Abstract

Using the γ‐glutamylcysteine synthetase inhibitor, l‐buthionine‐[S,R]‐sulphoximine (BSO), the role for phytochelatins (PCs) was evaluated in Cu, Cd, Zn, As, Ni, and Co tolerance in non‐metallicolous and metallicolous, hypertolerant populations of Silene vulgaris (Moench) Garcke, Thlaspi caerulescens J.&C. Presl., Holcus lanatus L., and Agrostis castellana Boiss. et Reuter. Based on plant‐internal PC‐thiol to metal molar ratios, the metals’ tendency to induce PC accumulation decreased in the order As/Cd/Cu > Zn > Ni/Co, and was consistently higher in non‐metallicolous plants than in hypertolerant ones, except for the case of As. The sensitivities to Cu, Zn, Ni, and Co were consistently unaffected by BSO treatment, both in non‐metallicolous and hypertolerant plants, suggesting that PC‐based sequestration is not essential for constitutive tolerance or hypertolerance to these metals. Cd sensitivity was considerably increased by BSO, though exclusively in plants lacking Cd hypertolerance, suggesting that adaptive cadmium hypertolerance is not dependent on PC‐mediated sequestration. BSO dramatically increased As sensitivity, both in non‐adapted and As‐hypertolerant plants, showing that PC‐based sequestration is essential for both normal constitutive tolerance and adaptive hypertolerance to this metalloid. The primary function of PC synthase in plants and algae remains elusive.

Received 17 May 2002; Accepted 2 August 2002

Introduction

Phytochelatins (PCs) are small metal‐binding peptides with the structure (γ‐glu‐cys)n‐gly, (γ‐glu‐cys)n‐β‐ala, (γ‐glu‐cys)n‐ser, (γ‐glu‐cys)n‐glu, (γ‐glu‐cys)n‐gln or (γ‐glu‐cys)n, in which n varies from 2 to 11 (Grill et al., 1985, 1986a; Mehra and Winge, 1988; Meuwly et al., 1993; Klapheck et al., 1994). Their synthesis from glutathione (Grill et al., 1989), homo‐glutathione, hydroxymethyl‐glutathione (Klapheck et al., 1995) or γ‐glutamylcysteine (Hayashi et al., 1991) is catalysed by a transpeptidase, named phytochelatin synthase, which is a constitutive enzyme requiring post‐translational activition by heavy metals (Grill et al., 1989; De Knecht et al., 1995; Klapheck et al., 1995; Chen et al., 1997). Phytochelatin synthase (PCS) has been shown to be activated by a broad range of metals and metalloids, in particular Cd, Ag, Pb, Cu, Hg, Zn, Sn, Au, and As, both in vivo and in vitro (Grill et al., 1987; Maitani et al., 1996; Chen et al., 1997). The capacity to synthesize PCs is supposed to be present in all higher plants (Gekeler et al., 1989) and the majority of algae (Ahner et al., 1995). They also have been detected in several fungi, including Schizosaccharomyces pombe, Candida glabrata, and Mucor racemosus (Grill et al., 1986b; Mehra et al., 1988; Miersch et al., 2001). In addition, the nematode wurm, Caenorhabdites elegans, appeared to possess a PCS gene, which restored PC synthesis and Cd tolerance in an S. pombe PCS knock‐out strain, suggesting that functional PCS genes may be present in certain animals too (Clemens et al., 2001; Vatamaniuk et al., 2001).

There has been considerable debate concerning the function of PCs. They have been assumed to function in the cellular homeostasis or trafficking of essential heavy metal nutrients, particularly Cu and Zn (Thumann et al., 1991). However, the Cu and Zn exposure levels that are minimally required to induce PCs at considerable concentrations in plant cells are often far above the normal nutritional requirements, or even close to the toxicity thresholds (Schat et al., 2000). Moreover, PC‐deficient mutants, such as Arabidopsiscad1 (Howden et al., 1995), have never been reported to exhibit increased requirements for essential metal nutrients, and did not show considerably increased sensitivities to Cu or Zn (Howden and Cobbett, 1992). Conforming with these observations, ecotypic differences in Cu tolerance in Arabidopsis thaliana were shown to be correlated with type‐2 metallothionein expression, rather than PC accumulation rates (Murphy and Taiz, 1995). However, S. pombe cells disrupted in their PCS gene were hypersensitive to Cu (Clemens et al., 1999), indicating that PC synthesis may be required for Cu detoxification in some organisms, at least.

On the other hand, there is convincing evidence that PCs are essential for normal constitutive tolerance to several non‐essential metals, particularly Cd. First, disruption of the PCS gene in S. pombe resulted in hypersensitivity to Cd (Clemens et al., 1999; Ha et al., 1999). Second, expression of PCS cDNAs from wheat, Arabidopsis, and S. pombe dramatically increased Cd tolerance in Saccharomyces cerevisiae, even in mutants deficient in vacuole formation or vacuolar acidification (Clemens et al., 1999). Third, a number of Cd‐hypersensitive Arabidopsis mutants appeared to be impaired in PC synthesis (Howden et al., 1995; Cobbett et al., 1998). In addition, tomato cell lines selected for hypertolerance to Cd exhibited enhanced PC synthesis under Cd exposure, due to increased γ‐glutamyl cysteine synthetase (γ‐ECS) activity (Chen and Goldsbrough, 1994). Furthermore, overexpression of bacterial γ‐ECS or glutathione synthetase (GS) in Brassica juncea enhanced PC synthesis and Cd tolerance (Zhu et al., 1999a, b). PCs might also be required for tolerance to non‐essential Hg and As. PC‐deficient Arabidopsiscad1 mutants were also hypersensitive to Hg (Howden and Cobbett, 1992). Inhibition of PC synthesis by treatment with the γ‐ECS inhibitor, buthionine sulphoximine (BSO), enhanced Hg sensitivity in Hydrilla verticillata and Vallisneria spiralis (Gupta et al., 1998). Likewise, BSO‐treated cell cultures of tobacco and Rauvolfia serpentina were found to be hypersensitive to As (Nakazawa et al., 2000; Schmöger et al., 2000).

