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It was shown that dissipation of condensed PAHs might be enhanced in the presence of AM fungi in the soil spiked with a mixture of different PAHs [127]. These results were consecutively verified in pot experiments conducted in two different industrially polluted soils [128]. In general, following explanations were suggested by the authors to elucidate the contribution of AM to PAH degradation: i) mycorrhiza modifies root enzyme activity, exudation and architecture in a manner that stimulates PAH degradation, either by root-derived enzymes or by rhizosphere microorganisms, ii) mycorrhizal colonisation affects root surface or rhizosphere soil properties that act on PAH availability through adsorption. Several examples of above mentioned capabilities of AM fungi can be found in the literature. Mycorrhiza was reported to enhance amount of hydrogen peroxide in the roots [129] and to stimulate oxidoreductase activities in the roots and the rhizosphere [130]. These mechanisms may contribute to increase of PAH dissipation associated with mycorrhizal roots. Mycorrhizal colonisation also modifies root exudation both quantitatively and qualitatively [131, 132], which could have further effects on the composition and activity of microbial communities in the rhizosphere [133, 134]. This mycorrhiza-associated microflora may be more effective in organics degradation in comparison with that related to non-mycorrhizal roots. Accumulation of phenolics in the roots or rhizosphere soil of mycorrhizal plants could induce degradation of more complex aromatic compounds [135].

Effects of AM on root longevity and proportion of higher order lateral roots with short life span was documented for several plant species [136]. Accumulation of root debris together with the ability of mycorrhizal plants to enrich the soil in organic matter [137] may contribute to enhanced PAH adsorption in plant rhizosphere. Finally, AM colonisation results in formation of an extensive network of extraradical mycelium, which can modify surrounding environment by extrusion of glycoproteins [138] and extracellular enzymes [139]. Hyphal biomass together with the two latter substances can serve as substrates for microbial growth. Nutrients derived from extraradical hyphae of AM fungi were hypothesised to drive co-metabolic degradation of PAHs within small soil pores where PAHs are spatially unavailable to roots [127].

Most of the ectomycorrhizal fungi have some limited ability to use polymers such as lignin and cellulose as substrates for their growth [2], however, their ability to degrade polymers of this kind was supposed to be much lower as compared to wood decomposers or even of some ericoid mycorrhizal fungi [140, 141, 142]. As it was recently reviewed, most of the EcM fungi screened for degradation of persistent organic pollutants are able to transform these compounds [143]. On the other hand, most reports concern degradation in pure cultures, rather than in symbiosis with plants. It has been demonstrated that EcM fungi can degrade e.g. trinitrotoluene [144], dichlorphenol [145], atrazine [146] and several 3-5 ring PAHs [147]. Only limited PAH degrading abilities of EcM fungi and decreased mineralisation of some PAHs in microcosms with EcM pine seedlings vs. unplanted microcosms inoculated with ectomycorrhizosphere soil were documented [148]. On the contrary, mineralisation of dichlorphenol was stimulated when EcM fungi were cultivated in symbiosis with pine than when grown in absence of the host [145].

Also degradation of lower chlorinated PCBs by eight of 13 studied species of EcM fungi was reported, however, only two species were able to degrade 4 and 5 chlorinated biphenyls [149]. While EcM fungi could sequentially hydroxylate a halogenated biphenyl ring (parent compound is then more polar and bioavailable), they were unable to cleave the ring [150]. However, hydroxylation of the biphenyl ring can be an important initial metabolic step facilitating degradation of PCBs by other rhizosphere organisms that have the capacity to degrade compounds further [143]. In this context it can be hypothesised that the presence of EcM fungi may negate the need for the presence of co-substrates. Although the tolerance of EcM fungi to different aromatic compounds is lower as compared to white rot fungi and depends on the compound type, its external concentration and the fungal species, the degradative capabilities of different fungi varied between species but not generally between the biotrophic and saprotrophic fungi [151].

6. Conclusion

Application of mycorrhizal fungi as supportive agents for phytoremediation can generally eliminate or reduce some known limitation of phytoremediation processes. Mycorrhizal fungi can help plants to acquire nutrients more effectively, increase their tolerance to edaphic stress and change accumulation of pollutants in plant tissues.

Reduced HM concentrations in plant tissues of mycorrhizal plants, together with an amplified barrier against metal translocation from plant roots to shoots, are favourable for phytostabilisation that is aimed at prevention of contamination spreading into the surroundings. The presence of AM fungi leading to decreased HM concentrations in plant shoots can be also an important factor for quality and safety of plants which enter food chains such as forage crops, medicinal herbs and vegetables. For example, tobacco was found to accumulate significantly less cadmium in the leaves when inoculated with selected AM fungi [8].

On the other hand, enhanced HM concentrations in the shoots of mycorrhizal plants induced by some mycorrhizal fungal isolates represent optimal conditions for phytoextraction technique. In some cases, elimination of present AM fungi populations (in particular if they involve strains decreasing HM translocation to the shoots) might be recommended before phytoextraction beginning. For example, application of the fungicide benomyl detrimental to mycorrhiza was shown to significantly decrease root colonisation and simultaneously to increase Pb concentrations in plant shoots [153]. Mycorrhizal fungi may be crucial also for re-vegetation efforts after heavy metal removal as the rate of site re-vegetation may be accelerated when AM fungi are present in soil. However, little is known about mycorrhiza functioning under conditions imposed by particular metal remediation protocols. First investigations have appeared showing that the quantity and species composition of glomalean propagules and the functioning of AM symbiosis could be significantly influenced by phytoextraction treatments (the choice of plant species, i.e. non-mycotrophic vs. mycotrophic, soil supplements etc.) [7]. Recently, negative effects of synthetic chelates used for chelate-induced HM phytoextraction such as EDTA on AM development were also described [102, 152].

Based on the intended phytoremediation strategy, the appropriate management of native fungi and/or application of artificial inocula should be chosen. For the introduction of artificial inocula, it seems to be essential to formulate specific products (mixtures of strains compatible with target plants and environment) rather than to use generic products all over the scale of edaphic conditions. However, it should considered that the effect of mycorrhizal inoculation may interact e.g. with fertilisation regime as demonstrated for the grasses grown in mine tailings containing high levels of zinc: plant growth was best after inoculation combined with nitrogen and phosphorus fertilisation, whereas neither mycorrhiza nor fertilisation alone had any effect on plant biomass [77]. Therefore, mycorrhizal inoculation cannot be considered as panacea and should be combined with other practices such as appropriate fertilisation or soil amendments to maximise re-vegetation success.

To conclude, an extension of knowledge on the involvement of mycorrhizal fungi in phytoremediation should still be achieved. Great attempt should be also undertaken to increase awareness of potential users of mycorrhizal inoculants regarding all possible functions and impacts of mycorrhiza applications in phytoremediation processes.

Acknowledgement

Presentation of this contribution was supported by the grant 526/04/0996 of the Grant Agency of the Czech Republic and by the Research Centre for Bioindication and Revitalisation funded by the Ministry of Education, Youth and Sports of the Czech Republic (grant 1M67985939).

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