory, United Kingdom. Stock cultures were grown in axenic conditions, in natural seawater enriched with the f/2 medium (Guillard 1975), at 21°C and fluorescent daylight (100-μmol photons/m2/s1) in a 16:8 light–dark cycle.

Incubation experiments were carried out by inoculating algae, from a stock culture, in the elutriates at an initial cell density of 1 × 102 cells/ml. Before inoculum, all media (150 ml) were enriched with the f/2 medium lacking the trace metal stock solution and sterilized by filtration (0.2-μm membrane filters). The cultures were allowed to grow for 6 days during the exponential phase, until a final cell density of (3–6) × 104 cells/ml. Cell counts were carried out using a Neubauer counting chamber under a microscope. A control culture was performed using natural seawater that had undergone the same procedure of the elutriate (without sediment). Experiments were carried out in triplicate and the results were analyzed using the Student’s t-test.

After incubation, the algae were collected by filtration onto 1.2-μm membrane filters, resuspended in 1.5 ml of 0.1 M HCl/5 mM diethylenetriaminepentacetic acid (DTPA) and then disrupted by sonication (Sonopuls ultrasonic homogenizer, Bandelin) for 3 min with a repeating duty cycle of 0.3 s, in an ice bath. The cellular homogenate was centrifuged (20,000×g, 45 min) and the supernatant was used to determine the PCs. The PCs were separated and quantified by high-performance liquid chromatography (HPLC) after derivatization with the fluorescent tag monobromobimane (mBrB), as described elsewhere (Morelli et al. 2009).

Genotoxicity Bioassay

Genotoxicity was investigated using the yeast S. cerevisiae D7 diploid strain obtained from Zimmermann (Zimmerman et al. 1975). Mitotic gene conversion (GC) and reverse point mutation (PM) were measured at the trp5 and at the ilv1 loci, respectively. Stock cultures were maintained by inoculating ~1 × 107 cells/ml in a liquid complete medium (2% glucose, 2% bactopeptone, 1% yeast extract) and incubating at 30°C for 48 h until the stationary phase.

Bioassays were performed using growing yeast cells in the presence of elutriates. Yeast cells from a maintenance culture were inoculated in a 4-ml liquid complete medium (20% glucose, 2% bactopeptone, 1% yeast extract) in order to obtain ~2 × 107 cells/ml. The cultures were supplemented with 500 μl of elutriate and incubated under shaking at 150 rpm at 30°C for 6 h to reach the logarithmic growth phase. Control cultures (yeast cells in liquid complete medium) were performed with the same procedure. Aliquots of these cultures were diluted and then plated on complete and selective media to evaluate cytotoxicity as well as to ascertain trp convertants and ilv revertants. The results are expressed as a mean of three independent experiments and were analyzed using the Student’s t-test. Differences were considered significant for p ≤ 0.05.

Results and Discussion

Chemical Analysis

Sediments were collected during the summers of 2007 and 2008 from three coastal sites of the Gulf of Follonica: Cala Violina (site V), the mouth of the river Pecora (site P), and the mouth of the river Cornia (site C). Cala Violina was chosen as the reference site, owing to its low anthropogenic impact, as proved in a study from 2004 (Frassinetti et al. 2006). Sites C and P were expected to be more polluted, as they flow through industrial, agricultural, and mining areas. The mean values of As, Cd, Cr, Cu, Ni, and Pb for each set of measurements in each site are reported in Table 1. As a comparison, quality standard values of metal concentration in marine sediments suggested by the Italian government (Italian Ministerial Decree 367/2003) are reported, along with Consensus Based-Threshold Effect Concentration (CB-TEC) and Consensus Based-Probable Effect Concentration (CB-PEC) suggested by MacDonald et al. (2000). Table 1 highlights that, in 2007, metal concentrations from site V were far higher than those of 2004. This suggests that some contamination occurred in that period in the area; therefore, these concentrations could not be used as reference values. Cd, Cu, and Ni levels of site V 2007 sediments were comparable to those measured in the other two polluted sites. On the other hand, Cr and Pb values were lower than those found in the other two sites, but were 1.7- and 5.5-fold higher than the 2004 values, respectively. Metal concentrations in site V had returned to the background levels in 2008, with the slight exception of Pb and Cr, suggesting that the cause of pollution had ceased. The Pecora River mouth appeared to be the most polluted site, apart from Cr, without appreciable changes over the 2 years. Metal concentrations in site C sediments were generally higher in 2007 than in 2008, thus indicating higher levels of contamination in 2007 for this site as well. It was the most affected by Cr contamination, with a peak of 121 mg/kg, which is more than twice the quality standard value of marine sediments suggested by the Italian government (50 mg/kg; Italian Ministerial Decree 367/2003) and by MacDonald et al. (2000).
Table 1

