Phytoextraction of Trace Metals: Principles and Applications



Trace elements (TEs) occur at minor concentration (>1 g kg−1) in the organisms, and some are essential nutrients (Cr, Mn, Fe, Co, Ni, Cu, Zn, Mo, B, and Cl) for animals and plants. As a consequence of human activities such as industrial production, mining, transport, and agriculture, they are released in the environment at high concentrations. TEs can accumulate over time under specific environmental conditions, thus becoming environmental contaminants (Cs, Cr, W, U, Cd, Hg, Tl, Pb, Sn, As, Sb, Se). The environmental risk of TEs is associated with the mobility and bioavailability of the metals more than their total concentration. When they become environmentally mobile and move between media (i.e. soil to water), they can enter the food chain by being taken up by plants and animals. TEs cannot be degraded or broken down and at high concentration are toxic to organisms and tend to bioaccumulate in the environment. For example, selenium (Se) is a naturally occurring element with a wide distribution in almost all parent materials on Earth. At low concentration, Se is an essential nutrient but at high concentration is toxic. In the western side of the San Joaquin Valley in California, soils contain significant quantities of soluble mineral salts and trace elements such as Se and boron (B) that have been leached into shallow groundwater and/or drainage waters because of irrigation practices at Kesterson Reservoir in California. Soluble Se bioaccumulated in the avian food chain and resulted in an environmental disaster with high mortality and reproduction failure of migratory birds (Letey et al. 2002; Ohlendorf et al. 1986).


Life Cycle Assessment Cover Crop Serpentine Soil Hyperaccumulator Plant Willow Plantation 
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1 Introduction

Trace elements (TEs) occur at minor concentration (>1 g kg−1) in the organisms, and some are essential nutrients (Cr, Mn, Fe, Co, Ni, Cu, Zn, Mo, B, and Cl) for animals and plants. As a consequence of human activities such as industrial production, mining, transport, and agriculture, they are released in the environment at high concentrations. TEs can accumulate over time under specific environmental conditions, thus becoming environmental contaminants (Cs, Cr, W, U, Cd, Hg, Tl, Pb, Sn, As, Sb, Se). The environmental risk of TEs is associated with the mobility and bioavailability of the metals more than their total concentration. When they become environmentally mobile and move between media (i.e. soil to water), they can enter the food chain by being taken up by plants and animals. TEs cannot be degraded or broken down and at high concentration are toxic to organisms and tend to bioaccumulate in the environment. For example, selenium (Se) is a naturally occurring element with a wide distribution in almost all parent materials on Earth. At low concentration, Se is an essential nutrient but at high concentration is toxic. In the western side of the San Joaquin Valley in California, soils contain significant quantities of soluble mineral salts and trace elements such as Se and boron (B) that have been leached into shallow groundwater and/or drainage waters because of irrigation practices. Soluble Se bioaccumulated in the avian food chain and resulted in an environmental disaster with high mortality and reproduction failure of migratory birds (Letey et al. 2002; Ohlendorf et al. 1986).

According to the German Advisory Council on Global Change, 22 million hectares of land are contaminated with TEs worldwide (GACGC 1994). Specifically, a comprehensive inventory of global soil contamination is lacking ( For example, at the European level, it is unclear what threshold values should be used to classify a soil as polluted and with regard to TEs at which locations can high natural background values be expected (Morvan et al. 2008).

For the government (i.e. USA and EU) to act, usually the contamination must be severe enough to cause barren soils, limit crop production, and cause groundwater contamination and unsafe living conditions. Lack of global inventory of soil contamination and of consequent public awareness of the health risks is one of the reasons why governments fail to require and fund remediation in areas that are of low economic value such as marginal agricultural land, former mining areas, landfills, and postindustrial sites. Knowledge of the extent and the location of the contaminated areas and assessment of the environmental risk associated with the contamination would trigger reaction and concerns in the local community near or close the contaminated area. Ecosystem disruption is not enough to trigger action by the government. In the majority of cases when contamination is ‘mild’ and does not cause evident unsafe living conditions, the economic and social responsibility of cleaning up the site or setting it aside from food production is in the hands of the landowner.

