Remediation of Heavy Metal Contaminated Soils: Phytoremediation ...

Critical Reviews in Environmental Science and Technology, 39:622–654, 2009 Copyright © Taylor & Francis Group, LLC ISSN: 1064-3389 print / 1547-6537 online DOI: 10.1080/10643380701798272

Remediation of Heavy Metal Contaminated Soils: Phytoremediation as a Potentially Promising Clean-Up Technology

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´ ANA P. G. C. MARQUES, ANTONIO O. S. S. RANGEL, and PAULA M. L. CASTRO Escola Superior de Biotecnologia, Universidade Católica Portuguesa, Rua Dr. António Bernardino de Almeida, 4200-072 Porto, Portugal

Increased soil pollution with heavy metals due to various human and natural activities has led to a growing need to address environmental contamination. Some remediation technologies have been developed to treat contaminated soil, but a biology-based technology, phytoremediation, is emerging. Phytoremediation includes phytovolatilization, phytostabilization, and phytoextraction using hyperaccumulator species or a chelate-enhancement strategy. To enhance phytoremediation as a viable strategy, microbiota from the rhizosphere can play an important role, but the use of genetic engineering can also increase the success of the technique. Here we review the key information on phytoremediation, addressing both potential and limitations, resulting from the research established on this topic. KEY WORDS: phytoremediation, heavy metals, rhizosphere community, genetic engineering

INTRODUCTION Because it is at the interface between the atmosphere and the earth’s crust, as well as being the substrate for natural and agricultural ecosystems, the soil is open to inputs of heavy metals from many sources (Alloway, 1990). Due to their immutable nature, metals are a group of pollutants of much concern. The danger of toxic metals is aggravated by their almost indefinite Address correspondence to Paula M. L. Castro, Escola Superior de Biotecnologia, Univer´ ´ sidade Catolica Portuguesa, Rua Dr. Antonio Bernardino de Almeida, 4200-072 Porto, Portugal; Tel.: + 351 22 558 00 59; Fax: + 351 22 509 03 51; E-mail: [email protected] 622


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persistence in the environment (Garbisu & Alkorta, 2001). Heavy metals cannot be destroyed but can only be transformed from one oxidation stage or organic complex to another. Pollution of the biosphere with toxic metals has accelerated dramatically since the beginning of the industrial revolution (Nriagu, 1979). The primary sources of this pollution are the burning of fossil fuels, mining and smelting of metalliferous ores, metallurgical industries, municipal wastes, fertilizers, pesticides, and sewage (Alloway, 1990). In addition to sites contaminated by human activity, natural mineral deposits containing particularly large quantities of heavy metals are present in many regions of the globe (Memon et al., 2000). In response to a growing need to address environmental contamination, many remediation technologies have been developed to treat contaminated soil (Riser-Roberts, 1998), mainly mechanically or physio-chemically based remediation methods. The most commonly used techniques are listed in Table 1. However, these technologies are usually expensive and soil disturbing, sometimes rendering the land useless as a medium for further activities such as plant growth. Consequently, a biologly-based emerging technology is gaining the attention of both soil remediation scientists and the general public—phytoremediation. Phytoremediation makes use of the naturally occurring processes by which plants and their microbial rhizosphere organisms sequester, degrade or immobilize pollutants for cleaning not only soils but also water matrices contaminated with heavy metals or organic pollutants (Pilon-Smits, 2005).

PHYTOREMEDIATION OF HEAVY METAL CONTAMINATED SOILS The basic idea that plants can be used for environmental remediation is very old and cannot be traced to any particular source (Raskin et al., 1997). Nevertheless, an interdisciplinary research approach combined with a series of fascinating scientific discoveries have allowed the development of this idea into an emerging technology, phytoremediation, which uses plants and their associated rhizospheric microorganisms to remove, degrade, or immobilize various contaminants from polluted soils, but also from sediments, groundwater, or surface water. Early research indicates that phytoremediation is a promising clean-up solution for a wide variety of contaminated sites, although it has its restrictions. Many of the limitations and advantages of phytoremediation are a direct result of the biological aspect of this type of treatment system (Singh et al., 2003). Plant-based remediation technologies can function with minimal maintenance after its establishment, as the costs of growing a crop are minimal compared to those of soil removal and replacement. Because biological processes are ultimately solar-driven, phytoremediation is on average ten-fold cheaper than engineering-based

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A. P. G. C. Marques et al.

TABLE 1. Technologies for remediation of heavy metal contaminated soil (Hamby, 1996; Khan et al., 2004; Mulligan et al., 2001; Ottosen and Jensen, 2005) Soil washing

Soil vapor extraction


Soil flushing

Solidification Stabilization/immobilization Vitrification

Electrokinetics Thermal desorption

Encapsulation

The process separates coarse soil (sand and gravel) from fine soil (silt and clay), where contaminants tend to bind and sorb. This soil fraction must be further treated with other technologies. Involves the installation of wells in the area of contamination. Vacuum is applied through the wells to evaporate the volatile constituents of the contaminated mass, which are subsequently withdrawn through an extraction well. Afterward, the extracted vapors are adequately treated. “Floods” contaminated soils with a solution that moves the contaminant to an area where they can be removed. Soil flushing is accomplished by passing an extraction fluid through soils using an injection or infiltration process. Recovered fluids with the absorbed contaminants may need further treatment. Encapsulates the waste materials in a monolithic solid of high structural integrity. Reduces the risk posed by a waste by converting the contaminant into a less soluble, immobile, and toxic form. Uses a powerful source of energy to “melt” soil at extremely high temperatures (1600–2000o C), immobilizing most inorganics into a chemically inert, stable glass product and destroying organic pollutants by pyrolysis. Removes contaminants from soil by application of an electric field. Contaminated soil is excavated, screened, and heated to temperatures such that the boiling point of the contaminants is reached, and they are released from the soil. The vaporized contaminants are often collected and treated by other means. Physical isolation and containment of the contaminated material. The impacted soils are isolated by low permeability caps or walls to limit the infiltration of precipitation.