Normal constitutive tolerance to Cd and, possibly, Hg and As, is apparently not entirely explained by the mere chelation of Cd by PCs in the cytosol. Several Cd‐hypersensitive S. pombe mutants showed normal PC synthesis under Cd exposure, but appeared to be deficient in functional HMT1, an ABC‐type transporter mediating the transport of Cd–PC complexes into vacuoles (Ortiz et al., 1992, 1995), or impaired in a further stabilization of vacuolar Cd‐PC complexes through the incorporation of acid‐labile sulphide (Speiser et al., 1992b; Juang et al., 1993). MgATP‐dependent tonoplast transport and vacuolar accumulation of Cd–PC complexes have also been demonstrated in oat and tobacco, respectively (Vögeli‐Lange and Wagner, 1989; Salt and Rauser, 1995). Also, acid‐labile sulphide incorporation in Cd–PC complexes has been demonstrated in B. juncea and Silene vulgaris (Speiser et al., 1992a; De Knecht et al., 1994), suggesting that vacuolar compartmentalization and further stabilization of Cd–PC may be essential for normal Cd tolerance in plants too.

Naturally selected heavy metal hypertolerance, which is commonly found in plant populations from strongly metal‐enriched soils, does not seem to be associated with enhanced PC synthesis. De Knecht et al. (1995) obtained equal capacities and activation constants for Cd‐induced PC synthesis in crude protein extracts prepared from roots of Cd/Zn‐hypertolerant and non‐metallicolous S. vulgaris. The root PC concentrations measured in vivo, however, were much lower in the hypertolerant plants, even when compared at equal rates of Cd uptake (De Knecht et al., 1994). The in vivo acid‐labile sulphide contents of the Cd–PC complexes and PC chain length distributions were identical, suggesting that possible differences in the stabilities of the complexes formed in both plant types were absent (De Knecht et al., 1994). Also, the reduced glutathione (GSH) concentrations in the roots responded similarly to Cd exposure, and the rates of PC degradation and GSH recovery after arresting the exposure were identical, suggesting that the lower PC accumulation in the hypertolerant plants resulted neither from a lower GSH availability, nor from a higher PC turnover rate (De Knecht et al., 1995). Finally, the fraction of exclusively acid‐extractable root Cd was consistently higher in the hypertolerant plants (De Knecht et al., 1994). In the same species, Cu‐induced and Zn‐induced accumulation of PCs in roots also appeared to be much higher in non‐metallicolous plants than in Cu‐hypertolerant and Zn‐hypertolerant plants respectively, both when compared at equal metal exposure levels and at equal rates of metal accumulation in the roots (De Vos et al., 1992; Harmens et al., 1993). Moreover, decreased PC accumulation was shown to co‐segregate with Cu hypertolerance in crosses between non‐metallicolous and Cu‐hypertolerant plants (Schat and Kalff, 1992). Thus, although artificial overexpression of enzymes and transporters involved in the PC‐based metal sequestration machinery, such as γ‐ECS, GS, PCS or HMT1, has been shown to increase metal tolerance or Cd tolerance, at least (Ortiz et al., 1995; Clemens et al., 1999; Zhu et al., 1999a, b), and although several examples of enhanced PC synthesis in cell lines artificially selected for Cd hypertolerance have been reported (Chen and Goldsbrough, 1994), there is no evidence of naturally selected enhanced PC synthesis in hypertolerant plant populations from Cd‐, Zn‐, or Cu‐toxic environments. Moreover, BSO did not detectably enhance the response to a 40 µM Cd treatment in a hypertolerant ecotype of S. vulgaris, although it dramatically sensitized a non‐metallicolous ecotype under identical conditions, suggesting that PC synthesis might not be required for naturally selected Cd hypertolerance in this species (De Knecht et al., 1992). In general, the strongly decreased rates of PC accumulation in Cu‐, Cd‐, and Zn‐hypertolerant S. vulgaris mine populations (see above) might, in fact, result from increased activities of alternative PC‐independent sequestration mechanisms leading to decreased cytoplasmic metal availabilities for PCS activation (De Knecht et al., 1995). On the other hand, naturally selected As hypertolerance in Holcus lanatus was found to be associated with enhanced rates of PC accumulation and increased PC‐thiol to As molar ratios in roots, suggesting that PC synthesis might be essential for hypertolerance to As, at least (Hartley‐Whitaker et al., 2001).

The present evidence with regard to the precise role for PCs in constitutive and naturally selected high‐level metal tolerances is often fragmentory and ambiguous. Much of the evidence is based on single‐concentration exposures. In this study, the complete dose–response curves were compared for root growth inhibition and PC accumulation imposed by Cu, Cd, Zn, Ni, Co, and As in metallicolous and non‐metallicolous ecotypes of the pseudometallophytes Silene vulgaris, Holcus lanatus, Agrostis castellana, and the Zn hyperaccumulator, Thlaspi caerulescens. To assess the possible role for PCs in metal tolerance, the exposures were done with and without BSO in the nutrient solution.

Materials and methods

Plant materials

S. vulgaris seeds were collected from a copper mine near Marsberg (Germany), a zinc smelter waste deposit at Plombières (Belgium), and a non‐metalliferous site at the Free University Campus (Amsterdam, The Netherlands). The population from Marsberg is Cu‐hypertolerant, and shows low degrees of hypertolerance to Zn and Cd. The population from Plombières is hypertolerant to Zn and Cd, and shows pleiotropic hypertolerance to Ni and Co (Schat and Vooijs, 1997). More detailed site and population characteristics have been given in Schat et al. (1996). Seeds of T. caerulescens were collected from a Zn ore waste deposit near La Calamine (Belgium), a non‐metalliferous site at Willerwiltz (Luxemburg), and from a serpentine hill (Monte Prinzera, Italy). From previous studies it appeared that the Zn tolerance of these populations, as estimated from threshold exposure levels for leaf chlorosis, varied strongly, decreasing in the order La Calamine (LC) > Monte Prinzera (MP) > Willerwiltz (W). Likewise, Ni tolerance decreased in the order MP >> LC > W, and Cd tolerance in the order LC >> W > MP. Zn, Ni, and Cd accumulation were much higher in MP and W than in LC (Assunção et al., 2001; AGL Assunção and H Schat, unpublished results). Seeds of A. castellana and H. lanatus were collected at an As‐enriched gold mine waste dump near Jales (Portugal). Non‐metallicolous H. lanatus was collected from the botanical garden of the Free University Campus (Amsterdam, The Netherlands).

