Variations in sensitivity to copper and zinc among three isolated populations of the seagrass, Zostera capricorni

Journal of Experimental Marine Biology and Ecology 302 (2004) 63 – 83 www.elsevier.com/locate/jembe

Variations in sensitivity to copper and zinc among three isolated populations of the seagrass, Zostera capricorni Catriona M.O. Macinnis-Ng *, Peter J. Ralph Department of Environmental Sciences, Institute of Water and Environmental Resource Management, University of Technology, Westbourne Street, Sydney, Gore Hill, NSW 2065, Australia Received 29 May 2003; received in revised form 7 September 2003; accepted 1 October 2003

Abstract Metal accumulation in seagrass is well documented, but toxic impacts and mechanisms of tolerance in seagrass are not well understood. We looked at the impacts of 10 h exposure to copper and zinc for three isolated populations of Zostera capricorni in the Sydney (Australia) region. Photosynthetic efficiency (measured as the effective quantum yield, DF/FmV) and chlorophyll pigment concentrations showed different sensitivities to metal impacts at the three geographically isolated sites. Seagrasses from the least developed estuary were the most sensitive to metals and the two more developed estuaries had more tolerant populations. Determination of metal concentrations in the leaves showed that there was no difference in metal exclusion as the sensitive seagrass accumulated no more metal than the tolerant seagrass. Equally, background levels of copper and zinc in the sediments and seagrass tissue could not explain the differences in tolerance. We discuss some other possible mechanisms of tolerance. The outcomes suggest that assessing metal content in seagrass tissue may not demonstrate degree of photosynthetic impact. D 2003 Elsevier B.V. All rights reserved. Keywords: Chlorophyll a fluorescence; Metal tolerance; Metal toxicity; Photosynthetic impacts; Seagrass

1. Introduction Metal accumulation has been extensively studied in a variety of seagrass species (e.g. Fabris et al., 1982; Brix et al., 1983; Nienhuis, 1986; Malea et al., 1995; Schlacher-Hoenlinger and Schlacher, 1998; Pergent-Martini and Pergent, 2000). Other * Corresponding author. Tel.: +61-2-9514-4072; fax: +61-2-9514-4003. E-mail address: [email protected] (C.M.O. Macinnis-Ng). 0022-0981/$ - see front matter D 2003 Elsevier B.V. All rights reserved. doi:10.1016/j.jembe.2003.10.002

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studies have examined the impact of metals on seagrass physiology (e.g. Lyngby and Brix, 1989; Brackup and Capone, 1985; Hamoutene et al., 1996; Ralph and Burchett, 1998; Prange and Dennison, 2000; Macinnis-Ng and Ralph, 2002) but none of these studies have linked the concentrations of metals in seagrass tissue to physiological impacts. Thus, we do not know what impact a certain concentration of a particular metal will have on a particular seagrass or if impacts are consistent across different populations and different species. Furthermore, the issue of metal tolerance in seagrasses has not been addressed. Plants which are tolerant to metal exposure occur as a result of natural selection, where selection pressure favours the tolerant genotype in a genetically diverse population (Turner, 1994). The development of tolerance in short-lived plants by natural selection of tolerant genotypes is well understood and cultivars tolerant to metal exposure have been specifically bred for growth at contaminated sites (Turner, 1994; Ciscato et al., 1997). The longevity of some species means that genetic mechanisms of tolerance have only limited bearing on the survival of the population and the concept of plant acclimation to incidence of pollution has received more attention (Turner, 1994; Shaw, 1999). Indeed, phenotypic plasticity leading to metal tolerance has been linked to the nature of contamination in terrestrial plants and inducible tolerance by pre-treatment to cadmium was demonstrated by Turner (1994). Experiments using pre-treatment of clonal plant materials have shown that metal tolerance can be induced and this tolerance is also ‘lost’ when the metal influence is removed (Macnair and Baker, 1994). In the marine environment, a similar result was found by Harland and Brown (1989) where corals exposed to daily run-off enriched with iron showed a greater tolerance to this metal. For seagrasses in particular, Alberte et al. (1994) suggested that losses of biomass and distribution due to exposure to anthropogenic impacts may lead to loss of genetic diversity since certain genotypes are less able to compete in the presence of pollutants. This has serious implications for the management of different populations of seagrass since tolerances to environmental and anthropogenic perturbations may be different for different populations (Waycott, 1998). Understanding differential sensitivity in different genotypes is seen as an area requiring considerable attention in the field of ecotoxicology (Moore, 2002). Furthermore, studies of physiological costs associated with tolerance in metal tolerant plants on contaminated sites suggest that tolerant populations are unable to compete with intolerant populations on uncontaminated sites (Shaw, 1999). While direct experimental investigations are mostly inconclusive, the possible ramifications for seagrass management are serious enough to warrant further investigation. Hormesis, a concept receiving more and more attention in general toxicology studies, may also play a role in tolerance to metals in seagrasses. Hormesis is an adaptive response involving compensatory biological processes following an initial disruption in homeostasis (Calabrese and Baldwin, 2002). In particular, overcompensation/stimulation hormesis is an adaptive response to low levels of stress which can result in enhanced fitness, due to a modest overcompensation to a disruption in homeostasis. When this occurs, a low dose of heavy metals administered prior to a more catastrophic dose may reduce the toxic impact of the subsequent higher concentration (Calabrese and Baldwin, 2002). This concept was