Most commercial soil remediation strategies rely on the use of engineered methods such as soil excavation (dig and haul), pump-and-treat systems, soil washing, leaching, and soil capping. These methods are effective and are usually applied when the contaminated area has commercial value (i.e. growth of urban areas) and the soil needs to be remediated in short time. However, these methods are prohibitively expensive and destructive of the soil ecology and fertility. They also generate high amounts of waste that needs to be disposed of (Conesa et al. 2012).

Phytotechnologies represent a green alternative to conventional remediation methods because they are based on the use of solar-driven biological processes to remediate or reduce the risk of contamination. Phytotechnologies are low cost and gain wide public acceptance because of the environmental benefit they provide to the remediated areas, i.e. revegetation, decreased formation of soil dust, reduced soil erosion, carbon sequestration, etc. They comprise a number of different methods that aim at removing, extracting, transforming, and immobilising contaminants using plants and their associated root microorganism (Pilon-Smits 2005).

A number of authors have described in detail the different phytoremediation technologies and their advantages and disadvantages and their applications (Chaney et al. 2007; Dickinson et al. 2009; Pilon-Smits 2005). In this chapter, we focus on cost-effective phytoextraction and on the economic viability and environmental benefits of phytoextraction as successful commercial phytotechnology.

2 Limits of Phytoextraction

Phytoextraction is the removal of TEs from the soil by growing plants that have the ability to take up TEs in their above-ground biomass at high concentrations. Harvest of the TE-rich biomass and multiyear growth cycles of the plants may allow the removal of the TEs from the soil to a concentration level acceptable by the environmental regulatory authority. The removed biomass that has no value as bio-ore is usually incinerated, composted, or digested to reduce the volume and disposed in landfills or in hazardous waste landfills. Chaney et al. (2010) note that disposal of biomass represent only a disposal cost rather than a problem in most cases except for radionuclides.

The development of phytoextraction has begun a few decades ago with the discovery of hyperaccumulator plants by pioneering studies by Robert Brooks (Brooks et al. 1977), Alan Baker (1981), and Rufus Chaney (1983). TE concentrations in the shoots of hyperaccumulator are about 100–1,000 times higher than that found in normal plants under most circumstances. Specifically, the concentration values (in their dried foliage) to define a hyperaccumulator are as follows: 100 mg kg−1 of Cd, Se, and Tl; 300 mg kg−1 of Co, Cu, and Cr; 1,000 mg kg−1 of Ni, Pb, and As; 3,000 mg kg−1 of Zn; and 10,000 mg kg−1 of Mn when grown in its natural habitat (van der Ent et al. 2013). Based on this criteria, more than 500 plant taxa have been cited as ‘hyperaccumulators’ of one or more elements including As, Co, Cd, Cu, Mn, Ni, Pb, Se, Tl, and Zn. At present, the approximate number of hyperaccumulators for various elements is as follows: Ni (450), Cu (32), Co (30), Se (20), Pb (14), Zn (12), Mn (12), As (5), Cd (2), and Tl (2) (van der Ent et al. 2013). Nearly 25 % of hyperaccumulators belong to the family of Brassicaceae and the genera Thlaspi and Alyssum.

The benefit and adaptive advantage of hyperaccumulators have not yet been explained, but a variety of hypotheses have been proposed. The most popular one is the ‘elemental defence’ hypothesis (Boyd 2007) which suggests that the high amounts of TEs in the shoots of hyperaccumulator plants render them less palatable to pathogen and herbivores, thus reducing the possibility of their attacks and stimulating a defence against them. Despite the numerous studies supporting this hypothesis, more information is required since only a few taxa and a limited number of TEs have been analysed.