Tolerance testing

Seeds were germinated on moist peat and 8‐d‐old seedlings were transferred to 1.0 l polyethylene pots (three plants per pot) with aerated MES‐buffered nutrient solution composed as in Schat et al. (1996), or, in case of the grasses, with half‐strength macronutrient concentrations. After 5 d of hydroponic culture, the solution was replaced by a fresh one of the same composition. Half of the pots were supplied with l‐BSO at a 250 µM concentration (higher concentrations did not produce a further decrease of root GSH levels, as demonstrated in pilot experiments). After another 5 d period, the solutions were replaced again, and the metals were added at appropriate concentrations. All the metals were added as sulphate salts, except As, which was supplied as sodium arsenate. The test solutions were the same as during preculture, except for the Cu‐spiked solutions, from which Fe‐EDTA was omitted to prevent Cu‐EDTA formation. Also, in the As‐spiked solutions, the NH4H2PO4 concentration was reduced to a 10 µM level, to prevent excessive competitive inhibition of arsenate uptake. The BSO treatment was maintained during metal exposure, except in the case of Cu, because Cu, by contrast with the other metals, appeared to be complexed by BSO in the nutrient solution, as shown by measurements with ion‐specific electrodes (H Schat, unpublished results). The root elongation was measured after 4 d of exposure, using the charcoal staining method (Schat and Ten Bookum, 1992) in the case of S. vulgaris and T. caerulescens. Part of the plants were left unstained and used for PC and metal analysis. In the case of the grasses, the roots were cut off at the start of the metal treatment, and the length of the longest new root was measured after 4 d.

Growth chamber conditions were exactly as in Schat et al. (1996).

PC and metal analysis

Prior to harvest, the root systems were desorbed in an ice‐cold 5 mM Pb(NO3)2 solution for 30 min. Roots and shoots were separated, immediately frozen in liquid nitrogen, lyophilized, and stored under vacuum until analysis. Twenty to 100 mg aliquots of ground dry material were digested in 2 ml of a 1:4 (v/v) mixture of 37% (v/v) HCl and 65% (v/v) HNO3, in closed Teflon cylinders for 6 h at 140 °C. Metals in the digests were measured using a flame atomic absorption spectrophotometer (Perkin Elmer 2100), in the case of As with a coupled MHS‐10 hydride system. Phytochelatins were extracted and measured by HPLC, using post‐column derivatization with Ellman’s reagent, exactly as described in De Knecht et al. (1994), except for samples of As‐treated plants. The latter were analysed after pre‐column derivatization with monobromobimane, exactly as in Sneller et al. (1999). Corrections for differential derivatization efficiences were made according to Sneller et al. (2000).

Statistics

The data were statistically analysed using two‐way ANOVA after log‐transformation of the data. Significance of the BSO×metal concentration interaction was used as a criterion for BSO‐imposed hypersensitivity.

Results

Copper

PC‐mediated Cu tolerance was tested in non‐metallicolous and cupricolous S. vulgaris ecotypes (populations Amsterdam and Marsberg, respectively). Copper induced the accumulation of PCs in roots of both ecotypes. However, the threshold Cu exposure level required to induce significant PC accumulation was much higher in the cupricolous ecotype than in the non‐metallicolous ecotype. Moreover, the cupricolous ecotype exhibited lower root‐internal PC‐thiol (PC‐SH) to Cu molar ratios (Table 1). Shoot PC concentrations were negligible in both ecotypes (data not shown). In the roots, however, PCs were strongly decreased by BSO in both ecotypes (Table 1). Usually, the root copper concentrations were also decreased, though to a much lower degree, resulting in lower PC‐SH to Cu molar ratios in the BSO‐treated plants (Table 1). The root growth response to Cu, however, was completely unaffected by BSO, both in the non‐metallicolous and the cupricolous ecotype (Fig. 1).

Cadmium

PC‐mediated Cd tolerance was established in non‐metallicolous and Cd/Zn‐hypertolerant Silene vulgaris (populations Amsterdam and Plombières, respectively). Cd, like Cu, strongly induced the accumulation of PCs in roots (Table 2), but barely or not in shoots (data not shown). Again, the PC concentrations and the root‐internal PC‐SH to Cd ratios were higher in the non‐metallicolous ecotype than they were in the hypertolerant calamine ecotype (Table 2). BSO strongly decreased root GSH and PC concentrations and, though to a much lower degree, root Cd concentrations, resulting in considerably decreased PC‐SH to Cd molar ratios (Table 2). BSO significantly enhanced Cd‐imposed root growth inhibition in the non‐metallicolous ecotype (P <0.001), but did not detectably effect the root growth response of the hypertolerant ecotype (Fig. 2).

Using a similar experimental design, three T. caerulescens ecotypes with varying degrees of Cd tolerance and accumulation, originating from serpentine, calamine, and non‐metalliferous soil (populations Monte Prinzera, La Calamine, and Willerwiltz, respectively) were compared. Cd induced considerable PC accumulation in the roots and shoots of all these ecotypes, though to different degrees. When compared at similar root‐internal Cd concentrations, the root PC concentrations decreased in the order Monte Prinzera > Willerwiltz > La Calamine (Fig. 3), which is also the order of decreasing sensitivity to Cd (see Materials and methods). The same pattern was also found in shoots (data not shown). Cd‐imposed root growth inhibition was not enhanced by BSO in either of the ecotypes (Fig. 4). At the end of the experiment, i.e. 4 d after Cd supply, there were no visible effects of the treatments on shoot performance. However, the major sink for Cd in T. caerulescens is the shoot, and previous experiments (Assunção et al., 2001) clearly showed that after longer periods of exposure (>1 week) shoot performance responded much more sensitively to metal exposure than did root growth. Therefore, an additional longer‐term experiment was performed with the serpentine and the calamine ecotype. After 2 weeks of exposure, the sensitizing effect of BSO was clearly apparent from the shoot performance of the serpentine ecotype. At 1 µM external Cd, the shoots of the BSO‐treated plants were almost entirely bright yellow or white with necrotic parts, whereas those of the untreated plants were still more or less green. In addition, as indicated by the significance of the BSO×Cd interaction (P <0.01), BSO increased the shoot fresh weight response to the Cd treatment (Fig. 5). In the Cd‐hypertolerant calamine ecotype, shoot fresh weight was only significantly decreased at the highest treatment level, i.e. 125 µM Cd, and the BSO×Cd interaction was not significant. Chlorosis was only apparent at the 25 µM and the 125 µM treatments, both with and without BSO. In both ecotypes, the root elongation response was unaffected by BSO (data not shown).

