Plants Copper

In undisturbed soils, it appears that most of the plant copper is assimilated by root interception of ions without much movement by mass flow or diffusion in the soil solution. A Consequence of this is that copper absorption decreases markedly when root growth in the topsoil declines and stops. On the other hand, the uptake of copper by plants is enhanced by fungi (mycorrhizae) associated with the roots of certain species. Almost all of the copper in the root environment (the rhizosphere) is organically complexed by exudates from the roots or by soil humus compounds, but it is generally understood that the metal dissociates from these organic molecules before being absorbed as Cu2+. The root absorption mechanism is not entirely understood, but it obviously includes a great number of steps and may be influenced by a number of external as well as internal factors (Allen, 1999).

At the normal copper concentrations in soils reported (0.3 - 250 mg Cu/kg DM), plants rarely if ever show symptoms of toxicity or of adverse growth effects. Crops are often more sensitive to copper than the native flora, so protection levels for agricultural crops, ranging from 25 mg Cu/kg DM to >100 mg/kg DM, in general seem to be adequate for plant protection in general (IPCS, 1998). However, chronic or acute effects on sensitive plant species may occur in some soil types at copper levels in the range 50-150 mg Cu/kg DM. When copper levels in soils rise above 150 mg/kg, we begin to find more and more native and agricultural species showing chronic effects.

Generally speaking it is, nonetheless, very difficult to establish a clear threshold level for copper toxicity to plants, expressed in terms of total copper concentration in the soil. It has been suggested that a possible solution to this dilemma might be to express toxicity threshold values for copper (and other trace metals) in terms of concentrations in plant tissues, preferably in young shoots, rather than as concentration in soil. The most likely "critical" copper concentration in plant tissue appears to be in the range 10 - 30 mg Cu/kg dry tissue, but more exact values have to be determined for each of the important crops and soil categories (Landner et al., 2000).

The most sensitive tests with soil invertebrates seem to justify a toxicity threshold level in the range 30 - 40 mg Cu/kg DM (as total soil copper) for this community, at least in sandy soils.

Dumestre et al. (1999) studied the effects of copper on microorganisms in relation to its speciation. They found that the soil solution free Cu2+ activitiy (pCu2+) proved to be the best predictor of soil Cu toxicity determined as microbial respiration lag period (LP), and hence of the soil quality, and that pCu2+ could integrate the existing soil physicochemical variability. Although correlating well with pCu2+, maximum mineralization rate was found not to be a good effect, because of its high sensitivity to other soil charcteristics, like soil organic matter (SOM).

For most practical purposes, direct determination (ion-specific electrode) or calculation of the free cupric ion activity appears to be a sufficiently good measure of the bioavailable copper fraction in a soil. Recent data support the idea that the predominant copper species being absorbed by plant roots is the free cupric ion, and good correlations have usually been obtained between the free ion activity and results of toxicity tests with soil organisms (Sauvé et al., 1998). Based on an analysis of 31 different laboratory bioassays of copper toxicity in the soil ecosystem and a recalculation of the soil total copper concentrations corresponding to free ion activities that caused 25% inhibition in the various bioassay endpoints, it was found that the effective concentration was 20 mg Cu/kg DM in soils with pH 6.0, and 140 mg Cu/kg DM in soils with pH 7.0.

The calculated toxic threshold copper concentrations for soil invertebrates, soil microbial activities or soil microorganisms , are usually based on laboratory assays, and must therefore be utilized with caution. The threshold levels cannot be directly translated to inhibitory concentrations in the field, because laboratory bioassays tend to overestimate toxicity by up to 5 times (see above). However, the data given by Sauvé et al (1998) are useful to illustrate the strong influence of pH on toxicity threshold in the soil solution (Landner et al., 2000).

An important observation in this context is the existence of a relationship between the original, background concentration of total copper in the soil and the permissible addition of copper up to the level where the "critical" concentration is reached. Thus, the "critical" number of times the total copper concentration in agricultural soils can be increased was estimated to be 4 times (Witter, 1992). The corresponding "critical" elevation of copper in the mor (surface layer) of forest soils was estimated at 3 times, i.e. the number of times the copper concentration can be enhanced compared to the background concentration before adverse effects on the soil ecosystem begin to appear (Tyler, 1992).

The toxicity of zinc to terrestrial organisms is, or course, dependent upon its bioavailability, which in turn is determined by various factors such as the speciation of zinc and the physico-chemical characteristics of the soil. The bioavailable fraction of zinc in soil has been calculated to range from <1% to 10% of the total zinc concentration. Zinc has to be in a soluble form to be taken up by plants. In the case of zinc toxicity, zinc replaces other metals (e.g. iron and manganese) in the active centres of enzymes (e.g. hydrolases and haem enzymes) (IPCS, 2001).