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demonstrated in fish by Allin and Wilson (2000). They found that pre-acclimation to aluminium significantly reduced the mortality of rainbow trout (Oncorhynchus mykiss) when compared to aluminium-naı¨ve fish and they concluded that previous exposure to low levels of aluminium may be an important factor in negating the impact of this metal on fish. No pre-acclimation metal studies have been conducted on seagrass. However, different estuaries in the Sydney region have had different degrees of industrial, agricultural and urban activity in their catchments (Birch and Taylor, 1999). Thus, the seagrasses growing in Botany Bay, Sydney Harbour and Pittwater are likely to be acclimatised to different levels of different pollutants, leading to phenotypic variation. Furthermore, there may be some population differences due to genotypic variation which could influence the impact of pollutants; however, this is currently unknown. From the management perspective, whether any differences are phenotypic or genotypic does not matter, the important point is to determine whether differences in tolerance do or do not occur. If there are differences, the toxicological data from one estuary should not be used in the management of another, but if the results are similar for different areas of seagrass colonisation, data from one estuary can be applied to another with caution. The aim of the present experiment was to determine the consistency of metal impacts on three geographically isolated populations of Zostera capricorni. To this end, three pairs of identical experiments using the in situ technique developed by Macinnis-Ng and Ralph (2002) were carried out in three estuaries around Sydney. In addition to collecting chlorophyll a fluorescence and chlorophyll pigment data, water, sediment and seagrass samples were collected to assess the concentrations of metals present at the sites before, during and after the exposure period and to also assess any differences in uptake concentrations.

2. Methods 2.1. Study sites The in situ chamber experiments were carried out in three separate but adjacent estuaries in the Sydney region: Pittwater, Sydney Harbour and Botany Bay (Fig. 1). Pittwater is a semi-pristine embayment, the Sydney Harbour site is adjacent to an arterial road, a marina and heavy recreational boating activity and Botany Bay is renowned for its heavy industry. 2.2. Field experiments Perspex chambers were used to enclose and dose seagrass in the meadow. For details on the in situ exposure chamber design, fibre optics and stock solutions refer to Macinnis-Ng and Ralph (2002). The exposure regime is briefly outlined below. In situ samples were exposed to copper or zinc for 10 h (0800 to 1800 h), during which time hourly effective quantum yield (DF/FmV) measurements were taken. At the end of the exposure period, the sample leaves were marked with a wire peg and monitored for the

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Fig. 1. The location of sites visited for this experiment.

following 4-day period. One-way ANOVA was used at 2, 10 and 96 h to detect differences between treatments. Where a difference was found, Tukey’s HSD post-hoc comparison was used to determine which treatments were different (Statistica, Statsoft v5.5). There were four treatments, 1 mg l 1, 0.1 mg l 1 copper or zinc, chamber control and external control.

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2.3. Photosynthetic pigment analysis Four leaf samples for each treatment were collected after 10 and 96 h. Pigment extraction was performed using the solvent N,N-dimethylformamide (DMF). Refer to Macinnis-Ng and Ralph (2002) for more details on pigment extraction. Total chlorophyll concentrations were representative of pigment impacts so the other pigment concentrations and ratios are not reported. Total chlorophyll concentrations were compared using ANOVA to identify any differences between treatments and Tukey’s HSD post-hoc comparison as above. 2.4. Analysis of metal concentrations in seagrass, water and sediment samples In order to determine the true concentration of metals in the chambers, water samples were collected from each chamber and the external control at the beginning and end of the exposure period. A 60 ml sample was removed from each chamber using a plastic 20 ml syringe and placed in plastic sample jars with 1 ml concentrated nitric acid. Sediment samples 27 cm3 were collected at the end of the exposure period and the end of the experiment to determine the current concentration of metals at the site and also to determine whether any adsorption occurred in the sediment due to experimental exposure. Leaf samples were collected at the conclusion of the exposure period and at the end of the recovery period. These samples were cleaned of epiphytes, placed in paper bags and on return to the laboratory, shaken in 50 ml 1 mM EDTA solution for 1 min to remove metals adsorbed to the surface of the leaves (Lyngby and Brix, 1984). Epiphyte loads at the three sites were low, at less than 3% leaf surface area. Roots were only collected at the end of the experiment to avoid disrupting the seagrasses during the experiment. A summary of the samples collected is shown in Table 1. 2.5. Acid digestion of leaf, root and sediment samples Digestion of sediment and plant material followed the method of Krishnamurty et al. (1976). Approximately 0.3 g dried leaf material, 1 g dried root material and 0.5 g dried sediment were placed into 50 ml beakers. Sediment samples were prepared in

Table 1 Samples collected at different stages of the experiment for metal analysis Stage of experiment t = 1/2 t = 10 t = 96

Sample Water

Sediment

Leaves (in and out)

Roots

4 treatments 4 treatments 1 treatment

4 treatments 4 treatments

4 treatments 4 treatments

4 treatments

In all cases, samples were collected for all four treatments (external control, chamber control, 0.1 and 1 mg l 1), except for the 96 h water sample which was assumed to be similar for all treatments, since the chambers had been removed.