remediation methods, such as soil excavation, soil washing or burning, or pump-and-treat systems (Glass, 1999). The fact that phytoremediation is carried out in situ contributes to its cost-effectiveness and may reduce exposure of the polluted substrate to humans, wildlife, and the environment (PilonSmits, 2005). However, it is not always the best solution to a contamination problem. The use of phytoremediation is limited by the climatic and geological conditions of the site to be cleaned, such as temperature, altitude, soil type, and the accessibility for agricultural equipment (Schmoger et al., 2000). On one hand, phytoremediation is far less disruptive to the environment, but, on the other, other problems may arise (e.g., contaminants can be accumulated in wood that can be used as fuel; the pollutants collected in leaves can be released again into the environment during litter fall; see Schmoger et al., 2000). One way to summarize many of the limitations of



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phytoremediation is that contaminants must be available to a plant and its root systems (i.e., the plants that mediate the clean-up have to be where the pollutant is and have to be able to act on it; see Pilon-Smits, 2005). Plants have a range of potential mechanisms at the cellular level that might be involved in the detoxification and tolerance to heavy metal stress. These all appear to be involved primarily in avoiding the build-up of toxic concentrations at sensitive sites within the cell, thus preventing the damaging effects (Hall, 2002). When metals accumulate in tissues they often cause toxicity, both directly by damaging cell structure and indirectly via replacement of other essential nutrients (Taiz & Zeiger, 2002). The strategies for avoiding heavy metal build-up are diverse (Hall, 2002; Marschner, 1995; Mejáre & Bülow, 2001). One way of avoiding metal accumulation can be the restriction of its movement to roots with the help of mycorrhizal fungi. As an example, Huang et al. (2002) reported an exclusion strategy of Zn in arbuscular mycorrhizal Zea mays. Reduction of the influx across plasma membrane as well as binding to cell wall and root exudates can also be possible avoidance strategies, as shown by Marques et al. (2007b) for retention of Zn in the cell walls of Solanum nigrum. Other mechanisms used by plants to avoid metal build-up can be the stimulation of the efflux of metals into the apoplast. As an example, Benaroya et al. (2004) demonstrated that this stimulation occurred, and that the apoplastic accumulation of Pb was very significant in Azolla filiculoides-, or the chelation in cytosol by various ligands. Ligands such as phytochelatins and metalotheins promote the detoxification abilities of metals in the plant, as shown for the engineered Nicotiana tabacum (Mejáre & Bülow, 2001). A possible avoidance strategy is that transport and accumulation of metals in the vacuole—the Ni hyperaccumulator Thlaspi goesingense—enhances its Ni tolerance by transporting and compartmentalizing most of the intracellular leaf Ni into the vacuole (Kramer et al., 2000) in order to restrict metal accumulation in areas of the cell where the occurrence of metals will be damaging to cell functions. Various phytoremediation strategies are possible for the remediation of heavy metal contaminated soils (Salt et al., 1998). Different phytotechnologies make use of different plant properties (Pilon-Smits, 2005). The main treatment streamlines are described in Figure 1 and can be regarded as one of the following:

r phytovolatilization: contaminants taken up by the roots pass through the plants to the leaves and are volatized through stomata, where gas exchange occurs (Vroblesky et al., 1999); r phytostabilization: plants are used to reduce the mobility and bioavailability of environmental pollutants (Vangronsveld et al., 1995); or r phytoextraction: plant roots take up contaminants and store them in stems and leaves (harvestable regions) (Kumar et al., 1995).


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FIGURE 1. Types of soil phytoremediation (adapted from Sing et al., 2002; Suresh and Ravishankar, 2004).

Phytovolatilization of Heavy Metals The chemical conversion of toxic elements into less toxic and volatile compounds is a possible strategy for detoxification of metal ion contaminants, resulting in the removal of specific harmful volatile elements (e.g., Hg and Se) from soil and plant foliage to the atmosphere (Raskin et al., 1997). For example, the volatilization of Se involves the assimilation of inorganic Se into the organic selenoaminoacids selenocysteine and selenomethionine. The latter can be biomethylated to form dimethylselenide, which is volatile and can be lost to the atmosphere (Terry et al., 2000). Brassica juncea was identified as a valuable plant for removing Se from soils (Bañuelos & Meek, 1990; Bañuelos et al., 1993) via Se volatilization. Mercury in its elemental form is also easily volatilized, as it is liquid at room temperature. However, because of its high reactivity, Hg in the environment exists mainly as a divalent cation Hg2+ ; bacteria can catalyze the reduction of the mercuric ion to elemental Hg and enhance the volatilization abilities of associated plants (Fox & Walsh, 1982). The volatilization of As has also been demonstrated for microorganisms but, as for Hg, this does not appear to be a significant process in plants (Rugh et al., 1996). The practicality of using plants able to volatilize metals for environmental remediation seems questionable, however: if a toxic volatile compound is emitted by plants during phytoremediation, the fate of the gas in the


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atmosphere should be determined as part of risk assessment (Pilon-Smits, 2005). Such a study was done for volatile Se and Hg, and it was reported that the pollutants were dispersed and diluted to such an extent that volatilization did not pose a threat (Lin et al., 2000; Meagher et al., 2000). Therefore, the work developed in this area points to a new environmental use of plants.


Phytoextraction of Heavy Metals The term “phytoextraction” mainly concerns the removal of heavy metals from soil by means of plant uptake. This technology is based on the capacity of the roots of plants to absorb, translocate, and concentrate toxic metals from soil to the aboveground harvestable plant tissues. The concentration process results in a reduction of the contaminated mass and also in the transfer of the metal from an aluminosilicate-based matrix (soil) to a carbon-based matrix (plants). The carbon in the plant material can be oxidized to carbon dioxide, further decreasing (and concentrating) the mass of material to be treated, disposed, or recycled (Blaylock & Huang, 2000). Metals can exist in the soil as discrete particles or can be associated with different soil components, including free metal ions and soluble metal compounds in the soil solution, exchangeable ions sorbed onto inorganic solid phase surfaces, nonexchangeable ions and precipitated or insoluble inorganic metal compounds (carbonates, phosphates, etc.), metals complexed by soluble or insoluble organic material, and metals bound in silicate minerals. Contamination events are usually indicated by discrete particles or high concentrations found in the first four components; the fifth component is indicative of background or indigenous soil metal concentrations (Ramos et al., 1994). The metals considered available for plant uptake are those that exist as soluble components in the soil solution or are easily desorbed or solubilized by root exudates or other components of the soil solution, these portions representing often only a small part of the total metal content of the soil (Blaylock & Huang, 2000). A major factor driving up the availability of metallic ions, solubility (Petrangeli et al., 2001), depends on various soil physicochemical factors, such as the degree of complexation with soluble ligands, the type and density of the charge on soil colloids, the reactive surface area (Norwell, 1984), and also the soil pH (Harter, 1983). Soil colloidal particles provide large interface and specific surface areas, which play an important role in regulating the concentrations of many trace elements and heavy metals in natural soils. In the soil, metal availability to plant roots decreases as the soil pH increases, as shown by Wang et al. (2005) for Thaspi caerulescens growing in a Cd- and Zn-contaminated soil. Additionally, the presence in the soil of particles with a high specific surface area may also reduce the soluble concentration of specific metals in the contaminated soil. However, this seems to be metal-specific—McBride and Martinez (2000) have


