According to modelling results, the uncomplexed free Cd2+ pool was the predominant (50-74%) in acid (pH < 5.0) conditions, but its activity linearly decreased with pH increasing, it was almost fully depleted (<1%) at pH 9.0 (Fig. 22.4). In the very narrow range of slight acidity (pH 6.0-6.5) Cd-organocomplexation was the most pronounced (44-45%), especially in the LMW pool (>35%), whereas inorganic complexation of Cd dominated (38-97%) in basic (pH > 8) conditions (Fig. 22.4). In contaminated rhizosphere conditions (vs. non-contaminated), distribution in Cd2+ and Cd-inorganic pools stayed rather stable, whereas in other pools Cd was redistributed from the HMW to the LMW fraction. The main cause of such Cd redistribution inside organic pools is that in contaminated conditions certain LMW organic acids were increased in the rhizosphere many times (Table 22.2) as a consequence of pronounced rhizodeposition (see the next).

Although Cd-inorganic complexation started to increase from pH 5.5 (Fig. 22.4a), CdCl+ activity from the Cd-inorganic pool continued decreasing throughout the pH range (Fig. 22.4b). The main reason is that Cd sorption with other ligands (HCO3-, CO32-, HPO42-, etc.) dominated over Cd-chlorocomplexation (data not shown). The most dominant LMW- and HMW-Cd forms were Cd-malate and FA1-Cd(6) (i.e. Cd bound to FAs via carboxylic groups) over the whole pH range (data not shown), with Cd-malate concentrations being the highest in slightly acidic pH (5.5-6.0), and those of FA1-Cd(6) in neutral to slightly basic pH (7-7.5) (Fig. 22.4b) . In contaminated conditions, the curves showing activity of Cd forms retained the same shape but the activities were an order of magnitude higher than in non-contaminated model (data not shown).

Calculated saturation indices for over 200 possible mineral phases included in the Visual Minteq database were checked, and several of observed TE minerals were found to be oversatu-rated (i.e. SI > 0), but only in contaminated and basic conditions: at pH 8 malachite (Cu mineral), hydrozincite and smithsonite (Zn minerals) and otavite (Cd mineral) and at pH 9 all Zn and Cd minerals as at pH 8 plus zincite (Zn mineral) (data not shown).

The mobility and uptake of Cd, Zn and Cu is strongly dependent on chemical speciation/distri-bution and concentrations (activities) of TEs in

Fig. 22.4 Distribution of Cd species (a) among four and activities of some of the most abundant Cd species pools [non-contaminated conditions = L (left bars); (b) inside each pool (non-contaminated conditions only) contaminated conditions = R (right bars with bold %)]

the rhizosphere solution. Although it is believed that only uncomplexed i.e. free cationic metal form may be taken up by roots, there is increasing evidence that Cd (Zn, Cu) can be mobilised in soil/nutrient solution and then taken up/phytoac-cumulated complexed with inorganic ligands such as chlorides (Smolders and McLaughlin 1996; Khoshgoftar et al. 2004; Weggler et al. 2004; Khoshgoftarmanesh et al. 2006; Ondrasek et al. 2009a) or sulphates (McLaughlin et al. 1998b). Indeed, Weggler et al. (2004) observed that shoot Cd concentrations of wheat grown in a biosolid-amended soil were most closely correlated with

CdCl+ activity in soil solution, whereas the correlation with the Cd2+ activity was weak. McLaughlin et al. (1998a) observed that Cd shoot/ root concentrations in Swish chard were unaffected by additions of sulphate to nutrient solution despite Cd2 + activities decreasing markedly in the rhizosphere. The above studies suggested that Cd-inorganic complexes (CdSO4, CdCl+) could be phytoavailable and enter root plasma membrane either directly as a metal-complex and/ or dissociating in the apoplast and entering cells as the free metal cation (Smolders and McLaughlin 1996).