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duplicate. Five-milliliter concentrated analytical grade nitric acid was added to each beaker before it was placed on a sand bath (Labec, Laboratory Equipment, Sydney) at 120 jC with a watch-glass on top of it. The samples were allowed to digest for several hours before 3 ml hydrogen peroxide was added step-wise, while the samples remained on the sand bath for several more hours. On completion of digestion, the samples were allowed to cool, then filtered into 25 ml volumetric flasks using acid washed fibre Whatman (5b, 10 cm) filter papers. Samples were analysed in the Atomic Absorption Spectrophotometer (AAS, VarianAA-1275, Australia). Standards were matrix-matched and international tissue/sediment samples were digested with each batch to check recovery rates. Recovery rates were approximately 80% for leaf material (NBS 1575 pine needles) and 90% for sediment samples (BCR 141 calcareous loam) for both metals. 2.6. Sediment physicochemical parameters Sediment physicochemical parameters were assessed to identify any major differences in the sediments at the three sites, which may influence the bioavailability of copper and zinc. Loss on ignition (LOI) was measured by combusting pre-dried, pre-weighed samples at 550 jC for 5 h as an indication of organic carbon content (Allen, 1989). Conductivity and pH were measured after mixing pre-dried samples with Milli-Q water on a 1:3 sediment to water ratio (Allen, 1989). Sediment size fractions were determined using a mechanical sieve system with 2 mm and 100 Am meshes. Sand was defined as the 2 mm – 100 Am class and the silt/clay fraction was < 100 Am (Allen, 1989). Finally, the CaCO3 or shell content was determined by adding dilute HCl to pre-weighed, pre-dried samples until effervescence no longer occurred and then redrying and reweighing the samples. The difference between the original and final masses gives an indication of the shell content of the samples (Morrisey et al., 1998).

3. Results 3.1. Copper photosynthetic parameters 3.1.1. Chlorophyll a fluorescence Samples from Pittwater were the most photosynthetically sensitive to copper exposure (Fig. 2), during both the exposure and recovery periods. Pittwater samples exposed to 1 mg l 1 Cu2 + had DF/FmV values as low as 0.2 units during the exposure period (Fig. 2a), while the DF/FmV of samples from Sydney Harbour and Botany Bay (Fig. 2b and c) did not drop below 0.3 units at the lowest point. Additionally, during the recovery period, the samples exposed to copper from Sydney Harbour and Botany Bay returned to preexposure levels of DF/FmV, while the Pittwater samples showed some recovery but the final measurement had declined substantially in some of the 1 mg l 1 Cu2 + samples. This is supported by the statistical analysis of the results, which showed that the Pittwater samples were significantly impacted by copper at 2 and 96 h. The 1 mg l 1 Cu2 + treatment had a significant impact on the Sydney Harbour samples at 2 h but the

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Fig. 2. Effective quantum yield values for samples exposed to copper at three different sites, (a) Pittwater, (b) Sydney Harbour and (c) Botany Bay. Treatments were: external control (q), chamber control (.), 0.1 mg l 1 copper (n) and 1 mg l 1 copper (E). The break on the abscissa represents the end of the exposure period and the beginning of ‘recovery’. All values are mean F standard error of the mean.

statistically significant difference at 96 h was not different to both controls so recovery was evident. There were no differences in the Botany Bay samples (Table 2). 3.1.2. Chlorophyll pigments Chlorophyll pigment determinations also indicated that the samples from Pittwater were more sensitive to copper exposure than those from Botany Bay and Sydney Harbour

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Table 2 Statistical analysis of DF/FmV data at 2, 10 and 96 h at the three sites for exposure to copper and zinc Treatment

2h

10 h

96 h

p

ext

cnt

0.1

1

p

ext

0.000 0.629

a ns

a

b

b

0.039 0.755

Sydney Harbour Copper 0.000 Zinc 0.694

a ns

a

a

b

Botany Bay Copper 0.131 Zinc 0.926

ns ns

Pittwater Copper Zinc

cnt

0.1

1

p

ext

cnt

0.1

1

ns ns

0.015 0.393

a ns

a

a

b

0.404 0.176

ns ns

0.007 0.027

a ns

ab

ab

b

0.951 0.622

ns ns

0.116 0.726

ns ns

Significance level: p < 0.02 (due to Bonferonni correction for multiple analyses); ns = not significant. Where a significant difference occurred, a Tukey’s HSD post-hoc comparison was used and the differences between treatments were reported as different letters in the table. Treatments were ext = external control, cnt = chamber control, 0.1 = 0.1 mg l 1 copper or zinc and 1 = 1 mg l 1 copper or zinc.

(Table 3). Total chlorophyll concentration was lower in the 1 mg l 1 treatment in comparison to controls at Pittwater. There was also a significant difference in the Sydney Harbour samples but these were not related to the copper treatments. The significant difference between the chamber control and external control indicated inherent variation may have been overriding any impacts due to copper exposure. There were no significant differences at Botany Bay.

Table 3 Total chlorophyll concentration (mean values F standard error, Ag cm capricorni to copper and zinc at three field sites Treatment

Pittwater

2

) after 10 and 96 h exposure of Z.