One of the key factors for the success of phytoextraction is the high uptake of TEs in the above-ground biomass, and that is why hyperaccumulators have been widely studied for the commercial development of phytoextraction technologies. The ability of hyperaccumulators to take up high amount of TEs depends on two unique traits, such as the constitutive up-regulation of transmembrane metal transporter, which confers faster and effective root-to-shoot translocation of TEs, and hypertolerance, which is mostly dependent on effective detoxification and storage of TEs in leaf cell vacuoles. Another very important characteristic of hyperaccumulators relative to normal crops is a greater ability to take up TEs from the soil.

It has been shown that hyperaccumulators absorb metals from the same labile pool in soils as normal plant species; however, crop plants cannot absorb high amounts of TEs to support phytoextraction. Despite the recent advancements in understanding the physiological mechanisms of metal uptake and translocation to shoot (Milner and Kochian 2008), no mechanisms are yet known where hyperaccumulator plants can attack the non-labile pool of metals in soils. Centofanti et al. (2012) investigated whether the Ni hyperaccumulator Alyssum corsicum possess distinct extraction mechanisms for different Ni species present in soils, as they have different solubility and potential bioavailability to roots. Their study showed that Ni uptake is related to Ni solubility and plant transpiration rate. The authors also suggested that Ni uptake is driven by convection, which depends on the initial concentration of Ni in solution and the plant transpiration rate. Metals enter the roots via uptake of the soil solution, which is then transferred to the stems and leaves and lost via transpiration. High metal concentration in the roots can result from plant water uptake inducing metal migration via mass flow (Zhao et al. 2000). Hyperaccumulators have the ability to translocate the absorbed metals from the roots to the shoot and store them in the leaf cell vacuoles (Broadhurst et al. 2004, 2009; Tappero et al. 2007).

The ability of plants, and of trees in particular, to pump large amount of water and solutes has been used to decrease the downward movement of solutes and leaching into the groundwater and to stabilise and break down contaminants in soil and groundwater. A successful example of boron (B) phytoextraction has been reported by Robinson et al. (2003a, 2007) where high water-use poplar trees were used to evapotranspire water, control leaching, and remove B from the site by coppicing the trees that accumulated significant amount of B in their leaves. Fast-growing and metal-resistant trees (i.e. Salix spp.) have two advantages relative to hyperaccumulator plants: (1) extracting more metals from the soil because of their large biomass in both above and below ground (Pulford and Watson 2003) and (2) stabilising the metals in the soil and reducing soil erosion by wind and water.

The low biomass production is the major limitation to the commercial development of phytoextraction using hyperaccumulator plants. Most hyperaccumulator species have a small size and usually small leaf area, thus producing little biomass compared to crop plant and trees. Robinson et al. (2003b) suggested a model to calculate the time needed for phytoextraction to lower the contaminant concentration in soil to the level required by environmental regulation. Phytoextraction is a time-consuming process because it is dependent on biomass production and ability of the plant to take up metals that is a function of root exposure to bioavailable TEs. Low biomass production results in low evapotranspiration, which affects the uptake and translocation of metals from the soil solution. In addition, the distribution of TEs in the soil profile is heterogeneous, and the plant roots might not have access to the TE ‘hot spots’. Furthermore, hyperaccumulator are metal specific and can only extract high concentrations of one contaminant. Therefore, remediation of a site polluted with more than one TE might require sequential phytoextraction with different species, a process that will lengthen the time needed to clean up the site. On average, the time required for phytoextraction to clean even a moderately contaminated soil is in the order of decades.

Induced phytoextraction to increase metal availability has been studied in the past. It consists in addition of solubilising metal chelators (i.e. EDTA, ethylenediaminetetraacetate) to the soil to increase the mobility and allowing the metals to be taken up more easily by the plant. However, induced phytoextraction has no application in situ because it poses an environmental risk through metal leaching to groundwater and it is cost prohibitive (Chaney et al. 2007).

A more promising approach to increase the biomass and obtain hyperaccumulation is being studied where all genes needed for hyperaccumulation and hypertolerance are cloned and expressed in a high biomass plant (Cherian and Oliveira 2005; Clemens et al. 2002; Dhankher et al. 2012; Rugh et al. 1998). However, the development of a high biomass bioengineered metal hypertolerant and hyperaccumulator plants is still in its infancy. Hyperaccumulator traits are expressed in several genes, and the engineering of transgenic plants might become impractical beyond a certain number of genes (Krämer 2010).