Arsenic

BSO‐imposed effects on As tolerance were investigated in non‐metallicolous S. vulgaris (Amsterdam) and H. lanatus (Amsterdam), as well as in H. lanatus and A. castellana from a strongly As‐enriched gold mine waste deposit (Jales) (see Materials and methods). Arsenate caused a strong accumulation of PCs in S. vulgaris, albeit exclusively in roots, the root PC‐SH to As molar ratios being above 3, except for the most toxic exposure levels (Table 3). The grasses also showed high PC accumulation rates, but the results were considered to be unreliable, owing to a variable recovery of the internal standard added to the samples, N‐acetyl‐l‐cysteine (data not given). However, the PC‐SH to As molar ratios obtained, though probably underestimated, were between 1.9 and 2.7, and did not vary with the degrees of As tolerance. BSO treatment dramatically decreased the root PC concentrations (Table 3), particularly in the grasses, where the root PC concentrations consistently remained below 0.1 µmol g–1 DW in the BSO treatment. In all the species tested, BSO strongly increased the root growth response to As, both in tolerant and non‐tolerant ecotypes. In the case of the grasses, BSO completely arrested root growth, even at As exposure levels that did not cause root growth inhibition in the absence of BSO (Figs 6, 7). In fact, the root sytems of most of the plants seemed to have died off completely, and their leaves were wilted or largely necrotic, except for the younger ones, which showed some chlorosis, but remained turgid.

Zinc, Nickel, Cobalt

PC accumulation induced by Zn, Ni, or Co and possible PC‐dependent tolerance to these metals were investigated in non‐metallicolous and Zn‐hypertolerant ecotypes of S. vulgaris (populations Amsterdam and Plombières, respectively) and T. caerulescens (populations Willerwiltz and La Calamine, respectively), as well as in a serpentine T. caerulescens ecotype (population Monte Prinzera) (see Materials and methods). Zn exposure caused a pronounced concentration‐dependent accumulation of PCs in roots of non‐metallicolous S. vulgaris, although the PS‐SH to Zn molar ratios were far below unity (Table 4). The calamine ecotype exhibited lower root PC concentrations, in spite of a much higher root Zn accumulation rate, leading to dramatically decreased PC‐SH to Zn molar ratios, as compared with the non‐metallicolous ecotype (Table 4). In both ecotypes, the root growth response was unaffected by BSO (Fig. 8), although the BSO treatment effectively decreased the PC‐SH to Zn molar ratios in both ecotypes (Table 4). PC accumulation was not apparent in shoots.

Zn‐induced PC accumulation was also found in each of the T. caerulescens ecotypes, particularly in roots, but also in shoots. The PC concentrations were inconsiderable, however (less than 2 µmol g–1 dry weight), albeit much higher than in control plants (about 0.04 µmol g–1 dry weight), and did not increase with exposure levels within the range tested (25–1250 µM Zn). There were no obvious interecotypic differences in PC‐SH to Zn molar ratio’s. BSO did not detectably affect the growth response to Zn or the PC‐SH to Zn molar ratios in either of the ecotypes (data not shown).

Ni and Co induced PC accumulation in both ecotypes of S. vulgaris, albeit exclusively in roots. The root PC concentrations, though being higher than in control plants (about 0.06 µmol g–1 dry weight), never exceeded 1 µmol g–1 dry weight, and tended to decrease with exposure level (5–80 µM Ni; 3–243 µM Co) in both ecotypes. The Zn‐hypertolerant ecotype, which exhibits some pleiotropic hypertolerance to Ni and Co (Schat and Vooijs, 1997), consistently displayed much lower PC concentrations and lower PC‐SH to metal molar ratios than did the non‐metallicolous ecotype. BSO neither affected the PC‐SH to metal molar ratios, nor the root growth response in either ecotype (data not shown).

PC accumulation under Ni exposure was also found in all of the three T. caerulescens ecotypes tested. The root PC concentrations were usually lower than 1 µmol g–1 dry weight in the non‐metallicolous and the calamine ecotypes, and between 1 and 2 µmol g–1 in the serpentine ecotype, more or less irrespective of the level of exposure (15–450 µM Ni). The PS‐SH to Ni molar ratios, however, were lower in the serpentine ecotype than in the other ones. Ni induced some PC accumulation in the shoot, up to 0.15 µmol g–1 dry weight, albeit exclusively in the serpentine ecotype. BSO did not affect the growth response to Ni in either of the ecotypes (data not shown). Cobalt‐induced PC accumulation was only investigated in the calamine ecotype. The root PC concentrations were below 1 µmol g–1 and decreased with increasing exposure level (3–243 µM Co). The shoot PC concentrations were below the detection limit. BSO did not affect the growth response to this metal (data not shown).