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reported that the solubility of Mo, As, Cd, Pb, and Cu was decreased by the addition of an amendment consisting of hydroxides with high reactive surface area, whereas the solubility of Zn and Ni remained unchanged. These physiochemical factors are dependent upon soil properties, including metal concentration and form, particle size distribution, quantity and reactivity of hydrous oxides, mineralogy, and degree of aeration and microbial activity (Magnuson et al., 2001). It is thus clear that the soil factors influencing the concentration, form, and plant availability of metals are highly complex. The supply of ions from the soil is controlled by the kinetics of solubilization of ions absorbed to its solid phase (Chaney et al., 1997). The limited bioavailability of various metallic ions, due to their low solubility in water and strong binding to soil particles, restricts their uptake/accumulation by plants. The plant itself can enhance metal bioavailability. For example, plants can extrude H+ via ATPases, which replace cations at soil cation exchange capacity (CEC) sites, making metal cations more bioavailable (Taiz & Zeiger, 2002). Plant species vary significantly in the ability of accumulating metals from contaminated soils, as a balance between the uptake of essential metal ions to maintain growth and development and the ability to protect sensitive cellular activity and structures from excessive levels of essential and non-essential metals is required (Garbisu & Alkorta, 2001). Generally, metals enter the plants primarily via absorption of the available metal ions from the soil solution into the root symplasm, driven by the electrical chemical potential gradient across the plasma membrane of root cells (Blaylock & Huang, 2000). Once inside the plant, most metals are too insoluble to move freely in the vascular system, so they usually form phosphate, sulphate, or carbonate precipitates. These precipitates are then immobilized in the apoplastic (extracellular)—cellular walls and intercellular spaces—and symplastic (intracellular) compartments, such as vacuoles. Unless the metal ion is transported as a non-cationic metal chelate, apoplastic transport is further limited by the high CEC of cell walls (Raskin et al., 1997). Some metals may be transported to the shoots by the transpiration stream complexed to organic acids, mainly citrate (Senden et al., 1992). Taking into account the features of the uptake and translocation mechanisms cited above, the ideal plant to be used in phytoextraction should have the following characteristics:

r r r r r

be tolerant to high levels of the metal; have a profuse root system; have a rapid growth rate; have the potential to produce a high biomass in the field; and accumulate high levels of the metal in the harvestable parts, as generally the harvestable portion of most plants is limited to the aboveground parts (although the roots of some crops may also be harvestable).


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However, the capacity of plants to concentrate metals in their harvestable tissues may be considered by some authors as a detrimental trait, as some plants can be directly (via plant consumption) or indirectly (via animal consumption and further contamination of the food chain) responsible for a portion of the dietary intake of toxic metals by humans, which can have damaging effects on human health (Brown et al., 1995b; Ow, 1996). Two different approaches have been generally proposed for the phytoextration of heavy metals, based on the different characteristics required for a plant to be useful for this application: the use of natural hyperaccumulator plants with exceptional metal-accumulating capacities, and the utilization of high biomass plants with a chemically (chelate) enhanced method of phytoextraction (Salt et al., 1998).

HYPERACCUMULATION

OF

HEAVY METALS

Some naturally occurring plants, termed metal hyperaccumulator plants, can accumulate in their harvestable tissues abnormally high levels of some metals. According to Reeves and Baker (2000), the term hyperaccumulator, describing a plant with a highly abnormal level of metal accumulation, appears to have been first applied by Jaffré et al. (1976), who reported high Ni concentrations in the New Caledonian plant Sebertia acuminate. The specific use of the term to denote a defined concentration (>1000 mg Ni/kg) was introduced by Brooks et al. (1977) in discussing Ni concentrations in species of Homalium and Hybanthus from various parts of the world— hyperaccumulation was used to describe accumulation of Ni to >1000 mg kg−1 in dry leaf tissue, because it was a level 100 to 1000 times higher than that normally found in plants growing on soils non-contaminated with Ni, and 10 to 100 times higher than that found for most other plants growing on Ni-rich soils. An attempt to give greater precision to the definition of hyperaccumulation was made by Reeves (1992) for Ni: “A hyperaccumulator of Ni is a plant in which a Ni concentration of at least 1000 mg kg −1 has been recorded in the dry matter of any aboveground tissue in at least one specimen growing in its natural habitat.” The definition of hyperaccumulation has extended to elements other than Ni. Brooks et al. (1980), Malaise et al. (1978), and Reeves and Brooks (1983) applied the 1000 mg kg−1 criterion for, respectively, Co, Cu, and Pb accumulation. For Zn, normally present at higher and more widely ranging concentrations, a 10,000 mg kg−1 threshold was suggested by Baker and Brooks (1989). The present definition of an hyperaccumulator is more extensive and should meet the following requirements:

r the concentration of the metal in the shoot must be higher than: 1.0% for Zn and Mn, 0.1% for Al, As, Se, Ni, Co, Cr, Cu, and Pb, and 0.01% for Cd (Baker and Brooks, 1989);

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r the shoot to root concentrations ratio must be invariably higher than 1


(McGrath & Zhao, 2003), indicating an efficient ability to transport metals from roots to shoots and, most likely, the existence of tolerance mechanisms to cope with high concentrations of metals; and r the shoot to soil concentration ratio must be higher than 1, indicating higher metal concentrations in the plant than in the soil, which emphasizes the degree of plant metal uptake (McGrath & Zhao, 2003).