Phytoextraction is a low-cost environmentally friendly technology, but the fact that it is ‘time consuming’ (Conesa et al. 2012) adds additional unpredictable costs and makes it less appealing relative to engineering methods or nonaction. One of the major drawbacks of phytoextraction that hinders the commercial development of the technology is the cost efficiency. Growing plants for cleanup of a contaminated soil is costly, and the cost of production (fertilisers, irrigation water, harvest machines, etc.) adds up to the cost of biomass disposal. Farmers and landowners may desire soil cleanup, but the decision to take action is dependent upon the economics of their farm operations.

The importance of agricultural management practices for producing high yield for phytoextraction crops has often been overlooked. Chaney et al. (2007) describe the agronomy of phytoextraction and point out the importance of practices such as fertilisation, pH optimising, weed control, and scheduling of harvest as critical field management practices for the success of phytoextraction. However, field management operation and agricultural practices need to be included in the cost analysis of phytoextraction and in models to estimate value and cost of phytoextraction products (Robinson et al. 2003b). When the phytoextraction is combined with a profit-making operation, then the time constraint may become less important. In addition, when the phytoextraction product (biomass) has commercial value (timber, bioenergy, fertiliser, food supplement, etc.), it can be sold to offset the cost of production and farm practices operation.

There are only few successful demonstrations of the possibility of using phytoextraction in field sites and production of secondary products that offset the costs of production. Perhaps the most economically viable example is the use of Se-enriched crops (i.e. Opuntia ficus-indica (cactus pear) and Brassica oleracea L. (broccoli)) (Bañuelos et al. 2012) for human consumption and production of forages (Brassica napus L., canola) with enough Se to replace Se supplements normally added to livestock feed (Bañuelos 2006). The oil and seed meal extracted from Brassica seeds grown in Se-rich soils of the San Joaquin Valley in Central California have been used as source of biofuels, green fertilisers, and bioherbicide (Bañuelos 2009; Bañuelos and Hanson 2010).

Two other important ways to combine phytoextraction with production of secondary products, such as biomass for bioenergy and recovery of metals from the plant as bio-ore, are described in the following sections.

3 Phytoextraction and Bioenergy Production

Biofuels are renewable fuels derived from biological feedstock and are largely carbon neutral because the CO2 released during biofuels combustion is offset by carbon fixation during plant growth. Biofuels are considered as a key to reducing reliance on foreign oil, lowering greenhouse gas emissions, and meeting rural development goals by developing local and sustainable energy sources. However, the political and public support for biofuels has been undermined by concerns related to food security because the conversion of croplands to produce biofuels may cause food shortages and associated increase in food prices (Koh and Ghazoul 2008).

Utilisation of poor-quality soils and contaminated land can extend the area available to grow energy crops, and it can avoid competition between energy crops and food products. Poplar (Populus spp.) and willow (Salix spp.) have been demonstrated to be the most successful tree crops that can be grown on contaminated land for biomass production and phytoextraction of TEs (Pulford and Watson 2003). They are fast to propagate, have many and deep roots, achieve high annual biomass production, take up large quantities of water, and generally possess high tolerance to trace metals (Cd, Cu, Zn, Pb) (Granel et al. 2002; Hu et al. 2013; Maxted et al. 2007; Meers et al. 2007; Mirck et al. 2005; Vervaeke et al. 2003). Salix and Populus spp. have an effective nutrient uptake and high evapotranspiration rate and a pronounced clone-specific capacity for heavy metal uptake. Success of Salix spp. (willow) as phytoextracting plants depends on its biomass production, metal accumulation capacity, and the site of metal accumulation in the plant. Willow is usually grown in short-rotation coppice (SRC) systems because plants have the ability to resprout after harvest. This characteristic makes willow very suitable to phytoextraction because the frequency and number of harvests will trigger higher metal removal. The estimated economic lifespan of a short-rotation willow coppice stand is 20–25 years, with 6–7 harvests (the time frame from planting to first harvest is typically 4 years).