Sydney Harbour

Botany Bay

10 h

96 h

10 h

96 h

10 h

96 h

Copper External control Control 0.1 mg l 1 Cu2 + 1 mg l 1 Cu2 + ANOVA F

10.4 F 1.4 15.4 F 0.6 12.5 F 1.4 10.3 F 1.8 2.986

15.0 F 0.5a 17.1 F 1.3a 14.2 F 1.2ab 9.9 F 1.1b 7.963*

19.8 F 0.3a 14.5 F 1.1b 16.1 F 0.8ab 18.6 F 1.4ab 5.829*

14.6 F 0.5 17.1 F 1.6 19.2 F 1.0 17.5 F 1.2 2.897

15.5 F 1.5 19.1 F 1.0 20.4 F 2.5 21.1 F 2.4 1.653

15.0 F 0.7 21.5 F 0.8 21.0 F 2.4 14.8 F 2.5 4.181

Zinc External control Control 0.1 mg l 1 Zn2 + 1 mg l 1 Zn2 + ANOVA F

14.5 F 1.2 13.0 F 1.0 13.9 F 1.0 15.3 F 0.7 0.999

16.6 F 1.3 15.1 F 1.2 17.8 F 2.2 15.9 F 1.0 0.570

15.3 F 0.7 18.8 F 0.8 18.4 F 1.7 17.2 F 0.1 2.101

17.5 F 1.1 19.4 F 1.5 18.4 F 1.6 18.6 F 1.1 0.350

24.8 F 2.7 25.0 F 0.4 24.9 F 1.5 24.0 F 2.7 0.052

26.1 F 2.5 30.7 F 1.8 26.3 F 1.5 27.7 F 2.4 1.061

Treatments with the same letter are not significantly different (one-way ANOVA, Tukey’s HSD multiple comparison [ p < 0.05] test) within each group of treatments.

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3.2. Zinc photosynthetic parameters 3.2.1. Chlorophyll a fluorescence Zinc had virtually no impact on the DF/FmV in all three estuaries (Fig. 3a –c). The DF/ FmV values remained steady throughout the experiments and there were no significant differences between the samples due to zinc exposure (Table 2).

Fig. 3. Effective quantum yield values for samples exposed to zinc at three different sites: (a) Pittwater, (b) Sydney Harbour and (c) Botany Bay. Treatments were: external control (q), chamber control (.), 0.1 mg l 1 copper (n) and 1 mg l 1 copper (E). The break on the abscissa represents the end of the exposure period and the beginning of ‘recovery’. All values are mean F standard error of the mean.

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3.2.2. Chlorophyll pigments There were no statistically significant differences in total chlorophyll concentrations for the zinc experiments (Table 3). 3.3. Metal determinations 3.3.1. Background concentrations The background concentration of metals varied between the three sites (Fig. 4). For the copper concentrations (Fig. 4a), each of the sites had significantly different values for the sediments ( F = 142, p = 0.000), inside leaves ( F = 92, p = 0.000), outside leaves ( F = 44, p = 0.000) and within the root fraction ( F = 87, p = 0.000). Concentrations of copper in the water were the same at Pittwater and Botany Bay, but both of these were different to Sydney Harbour ( F = 6.8, p = 0.004). In all of the fractions where all three sites had different concentrations, the highest concentration was in Sydney Harbour, then Pittwater and Botany Bay had the lowest concentrations of copper. Zinc concentrations (Fig. 4b) were more varied within sites than copper and the actual concentrations were also significantly higher. Outside the leaves ( F = 9.4, p = 0.001) and in the water ( F = 7.7, p = 0.002) were the only two fractions which had significant differences due to the sites. Outside the leaves, the lowest concentration occurred at Pittwater which was different to the highest concentration at Sydney Harbour while Botany Bay was an intermediary. For the water samples, zinc concentrations were similar at Pittwater and Botany Bay and both of these sites were different to the higher concentration in Sydney Harbour. 3.3.2. Experimental concentrations Concentrations of metals in the water column of exposure chambers are reported in Table 4. The nominal 1 mg l 1 treatment in both the copper and zinc experiments was closer to 0.5 and the 0.1 mg l 1 treatment ranged between 0.11 and 0.24 mg l 1 at 0.5 h. The background concentration already present at different sites may be contributing to the variation in these values. Due to the variable concentrations of copper and zinc in sediment, leaves and roots in control samples, we have calculated the percentage difference between control concentrations and treatments (Tables 5 and 6) to enable a clearer interpretation of the relative impact at the three sites. Due to the variation in these values, even within a site, the differences were only considered to be significant when the percentage difference was greater than two times lower or higher (that is, < 50% or >200%). For the copper experiments, all of the parameters which met this criterion had values greater than the controls (Table 5). For the Pittwater experiment, the concentration of copper inside the leaves reached about 220% of the control concentration after 10 h exposure to 1 mg l 1 Cu2 + but had dissipated this load by 96 h and the concentration outside the leaves reached about 290% after 10 h for this concentration but also dissipated by 96 h. In Sydney Harbour, the concentration of copper on the outside of leaves increased to 230% in the higher treatment. Concentrations of copper were most elevated inside and outside the Botany Bay leaves after 10 h exposure to 1 mg l 1 copper. These values reached 1120% and 1550%, respectively, and both values remained elevated at 96 h (290% and 250%).