Despite these requirements, hypertolerance seems to be the key property that makes hyperaccumulation possible. The apparent tolerance of plants to increasing levels of toxic elements can result from the exclusion of toxic elements or their metabolic tolerance to specific elements. The major mechanism in tolerant plant species appears to be compartmentalization of metal ions (i.e., sequestration in the vacuolar compartment or cell walls), which excludes them from cellular sites where processes such as cell division and respiration occur, thus providing an effective protective mechanism (Chaney et al., 1997). This is consistent in reports of histochemical localization of metals in several plants, namely A. maritime ssp. halleri (Heumann, 2002), Silene vulgaris (Harmens et al., 1993), Avicenia marina (MacFarlane & Burchett, 2002), and Solanum nigrum (Marques et al., 2007b). There are studies indicating that hypertolerance in known hyperaccumulators, such as the Znhyperaccumulator Thlaspi caerulescens, is due to an alteration of these mechanisms (Vázquez et al., 1994). The majority of hyperaccumulating species discovered so far are restricted to tropical areas (Baker & Brooks, 1989; Baker et al., 1993; Brooks et al., 1993; Ma et al., 2001). More than 430 taxa to date were reported to hyperaccumulate heavy metals, ranging from annual herbs to perennial shrubs and trees (Whiting et al., 2002), and some species, such as Sedum alfredii, show the capacity of accumulating two or more elements (He et al., 2002; Yang et al., 2002, 2004). For example, one of the best known Zn hyperaccumulators is Thlaspi caerulescens. While most plants exhibit toxicity symptoms at Zn concentrations of about 100 mg kg−1 , T. caerulescens was shown to accumulate up to 26,000 mg kg−1 without showing any damage to the plant (Brown et al., 1995b). In addition, this species extracted up to 22% of soil exchangeable Cd from a contaminated site (Gérard et al., 2000). Unfortunately, T. caerulescens can be described as a low biomass plant, as it typically produces 2–5 t ha−1 of shoot dry matter (McGrath et al., 2002). Indeed, hyperaccumulator species tend to grow slowly and to have low biomass yields (Chen et al., 2004; Raskin et al., 1997); the annual yields in biomass of hyperaccumulators are generally one to two orders of magnitude lower than those of robust crop plants (Ow, 1996). However, some authors (Chaney et al., 1997; McGrath & Zhao, 2003) keep defending that natural metal hyperaccumulator phenotype appears to be much more important

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Remediation of Heavy Metal Contaminated Soils TABLE 2. Abnormal metal accumulation levels registered for the aboveground sections of some plant species growing in contaminated soils


Plant species

Metal

Maximum concentration in the plant (mg kg−1 )

Ni Ni Ni Ni Ni Zn Zn Zn Zn Pb Pb Pb Mn Mn Se Se Cu Co

35720 10250 12780 7440 14800 20100 13620 30000 10000 20000 1600 13490 14000 33750 18200 14920 13700 10232

Psycotria vanhermanni Psycotria glomerata Psycotria osseana Garcinia bakeriana Streptanthus polygaloydes Thlaspi tatrense Cardaminopsis halleri Dichapetalum gelonioides Viola calaminaria Minuarti vernia Armeria maritime Agrostis tenuis Alyxia rubricalis Maytenus bureaviana Lecythis ollaria Astragalus racemosus Aeollanthus subacaulis Haumaniastrum robertii Adapted from Reeves and Baker (2000).

than high biomass yield when using plants to treat metal contaminated soils via phytoextraction. Nevertheless, Long et al. (2002) have recently reported a large biomass plant species, Sedum alfreddi, growing in some ancient Pb-Zn mine areas in Eastern China, which can also hyperaccumulate Zn. Examples of other plants accumulating metals and metalloids are diverse. Ma et al. (2001) reported the first As hyperaccumulator in terrestrial plants, the brake fern Pteris vittata, which can produce a relatively large biomass in favourable climates; it can accumulate up to 22,000 mg As kg−1 in the frond (dry weight), although phytotoxicity occurs when shoot arsenic level is higher than ca. 10,000 mg kg−1 (Wang et al., 2002). Hyperaccumulation of Mg has also been reported, namely in Vaccinium myrtillus (Denayer-De Smet, 1966). Other examples are listed in Table 2.

CHELATE-ASSISTED PHYTOEXTRACTION

OF

HEAVY METALS

Chelate-enhanced phytoextraction is based on the fact that the application of metal-chelating agents to a contaminated soil may enhance metal accumulation by plants (Garbisu & Alkorta, 2001). In the majority of cases, metal uptake into roots occurs from the aqueous phase (Lasat, 2002). In soil, some metals occur primarily as soluble or exchangeable, readily available form. Nevertheless, other metals occur as insoluble precipitates that are largely unavailable for plant uptake (Pitchel et al., 1999). Binding and immobilization within the soil matrix can significantly restrict the potential for soil phytoextraction (Lasat, 2002). In general, for any given heavy metal, only a fraction


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is bioavailable, and thus, potentially, it is only this fraction that can be taken up by the plants. More of the metal could be converted to the bioavailable fraction as it is gradually removed by the plant, but the extent to which this happens and the kinetics of such processes are soil-specific (Khan et al., 2000). The addition of chelating agents and the consequent formation of metal-chelate complexes prevents precipitation and sorption of the metals in the soil, thereby maintaining their availability for plant uptake. The addition of chelates to the soil can also bring metals into solution through desorption of sorbed species and dissolution of precipitated compounds (Norwell, 1984). Additionally, the application of certain chelates to the soil increases the translocation of heavy metals into the shoots (Blaylok et al., 1997). Luo et al. (2004) reported that the application of chelating agents increased the root-to-shoot ratios of the metals Cu, Pb, Zn, and Cd in Zea mays and Phaseolus vulgaris. Many studies concerning chelate-assisted phytoextraction have been reported, with the use of chelating agents such as CDTA (trans-1,2-diaminocyclohexane-N,N,N’,N’-tetraacetic acid), HEIDA [N-(2hydroxyethyl)iminodiacetic acid] (Chiu et al., 2006), HEDTA (Nhydroxyethylenediaminetriacetic acid) DTPA (diethylenetriaminepentaacetic acid) (Chiu et al., 2006; Huang et al., 1997), EGTA [ethyleneglycol-bis(β -aminoethyl ether),N,N,N , N-tetraacetic acid], or EDDHA (ethylenediaminedi-o-hydroxyphenylacetic acid) (Huang et al., 1997). Other components— namely, the malic (Chiu et al., 2006; Wu et al., 2004), citric (Chiu et al., 2006; Quartacci et al., 2006; Wu et al., 2004), and nitriloacetic acids (Chiu et al., 2006; Quartacci et al., 2006)—have been proposed as useful for enhanced-phytoextraction. Nevertheless, the majority of the reports indicate EDTA (ethylenediaminetetraacetic acid) and/or EDDS (SSethylenediaminedissucinic acid) as the main chelates applied in these studies, being those that more successfully improve heavy metal uptake by plants (Chen et al., 2004; Grˇcman et al., 2003; Huang et al., 1997; Lai & Chen, 2004, 2005; Luo et al. 2005, 2006; Marques et al, 2007c; Wu et al., 2004). As an example, Marques et al. (2007c) reported that the addition of EDTA to contaminated soils promoted an increase in the concentration of Zn accumulated by Solanum nigrum of up to 231% in the leaves, 93% in the stems, and 81% in the roots, while EDDS application enhanced the accumulation in leaves, stems, and roots up to 140, 124, and 104%, respectively, with the plants accumulating up to 8267 mg Zn kg−1 in the stems. Despite the possible usefulness of this technology, some concerns have been expressed regarding the potential inherent risk of leaching of metals to groundwater. The addition of chelates to a metal-contaminated matrix can increase the levels of the water extractable metals. For example, the application of EDTA to a metal-contaminated soil has been reported to increase significantly the concentrations of Cd, Zn, and Pb in the soil solution (Lai & Chen, 2004, 2005). High concentrations of heavy metals in soil solution could