Research on the environmental sustainability of willow production began in the mid-1980s by various groups in the USA and Europe (Gomes 2012; Rowe et al. 2009; Šyc et al. 2012; Volk et al. 2006). The willow cropping system utilises agricultural practices that are familiar to farmers, and after establishment, it is a relatively low-input crop with winter harvests, thus having a limiting effect on other farming operations. Willow biomass production systems involve intensive site preparation to control weeds, double-row mechanical planting of high density (15,300 plants ha−1), nitrogen inputs at the beginning of each rotation, and 3–4-year rotations. It has been demonstrated that the use of cover crops during the establishment phase of willow plantations and mechanical control of the cover crops (such as rolling, undercutting, or partial rototilling) will reduce the risk of erosion during tree establishment and will allow tree plantations on sloping farmland.

A common critique to the sustainability of willow SRC systems is the creation of ‘biological deserts’ across the landscape due to the monoculture of willow. However, long-term research on the above- and below-ground biodiversity in willow plantations has shown positive effects on avian biodiversity comparable to natural habitats including shrubland and successional habitats (abandoned fields, second-growth forests, regenerating clear-cuts) (Volk et al. 2006). Šyc et al. (2012) have suggested intercropping of fast-growing species such as willow and poplar with hyperaccumulators to increase the intake of metals for phytoremediation and to contribute to increases in biomass production and positive effects on biodiversity. The feasibility of intercropping hyperaccumulators with SRC needs to be studied in relation to impediment of mechanical operations for harvest and other agronomic practices (fertilisation and weed control).

It is generally expected that SRC will have higher water demand than arable crops due to the higher growth rates, high transpiration rates, and longer seasonal growth. In some European countries (i.e. UK), government guidelines require the plantation of SRC in areas where annual rainfall is at least 600 mm year−1 (Rowe at al. 2009). One environmental advantage of the high transpiration rate of willow is that the amount of water removed from the soil by the transpiration stream can decrease the downward flow through the soil and can reduce leaching losses. In addition, the perennial nature of SRC, their extensive root system, and the potential use of conservation tillage and cover crops will minimise nutrient outflow, strongly decrease the risk of soil erosion, and maintain good water quality (Abrahamson et al. 1998). Heller et al. (2003) carried out a life cycle assessment of willow plantation for bioenergy, and they showed that nitrogen fertilisation of willow accounts for the majority (37 %) of primary energy consumed over seven harvest rotations of willow biomass crops. The production of N fertilisers consumes large amount of non-renewable fossil fuel. The authors showed that substituting inorganic fertiliser with sewage sludge biosolids can increase the net energy ratio by more than 40 % (Heller et al. 2003). The net energy ratio for the production and conversion of short-rotation woody crops (SRWC) is 1:11, meaning that for every unit of non-renewable fossil fuel energy used to grow, harvest, and deliver SRWC, 11 units of usable energy are produced. In contrast, the net energy ratio for ethanol production from corn is 1:1.3 and for natural gas is 1: 0.4. In essence, willow crops are large solar collectors that capture the sun’s energy and store it in the woody biomass. In addition to the positive energy balances, willow crops can revitalise local economies and sustain rural development by diversifying farm crops and income.