Discussion

The accumulation of PCs under copper stress has been demonstrated in a large number of algae and plants. However, more or less precise quantitative information with regard to the dose–effect relationships of this phenomenon is scarce. The threshold exposure levels for PC accumulation appeared, in general, to be lower for Cd than for Cu (Ahner and Morel, 1995; Rijstenbil and Wijnholds, 1996). In several studies Cu‐induced PC accumulation was not apparent until the threshold exposure level for acute toxicity had been exceeded, suggesting that PCs are normally not involved in Cu sequestration under conditions of subtoxic exposure (De Vos et al., 1992; Rijstenbil et al., 1998; Rijstenbil and Gerringa, 2002). In the present study, the threshold Cu exposure levels for root growth inhibition and PC accumulation seemed to coincide, both in non‐tolerant and Cu‐tolerant S. vulgaris (Fig. 1, Table 1), suggesting that Cu did not induce PC accumulation until the capacity of the normal cellular homeostasis had been exhausted. The increased homeostatic capacity in the tolerant ecotype relies most probably on a combination of an enhanced capacity to efflux Cu from the root cells and a strong constitutive overexpression of a 2b‐type metallothionein, SvMT2b (Van Hoof et al., 2001a, b). Also, the latter properties would be expected to reduce the activity of cytosolic Cu available for PCS activation, which would then explain the lower PC‐SH to Cu molar ratios in the tolerant ecotype (Table 1). The absence of considerable PC accumulation under subtoxic exposure obviously explains the lack of effect of BSO on the threshold Cu exposure levels for root growth inhibition (Fig. 1). The absence of a clear effect of BSO on the slope of the dose–response curves is more difficult to explain. Cu–PC complexes seem to be highly stable (Mehra and Mulchandani, 1995; Leopold and Gunther, 1997), and intact Cu–PC complexes have been isolated from S. vulgaris roots (Verkleij et al., 1989), suggesting that PCs should contribute, at least to some extent, to Cu detoxification in Cu‐stressed plants. One might argue that the BSO‐imposed inhibition of PC synthesis was not complete, possibly through shoot‐to‐root transport of GSH (De Knecht et al., 1995). However, the PC‐SH to Cu molar ratios in the treated plants were much decreased (Table 1), showing that the root performance under Cu‐toxic conditions was not limited by the PC synthetic capacity. Moreover, the PCS‐deficient cad1 mutant of Arabidopsis thaliana did not exhibit considerably increased sensitivity to Cu (Howden and Cobbett, 1992), although Cu has been shown to induce the accumulation of PCs in this species (Murphy and Taiz, 1995). Thus, most of the evidence available thus far suggests that PCs may not effectively contribute to Cu detoxification in most algae and higher plants, although they appear to do so in fission yeast (Clemens et al., 1999). The reason for this might lie in the presence or absence of more effective efflux‐ or MT‐based alternative detoxification systems.

Cd has been shown to be a strong inducer of PC accumulation in a broad variety of algae and higher plants, as well as in several fungi, and PC‐based Cd sequestration is generally considered to be essential for normal Cd tolerance in organisms with functional PCS genes (Howden and Cobbett, 1992; Ortiz et al., 1992; Speiser et al., 1992b; Cobbett et al., 1998; Clemens et al., 1999; Vatamaniuk et al., 2001). In agreement with this viewpoint, a strong Cd‐induced PC accumulation and BSO‐imposed hypersensitivity to Cd in non‐metallicolous S. vulgaris was observed (Fig. 2). In the Cd‐hypertolerant ecotype, however, the PC‐SH to Cd molar ratios were much lower (Table 2), and BSO‐imposed hypersensitivity to Cd was not apparent, irrespective of the level of exposure (Fig. 2), suggesting that Cd hypertolerance is achieved through enhanced activity of an as yet unknown PC‐independent Cd sequestration mechanism decreasing the activity of cytoplasmic Cd available for PCS activation (De Knecht et al., 1992, 1995). The results obtained with T. caerulescens are basically in line with this. The degrees of Cd‐imposed PC accumulation (Fig. 3) in the different ecotypes were inversely related to the levels of Cd tolerance. Apparently, the Cd‐hypertolerant ecotype did not possess PC‐dependent Cd tolerance, as shown by the absence of any BSO‐imposed hypersensitivity (Fig. 4). In accordance with this, Ebbs et al. (2002) concluded that PC synthesis was not responsible for Cd tolerance in T. caerulescens from Prayon (Belgium), which exhibits a similar degree of Cd hypertolerance. However, BSO did significantly increase Cd sensitivity in the Cd‐sensitive serpentine ecotype (Fig. 4), suggesting that PC‐dependent constitutive Cd tolerance does occur in non‐metallicolous ecotypes of this species. The major differences with S. vulgaris were that T. caerulescens showed considerable accumulation of PCs in the shoots, though less than in the roots, and that the responses to toxic Cd exposure and BSO were primarily apparent from shoot performance, rather than from root elongation, which is most probably due to the much higher rate of Cd translocation to the shoot.

Arsenic has been shown to induce high levels of PC accumulation in a variety of plant species (Grill et al., 1987; Maitani et al., 1996; Nakazawa et al., 2000; Schmöger et al., 2000). Intact As–PC complexes have been isolated from S. vulgaris (Sneller et al., 1999), and PC synthesis is supposed to be essential for As detoxification in plants (Schmöger et al., 2000; Hartley‐Whitaker et al., 2001). In agreement with this, a strong As‐induced PC accumulation was found in all the species tested. Moreover, BSO consistently produced hypersensitivity to As (Figs 6, 7). The relatively small BSO effect in S. vulgaris, as compared with the grasses (Fig. 6), might be due to the fact that BSO decreased root GSH to a much lower degree in S. vulgaris (to about 30% in unexposed controls) than it did in the grasses (to less than 5%). Arsenate, being a phosphate analogue, is taken up by phosphate transporters, and arsenate hypertolerance has been shown to be achieved through constitutive suppression of the high‐affinity phosphate uptake system in a number of grass species, including H. lanatus (Meharg and Macnair, 1990, 1991a, b, 1992). This suppression would reduce the influx of As to a level that can be coped with by the constitutive PC‐based detoxification machinery (Hartley‐Whitaker et al., 2001). These results support this hypothesis. Even extreme As hypertolerance, such as observed in the Jales population of A. castellana, was largely lost under BSO exposure (Fig. 7), suggesting that PC synthesis is equally essential for both normal constitutive tolerance and hypertolerance to As. Also in agreement with Hartley‐Whitaker et al. (2001), tolerance‐correlated decreases of PC‐SH to As molar ratios were not observed in the mine ecotypes, suggesting that alternative, PC‐independent sequestration mechanisms do not play any significant role in As hypertolerance.