Metal concentration (mg/kg dry weight) in sediments sampled in summer 2007 and 2008

Site

[As]

[Cd]

[Cr]

[Cu]

[Ni]

[Pb]

Cala Violina

 Year 2004

11.8 (1.6)

0.009 (0.004)

18.6 (9.5)

0.77 (0.1)

8.0 (2.2)

0.52 (0.3)

 Year 2007

7.7 (2.6)

0.044 (0.001)

31.6 (2.3)

1.80 (0.47)

16.8 (4.1)

2.84 (1.03)

 Year 2008

9.1 (0.8)

0.006 (0.004)

27.9 (5.7)

0.85 (0.18)

9.8 (2.4)

1.06 (0.28)

Pecora

 Year 2007

14.2 (2.2)

0.061 (0.018)

66.4 (20.7)

1.95 (0.52)

15.1 (5.6)

11.20 (3.41)

 Year 2008

17.7 (1.4)

0.052 (0.005)

68.7 (6.7)

2.74 (0.51)

16.1 (2.1)

9.55 (1.26)

Cornia

 Year 2007

13.0 (2.3)

0.047 (0.005)

121.1 (20.2)

1.70 (0.61)

16.2 (1.7)

5.11 (0.68)

 Year 2008

14.2 (1.8)

0.037 (0.006)

84.5 (30.3)

1.50 (0.36)

11.1 (1.8)

3.36 (0.35)

D.M 363 2003a

12

0.3

50

30

30

CB-TECb

9.79

0.99

43.4

31.6

22.7

35.8

CB-PECc

33

4.98

111

149

48.6

128

Note: Each number is the average value of the 11 samples coming from the site (see Fig. 1), and the values in parentheses represent the standard deviation. Cala Violina 2004 values are reported as a comparison

aItalian Ministerial Decree n. 367 (2003), Quality standard of sediments of marine/coastal waters

bConsensus-based threshold effect concentration

cConsensus-based probable effect concentration (MacDonald et al. 2000)

The As concentration was generally higher than the average values (5–6 mg/kg) reported for uncontaminated sites (Focardi and Tiezzi, 2009) in all the samples, including those from site V, in accordance with what had already been observed in this area (Baroni et al. 2004; Frassinetti et al. 2006; Leoni and Sartori 1997). However, these values did not change significantly over time. As and Cr concentrations in sites C and P were always higher than the CB-TEC, although with some variability, whereas all the other metal concentrations were below them.

A comparison between metal analyses of sediments and their elutriates for one selected transect of each site is reported in Table 2. The data show that, in agreement with the higher metal concentrations in 2007 compared to 2008 sediments, higher Cd and Cu concentrations were measured in the elutriates in the first year. The Pb concentration in the elutriates did not change appreciably, despite the fact that the Pb concentration in sediments in 2007 was higher than in 2008, suggesting that this metal was present in a nonexchangeable chemical form. Concerning the 2007 samples, it is worth noting that higher Cd levels were measured in the site V elutriates than in sites P and C, although metal concentrations in sediments were similar in all the sites.
Table 2

Chemical analyses of metals in sediments and elutriates, carried out in one selected transect of each site

Site

Year

Sample

Sediment (mg/kg DW)

Elutriate (nM)

[As]

[Cd]

[Cr]

[Cu]

[Ni]

[Pb]

[Cd]

[Pb]

[Cu]

Cala Violina

2007

V 0 m

4.2 (0.4)

0.043 (0.001)

30.8 (0.7)

1.57 (0.18)

13.9 (0.9)