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pose an environmental risk in the form of groundwater contamination (Lombi et al., 2001). A similar trend is also reported for EDDS. However, EDDS has a reduced risk of metal leaching in the application to metal contaminated soils (Grˇcman et al., 2003; Luo et al., 2006; Marques et al., 2007c). Marques et al. (2007c) reported that the Zn concentration in water-extracts of soils collected at the time of plant harvest were significantly increased by the addition of EDTA or EDDS, by up to 4.0- and 3.1-fold, respectively. EDDS thus seems a safer option when the application of chelate-assisted phytoextraction is considered. Additionally, synthetic chelating agents at high concentrations can also be toxic to plants. Chlorosis, necrosis, and impairment of plant growth have been reported for plants growing in soils amended with EDTA (Chen et al., 2004; Luo et al. 2005, 2006; Wu et al., 2004), EDDS (Luo et al., 2005, 2006), NTA (Kulli et al., 1999; Quartacci et al., 2006), and citric acid (Quartacci et al., 2006). Moreover, the presence of these chelates can reduce the occurrence and number of microorganisms in the rhizosphere (Marques et al., 2007c). Chelates, especially EDTA (Grˇcman et al., 2003), can greatly reduce the number of microbivorous nematodes (Römkens et al., 2002) and increase the stress index of microbial populations. Marques et al. (2007c) reported that the addition of EDDS, and especially of EDTA, to Zn-contaminated soils promoted a decrease in the root colonization of Solanum nigrum by AMF. Grˇcman et al. (2003) showed that EDTA and EDDS addition to a metalcontaminated soil induced fungi stress. Chen et al. (2004) indicated lower AMF colonization of Zea mays when grown in EDTA-treated soil. Sudová et al. (2007) also reported a reduced AMF colonization of Nicotiana tabacum when growing in EDDS-treated soil. The ability of plants to accumulate metals in the harvestable tissues is not the only factor influencing or determining the ability of phytoextraction, either by using hyperaccumulators or by adding chelating agents, to effectively remediate a metal-contaminated site. Other important factors such as the adequate selection of a site conductive to phytoextraction, as well as metal solubility and availability for uptake, should be taken into account (Blaylock & Huang, 2000). Soil clean-up criteria are also important in considering phytoextraction as a remedial option. The regulatory goals and timeline must be indicative that phytoremediation is an applicable solution. Therefore, treatability studies should be conducted to evaluate a particular site, especially the suitability of the soil for a phytoextraction treatment, with the evaluation of metal solubility as an essential step in the study (Blaylock & Huang, 2000).

Phytostabilization of Heavy Metals Some soils are so heavily contaminated that the use of plants for removing metals would not be an adequate approach and would take an unrealistic


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amount of time. Nevertheless, without some remediation effort, these contaminated areas remain barren, and the contaminated soil remains exposed to human and animal contact and to erosion that may carry contaminants off site. An alternative means of decreasing the environmental risk posed by these metal-contaminated soils may be the use of plants to stabilize the surface, thus reducing erosion and leaching to the soil deeper layers. This option is called phytostabilization, and considers the use of metal-tolerant plant species to immobilize heavy metals belowground, decreasing metal mobility and reducing the likelihood of metals entering into the food chain (Wong, 2003). Phytostabilization is hence used where phytoextraction is not possible or desirable (McGrath & Zhao, 2003). Additionally, it can also be applied at sites where regulatory or technical constraints delay the selection of the most appropriate techniques for site recovery as a provisional strategy to reduce environmental risk, by protecting barren contaminated areas from continuous erosion or leaching (Berti & Cunningham, 2000). This technique can indeed be adapted to a variety of sites and situations, with different conditions (e.g., soil pH, salinity, soil texture, metal levels, and contaminant types) through the careful selection not only of the appropriate plant species but also of the applied amendments (Berti & Cunningham, 2000). There are thus two major components in the phytostabilization process: the plant itself and the amendments added to the system. Plants play an important role in phytostabilization, not only by protecting the soil surface from human contact and rain impact with a dense canopy, but also by physically stabilizing the soil with dense root systems to prevent erosion. Plant roots also help to minimize water percolation through the soil, further reducing contaminant leaching (Berti & Cunningham, 2000). In addition, plant roots can also provide surfaces for sorption or precipitation of metal contaminants (Laperche et al., 1997). Consequently, the selection of the adequate plant species for phytostabilization should take into consideration the following:

r r r r

plants should be tolerant to the soil conditions; plants must grow quickly to set up a ground cover; plants should have dense rooting systems; plants must be easy to establish and to maintain under field conditions; and r plants must have a relatively long life or be able to self propagate (Berti & Cunningham, 2000). As phytostabilization is similar to establishing a meadow, soil amendments similar to those used in agriculture can also be applied and assume a role of great importance by helping to inactivate metal contaminants, preventing plant uptake, and reducing biological activity. Ideally, soil amendments