Zn and, particularly, Ni and Co, are considered to be relatively weak activators of PCS, both in vivo and in vitro (Grill et al., 1988; Ahner and Morel, 1995; Klapheck et al., 1995). The stability of the Zn–PC complex is comparatively low (Maitani et al., 1996; Leopold and Gunther, 1997). The stabilities of Ni–PC and Co–PC complexes are unknown at present, but might be expected to be even lower, as suggested by the relatively low affinities of Ni and Co to other cysteine‐based ligands (Perrin, 1979). Davies et al. (1991) reported that BSO did not increase Zn‐imposed root growth inhibition in Festuca rubra. Also, the PC‐deficient Arabidopsiscad1 mutant did not exhibit considerably enhanced Zn‐sensitivity (Howden and Cobbett, 1992). In agreement with this, low, but detectable PC accumulation rates and low PC‐SH to metal molar ratios were observed in all the species and ecotypes under Zn, Ni, or Co exposure. BSO‐mediated hypersensensitivity was consistently absent, both in S. vulgaris and T. caerulescens, suggesting that PC‐based sequestration is not essential for the detoxification of either of these metals. Hypertolerance to Zn, Ni, and Co in S. vulgaris, and to Ni in T. caerulescens was again associated with decreased PC‐SH to metal molar ratios, which may be taken to indicate that these hypertolerances are achieved through sequestration mechanisms that decrease the metals’ availability for PCS activation in the cytoplasm, such as increased vacuolar transport in the case of Zn hypertolerance in S. vulgaris (Chardonnens et al., 1999). Comparable tolerance‐related differences in PC accumulation were not found among the three distinctly Zn‐tolerant T. caerulescens ecotypes, however. In particular, Zn, in contrast to Cd, induced very little PC accumulation in T. caerulescens, as compared to S. vulgaris. The reason for this is elusive, but might be related to the Zn hyperaccumulation trait.

To summarize, these results do not provide evidence in favour of a role for PCs in the detoxification of the essential metal micronutrients Zn, Ni, and Cu in plants, although Cu, when present at toxic concentrations, induced considerable PC accumulation and apparently formed stable complexes with PCs. Also, it is highly unlikely that PCs are essential in the detoxification of Fe, Mo, and Mn (Grill et al., 1989; Brune et al., 1995), suggesting that, in general, PCs might not be involved in the detoxification of excessively accumulated micronutrients in plants. On the other hand, this study confirms that PCs are required for the detoxification of certain non‐essential metals. Taking further evidence from the literature into account, it might be hypothesized that, in plants and algae at least, the primary function of PCS would lie in the detoxification of non‐essential metals and metalloids with relatively high affinities to sulphur, such as Cd, Hg (Howden and Cobbett, 1992; Gupta et al., 1998) and, particularly, As. Such metals, however, although being highly toxic and ubiquitous, are mostly present at negligible concentrations in the environment, which makes it difficult to believe that the conservation of functional PCS throughout the plant kingdom would be ultimately explained by toxic exposure to either of these metals. Moreover, in the case of Cd, these results clearly show a decreased dependency on PC‐based sequestration in hypertolerant S. vulgaris and T. caerulescens, suggesting that PC‐mediated detoxification might not be the most effective strategy to cope with toxic exposure to this metal, at least.

In view of the spread of significantly homologous PCS genes over the animal, plant, and fungal kingdoms (Clemens et al., 1999, 2001; Vatamaniuk et al., 2001), it seems likely that they must have evolved from an ancient ancestral gene. PCS genes seem to have been lost in a number of animal and fungal lineages, possibly as a consequence of the evolution of more effective and more specific MT‐based metal sequestration systems. One might argue that PCS originally functioned in Cu detoxification, as it still seems to do in S. pombe (Clemens et al., 1999). Its present function in plants and its ubiquitous occurrence throughout the plant kingdom are still enigmatic, however. Although the primary function of PCS does not seem to lie in the detoxification of excessively accumulated metal micronutrients, it cannot be excluded that it somehow functions in metal micronutrient homeostasis under non‐toxic physiological conditions, particularly because plants exposed to normal nutritional micronutrient exposure levels appear to contain PCs at low, but detectable concentrations.

Acknowledgements

This work was supported by a grant from the Catalonian Generality (CIRIT, project BEAI 300151). The authors are indepted to Dr Alessandra Lombini, University of Bologna, who kindly provided the seeds of the Monte Prinzera population of T. caerulescens.

Fig. 1. Mean root elongation (n=15) throughout 4 d of exposure to Cu in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicous (circles) and cupricolous (squares) S. vulgaris. Standard errors varied between 2% and 8% of the means.

Fig. 1. Mean root elongation (n=15) throughout 4 d of exposure to Cu in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicous (circles) and cupricolous (squares) S. vulgaris. Standard errors varied between 2% and 8% of the means.

Fig. 2. Mean root elongation (n=15) throughout 4 d of exposure to Cd in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous (circles) and Cd‐hypertolerant (squares) S. vulgaris. Standard errors varied between 3% and 14% of the means.

Fig. 2. Mean root elongation (n=15) throughout 4 d of exposure to Cd in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous (circles) and Cd‐hypertolerant (squares) S. vulgaris. Standard errors varied between 3% and 14% of the means.

Fig. 3. Mean PC‐thiol to Cd molar ratio (n=12) in roots, as a function of the root‐internal Cd concentration, in T. caerulescens from calamine (closed circles), serpentine (open circles), and non‐metalliferous soil (open squares), after 4 d of exposure to a series of Cd concentrations in the nutrient solution. Standard errors varied between 5% and 19% of the means.

Fig. 3. Mean PC‐thiol to Cd molar ratio (n=12) in roots, as a function of the root‐internal Cd concentration, in T. caerulescens from calamine (closed circles), serpentine (open circles), and non‐metalliferous soil (open squares), after 4 d of exposure to a series of Cd concentrations in the nutrient solution. Standard errors varied between 5% and 19% of the means.

Fig. 4. Mean root elongation (n=12) throughout 4 d of exposure to Cd in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous (diamonds), serpentine (circles), and calamine (squares) T. caerulescens. Standard errors varied between 3% and 16% of the means.

Fig. 4. Mean root elongation (n=12) throughout 4 d of exposure to Cd in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous (diamonds), serpentine (circles), and calamine (squares) T. caerulescens. Standard errors varied between 3% and 16% of the means.

Fig. 5. Median shoot biomass (n=12) after 2 weeks of exposure to Cd in BSO‐treated (open symbols) and untreated (closed symbols) serpentine (circles) and calamine (squares) T. caerulescens. Standard errors varied between 7% and 24% of the means.

Fig. 5. Median shoot biomass (n=12) after 2 weeks of exposure to Cd in BSO‐treated (open symbols) and untreated (closed symbols) serpentine (circles) and calamine (squares) T. caerulescens. Standard errors varied between 7% and 24% of the means.

Fig. 6. Mean root elongation (n=15) throughout 4 d of exposure to arsenate in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous S. vulgaris (circles) and H. lanatus (squares). Standard errors varied between 2% and 15% of the means.

