Heavy metal (HM) pollution is one of the main problems that negatively affects both human and environmental health. Some HMs are important as micronutrients (Fe, Mo and Mn). Some toxic HMs are trace elements (e.g., Zn, Ni and Cu): they are required in numerous enzyme catalyzed or redox reactions and in electron transfer as well as playing structural function in nucleic acid metabolism (Zenk 1996). Although some of these metals are essential for life, when they are present in excess, they induce macroscopic effects concerning plant growth and leaf morphology (Rout and Das 2003; Todeschini et al. 2011), as well as root development, via large alterations of the mitotic activity (Liu et al. 2009) and via genotoxic damage. Their toxicity is mainly related to oxida-tive and (or) genotoxic mechanisms (Gamalero et al. 2009) . They can be taken up in different ways: Cu uptake and translocation by plants is strictly regulated, resulting in very low leaf concentrations (Todeschini et al. 2007), whereas Zn is primarily accumulated in leaves, with concentrations in the order of hundreds of ppm (Rosselli et al. 2003) ; by contrast, two ferns, Polypodium cambricum and Pteris vittata differently respond to Zn pollution: the latter accumulates relatively high metal concentrations in the fronds, causing a progressive deterioration of anatomical structures and functions, while the former shows a saturation mechanism in the root/rhyzome, which is evident already at non-toxic Zn concentrations, so ensuring the preservation of frond anatomy and function even under the exposure to sub-lethal Zn doses (Roccotiello et al. 2010).
Finally, several HMs have no known nutritional function, but are toxic for plants and microorganisms (Hg, Ag, Cd, Pb and U) (Fusconi et al. 2007) . Non-essential metals are usually toxic at lower concentrations than essential ones (Clemens 2006). In addition to these, there are also a number of toxic metalloids, including arsenic.
4.1 Heavy Metals and Essential
AM fungi are a direct link between soil and roots, and they can therefore be very important for HM availability and toxicity to plants. The AM symbiotic status changes the chemical composition of root exudates (Barea et al. 2002), thus quantitatively and qualitatively affecting the microbial populations in the rhizosphere and/or in the rhizo-plane (Offre et al. 2008). All these factors, alone or in combinations, can influence metal mobility or availability; nevertheless, the role of AM fungi in the uptake and in the transfer of HMs to the plant is not yet completely understood and the literature is conflicting. Some reports indicate that AM fungi enhance plant accumulation and the tolerance of HMs while reduced HM concentrations were instead found in a number of myc-orrhizal associations (Citterio et al. 2005; Lingua et al. 2008). However, the bulk of evidence seems to suggest a species specific effect of AM associations on root metal uptake.
Spores and pre-symbiotic hyphae are generally sensitive to HMs in the absence of the host plants (Gohre and Paszkowski 2006). Anyway, AM symbiosis has been observed in plants growing in soils containing HMs (Vallino et al. 2006; Bothe et al. 2010; Massa et al. 2010). Specifically adapted HM tolerant plants grow in HM-contaminated soils, and some have been reported to harbour AM
fungi though with a low degree of colonization (Bothe et al. 2010) . Zinc violets are clear examples of the exploitation of AM fungi by plants for heavy metal tolerance as their roots are intensely colonized (Tonin et al. 2001).
It is well known that AM fungi can alter metal concentrations and induce increased tolerance in plants, and different mechanisms have been proposed to explain this: the binding of metals to fungal cell walls and subsequently being accumulated in the vacuoles; sequestration by sidero-phores, deposited in the root apoplasm or in the soil, and possibly taken up by plant ferrisidero-phore receptors, or complexing of metals to met-allothioneins or phytochelatins synthesized by the fungus or the plant, as well as organic acids, amino acids, and metal-specific chaperons (shown for plants, but assumed for AM fungi) (Miransari, 2010) or metal chelation by fungal compounds, such as glomalin (Hildebrandt et al. 2007; Bedini et al. 2009).
Different plant species (and even different clones of the same species) respond in different ways to metal stress and to AM colonization. AM symbiosis can be either beneficial or ineffective, under metal pollution conditions, in relation to the host plant, and have proved to alleviate HM stress in the more sensitive species, probably by improving P nutrition (Todeschini et al. 2007; Lingua et al. 2008; Castiglione et al. 2009). An example is that of the results obtained by Lingua et al. )2008) on two poplar clones (Villafranca and Jean Pourtet), which responded differently to Zn and Cu addition and to AM symbiosis.