Uptake of zinc in terrestrial plants is significantly increased at a low soil pH, but reduced when there is a high content of organic matter. Normal levels of zinc in most crops and pastures range from 10 mg/kg to 100 mg/kg (IPCS, 2001). Some plant species are zinc accumulators, but the extent of the accumulation in plant tissues varies with soil properties, plant organ and tissue age.

As a general rule, it has been found that plants from environments poor in zinc are characterized by low zinc concentrations, while those from zinc-enriched environments have high concentrations. The critical leaf tissue concentration of zinc at which growth is affected was found for many plant species to be between 200 and 300 mg/kg DM. However, zinc phytotoxicity in leaves can depend to a large extent on the plant species, the age of the leaf and other factors, such as exposure period and exposure concentration. Concentrations of zinc that are subtoxic or non-toxic to plants may have metabolic effects higher up the food chain. The disappearance of herbivorous insects on zinc-tolerant plants is one example of differences in species-specific tolerances. Similarly, the zinc-content of zinc-efficient plants may be insufficient for optimum performance of herbivorous animals, especially if the zinc is present in a form which is not readily bioavailable (IPCS, 2001).

Van Gestel et al. (1993) exposed earthworms (Eisena andrei) to zinc as zinc chloride at concentrations in dry artificial soil of 100 - 1000 mg/kg. Zinc significantly reduced reproduction at soil concentrations of 560 and 1000 mg/kg and induced the production of malformed cocoons. EC5o values for the effect of zinc on cocoon production and the number of juveniles per worm per week were 660 and 510 mg/kg DM soil, respectively. At the end of a 3-week recovery period, reproduction had returned to normal.

Smit et al. (2002) studied the effect of zinc on the nematode soil fauna in an experimentally contaminated outdoor field plot soil with regard to Zn speciation in soil porewater and CaCl2-exchangeable Zn, as a measure of Zn-bioavailability. Comparing the observed response with benchmark concentrations (like H5 and H50 values) as derived for Zn from the 'general species sensitivity distribution' (SSD) of soil organisms, predictions of the SSD model could be confirmed, i. e. the community NOEC was in agreement with benchmark values that should protect the integrity of the soil ecosystem.

Lock and Janssen (2001) compared a zinc-spiked artificial soil and a historically contaminated field soil to examine if chronic zinc toxicity (to the springtail Folsomia candida) can be predicted by a surface-response model based on soil pH, cation exchange capacity (CEC) and total zinc. However, the model could not adequately predict chronic zinc toxicity in field soils, as porewater, water- and CaCl2-extractable zinc were lower than predicted by the model developed from artificial soils. The authors suggested that effects of aging on the bioavailability of zinc should be taken into account in this context (cf. section 5.5.6). Also reproduction of F. candida in field soils was lower than predicted by models based on porewater, water and CaCl2-extractable zinc from artificial soils suggesting the existence of other bioavailable zinc (or other metal) fractions and/or a possible dietary route of uptake. For this reason, both ageing and dietary uptake should become included in effect-based risk assessment of metal-contaminated soils (see section 5.5.6).

From their experimental data, Lock and Janssen (2001) resumed that porewater, water and CaCl2-extractable zinc are predictable by pH, CEC and total zinc in artificial but not in field soils, and that porewater, water and CaCl2-extracted zinc - but not total zinc - predicted toxicity (reproduction) in these laboratory systems. In field soils, reproduction was lower (toxicity higher) than predicted. In contrast to the potworm E. albidus experiments, reproduction of F. candida was not influenced by soil characteristics (pH and OM), and hence, this species seems to be more appropriate when assessing the ecotoxicity of zinc in various soils.

It may be concluded that if porewater and extractable zinc concentrations were the only bioavailable fractions, zinc toxicity would be easily predictable. Indeed the authors succeeded to show that reproduction of F. candida was accurately predicted by PLS models ('partial least squares projection to latent structures'), which are based on those fractions in artificial soils that are independent of pH and CEC. However, this modelling approach also confirmed a lower reproduction in field soils than predicted by the surface-response model confirming again the existence of other uptake routes or bioavailable fractions.

It is once more pointed out that most terrestrial toxicity tests work with high added concentrations and short equilibrium times, which inevitably result in a high acute toxicity from artificially high porewater concentrations long before effects of dietary exposure become visible. This is, of course, obstructing the final data interpretation. However, the porewater approach still seems to be appropriate to protect organisms from acute toxicity in extremely contaminated soils, although major regulatory and scientific concern is focused on moderately contaminated soils. The need to increasingly consider these special types of ecosystems and also dietary routes of exposure is again obvious.

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