2.82 (0.13)

15.1 (2.0)

0.4 (0.1)

66 (23)

V 50 m

7.1 (0.6)

0.043 (0.002)

29.2 (0.1)

2.16 (0.04)

20.9 (0.7)

1.59 (0.16)

11.5 (3.2)

0.4 (0.2)

98 (12)

V 100 m

9.4 (0.3)

0.064 (0.002)

29.4 (0.0)

2.50 (0.24)

21.1 (0.2)

1.75 (0.16)

2.5 (1.2)

1.4 (0.3)

78 (10)

2008

V 0 m

9.0 (0.6)

0.001 (0.001)

25.4 (0.3)

0.59 (0.10)

7.6 (0.3)

0.91 (0.01)

0.2 (0.1)

0.7 (0.1)

31 (3)

V 50 m

8.6 (0.4)

0.007 (0.001)

22.4 (0.6)

0.70 (0.01)

10.0 (0.3)

1.00 (0.01)

0.1 (0.2)

0.5 (0.3)

42 (9)

V 100 m

8.8 (0.2)

0.017 (0.001)

34.3 (1.3)

0.72 (0.06)

10.6 (0.4)

1.15 (0.05)

0.5 (0.3)

0.9 (0.4)

20 (3)

Pecora

2007

P 0 m

11.6 (0.4)

0.050 (0.001)

65.2 (5.6)

2.24 (0.04)

10.5 (0.2)

12.10 (0.11)

3.1 (0.3)

0.9 (0.2)

162 (11)

P 50 m

14.7 (0.5)

0.104 (0.002)

69.7 (3.0)

1.86 (1.22)

14.4 (0.3)

12.80 (0.42)

4.3 (0.5)

1.1 (0.2)

48 (6)

P 100 m

16.3 (1.2)

0.066 (0.001)

110.3 (0.5)

2.25 (0.18)

15.2 (0.3)

14.81 (0.20)

0.1 (0.2)

1.3 (0.4)

63 (4)

2008

P 0 m

18.8 (0.1)

0.048 (0.005)

62.8 (1.8)

1.87 (0.10)

16.2 (0.3)

10.93 (0.01)

0.8 (0.5)

1.2 (0.2)

10 (4)

P 50 m

14.0 (0.9)

0.051 (0.001)

62.5 (2.7)

2.13 (0.04)

13.3 (0.1)

7.93 (0.07)

0.1 (0.1)

1.7 (0.3)

n.m.

P 100 m

19.0 (0.6)

0.061 (0.004)

69.6 (0.0)

2.60 (0.60)

17.4 (0.1)

9.47 (0.25)

0.1 (0.1)

1.5 (0.2)

n.m.

Cornia

2007

C 0 m

13.1 (0.6)

0.044 (0.002)

113.3 (8.8)

0.96 (0.02)

15.8 (0.2)

4.84 (0.04)

1.2 (0.3)

0.8 (0.2)

161 (20)

C 50 m

13.7 (0.1)

0.044 (0.001)

160.4 (0.6)

2.65 (0.64)

19.4 (0.8)

5.27 (0.48)

1.5 (0.5)

1.2 (0.1)

139 (5)

C 100 m

14.3 (0.4)

0.048 (0.002)

108.1 (3.8)

1.05 (0.04)

14.0 (0.4)

5.46 (0.22)

3.5 (1.0)

0.9 (0.2)

84 (10)

2008

C 0 m

18.6 (0.5)

0.038 (0.003)

84.3 (1.2)

1.34 (0.02)

12.4 (0.8)

4.13 (0.07)

0.1 (0.1)

0.6 (0.5)

5 (4)

C 50 m

14.7 (0.2)

0.036 (0.003)

93.7 (0.5)

2.31 (0.47)

11.0 (0.2)

3.54 (0.05)

0.1 (0.1)

0.3 (0.1)

3 (3)

C 100 m

14.7 (0.2)

0.035 (0.001)

50.0 (1.7)

1.44 (0.29)

11.9 (0.0)

3.58 (0.08)

0.3 (0.2)

0.6 (0.1)

40 (4)

Control seawater

       

0.1 (0.1)

0.3 (0.2)