Under HM stress, unfavourable oxidative effects adversely influence plant growth. However, AMs are able to enhance the production of antioxidant enzymes, including glutathione S-transferase, superoxide dismutase, cytochrome P450 and thi-oredoxin, which can alleviate the stress of HMs (Hildebrandt et al. 2007). The enhanced tolerance of AM plants is related to the simultaneous regulation of AM stress genes and plant tolerance genes (Hildebrandt et al. 2007; Gamalero et al. 2009). The improvement in plant tolerance to HMs may be related to changes in gene expression as well as protein synthesis induced by the symbiosis itself. As an example, the Zn transporter MtZIP2
from Medicago truncatula is up-regulated by the presence of Zn and down-regulated by AM colonization, leading to a lower content of Zn within the host plant tissues (Burleigh et al. 2003). HM stress increases the transcript levels of some LeNramp2 (encoding a broad-range HM transporters) and Lemt1, Lemt3 and Lemt4 genes (encoding metal-lothioneins). On the other hand, AM fungal colonization results in the down-regulation of other HM transporter genes, presumably because the content of HM is lower in AM plants than in non-mycorrhizal ones. However, the down-regulation of plant mRNA (Ouziad et al. 2005; Burleigh et al. 2003) may be related to the "dilutive effect" of HM that occurs when plant growth improves as a result of AM colonization (Burleigh et al. 2003).
Among the HMs, cadmium (Cd) is of great environmental concern. Even in trace concentrations, this metal can cause serious health hazards to most living organisms of both the eukariotic and prokariotic kingdoms (Fusconi et al. 2007). Cd interacts with various functional groups of proteins, mainly with SH groups, which results in the alteration of the reactive centre of many enzymes, the reduction of the photosynthetic rate and chlorophyll content, alterations of membrane permeability, oxidative damage, increases in the cell polyamine pool (Sharma and Dietz 2006; Lingua et al. 2008) . In addition, Cd influences protein-protein and protein-DNA interactions (Freedman et al. 1988). Different mechanisms have been proposed to explain mycorrhiza alleviation of Cd stress (Rivera-Becerril et al. 2002; Aloui et al. 2009) . In pea, the buffering effect of AM symbiosis vis-à-vis Cd pollution has been linked to the modulation of root protein profiles. A protein band of about 30 kDa, a short-chain alcohol dehydrogenase (ADH), a UTP-1-phosphate uridylyltransferase (UDP-glucose pyrophosphorylase, UDPGP or UGPase), and a protein with a high homology to subunit B from a vacuolar H + -ATP synthase (V-ATPase), were all induced in pea plants by Cd treatment but down-regulated by inoculation with G. mosseae (Repetto et al. 2003). An increase in the highest ploidy nucleus populations which is possibly related to an increased transcription of genes, leading to the synthesis of the proteins involved in response/detoxification mechanisms to Cd toxicity, has also been observed (Repetto et al. 2007).
A more recent study has again shown that AM colonization in plants exposed to Cd stress can modulate the pattern of protein expression (Aloui et al. 2009). This study has shown that Cd-induced root proteome changes in M. truncatula plants are buffered by AM symbiosis. There is evidence of down accumulation of Cd stress-plant responsive proteins and the concomitant accumulation of mycorrhiza-related proteins putatively involved in reducing Cd toxicity in AM plants. More precisely, seven specifically up-accumulated proteins in AM roots, and whose expression was not modified upon Cd exposure, were detected and identified. All these proteins, which corresponded to a cyclophilin (s1), a guanine nucleotide-binding protein (s2), a ubiquitin car-boxyl-terminal hydrolase (s3), a thiazole biosyn-thetic enzyme (s4), an annexin (s8), a GST-like protein (s13) and an SAM synthase (s14), had functions putatively relevant in alleviating Cd toxicity (Aloui et al. 2009)
Increased accumulation has also been found at the proteome level of antioxidant enzymes and non-enzymatic antioxidants, and this accumulation is probably involved in the protection against oxidative damage, thus reducing Cd toxicity (Aloui et al. 2009).