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Fig. 4. Concentrations of copper (a) and zinc (b) in controls at Pittwater, Sydney Harbour and Botany Bay. Concentrations are in Ag g 1 for sediments, inside and outside leaves and in roots and ng l 1 in the water. n = 8 for sediments and leaves, n = 4 for roots and n = 10 for water samples. All values are mean F standard error of the mean.

The 0.1 mg l 1 treatment also led to high copper levels inside and outside of the leaves (330% and 530%), while the concentration of copper remained high inside the leaves after the recovery period, the concentration outside leaves had returned to normal levels. In terms of absolute values, however, the samples with the highest concentrations of copper inside and outside leaves were found at Pittwater and Sydney Harbour, while the

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1

) in water at the beginning of the experiment (0 h), at the end of the exposure period (10 h) and the end of the experiment (96 h)

Nominal treatment

Pittwater

Sydney Harbour

0h

10 h

96 h

0h

10 h

96 h

0h

10 h

96 h

External control Chamber control 0.1 mg l 1 Cu2 + 1 mg l 1 Cu2 +

0.10 F 0.00 0.10 F 0.00 0.15 F 0.04 0.39 F 0.15

0.10 F 0.00 0.10 F 0.00 0.11 F 0.01 0.12 F 0.00

0.10 F 0.00 – – –

0.10 F 0.00 0.10 F 0.00 0.24 F 0.02 0.50 F 0.00

0.13 F 0.01 0.11 F 0.00 0.12 F 0.00 0.11 F 0.00

0.10 F 0.00 – – –

0.10 F 0.00 0.10 F 0.00 0.17 F 0.02 0.51 F 0.31

0.10 F 0.00 0.10 F 0.00 0.10 F 0.00 0.10 F 0.00

0.10 F 0.00 – – –

External control Chamber control 0.1 mg l 1 Zn2 + 1 mg l 1 Zn2 +

0.00 F 0.00 0.00 F 0.00 0.14 F 0.05 0.33 F 0.04

0.00 F 0.00 0.00 F 0.00 0.00 F 0.00 0.00 F 0.00

0.00 F 0.00 – – –

0.03 F 0.03 0.03 F 0.00 0.14 F 0.05 0.44 F 0.29

0.02 F 0.00 0.01 F 0.00 0.01 F 0.00 0.01 F 0.00

0.00 F 0.00 – – –

0.00 F 0.00 0.00 F 0.00 0.11 F 0.03 0.60 F 0.27

0.00 F 0.00 0.00 F 0.00 0.00 F 0.00 0.02 F 0.02

0.00 F 0.00 – – –

Values are mean F standard error of the mean.

Botany Bay

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Table 4 Metal concentrations (mg l

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Table 5 Absolute concentrations of copper in sediment, leaves, roots and on the outside of leaves (Ag g 1 dry weight) for each treatment, differences between these concentrations and the appropriate control and concentrations relative to control concentrations at each site, as a percentage Pittwater

Sydney Harbour

Botany Bay

0.1

1

0.1

1

0.1

1

Sediment 10 h Concentration (Ag g 1) Difference to control % of control

8F1 0 100 F 10

6F0 2.2 72 F 8

35 F 3 6.3 122 F 10

31 F 2 2.7 109 F 8

2F0 0.5 123 F 2

2F0 0.2 109 F 16

Sediment 96 h Concentration (Ag g 1) Difference to control % of control

6F1 2.3 71 F 5

6F0 1.6 74 F 17

27 F 4 1.8 94 F 15

28 F 1 0.4 99 F 4

1F0 0.1 104 F 2

2F0 0 99 F 5

Leaf (internal) 10 h Concentration (Ag g 1) Difference to control % of control

90 F 5 37.0 170 F 5

117 F 16 63.5 219 F 14*

119 F 11 35.1 142 F 9

111 F 12 27.0 132 F 11

25 F 4 17.4 334 F 18*

89 F 48 81.6 1120 F 53*

Leaf (internal) 96 h Concentration (Ag g 1) Difference to control % of control

63 F 5 10.0 118 F 7

72 F 11 18.4 135 F 15

76 F 8 7.4 91 F 10

85 F 7 1.0 101 F 8

17 F 1 9.4 226 F 5*

21 F 1 13.8 286 F 6*

Leaf (external) 10 h Concentration (Ag g 1) Difference to control % of control

11 F 1 4.8 172 F 6

19 F 4 12.4 288 F 21*

17 F 5 7.0 172 F 28

23 F 2 13.0 232 F 9*

5F1 3.8 527 F 65*

14 F 9 12.8 1550 F 27*

Leaf (external) 96 h Concentration (Ag g 1) Difference to control % of control

9F1 2.3 135 F 16

7F1 0.8 112 F 19

9F0 1.1 88 F 3

10 F 3 0.6 106 F 26

1F0 0.3 131 F 4

2F0 1.3 249 F 1*

Roots 96 h Concentration (Ag g 1) Difference to control % of control

20 F 3 1.9 110 F 13

19 F 1 1.0 105 F 6

41 F 2 2.6 94 F 4

45 F 1 2.0 105 F 1

6F0 1.2 125 F 7

6F1 1.2 124 F 20

Different times of sampling are also indicated. Values are mean F standard error of the mean, except for the ‘difference to control’ which is the difference between the mean for that treatment and the mean of the site control. * Difference between treatment and control > 200%.