Fig. 6. Mean root elongation (n=15) throughout 4 d of exposure to arsenate in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous S. vulgaris (circles) and H. lanatus (squares). Standard errors varied between 2% and 15% of the means.

Fig. 7. Mean root elongation (n=15) throughout 4 d of exposure to arsenate in BSO‐treated (open symbols) and untreated (closed symbols) metallicolous A. castellana (circles) and H. lanatus (squares). Standard errors varied between 4% and 18% of the means.

Fig. 7. Mean root elongation (n=15) throughout 4 d of exposure to arsenate in BSO‐treated (open symbols) and untreated (closed symbols) metallicolous A. castellana (circles) and H. lanatus (squares). Standard errors varied between 4% and 18% of the means.

Fig. 8. Mean root elongation (n=15) throughout 4 d of exposure to Zn in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous (circles) and calamine (squares) S. vulgaris. Standard errors varied between 6% and 17% of the means.

Fig. 8. Mean root elongation (n=15) throughout 4 d of exposure to Zn in BSO‐treated (open symbols) and untreated (closed symbols) non‐metallicolous (circles) and calamine (squares) S. vulgaris. Standard errors varied between 6% and 17% of the means.

Table 1.

Total phytochelatin thiol (PC‐SH) concentrations and PC‐SH to Cu molar ratios (means of three samples of three plants each; SE in parentheses) in desorbed roots of unpretreated and BSO‐pretreated non‐metallicolous (Amsterdam) and cupricolous (Marsberg) S. vulgaris, after a 4 d exposure to increasing Cu concentrations in the nutrient solution (nd=not determined)

Population Exposure (µM Cu) PC‐SH (µmol g–1 DW) PC‐SH:Cu (mol mol–1
  +BSO –BSO +BSO –BSO 
Amsterdam 0.1 <0.1 <0.1 nd nd 
 0.7 0.2 (0.07) 0.4 (0.11) 0.8 (0.23) 1.2 (0.40) 
 2.1 3.2 (0.63) 9.7 (2.35) 1.4 (0.21) 3.7 (0.65) 
 6.3 4.6 (0.41) 12.8 (1.98) 1.5 (0.47) 3.6 (0.15) 
 18.9 1.8 (0.16) 2.8 (0.61) nd nd 
Marsberg 0.1 <0.1 <0.1 nd nd 
 1.0 0.2 (0.08) 0.3 (0.08) 0.4 (0.04) 0.3 (0.07) 
 7.0 0.5 (0.04) 1.1 (0.23) 0.8 (0.23) 1.4 (0.12) 
 21.0 1.3 (0.24) 2.4 (0.37) 0.3 (0.06) 1.4 (0.16) 
 63.0 2.7 (0.19) 7.9 (1.46) 0.3 (0.10) 1.3 (0.08) 
 189.0 1.9 (0.45) 2.6 (0.09) nd nd 
Population Exposure (µM Cu) PC‐SH (µmol g–1 DW) PC‐SH:Cu (mol mol–1
  +BSO –BSO +BSO –BSO 
Amsterdam 0.1 <0.1 <0.1 nd nd 
 0.7 0.2 (0.07) 0.4 (0.11) 0.8 (0.23) 1.2 (0.40) 
 2.1 3.2 (0.63) 9.7 (2.35) 1.4 (0.21) 3.7 (0.65) 
 6.3 4.6 (0.41) 12.8 (1.98) 1.5 (0.47) 3.6 (0.15) 
 18.9 1.8 (0.16) 2.8 (0.61) nd nd 
Marsberg 0.1 <0.1 <0.1 nd nd 
 1.0 0.2 (0.08) 0.3 (0.08) 0.4 (0.04) 0.3 (0.07) 
 7.0 0.5 (0.04) 1.1 (0.23) 0.8 (0.23) 1.4 (0.12) 
 21.0 1.3 (0.24) 2.4 (0.37) 0.3 (0.06) 1.4 (0.16) 
 63.0 2.7 (0.19) 7.9 (1.46) 0.3 (0.10) 1.3 (0.08) 
 189.0 1.9 (0.45) 2.6 (0.09) nd nd 
Table 2.

Total phytochelatin thiol (PC‐SH) concentrations and PC‐SH to Cd molar ratios (means of three samples of three plants each; SE in parentheses) in desorbed roots of non‐metallicolous (Amsterdam) and calamine (Plombières) S. vulgaris after 4 d of exposure, with and without BSO, to increasing Cd concentrations in the nutrient solution (nd=not determined)

Population Exposure (µM Cd) PC‐SH (µmol g–1 DW) PC‐SH:Cd (mol mol–1
  +BSO –BSO +BSO –BSO 
Amsterdam 7.5 5.0 (1.07) 21.2 (2.07) 0.9 (0.08) 3.4 (0.40) 
 15.0 4.1 (0.43) 24.7 (4.72) 0.6 (0.12) 2.7 (0.12) 
 30.0 6.7 (1.24) 35.3 (2.91) 0.6 (0.03) 2.5 (0.31) 
 60.0 9.2 (0.89) 40.6 (5.33) 0.7 (0.15) 2.6 (0.45) 
 120.0 4.6 (0.39) 42.7 (2.71) 0.4 (0.10) 1.8 (0.06) 
 240.0 nd nd nd nd 
Plombières 7.5 2.8 (0.20) 7.3 (0.83) 0.7 (0.02) 1.8 (0.25) 
 15.0 3.0 (0.08) 9.3 (2.07) 0.5 (0.09) 1.7 (0.29) 
 30.0 3.5 (0.53) 11.2 (0.99) 0.6 (0.05) 1.6 (0.08) 
 60.0 6.1 (0.36) 16.(1.74) 0.6 (0.11) 1.7 (0.31) 
 120.0 5.4 (1.14) 17.(3.01) 0.3 (0.07) 1.1 (0.16) 
 240.0 9.3 (2.79) 26.(2.61) 0.3 (0.10) 0.7 (0.09) 
Population Exposure (µM Cd) PC‐SH (µmol g–1 DW) PC‐SH:Cd (mol mol–1
  +BSO –BSO +BSO –BSO 
Amsterdam 7.5 5.0 (1.07) 21.2 (2.07) 0.9 (0.08) 3.4 (0.40) 
 15.0 4.1 (0.43) 24.7 (4.72) 0.6 (0.12) 2.7 (0.12) 
 30.0 6.7 (1.24) 35.3 (2.91) 0.6 (0.03) 2.5 (0.31) 
 60.0 9.2 (0.89) 40.6 (5.33) 0.7 (0.15) 2.6 (0.45) 
 120.0 4.6 (0.39) 42.7 (2.71) 0.4 (0.10) 1.8 (0.06) 
 240.0 nd nd nd nd 
Plombières 7.5 2.8 (0.20) 7.3 (0.83) 0.7 (0.02) 1.8 (0.25) 
 15.0 3.0 (0.08) 9.3 (2.07) 0.5 (0.09) 1.7 (0.29) 
 30.0 3.5 (0.53) 11.2 (0.99) 0.6 (0.05) 1.6 (0.08) 
 60.0 6.1 (0.36) 16.(1.74) 0.6 (0.11) 1.7 (0.31) 
 120.0 5.4 (1.14) 17.(3.01) 0.3 (0.07) 1.1 (0.16) 
 240.0 9.3 (2.79) 26.(2.61) 0.3 (0.10) 0.7 (0.09) 
Table 3.