Among the essential metals, Zn and Cu have been particularly investigated. In particular, the negative effects of Cu on plant development (Wang et al. 2002) and modifications in the protein profile have been described (Bona et al. 2007) . The Zn toxicity mechanisms take place through a number of biochemical processes, as described for Cd.
Defence mechanisms, based on antioxidant enzymes and on small antioxidant molecules, including proline, may protect the plant cell from ROS (Sharma and Dietz 2006). Although metal-induced proline accumulation in plant tissues has been observed (Andrade et al. 2009; Fariduddin et al. 2009), reports on the effects of mycorrhizal symbiosis in proline or soluble amino acid contents are scarce. Recently, however, high proline accumulation has been shown in response to Cu in
AM jack bean leaves. One of the proposed roles of proline is to reduce the level of free radicals generated as a result of toxicity in a similar manner to other molecules like glutathione, ascorbic acid or tocopherol (Andrade et al. 2010).
Other defence- or stress-related compounds are polyamines (PAs), which are present in all living organisms and which are essential in higher plants for growth and development (Bagni et al. 1993) . The up-regulation of the PA metabolism has been reported in response to several environmental stress conditions (Urano et al. 2003), including HMs, in a number of plant species (Pirintsos et al. 2004; Scoccianti et al. 2006). Modifications in the content of a precursor of PAs in plants (arginine) have also been observed. The arginine contents were the most striking difference in the amino acid composition of mycor-rhizal and non-mycorrhizal jack bean plants, grown in the presence of high Cu concentrations: the latter consistently exhibited arginine concentrations of between 35 and 50% of the total amino acid pool (Andrade et al. 2010). Lingua et al. (2008) observed levels of both free and conjugated PAs in poplar colonized by G. mosseae and grown in a Zn-polluted soil, similar to those that occur in plants grown without Zn pollution, suggesting that, in the presence of this AM fungus, given that the amount of zinc accumulated was very high, the toxicity of the metal was reduced. The same effect was not observed in G. intrara-dices inoculated plants, in which the growth inhibition due to zinc was not alleviated and the PA profile was altered in comparison with that of the controls.
Metalloids can also be extremely toxic for plants. An example is arsenic (As), which is heavily toxic for all living organisms and the environment, where it is released by natural and human activities. Arsenic is mainly present in the soil as arsenate (AsV) and arsenite (AsIII). AsV is an analogue of phosphate (Pi), and it competes with the latter for plant uptake by Pi transporters (Smith et al. 2010a). AsIII is more mobile and its uptake is believed to occur passively through membrane aquaporins (Ma et al. 2008). Once inside the plant, AsV can interfere with the phosphate metabolism, substituting it in the ATP, while AsIII, due to its high affinity for thiols, can inactivate several enzymes. Plants contrast As toxicity by reducing AsV to AsIII, and the latter is then eliminated from free cell circulation by complexation with thiolic peptides, such as GSH, and phytochelatins. The AsIII-thiol complexes can be segregated into the vacuole, by means of some glutathione-conjugated transporters (Smith et al. 2010a).
Plants from As contaminated soils are generally mycorrhizal (Cairney and Meharg 1999; Leung et al. 2007), indicating that fungal symbi-onts can evolve arsenic tolerance. As arsenate As(V) is an analogue of Pi, it might be expected that AMs would enhance the uptake of both. Because of the As(V)/Pi analogy, it could also be expected that the role of AMs in As tolerance would be different than that in HM tolerance (Smith et al. 2010a).
As mentioned in Sect. 2, it is well recognized that AM plants have two pathways through which Pi is absorbed from the soil solution (1) the direct pathway, in which Pi is taken up by roots as in non-AM plants, and can result in depletion of Pi in the soil solution close to the root system. The low concentration in the rhizosphere reduces subsequent Pi influx and may increase competition from As(V); (2) the AM pathway that involves uptake by the external mycelium and translocation to the plant through the fungal hyphae and transfer across the arbuscules (Grace et al. 2009; Smith and Read 2008). This pathway overcomes severe diffusion limitation of Pi uptake, as the external hyphae scavenge Pi at long distances from the roots (Smith et al. 2010a). AM and direct Pi uptake pathways are integrated and controlled, and, although data are not always consistent, it appears that the AM plants, compared to non-AM, take up relatively more Pi than As(V), and this results in a higher P/As ratio in AM plants (Smith et al. 2010a).