highest sediment and root concentrations were found at Sydney Harbour (Fig. 4). To determine the amount of metal in the different fractions due to metal addition, rather than ambient concentrations, the difference between the mean concentration in each particular fraction at each site and the concentration in the treated samples was calculated. For copper exposure, sediment and root values were reasonably similar across the sites at both 10 and 96 h (Table 5). Concentrations of metals in the leaves were highest at Botany Bay,

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closely followed by Pittwater, in 1 mg l 1 Cu2 +-exposed samples, at 10 h but the values at different sites were much more similar at 96 h. Concentrations of copper absorbed to the outside of leaves were similar in the 1 mg l 1 copper treatments since all three sites had values of around 13 Ag g 1 at 10 h (Table 5).

Table 6 Absolute concentrations of zinc in sediment, leaves, roots and on the outside of leaves (Ag g 1 dry weight) for each treatment, differences between these concentrations and the appropriate control and concentrations relative to control concentrations at each site, as a percentage Pittwater

Sydney Harbour

Botany Bay

0.1

1

0.1

1

0.1

1

Sediment 10 h Concentration (Ag g 1) Difference to control % of control

118 F 15 8.3 93 F 13

142 F 30 15.1 112 F 21

95 F 44 3.5 104 F 46

131 F 38 39.6 144 F 29

91 F 45 16.0 85 F 50

39 F 14 9.2 91 F 4

Sediment 96 h Concentration (Ag g 1) Difference to control % of control

158 F 107 30.8 124 F 68

176 F 67 49.5 139 F 38

45 F 2 45.8 50 F 6#

93 F 22 1.9 102 F 24

124 F 5 16.5 115 F 4

38 F 1 53.5 50 F 3#

Leaf (internal) 10 h Concentration (Ag g 1) Difference to control % of control

261 F 10 93.1 155 F 4

713 F 209 545.2 425 F 29*

140 F 4 82.0 63 F 3

233 F 43 11.6 105 F 18

211 F 100 37.3 85 F 48

669 F 368 111.2 145 F 17

Leaf (internal) 96 h Concentration (Ag g 1) Difference to control % of control

374 F 66 205.9 223 F 18*

178 F 70 9.8 106 F 39

240 F 8 18.6 108 F 3

158 F 46 64.1 71 F 29

186 F 7 62.2 75 F 4

216 F 66 20.8 92 F 49

Leaf (external) 10 h Concentration (Ag g 1) Difference to control % of control

13 F 5 3.5 137 F 35

14 F 2 4.2 144 F 17

25 F 15 8.7 74 F 62

49 F 1 16.0 148 F 3

17 F 0 3.4 83 F 1

26 F 10 11.5 156 F 32

Leaf (external) 96 h Concentration (Ag g 1) Difference to control % of control

5F0 4.3 56 F 3

11 F 2 1.3 114 F 17

19 F 10 14.1 57 F 51

27 F 8 6.1 82 F 29

15 F 2 5.6 73 F 12

17 F 3 2.3 89 F 16

Roots 96 h Concentration (Ag g 1) Difference to control % of control

91 F 45 32.9 73 F 27

65 F 17 59.0 52 F 50

187 F 20 91.7 196 F 73

292 F 213 196.6 306 F 11

84 F 28 105.3 44 F 34#

167 F 9 3.4 102 F 5

Different times of sampling are also indicated. Values are mean F standard error of the mean, except for the ‘difference to control’ which is the difference between the mean for that treatment and the mean of the site control. * Difference between treatment and control > 200%. # Difference between treatment and control < 50%.

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Table 7 Sediment physicochemical parameters at the three experimental sites (mean F standard error, n = 5 in all cases) Parameter

Pittwater

Sydney Harbour

Botany Bay

F-value

LOI (%) pH Conductivity (mS cm Sand (%) Silt/clay (%) CaCO3 (%)

1.22 F 0.14a 7.43 F 0.08a 7.82 F 1.11a 90.9 F 1.3a 9.1 F 1.3a 0.77 F 0.41a

3.19 F 0.22b 8.09 F 0.00b 11.71 F 0.28b 96.3 F 0.4b 3.7 F 0.4b 6.91 F 0.52b

1.36 F 0.06a 8.15 F 0.03b 8.99 F 0.41b 97.2 F 0.2b 2.8 F 0.2b 0.25 F 0.11a

112.7* 163.3* 20.1* 19.6* 19.6* 31.8*

1

)

Different values (determined with Tukey’s HSD post-hoc pair wise comparison) are marked with different letters. * All parameters had a statistically significant difference at p < 0.000.