Despite the numerous environmental and rural development benefits of SWRC, their use and cultivation have not been widely adopted in the USA and Europe. the main reason being the high cost of production ($2.60–3.00 GJ−1 vs. $1.40–1.90 GJ−1 for fossil fuels) due to high operating costs (low harvesting efficiency, high transportation costs), relatively low yield, and low energy conversion efficiency. One additional disadvantage of SRC biomass for bioenergy is the inability of SRC systems to continuously supply biomass throughout the year; the biomass is usually insufficient as a sole source of fuel for a combustion plant. Co-firing with coal and other fuel products could be envisaged if regulation is respected. To avoid air pollution, combustion of Salix wood grown on contaminated land should occur only in industrial and collective boilers equipped with efficient filters that trap the volatile particles containing the contaminant (i.e. Cd). The bottom ash can be used as basic mineral amendment or inorganic fertiliser. Co-firing a 100 MW power plant with 10 MW of willow biomass would require approximately 4,000 ha (10,000 acres) of biomass crop establishment, which corresponds to 1 % of the area in 80 km transport radius around the power plant (Abrahamson et al. 1998). The low energy conversion efficiency of SRWC is an additional problem that hinders the development of the technology on large scale. Alternative technologies including gasification and pyrolysis that can increase the overall conversion efficiency of biomass from SRWC are currently in various stages of research (Volk et al. 2006).

Lewandowski et al. (2006) quantified the economic value of combined cadmium remediation and biomass production by willow in a cadmium-contaminated case study area in the Rhine Valley (Germany). They concluded that the value of the phytoremediation function to farmers assessed by the substitution cost (alternative cost of soil cleanup by use of hyperaccumulator T. caerulescens) and hedonic price analysis (economic loss if the farmer cannot apply phytoremediation and has to set aside cadmium-contaminated land) delivers similar results and it is about 14,000 € ha−1 over a period of 20 years. However, farmers in the Rhine Valley were only willing to pay 0–1,500 € ha−1 mainly because they consider soil cleanup government’s duty. In addition, farmers were negatively influenced by the fact that they considered contamination not being their fault, and none of the farmers interviewed tried to calculate the benefit of phytoremediation.

Ongoing research and development activities to promote commercialisation and development of biomass power are focused on improving socio-economic sustainability of SRC and promoting government actions that remove constraints and provide incentive for biomass fuel use. Biomass fuel can contribute to the economic longevity of local/rural coal-fired power plants and provide a carbon-neutral fuel for power producers to help them meet the goals of governmental climate change actions.

4 Phytomining

Commercial mining is usually performed on ores with high concentration of the target metal (for Ni, at least 30 g kg−1) and are environmentally costly and energy- and capital-intensive practice of mining. Few ore bodies of this kind occur on the Earth’s surface and are present in small localised areas, and some of these are becoming exhausted due to expanding economies and industrialisation.

Sub- or low-grade ores contain concentration of target metal below the content required to be economically extracted and smelted by conventional methods. Most of these ore bodies are associated with ultramafic deposits that generate serpentine soils after weathering of the ultramafic rocks. Serpentine soils are characterised by pH of 6–8; low Ca/Mg ratio and low levels of N, P, and K; and potentially toxic concentrations of Ni (Li et al. 2003a; Sheoran et al. 2009); they are not economic to mine and are unsuitable for agriculture due to high trace metal content (especially Ni). These deposits are scattered around the world and usually support a characteristic flora of endemic plants (Brooks 1987) that are able to tolerate and/or (hyper) accumulate the metals present in the serpentine soil. The use of plants to extract Ni and few other metals to produce a bio-ore is called phytomining. Dried plant material of hyperaccumulators grown on Ni-rich serpentine soil is reduced to an ash, and the metal is recovered using conventional metal refining methods such as acid dissolution or electrowinning.

The first field trial on Ni phytomining used a naturally occurring stand of Streptanthus polygaloides grown on serpentine soils in CA, and it extracted up to 100 Ni kg ha−1, worth $550 ha−1 at the prices of Ni in 1994 (Nicks and Chambers 1998). In a second phytomining field trial, Alyssum bertolonii (Robinson et al. 1997a) was used on a serpentine soil in Italy, and plants were fertilised with N, P, and K over a 2-year period. Fertilisation induced a threefold increase in dry biomass, indicating that agricultural practices similar to those applied to crops are important to increase yield in hyperaccumulator plants used for phytomining. Li et al. (2003b) demonstrated the feasibility of Ni recovery from the ash (bio-ore) of genetically improved Alyssum spp. Bio-ore contains higher Ni concentration (6–16 %) than normal Ni ores and are free of Mn, Fe, and Si oxides as in conventional ores (Robinson et al. 2009). Chaney et al. (2007) discuss the importance of fertilisation, pH optima, plant density, and development of improved cultivars for increasing shoot Ni concentration and yield of hyperaccumulator Alyssum murale. Weed management is also necessary when serpentine soils are fertilised, as other plants can compete with the hyperaccumulator grown for phytomining. In addition, in wet climate when soils are poorly drained, ridge tilling helped reduce adverse effect of heavy spring rainfall and consequent flooding on the survival and yield potential of Alyssum grown on a field trial in Canada (Chaney et al. 2007).