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should be easy to handle and to apply, safe to workers handling the amendment, non-toxic to the plants, easy to produce, and inexpensive—soil amendments that have little to no economic value, such as oyster shells for pH correction or manure for organic matter supplementation (Marques et al., 2007a), are preferred to more expensive materials (Berti & Cunningham, 2000). Attention should also be given to the capacity of the amendments to reduce the leaching of metals, as this could be an important advantage in an in situ stabilization process and play an important role in groundwater protection and reduction of metal dispersion (Ruttens et al., 2006a). Marques et al. (2007a) have shown that the sole application of organic matter amendments, such as manure or compost, to metal contaminated soil led to a significant reduction in the amount of Zn leached through the soil; in combination with plants, the reduction in metal percolation ascended up to 80%. A range of organic and inorganic compounds (Adriano et al., 2004), such as lime, phosphate, and other low economical value organic materials like biosolids, litter, compost, and manure, can be used. Liming has been considered as an important management tool in reducing the toxicity of metals in soils (Gray et al., 2006; Madejón et al., 2006). There is conclusive evidence for the mitigative value of both water-soluble (e.g., diammonium phosphate) and water-insoluble (e.g., apatite) phosphate to immobilize some metals in soils, thereby reducing their bioavailability for plant uptake (Brown et al., 1995a). Phosphate enhances the immobilization of metals in soils through various processes, including direct metal adsorption, phosphate anion-induced metal adsorption, and precipitation of metals with solution phosphate as metal phosphates (Adriano et al., 2004). In fact, Bolan et al. (2003) reported that the sole application of lime or phosphate is effective in reducing Cd in contaminated soils. The use of organic amendments, such as manure (Chiu et al., 2006; Clemente et al., 2006; Marques et al., 2007a; Walker et al., 2004; Ye et al., 1999), compost (Cao & Ma, 2004; Clemente et al., 2006; Marques et al., 2007a), and other bio-wastes ´ et al., 2006) is a standing practice used for restora(Karaca, 2004; Madejon tion of contaminated sites (Sopper, 1993). As examples, Walker et al. (2004) reported lower Zn tissue concentration in Chenopodium album L. plants when grown in compost or manure amended soils; Marques et al. (2007a) showed that the addition of manure or compost to the soils induced reductions in the Zn accumulation of Solanum nigrum of up to 80 and 40% while enhancing plant biomass yields; Ye et al. (1999) observed that Trifolium repens tended to accumulate less Pb in the shoots when manure was added to the growing matrix. In fact, organic matter amendments are among the most promising additives, especially due to their low commercial cost and consequent added value of their application for soil remediation purposes. Their application provides organic matter to improve soil physical properties, water infiltration, and water-holding capacity. They also contain


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essential nutrients for plant growth. Immobilization of metals by such amendments is achieved through adsorption, complexation, and redox reactions (Adriano et al., 2004)—organic matter makes strong complexes with heavy metals (Krogstad, 1983). The addition of organic amendments has often been shown to increase the CEC of soils (Marques et al., 2007a), increasing cation adsorption caused by the dissociation of H+ from the functional groups in organic matter (Zhu et al., 1991). The presence of phosphates, Al compounds, and other inorganic minerals in some organic amendments is also believed to be responsible for the retention of metals (Adriano et al., 2004). Additionally, amendment with organic matter and its resulting degradation may change the soil pH and thereby indirectly affect the bioavailability of metals (Karaca, 2004), as it is well known that metal solubility is greatly determined by the pH (Yoo & James, 2002). The research in soil amelioration using metal immobilizing amendments is now also focusing on the application of other type of compounds, such as cyclonic ashes (Ruttens et al., 2006b), calcium carbonate (Lee et al., 2004), zeolites (Chlopecka & Adriano, 1996), steel shots (Ruttens et al., 2006b), beringite (Mench et al., 1994), red mud ´ et (also known as bauxite residue; Gray et al., 2006), or leonardite (Madejon al., 2006), with positive effects on the reduction of soluble concentrations of heavy metals in soils.

TOOLS TO IMPROVE THE POTENTIAL OF PHYTOREMEDIATION OF HEAVY METALS The Role of the Microbial Community of the Rhizosphere in the Phytoremediation of Heavy Metals Metal uptake by plants can be influenced by soil microorganisms that associate with the plant roots to form the rhizosphere community (Shilev et al., 2001). It is well known that mycorrhizal fungi are a major component of the rhizosphere and form mutualistic associations with most plant species (Azcón-Aguillar & Barea, 1992). In all, 90–95% of all land plants form some type of mycorrhizal associations so that the symbiotic association, the mycorrhiza, seems to be the chief organ of nutrient uptake in the majority of plants (Bago et al., 2000; Entry et al., 2002). Of the existent mycorrhizal associations—ectomycorrhizas, arbuscular mycorrhizas, ericaceous mycorrhizas, and orchid mycorrhizas (Entry et al., 2000)—the arbuscular mycorrhizas (AM) associations between arbuscular mycorrhizal fungi (AMF) and the roots of terrestrial plant species are by far the most widespread (Smith & Read, 1997). Arbuscular mycorrhizal fungi can benefit plants in numerous ways. Benefits to plants include improved nutrition (Clark & Zeto, 2000), through extensive extraradical hyphal networks, which explore the soil, absorb nutrients,