2 (2)

Control seawater was used to extract elutriates

Note: n.m. = not measurable due to interferences. Values in parentheses represent the standard deviation (n = 3 analytical measurements)

The interannual variation in metal levels, detected in sediments from 2007 to 2008, suggests that some of the metals bound to the 2007 sediments had subsequently been released into the water column, thus confirming their higher content of water-soluble metal forms. This was more evident for the site V samples, which exhibited higher interannual variations in metal concentrations, especially for Cd. Consequently, we hypothesized that higher damage to living organisms could occur. To obtain a toxicological evaluation of the three sites, two separate bioassays, using the biomarkers of genotoxicity and metal bioavailability, were carried out in the elutriates.

PC-Based Bioassay

The bioavailability of metals measured in the elutriates was tested by applying a bioassay using the presence of PCs in T. weissflogii as a biomarker of metal exposure (Morelli et al. 2009). The results of the PC-induction test (Fig. 2) highlighted a higher production of PCs in the T. weissflogii grown in the elutriates from sediments sampled in 2007 than in those in 2008. This was in agreement with the general pattern of metal concentrations in the sediments and in the elutriates. The highest PC response was detected in algae grown in the site V elutriates (2007). These results suggest that the water-soluble fraction of the metals adsorbed by these sediments contained bioavailable forms of metals, which are potentially toxic. The fact that PC induction was higher in samples from the reference site suggests that occasional pollution events might be more hazardous in uncontaminated areas where the newly sediment-bound metals are more easily released. In addition, no significant synthesis of PCs was detected in algae grown in the 2008 elutriates (compared to the control), despite the fact that metals concentrations in the site C and site P sediments were 2–10 times higher than the reference site.
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Fig. 2

Phytochelatin concentration in cells of T. weissflogii grown for 6 days in the elutriates of selected sediment samples described in Table 2. Error bars correspond to the standard deviation (n = 3). Significant differences from the control (p ≤ 0.05) are indicated with an asterisk. PCs are expressed as moles of γ-Glu-Cys units per cell (amol = attomoles = 10−18 moles)

Of the most common PC-inducing metal ions, ASV measurements revealed that Cd and Cu were mainly released through sediment resuspension, whereas the release of Pb was negligible (see Table 2). To assess which of the two metal ions could have triggered the PC synthesis, we examined the relationship between the total dissolved metal concentration in the elutriates and the intracellular PC concentration in T. weissflogii. Figure 3 shows a positive and significant correlation (95% level) with [Cd] (r = 0.969) but not with [Cu] (r = 0.169), suggesting that Cd could be the element that was most responsible for PC induction. In addition, it is well known that Cd is the best activator of PC synthesis both in vivo and in vitro incubations (Ahner and Morel 1995; Grill et al. 1987). The absence of a relationship between Cu and PCs could be explained by a low bioavailability of Cu, as it is commonly reported that in coastal seawater, Cu is almost entirely complexed by organic ligands (Buck et al. 2007).
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Fig. 3

Correlation between PC concentration in cells of T. weissflogii and total dissolved concentration of Cd (a) or Cu (b) in the elutriates. Significant correlations (p ≤ 0.05) are indicated with an asterisk

In order to determine the level of toxicity, the results on the PC response should be integrated with other biological end points. Indeed, it is possible that the presence of non-PC-inducing metal ions or some other unidentified environmental contaminants might pose a threat to living organisms at various levels of biological complexity.

Genotoxicity Bioassay

The genotoxic effects of sediment elutriates were evaluated in exponentially growing cultures of the yeast S. cerevisiae by analyzing frequencies of mitotic gene conversion at the trp5 locus and the reversion frequencies at the ilv1 locus. Cytotoxic effects were evaluated as the percentage of cell survival. Genotoxic and cytotoxic effects were observed in two of three sites examined in 2007 (Fig. 4a–c). The results showed significant genotoxic effects in yeast cultures that had been supplemented with the site V and site C elutriates but not with the site P elutriates. Yeast cells exposed to site V elutriates showed almost twice the level of mitotic gene conversion than the control, whereas the highest mutagenic effect was evident in site C elutriates (Fig. 4a). Similarly, the PM values were significantly higher than the controls in yeast cells treated with site V and site C elutriates. No significant effect was found in the site P elutriates (Fig. 4b).
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Fig. 4