Further studies (Anderson et al. 1999; Harris et al. 2009; Keller 2004; Robinson et al. 1997b) have shown that other species could be grown for phytomining and that other metals could be extracted, such as thallium (Tl) and gold (Au). There are only a few Tl hyperaccumulators, namely, Iberis intermedia and Biscutella laevigata, from southern France (Leblanc et al. 1999) that can take up high level of Tl in dry matter, 0.4 mg kg−1 and 1.4 mg kg−1, respectively. Tl is quite rare in nature, representing only 0.7 mg kg−1 in the Earth’s crust. It is very toxic and is used in rat poison and for the control of ants. There is clearly potential for Tl phytomining if large areas are contaminated with Tl to obtain the advantage of large-scale operations. I. intermedia can produce 10 t ha−1, as determined from field observation in France, and it should produce about 700 kg ha−1 of bio-ore containing 8 kg Tl (Anderson at al. 1999). Plants do not normally accumulate Au, and the metal must be made soluble before uptake can occur. Au is extracted by induced phytoextraction, which consists of using a chelating agent, usually thiocyanate, to make Au available to plant roots. It requires standard ore mining and grinding of the ore and then placing the ground ore over a plastic surface to avoid leaching of the cyanide and thiocyanide used to induce phytoextraction.

To be economically viable, phytomining should be able to produce about $500 ha−1, and most phytomining operation can gain additional revenue from incineration of biomass to generate electricity. The vicinity of the phytomining farm to the power plant is important to reduce transportation costs and to facilitate the secondary use of biomass for bioenergy production in addition to the recovery of the target metal. Furthermore, as noted by many authors, the lack of continuity of biomass supply from phytomining operation where harvest occurs once or twice a year is an impediment to the use of biomass for bioenergy unless other sources of biomass are available locally to provide material for the continuity of combustion operation throughout the year. One way to solve this problem is the co-firing with other sources of biomass as explained in Sect. 3.

A critique related to the environmental sustainability of phytomining addresses the efficiency of phytomining relative to conventional mining. Robinson et al. (2009) pointed out that phytomining requires a large area and more time to produce a ton of Ni than conventional mining (2.5 ha year−1 vs. 22 m3 in a few hours). In addition, Robinson et al. (2009) note that phytomining of surface serpentine soil will take from 3 to 18 crop cycles before the surface Ni is depleted, and the authors argue that the surface soil needs to be removed to continue phytomining of the deeper soil layers. The authors are also concerned with the introduction of a monoculture of possibly exotic species and disruption of serpentine ecosystem and native endemic flora.

Other authors (Anderson et al. 1999; Chaney et al. 2010; Harris et al. 2009; Sheoran et al. 2009) are supportive of phytomining and consider its environmental impact similar to that of commercial farming. Phytomining is considered to have a positive effect on soil erosion by the effect of plant roots relative to the open pit and the ‘desert-like’ landscape left after conventional mining operation has ceased and substantial site remediation is required at the end of the mine. Phytomining, instead, can improve the quality of the soil for post-mining operation over the duration of the phytomine. It is necessary, however, to study the potential environmental impacts of the phytomine during the planning phase and before the phytomining operation is set in place.

Hence, this technology requires a team of expert agronomists, ecologists, and soil scientists to carefully choose the proper species for phytoextraction of a particular metal. Various steps can be taken to reduce the ecological risks associated with monoculture and introduction of potential exotic species, for example, (1) analysis of phytoextraction potential of native vegetation and selection of the best performing native species for metal extraction efficiency, (2) adoption of principles of agro-ecology such as cultivation of combination of various native species, and (3) use of alternative fertilisers (sewage sludge biosolids) and bioherbicides.