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and translocate them to the roots (Giovannetti et al., 2002), and root system modifications, generally resulting in a more extensive length and increased branching and therefore in a more efficient nutrient absorption (Berta et al., 2002). The increase in the uptake of inorganic P (Harrison & van Buuren, 1995; Smith & Read, 1997), K, Ca, S, Cu, Zn, Mg, Co, Ni, and N (Koide, 1991; Marschner & Dell, 1994; Smith & Read, 1997) are well-documented nutritional beneficial effects of AMF. In addition, the arbuscular mycorrhizal symbiotic status changes the chemical composition of root exudates (Laheurte et al., 1990) and influences soil pH (Li et al., 1991), thus quantitatively affecting ´ the microbial populations in the rhizosphere (Azcon-Aguilar & Barea, 1992; ´ Barea, 1997), protecting against soil-borne plant pathogens (Azcon-Aguilar & Barea, 1996; De la Peña et al., 2006), and improving soil structure (Rillig & Mummey, 2006). Other benefits include protection against insect herbivores (Gange & Brown, 2002), hormone regulation (Ludwig-Müller, 2000), and drought tolerance (Augé, 2001; Ruiz-Lozano et al., 2001). Arbuscular mycorrhizal fungi have also been shown to enhance plant tolerance to biotic and abiotic stresses, including the presence of high levels of heavy metals (Leyval et al., 2002): as they are a direct link between soil and roots, they can be very important for heavy metal availability and toxicity to plants (Leyval et al., 1997). It has been reported that mycorrhizal fungi can impact plant uptake or translocation of soil metals (Khan et al., 2000). When the host is exposed to metal stress, the role of AMF in the plant stress response is variable. Some studies indicate reduced metal concentrations in plants due to mycorrhizal colonization (Heggo et al., 1990; Jentschke et al., 1998). Huang et al. (2002) reported an exclusion strategy, showing lower Zn accumulation by AMF colonized Zea mays. However, other reports indicate enhanced metal uptake and accumulation in plants due to AMF colonization (Ahonen-Jonnarth & Finlay, 2001; Jamal et al., 2002; Joner & Leyval, 2001; Marques et al., 2006, 2007b). Citterio et al. (2005) have shown an enhanced growth and metal root to stem translocation on Cannabis sativa plants inoculated with the AMF Glomus mosseae, while Chen et al. (2005) observed that a mixed AMF inoculum enhanced Pb uptake and growth of Kummerowia striata, Ixeris denticulate, and Echinochloa crusgalli var mitis-, even resulting in metal levels toxic to plants (Weissenhorn & Leyval, 1995). Other reports indicate that both effects can occur or even show no effects exerted by AMF on the contaminant uptake and accumulation in the host plants (Joner et al., 2004). The bulk of evidence seems thus to suggest a species-specific effect of AM associations on plant metal uptake and accumulation. As examples, Marques et al. (2007c) have shown that the inoculation with the AMF G. intraradices or G. claroideum protected the host plant Solanum nigrum of excessive Zn, which was translated in a decrease in metal accumulation in AMF inoculated plants, whereas at lower Zn levels in the growing matrix, there was an increase in the metal accumulation. Diaz et al. (1996) have also reported similar tendencies for the uptake of Zn and


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Pb by Lygeum spartum and Anthylis cytisoides inoculated with G. mosseae in soils with different levels of theses metals: at low doses, mycorrhizal plants had equal or higher Zn or Pb concentrations than non-inoculated controls; at higher doses, however, metal concentrations in the plants inoculated with G. mosseae were lower than those found in the corresponding controls. Plant growth-promoting bacteria (PGPR) communities naturally existing in the rhizosphere can also be an important tool in the decontamination of metal-contaminated soils through plant use. PGPR can be divided into two groups according to their relationship with the plants: symbiotic bacteria and free-living rhizobacteria (Khan, 2005). The enhancement of crop plant growth using PGPR is documented (Reed & Glick, 2004); more recently, these organisms have been used to reduce plant stress associated with phytoremediation strategies for metal contaminated soils (Reed & Glick, 2005). The PGPR are able to enhance plant growth through various mechanisms, such as reduction of ethylene production (thus allowing plants to develop longer roots and better establish during early stages of growth; see Glick et al., 1998), nitrogen fixation, specific enzymatic activity (Khan, 2005), supply of bioavailable phosphorous and other trace elements for plant uptake, and production of phytohormones such as auxins, cytokinins, and gibberelins (Glick et al., 1995). These microorganisms can also produce antibiotic and other pathogen-depressing substances such as siderophores and chelating agents that protect plants from diseases (Kamnev & Lelie, 2000) and can also increase plant tolerance to flooding (Grichko & Glick, 2001), salt stress (Mayak et al., 2004a), and water deprivation (Mayak et al., 2004b). Plant growth promoting bacteria have also shown to reduce plant stress at metal exposure. Brassica napus has shown reduced accumulation of Cu when inoculated with Pseudomonas putida UW4 (Reed & Glick, 2005; Reed et al., 2005) and lower Ni toxicity when in the presence of the bacteria Kluyvera ascorbata SUD165 (Burd et al., 1998). Nevertheless, other studies indicate a PGPR-driven increase in the availability of heavy metals in soil, thus enhancing metal accumulation by plants, as reported for Zn accumulation by Thlaspi caerulescens (Whiting et al., 2001) and uptake of Ni by Alyssum murale and Thlaspi goesingense (Abou-Shanab et al., 2003; Idris et al., 2004) and of Se by Brassica juncea (De Souza et al., 1999). Although the employment of PGPR is potentially important in phytoremediation experiments, research in this area is not as extensive as for the AMF use, and further investigation is needed to better understand the prospects of PGPR application in such strategies. Different microorganisms may play assorted roles in plant growth and/or metal tolerance via different mechanisms, so it can be beneficial for the design of a phytoremediation plan to select appropriate multifunctional microbial combinations, which may include AMF and PGPR. Further examples of the contribution of selected microorganisms in phytoremediation—namely, bacteria, fungi, and a combination of the two—are described in Table 3.

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Ni, Cd

Cu Pb

Metylobacterium oryzal CBMB20 Burkohderia sp. OBMB40

Glomus mosseae

Glomus intraradices

Glomus manihotis

Glomus intraradices PH5

Lycopersicon esculentum

Trifolium pratense

Helianthus annuus

Zea mays

Trifolium repens

Glomus mosseae Brevibacillus sp. (B-I)

Cd

Bacillus sp. RJ16

Zn

Cr

Zn

Ni

Pseudomonas putida UW4

Brassica napus

- Stimulation of plant growth - Increase of Ni accumulation by plants - Stimulation of shoot biomass formation - Increase of tolerance to Ni - Stimulation of growth - Increase of metal accumulation in shoot tissues - Stimulation of plant growth - Decrease of metal accumulation in shoot and root tissues - Decrease of metal accumulation in shoot and root tissues - Increase of metal accumulation in roots - Increase of metal accumulation in root and shoot tissues - Increase of metal accumulation in root and shoot tissues - Stimulation of plant growth - Decrease of metal accumulation

Ni

Vivas et al., 2006

Malcova et al., 2003

Liao et al., 2003

Bi et al., 2003 Chen et al., 2003 Davies et al., 2001

Madhaiyan et al., 2007

Sheng and Xia, 2006

Farwell et al., 2007

Zaidi et al., 2006

Belimov et al., 2005

- Stimulation of root elongation

Wu et al., 2006

Reference

Cd

- Stimulation of plant growth - Protection of plant from metal toxicity

Effects

Rajkuman et al., 2006

Brassica juncea

Pb, Zn

Heavy metals

Cr

Microorganisms

Azotobacter chroococum HKN-5 Bacillus megaterium HKP-1 Bacillus mucilaginosus HKK-1 Pseudomonas sp. PsA4 Bacillus sp. Ba32 Variovorax paradoxus Rodococcus sp. Flavobacterium sp. Bacillus subtilis SJ-101