Total phytochelatin thiol (PC‐SH) concentrations and PC‐SH to As molar ratios (means of three samples of three plants each; SE in parentheses) in desorbed roots of non‐metallicous S. vulgaris (population Amsterdam), after 4 d of exposure, with and without BSO, to increasing As concentrations in the nutrient solution (nd=not determined)

Species (population) Exposure (µM As) PC‐SH (µmol g–1 DW) PC‐SH:As (mol mol–1
  +BSO –BSO +BSO –BSO 
S. vulgaris 1.7 (0.32) 3.4 (0.19) 2.7 (0.40) 4.1 (0.27) 
(Amsterdam) 3.7 (0.21) 8.6 (1.26) 1.9 (0.08) 3.1 (0.27) 
 16 7.5 (0.51) 17.3 (2.50) 2.7 (0.12) 3.8 (0.19) 
 32 6.8 (0.79) 27 4 (4.68) 1.1 (0.32) 2.7 (0.35) 
 64 3.4 (0.64) 31.7 (6.07) 0.6 (0.01) 3.0 (0.58) 
 128 <0.1 <0.1 nd nd 
Species (population) Exposure (µM As) PC‐SH (µmol g–1 DW) PC‐SH:As (mol mol–1
  +BSO –BSO +BSO –BSO 
S. vulgaris 1.7 (0.32) 3.4 (0.19) 2.7 (0.40) 4.1 (0.27) 
(Amsterdam) 3.7 (0.21) 8.6 (1.26) 1.9 (0.08) 3.1 (0.27) 
 16 7.5 (0.51) 17.3 (2.50) 2.7 (0.12) 3.8 (0.19) 
 32 6.8 (0.79) 27 4 (4.68) 1.1 (0.32) 2.7 (0.35) 
 64 3.4 (0.64) 31.7 (6.07) 0.6 (0.01) 3.0 (0.58) 
 128 <0.1 <0.1 nd nd 
Table 4.

Total phytochelatin thiol (PC‐SH) concentrations and PC‐SH to Zn molar ratios (means of three samples of three plants each) in desorbed roots of non‐metallicolous (Amsterdam) and calamine (Plombières) S. vulgaris, after a 4 d exposure, with and without BSO, to increasing Zn concentrations in the nutrient solution (nd=not determined)

Population Exposure (µM Zn) PC‐SH (µmol g–1 DW) PC‐SH:Zn (mol mol–1
  +BSO –BSO +BSO –BSO 
Amsterdam <0.1 <0.1 nd nd 
 25 2.1 (0.12) 6.1 (0.29) 0.21 (0.05) 0.34 (0.01) 
 50 3.4 (0.23) 8.4 (0.13) 0.12 (0.03) 0.23 (0.03) 
 100 3.7 (0.42) 8.6 (0.99) 0.08 (0.01) 0.22 (0.04) 
 200 6.0 (1.41) 11.9 (1.15) 0.09 (0.02) 0.17 (0.02) 
 400 5.8 (1.01) 14.9 (0.98) 0.11 (0.03) 0.20 (0.02) 
Plombières <0.1 <0.1 nd nd 
 200 0.3 (0.02) 0.5 (0.01) 0.008 (0.002) 0.014 (0.003) 
 400 0.5 (0.12) 1.3 (0.11) 0.009 (0.003) 0.019 (0.002) 
 800 2.1 (0.03) 3.4 (0.40) 0.004 (0.001) 0.017 (0.004) 
 1600 3.4 (0.38) 5.0 (1.02) 0.008 (0.001) 0.014 (0.001) 
 3200 2.1 (0.19) 2.7 (0.34) 0.005 (0.002) 0.008 (0.002) 
Population Exposure (µM Zn) PC‐SH (µmol g–1 DW) PC‐SH:Zn (mol mol–1
  +BSO –BSO +BSO –BSO 
Amsterdam <0.1 <0.1 nd nd 
 25 2.1 (0.12) 6.1 (0.29) 0.21 (0.05) 0.34 (0.01) 
 50 3.4 (0.23) 8.4 (0.13) 0.12 (0.03) 0.23 (0.03) 
 100 3.7 (0.42) 8.6 (0.99) 0.08 (0.01) 0.22 (0.04) 
 200 6.0 (1.41) 11.9 (1.15) 0.09 (0.02) 0.17 (0.02) 
 400 5.8 (1.01) 14.9 (0.98) 0.11 (0.03) 0.20 (0.02) 
Plombières <0.1 <0.1 nd nd 
 200 0.3 (0.02) 0.5 (0.01) 0.008 (0.002) 0.014 (0.003) 
 400 0.5 (0.12) 1.3 (0.11) 0.009 (0.003) 0.019 (0.002) 
 800 2.1 (0.03) 3.4 (0.40) 0.004 (0.001) 0.017 (0.004) 
 1600 3.4 (0.38) 5.0 (1.02) 0.008 (0.001) 0.014 (0.001) 
 3200 2.1 (0.19) 2.7 (0.34) 0.005 (0.002) 0.008 (0.002) 

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