The zinc concentrations were higher and lower than the controls in different situations (Table 6). At Pittwater, the concentration of Zn2 + in the leaves was 425% of the control concentration in the leaves at 10 h after exposure to 1 mg l 1 zinc and at 96 h, the 0.1 mg l 1 treatment caused an increase in zinc concentration to 223%, while the 1 mg l 1 samples had returned to normal levels. The samples from Sydney Harbour and Botany Bay had lowered zinc concentrations in the sediment at 96 h for the 0.1 and 1 mg l 1 treatments, respectively. Zinc concentration was also low in the roots at 96 h for the 0.1 mg l 1 treatment in Botany Bay, while it was high in the roots for the 1 mg l 1 treatment in Sydney Harbour. In terms of concentrations elevated above background levels, zinc was most accumulated in the leaves of the samples from Pittwater at 10 h, after exposure to 1 mg l 1 zinc (Table 6). Concentrations on the outside of leaves were slightly higher at Sydney Harbour and Botany Bay than Pittwater though, similarly, a substantial amount of zinc above background levels was detected in the roots of the Sydney Harbour samples. 3.4. Sediment physicochemical parameters Physicochemical parameters were slightly different at the three sites (Table 7). The highest loss on ignition, conductivity and CaCO3 contents were at the Sydney Harbour site. Silt/clay content was highest at the Pittwater site and that site also had the lowest pH. The Pittwater and Sydney Harbour site had significantly different values for all the parameters.

4. Discussion 4.1. Differences between background concentrations at the three sites The highest concentrations of background metals were at the Sydney Harbour site, then Pittwater and finally Botany Bay. A strong relationship between metal concentrations in tissue and availability in the sediment and water column was found across 40 sites (Lyngby and Brix, 1984). Furthermore, there is a relationship between the distance from the source of pollution and it’s concentration in the seagrass tissue (Ward, 1987; Haynes

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and Johnson, 2000). In the present study, the Sydney Harbour site is adjacent to a main arterial road and a large marina, both of which are known sources of metal contamination (Birch et al., 1996; Birch and Taylor, 1999). Metal sources at Pittwater are boating activity and stormwater (Birch and Taylor, 1999; Prange and Dennison, 2000). Botany Bay and Cooks River have been used extensively for industrial purposes. However, the sample site at Kurnell is on the other side of the bay, leading to lower sediment metal concentrations (Birch, 1996; Birch et al., 1996). In terms of distance from pollution sources, the Botany Bay site is the furthest away from intense pollution, then Pittwater and finally The Spit, following the order of baseline Cu2 + concentrations from lowest to highest. Zinc concentrations were much higher than copper concentrations at all three sites, in accordance with the findings of Birch (1996), Birch et al. (1996), Hayes et al. (1998) and Birch and Taylor (1999). The zinc concentrations were much more variable and there were no clear patterns in concentrations between the sites. Many physicochemical parameters of sediment and water influence the accumulation and toxicity of copper (Allen and Hansen, 1996), including sediment type (Ward et al., 1986) and particle size which is generally inversely related to metal concentration in the sediment (Waldichuk, 1985). Thus, high concentrations of silt at Pittwater may explain the reasonably high background concentrations of metals. Surprisingly, concentrations of copper in seagrass tissue from Pittwater and Sydney Harbour were higher than those detected in the same species in Lake Macquarie by Batley (1987), despite the degraded state of sediments within that lake. Other factors such as low pH at Pittwater would lead to slightly higher concentrations of the oxidised states of copper, which are the oxide and hydroxide forms (Waldichuk, 1985; Simpson et al., 2002). These forms are usually much more soluble than the reduced sulfide state (Waldichuk, 1985) and thus more prevalent in the water. Assessing the concentrations of bioavailable metal (Waldichuk, 1985; Allen and Hansen, 1996) would provide a clearer picture of copper fate. 4.2. Photosynthetic impacts of metals at the three sites The lack of response due to zinc exposure was not unexpected since zinc is less toxic to Z. capricorni than copper (Macinnis-Ng and Ralph, 2002). The greater impact of copper on samples at Pittwater was expected since the Pittwater samples may be more naı¨ve than those at Botany Bay and Pittwater (Handy, 1994). Heavy metals act on the photosynthetic apparatus by affecting CO2 fixation at several levels (Clijsters and Van Assche, 1985). Within photosystem II (PSII), photophosporylation or enzyme activity can be inhibited as can the oxidative site of PSII. This causes a decline in DF/FmV, particularly after copper exposure (Ralph and Burchett, 1998; Prange and Dennison, 2000; Macinnis-Ng and Ralph, 2002). This occurred in the Pittwater samples and the Sydney Harbour samples during exposure but not at Botany Bay. Chlorophyll content can be reduced by copper, because it interferes with pigment biosynthesis and causes chlorophyll degradation (Clijsters and Van Assche, 1985; Prasad and Strzalka, 1999; Macinnis-Ng and Ralph, 2002), as seen in the Pittwater samples. The interaction of pollutants in the environment, pollutant concentrations within an organism and the impacts of that pollutant on the organism have been rarely studied