Plants

TABLE 3. Examples of phytoremediation experiments using bacteria, AMF, and the conjugation of both microorganisms


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Genetic Engineering of Plants for the Phytoremediation of Heavy Metals Phytoremediation by natural plant species can be limited in several ways. These limitations could be overcome by using conventional plant breeding practices; however, conventional approaches can take decades. Genetic engineering, on the other hand, has the potential to produce plant populations with superior traits for phytoremediation in a relatively short time and even transfer genes form organism that can not be crossed by conventional breeding methods (Berken et al., 2002). Many genes are involved in metal uptake, translocation, and sequestration, and the transfer of any of these genes into candidate plants is a possible strategy for genetic engineering of plants for improved phytoremediation traits (Eapen & D’Souza, 2005). In genetic engineering of plants, a foreign piece of DNA is stably inserted into the genome of a cell, which is regenerated into a mature transgenic plant; the piece of DNA can come from any organism, from bacteria to mammals, or other plants. When the transformed plant is propagated, the foreign gene is inherited by its offspring (Pilon-Smits & Pilon, 2002). The ideal plant species to engineer for phytoremediation purposes is one that has high biomass production, is sufficiently hardy and competitive in the climate where it is to be used, has a good phytoremediation capacity to start with (Pilon-Smits & Pilon, 2002), has the ability to accumulate metals preferably in the aboveground parts, has a widespread, highly branched root system, is easy to harvest, and is amicable for genetic transformation (Eapen & D’Souza, 2005). Preferably, crop plants should not be used. Classic genetic studies have shown that only a few genes (up to three) are responsible for metal tolerance (Macnair et al., 2000). According to Eapen and D’Souza (2005), the possible areas of genetic manipulation as follows:

r metallotioneins: the transfer of human metallotionein gene in tobacco resulted in plants with enhanced Cd tolerance (Misra & Gedamu, 1989), and pea metallotionein gene transfer to Arabidopsis thaliana resulted in increased Cu accumulation (Evans et al., 1992); r phytochelatins: transgenic Brassica juncea overexpressing different enzymes involved in phytochelatin synthesis were shown to extract more Cd, Cr, Cu, Pb, and Zn than wild plants (Zhu et al., 1999a, 1999b); r organic acids: the overexpression of citrate synthase has shown to promote enhanced Al tolerance; r phytosiderophores: the overexpression of nicotianamine aminotransferase (NAAT) in rice resulted in the overproduction of the iron-chelator deoxymugineic acid, a phytosiderophore, and consequently promoted a more efficient growth in iron-deficient soils (Takahashi et al., 2001);


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r ferritin: the overexpression of the iron-binding protein ferritin has shown r r


r r

to increase up to 1.3-fold higher the iron level in tobacco leaves (Goto et al., 1998); metal transporters: transfer of Zn transporter-ZAT gene from Thlaspi goesingense to Arabidopsis thaliana resulted in two-fold higher Zn accumulation in its roots (Van der Zaal et al., 1999); alteration of metabolic pathways: transfer of Escherichia coli ars C and γ -ECS genes to Arabidopsis plants resulted in individuals that could transport oxyanion arsenate to aboveground tissues, reduce to arsenite, and sequester it to thiol peptide complexes (Dhankher et al., 2002); alteration of oxidative stress mechanisms: overexpression of glutathioneS-transferase and peroxidase in Arabidopsis plants resulted in enhanced Al tolerance (Ezaki et al., 2000); and alteration in biomass: increasing phytohormones synthesis can increase biomass of transgenic plants, as reported by Eriksson et al. (2000) for trees with genetically induced increase in giberellin biosynthesis that presented enhanced growth and biomass production.

Although no practical applications of transgenic plants are reported and the theoretical risk of escape of the genes from the transgenic plants has been calculated as negligible (Meagher et al., 2000), risk assessment of any use of these transgenic species should be carefully undertaken before any field testing or further application is to be planned (Wolfenbarger & Phifer, 2000). Some of the possible risks involved are biological transformation of metals into forms that are more bioavailable, enhanced exposure of wildlife and humans to metals, uncontrolled spread of transgenic plants due to higher fitness or general weedy nature, and/or uncontrolled spread of the transgenic plants by interbreeding with populations of wild relatives (Pilon-Smits & Pilon, 2002). These risks have to be assessed and weighed not only against the benefits of the technique, but also against the risks of doing nothing or using other methods. If the adequate prevention measures are taken, these new developments in plant genetic engineering may lead to fruitful applications in environmental cleanup.

CONCLUSIONS Phytoremediation is emerging as a bio-based and low-cost alternative in the cleanup of heavy metal-contaminated soils. The application of a vegetation cover can limit the local effects and the spreading of the contamination, or even remove via phytoextraction or phytovolatilization the metals from the polluted soil. The future of this technique is still mainly in the research phase, and the optimization and greater understanding of the process by which plants absorb, translocate, and metabolize heavy metals needs to be

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addressed. There is still much fundamental and applied and field research needed. The potential role of both free living and symbiotic soil microbes in the rhizosphere of plants growing in metal-contaminated soils in enhancing the phytoremediation process can be an important tool to support the technology. The outcome of undergoing genetic engineering investigation concerning plants applicable in phytoremediation may also lead to a better understanding of metal metabolism in plants, which can result in important contributions for the implementation of phytoremediation as a feasible soil remediation technology. A multidisciplinary research effort that integrates the work of plant biologist, soil chemists, microbiologists, geneticists, and environmental engineers thus seems essential for the success of phytoremediation as a soil cleanup technology.

ACKNOWLEDGMENTS This work was supported by Fundaça˜ o para a Ciência e a Tecnologia and Fundo Social Europeu (III Quadro Comunitário de apoio), a research grant of Ana Marques (SFRH/BPD/34585/2007), and funded by the Project MICOMETA - POCI/AMB/60131/2004, financed by Medida V.4-Acça˜ o V.4.1 of Programa Operacional Ciência e Inovaçaõ 2010 (Fundaça˜ o para a Ciência e Tecnologia).

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