Effects of the sediment elutriates on: mitotic gene conversion (a, d), reverse point mutation (b, e), and survival (%) (c, f) in growing cells of the D7 strain of S. cerevisiae. Sediment samples are described in Table 2. Results are derived from triplicate determinations, means ± standard deviations. Significant differences from the control (p ≤ 0.05) are indicated with an asterisk

In agreement with the genotoxic effects, our data also highlighted a significant cytotoxic effect. In fact, cell survival in cultures inoculated with the site V and site C elutriates was about 75 and 45% of the controls, respectively (Fig. 4c). On the other hand, no significant mutagenic and cytotoxic effects were observed in yeast cells treated with sediment elutriates in 2008 (Fig. 4d–f). The fact that we found significant genotoxicity in yeast cells supplemented with elutriates in 2007 but not in 2008 is consistent with the generally higher metal concentrations measured in the 2007 sediments. Several heavy metals are known to be mutagenic or carcinogenic compounds (Ohe et al. 2004; Vargas et al. 2001); thus, the metal contamination might explain some of the genotoxic results. In a complex mixture, such as sediment extracts, metals might also act as co-mutagens, thus increasing the genotoxic activity of other compounds (Magdaleno et al. 2008). Of the metals monitored in this work, Cr and Cd exhibited the highest mutagenic potential (Codina et al. 2000). It has been reported that Cd interferes with DNA repair processes (Bertin and Ayerbach 2006) and induces oxidative stress and DNA damage in S. cerevisiae (Brennan and Schestl 1996). Our results show significant genotoxic effects in the two sites with higher interannual variations in Cd or Cr concentrations in the sediments. Cd could be the main metal responsible for the mutagenic effects in the site V 2007 elutriates. The genotoxic effect found in the site V elutriates is in agreement with the presence of bioavailable metal forms, mainly Cd, as revealed by the PC bioassay. In addition, our results seem to suggest that the high concentrations of Cr, particularly in the site C sediments, might also be responsible for the genotoxic effects. Although the Cr concentration was not measured in these elutriates, the Cr concentration in the 2008 sediments underwent a mean decrease of 41 ± 14% compared to 2007 (calculated from the site C samples in Table 2), thus indicating the presence of water-soluble metal forms in the 2007 sediments. The absence of significant mutagenic effects in the site P elutriates seems to be in contrast with the metal content in these sediments. This could be explained by a low concentration of bioavailable metal forms in the elutriates or the absence of other mutagenic compounds in that specific site.

Conclusions

The present study coupled chemical analyses and biological responses in a coastal area that is subject to anthropogenic inputs. A metal analysis of the sediments, carried out in selected sites during a 2-year sampling period, showed higher metal concentrations in the 2007 sediments than in those of 2008. In agreement with chemical data, the use of a biomarker of metal bioavailability revealed the induction of PCs in T. weissflogii cells exposed to the elutriates of sediments collected in 2007. A further analysis suggested that Cd might be the metal mainly responsible for PC induction. However, elutriates are complex mixtures and can contain other metals, which might also be toxic because of their mutagenic potential, but they are not able to stimulate PC synthesis as well as other unmeasured organic pollutants. Alternative bioassays should thus be applied. In this article we used the D7 strain of S. cerevisiae to predict the genotoxicity of marine sediment resuspensions. The results showed DNA damage in yeast supplemented with 2007 elutriates, which also induced the highest PC synthesis in T. weissflogii. In conclusion, we believe that by integrating the biomarkers of metal bioavailability and genotoxicity, coupled with chemical analyses, the potential hazards and sublethal impacts on biota can be identified. This consequently provides a valuable contribution to the ecotoxicological characterization of marine sediments.

Acknowledgments

The authors would like to thank Dr. Roberto Bedini (Institute of Biology and Marine Ecology, Piombino–LI) for providing the sediment samples and Alessandro Puntoni (Institute of Biophysics, CNR–PI) for technical assistance.

Copyright information

© Springer Science+Business Media, LLC 2011