Some data have been reported on the influence of G. mosseae on As acquisition in Medicago sativa (Chen et al. 2007) , on the effect of AM
fungi of the Glomus spp. on biomass production and As accumulation in Pityrogramma calomelanos, Tagetes erecta and Melastoma malabathricum (Jankong and Visoottiviseth 2008) and on the As hyperaccumulation in the fern Pteris vittata (Liu et al. 2005; Trotta et al. 2006). Besides, Gonzalez-Chavez et al. (2002) reported that AM enhanced As resistance (through As exclusion) in Holcus lanatus.
In a study on As hyperaccumulation, in the absence and presence of G. mosseae and Gigaspora margarita in P. vittata, the expression in the fronds of the enzymes involved in photosynthesis and carbon fixation (i.e. RuBisCO, RuBisCO activase and ATP synthase) and sugar metabolism and bioenergetics (e.g. glyceralde-hyde-phosphate dehydrogenase and triosephos-phate isomerase) were especially affected. AM symbiosis also modulated the enzymes involved in the biosynthesis of S compounds (considering the role of some sulphurous compounds as non-toxic osmolytes or protective antioxidant agents) and some antioxidant enzymes, such as thiore-doxin peroxidase and glutathione peroxidase (Berta et al. 2008; Bona et al. 2010). Two proteins in the roots of P. vittata showed a specific expression pattern in response to As and mycor-rhization: glutamine synthetase, the key enzyme controlling the use of nitrogen inside the cells, and S-adenosylmethionine (SAM) synthase, which catalyzes SAM formation from methion-ine and ATP. These two proteins increased when non-AM plants were treated with As, while the same arsenic up-regulation did not occur in the presence of G. mosseae colonization; mycor-rhization alleviated the metalloid effect and a decrease in expression was detected when the AM plants treated with As were compared with the non-AM As plants (Bona et al. 2010). Since P. vittata is an arsenic hyperaccumulator, proteomic analysis did not detect any enzymes involved in ROS scavenging, and the only response to oxida-tive stress was the up-regulation of aldehyde dehydrogenase (ALDH). ALDHs have been considered as general detoxifying enzymes that eliminate toxic biogenic and xenobiotic aldehydes (Gao and Han 2009). The up-regulation of ALDH was detected in cadmium-exposed poplar plants
(Kieffer et al. 2008) and in aluminium-stressed tomato roots (Zhou et al. 2009).
Most of the information available in literature concerning organic pollutants examines the effect of AMs on polycyclic aromatic hydrocarbons (PAHs) in polluted soils (Leyval and Binet 1998; Binet et al. 2001; Joner et al. 2001; Liu and Dalpe 2009) . The phytotoxic effect of these contaminants is often due to their hydrophobicity, which compromises the uptake of water by plant roots. The uptake of nutrients, which are usually dissolved in the aqueous phase of the soil, is also considerably decreased (Volante et al. 2005). A study by Joner et al. I2001), on clover and ryegrass colonized by G. mosseae, has shown the beneficial effect of the mycorrhization on plant growth in a soil artificially polluted with anthracene, crysene and benzanthracene. The decrease in PAHs in the soil was mainly attributed to the enhanced nutrient uptake by AMF (Leyval et al. 2002), which leads to improved plant growth, which, in turn, may stimulate soil microbial activity (Rabie 2005; Liu and Dalpe 2009) . The contribution of the mycorrhizosphere to PAH biodegradation in the presence of ryegrass inoculated with G. mosseae has been studied by a number of authors (e.g. Corgie et al. 2006; Korade and Fulekar 2008; Liu and Dalpe 2009) .
While there is increasing interest in AM fungi as bioremediators for PAH polluted soil, little is known about their possible use for the reclamation of sites contaminated by aromatic hydrocarbons as benzene, toluene, ethylbenzene, meta-para- e ortho-xylene (BTEXs), which have mutagenic and carcinogenic properties as well as relatively high hydrosolubility. The effect of three AM fungal species on the persistence of BTEXs in artificially contaminated substrates was evaluated using leek as the host plant. A specifically designed mesocosm system, in which the internal air and substrate samples were analyzed for the BTEX content by means of gas chromatography, was used. Important reductions were observed in the BTEX concentration in the substrates in the presence of AM plants. The residual BTEX content ranged between almost total disappearance (<2%) and 40% of the original concentration (Volante et al. 2005).
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