With pH decreasing, the acid functional groups of LMW/HMW organic substances deprotonate, influencing solubility and formation of metallo-organo complexes. In many field/laboratory experiments with a range of species (including hyperaccumulators), increased mobility and improved uptake of Cd (Zn, Cu, Pb, etc.) was elicited by application of synthetic LMW chelat-ing agents (e.g. EDTA, NTA, etc. See review by Schmidt 2003). In the models proposed here, we did not consider synthetic but only naturally occurring LMW-OAs that markedly impact TEs phytoextraction. LMW-OAs are common constituents of root/microbial exudates (see review by Jones 1998) , which are present in relatively lower concentrations in non-contaminated vs. contaminated conditions (Table 22.2). Evangelou et al. (2006) observed improved Cu accumulation (up to 2.3-fold) in tobacco shoots with the addition of LMW-OAs (citric, tartaric and oxalic) to soil, similarly to Nigam et al. (2001) who recorded enhanced Cd concentrations in maize shoots (up to greater than twofold) after soil application of citric and malic acid. Rhizosphere contamination with Cd induced higher root exudation of certain LMW-OAs (propionic by 11.2-fold, butyric by 7.6-fold, acetic by 4.7-fold), with their concentrations in the rhizosphere positively correlating with the amount of Cd accumulated in millet shoots (r2=0.96 P < 0.001) and roots (r2=0.98 P < 0.001) (Chiang et al. 2011).

Studying the leaching of TEs from highly contaminated soils, Fischer et al. (1998) depleted total Cd by 75% (Cu by 54% and Zn by 56%) with the application of acidic (pH 4.4) grass silage effluent. Such metal removal efficiency was attributed to effluent's main component i.e. lactic acid, which under acid pHs acts as a proton-donor and strong metallo-complexing agent. According to the same authors, there are several crucial mechanisms that explain TE mobilisation by LMW-OAs: (1) chelation, (2) proton-initiated solubilisation, (3) surface complexation, (4) ion exchange and (5) reductive dissolution of metal binding substrates i.e. hydr/oxides.

Complexation of free metals ions (the most bioavailable and potentially the most toxic forms) with certain HMW SOM substances could be important in safe food production on metal-contaminated soils. In recent studies with horticultural crops (melon, strawberry, lettuce and radish) in slightly acidic pH (~6.0) and organically enriched (DOC > 76 mg/L) rhizosphere solution (Ondrasek 2008; Ondrasek et al. 2009a, submitted) using the same chemical modelling approach as described above, dissolved FAs dominated in Cd (Cu) sorption processes and formation of soluble metallo-organic complexes with relatively poor bioavailability (Otto et al. 2001). Similar findings were reported in maize (Shuman et al. 2002) , mustard grown in organic soil contaminated by Cd (Bolan et al. 2003), sorghum grown in OM-enriched nutrient solution (Pinto et al. 2004), whereby a decrease in phytotoxicity and Cd content in the shoots caused by the addition of OM was attributed to redistribution of Cd from the water-soluble/exchangeable to the organically bound fraction. Similarly to FAs, HAs may also influence TE bioavailability given they also contain acid functional groups. However, in most natural soil conditions, HAs are insoluble compared to naturally more soluble FAs.

However, when added to soil, FAs and HAs may cause contrasting effects on TEs, i.e. the formation of soluble metallo-organic complexes may enhance metal bioavailability and leaching, whereas interaction of metals with the solid phase of humics may lead to their immobilisation and thus decrease their potential hazardous environmental influence (Gondar and Bernal 2009). The phytoavailable forms of Cd (Zn, Cu, Pb) in soils were increased by adding HAs to two metal contaminated mineral soils (Halim et al. 2003) . In contrast, application of HAs to Cd-contaminated mineral soil did not change phytoavailable soil Cd, but phytoaccumulation of Cd in tobacco increased by up to 65% accompanied by toxicity.

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