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(Hopkin, 1993), particularly in seagrasses. Some relationships between metal levels in tissues and toxicological impacts have been demonstrated (Pergent-Martini and Pergent, 2000), yet the concentrations of copper in Z. capricorni found in this study do not explain the differences in toxicity demonstrated by chlorophyll a fluorescence and chlorophyll pigment determination. This is particularly evident in a comparison of the concentrations of copper (above baseline levels) in the leaves from Pittwater and Botany Bay. After 10 h exposure to 1 mg l 1 copper, there was about 80 Ag g 1 copper in the leaves of Botany Bay samples and about 60 Ag g 1 copper in the Pittwater samples (Table 5). Yet the Pittwater samples showed a physiological impact while the Botany Bay samples did not. The Sydney Harbour samples also did not show an impact but there was only about 30 Ag g 1 copper above baseline levels in these leaves. Exclusion of copper was not apparent since the levels of copper absorbed to the outside of leaves was reasonably similar in the three populations. The evolution of metal tolerant ecotypes within plant species is well known (Foy et al., 1978). There are seven different mechanisms of metal tolerance in higher plants according to Macnair and Baker (1994). Metals can be bound to the cell wall, the cell membrane can be metal tolerant, membrane transport can be reduced, metals may be removed from the cell through active efflux, metal tolerant enzymes can alleviate metal impacts, metals can be trapped in vacuoles through compartmentalisation and processes of chelation can bind metals either inside or outside the cells. In the current case, it is unclear which of these mechanisms (or combination of mechanisms) has contributed to the physiological differences between copper sensitivity in the three populations but some possibilities are considered below. Generation of active oxygen species within cellular tissues is triggered by metal toxicity and, to avoid accumulation of these toxic intermediates, plant tissues have a series of detoxifying agents which involve enzymatic and non-enzymatic mechanisms (Ciscato et al., 1997). Enzyme activity encourages biotransformation of toxins (Pergent-Martini and Pergent, 2000). Peroxidase activity has been found to increase with the onset of metal toxicity as a detoxifying mechanism (MacFarlane and Burchett, 2001). This detoxification process may be more efficient in the Botany Bay samples, leading to less toxic impact. Furthermore, metal-binding polypeptides, such as phytochelatins (Verkleij and Schat, 1990; Turner, 1994), may aid tolerance in the Botany Bay population. Restriction of transport or compartmentalisation of metals in root tissue (Baker and Walker, 1989) was not apparent here since the concentrations of copper in the roots were also similar across sites. Yet compartmentalisation of metals in vacuoles within the tissue or other chelation would not have been detected in the current study. To further understand mechanisms of increased tolerance in the Botany Bay samples, we suggest processes of compartmentalisation, formation of oxidative stress enzymes and the role of phytochelatins are investigated. 4.3. Influence of background concentrations on sensitivity Corals exposed to run-off enriched with iron were more tolerant to iron exposure than naı¨ve corals (Harland and Brown, 1989). So we expected that tolerant seagrass populations would have high background concentrations of metals, yet this was not the case. The

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samples from Pittwater were significantly more sensitive to copper exposure than the samples in Sydney Harbour or Botany Bay, counter-intuitively to the background concentrations. The lowest levels of these two metals were found at Botany Bay, yet these samples (along with the Sydney Harbour samples) were the least impacted when exposed to copper. Macnair and Baker (1994) reported that tolerance may not be metalspecific. Tolerance to one metal at a rehabilitated mine site may induce tolerance to other metals. This could be investigated by analysing other metal concentrations at Botany Bay, Sydney Harbour and Pittwater. Further experimentation is required to determine the influence of physicochemical parameters and genetic studies of the populations would certainly help determine the influence of genetic variation on tolerance. 4.4. Other reasons for differences in sensitivity to copper Since accumulation of metals is surface area dependant (Ward, 1989), the surface area of leaf tissue in the chambers should have been standardised. The leaves at Pittwater are shorter than those at Botany Bay and Sydney Harbour, while the densities of shoots were quite similar, leading to a greater surface area of leaves in Botany Bay and Sydney Harbour. This may not have had an impact in Botany Bay however, as the concentration of copper in the leaves above background levels after 10 h was similar to that at Pittwater. Similarly, epiphytic growth may also influence uptake of copper through the leaves where prolific epiphytes may absorb the copper themselves or even reduce the surface area of seagrass leaf able to absorb the metal. Either way, more epiphytes would lead to less uptake of copper. While this was not quantified, visual surveys suggested that there was the greatest amount of epiphyte biomass at the Sydney Harbour site, then at Botany Bay and finally at Pittwater. This may account for the reduced concentrations of copper in the leaves at Sydney Harbour.

5. Conclusion This study has demonstrated the variability in sensitivity of different populations of Z. capricorni to copper exposure. By testing three populations in situ, the importance of considering the natural environment has been addressed to some degree. Jepson and Sherratt (1996) suggested that assessment of ecotoxicological risks can be highly site specific. This study demonstrates that resource managers need information specific to their jurisdiction for effective decision making, regardless of the chemical mechanisms involved. This study also shows that simply assessing the metal content of seagrasses (such as studies by Lyngby and Brix, 1982, 1984; Brix et al., 1983; Ward, 1987, 1989; PergentMartini and Pergent, 2000) will not give a good indication of metal impacts, since some populations could be more tolerant than others. This has major implications for monitoring of heavy metal impacts on seagrasses in situ since determining concentrations of metals in seagrass tissue has been the traditional technique. Clearly, techniques based on a physiological assessment are preferable for assessing the risks posed by metal exposure. These findings also have implications for rehabilitation work since a sensitive population will not be appropriate to transplant to a contaminated site.

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Acknowledgements N. Ralph constructed the chambers. Field assistance was provided by D. Macinnis and R Luff. Macinnis-Ng was supported by an APA scholarship while carrying out this research. The project was conducted under NSW Fisheries Scientific Research Permit no. F99/363. [SS]

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