Finally, conventional mining is not economically viable on low-grade ore, and hence a comparison of the efficiency between phytomining and conventional mining is not appropriate.

If phytomining proceeds beyond the theoretical and trail stage, the most likely scenario that can be envisaged is that phytomining can be farmed out to small-scale landholders and farmers throughout the region where low-grade metalliferous soils are present. Small-scale operation can be environmentally sustainable because usually small farms are farmed directly by the landholder, which is therefore interested to maintain or increase the fertility of his/her land. Higher revenues can be obtained if farmers unite in co-op to reduce the costs of mechanisation and transportation by economy of scale.

Fertility of land is likely to be increased after phytomining operation because of increased root and microbial activity in the soil profile, Ni removal, and pH and fertility management. Therefore, landholders may want to use land for food production after phytomining has reached Ni depletion of the top layers and therefore reduced Ni toxicity to crops occurs. The possibility of restoring serpentine soils post-phytomining and convert those soils to food production is an important aspect of environmental sustainability of phytomining and needs to be addressed in future studies.

5 Conclusions

The study and development of phytotechnologies in the past 30 years have produced important advances in understanding fundamental aspects of physiology, ecology, and agronomy of hyperaccumulator plants and biogeochemistry and ecotoxicology of TEs. However, the application of the phytoextraction technology to real-world situations has been deterred by the lack of understanding that the complexity of biological systems, which are put to ‘work’ to provide a service to humans, can hardly fit the constraints of market economy (i.e. time efficiency and high revenues). As discussed above, the main constraints to the commercialisation of phytotechnologies and in particular phytoextraction are the time needed to clean up the soil and the costs associated with production and disposal of biomass.

A more interdisciplinary approach is needed to understand the multi-facets of introducing a designed biological system, made of specialised plants and trees, into an area of land that has been environmentally damaged. Studies on the ecological assessment, environmental sustainability, socio-economic aspects of rural development and improvement of local economies, and life cycle assessment are very important and need to be performed in any revegetation and phytoextraction plan. A complete analysis of the direct and indirect costs and benefits of the application of phytotechnologies is very important to gain a deeper understanding of the complexity of such a system. Consequently, the cost/benefit analyses of phytotechnologies will not only rely on principles of market economy (i.e. time efficiency and high revenues) but also on indirect environmental and socio-economic benefits to society. When future studies are able to provide a description of the multi-facets of phytotechnologies, a stronger support will be gained from local communities and the public at large, which will then trigger government support and financial aid.


  1. Abrahamson LP, Robison DJ, Volk TA, White EH, Neuhauser EF, Benjamin WH, Peterson JM (1998) Sustainability and environmental issues associated with willow bioenergy development in New York (U.S.A.). Biomass Bioenergy 15:17–22CrossRefGoogle Scholar
  2. Anderson CWN, Brooks RR, Chiarucci A, LaCoste CJ, Leblanc M, Robinson BH, Simcock R, Stewart RB (1999) Phytomining for nickel, thallium and gold. J Geochem Explor 67:407–415CrossRefGoogle Scholar
  3. Baker AJM (1981) Accumulators and excluders -strategies in the response of plants to heavy metals. J Plant Nutr 3:643–654CrossRefGoogle Scholar
  4. Bañuelos GS (2006) Phyto-products may be essential for sustainability and implementation of phytoremediation. Environ Pollut 144:19–23CrossRefGoogle Scholar
  5. Bañuelos G (2009) Phytoremediation of selenium contaminated soil and water produces biofortified products and new agricultural bioproducts. In: Bañuelos G, Lin ZQ (eds) Biofortification and development of new agricultural products. CRC Press, Boca Raton, pp 57–70Google Scholar
  6. Bañuelos GS, Hanson BD (2010) Use of selenium-enriched mustard and canola seed meals as potential bioherbicides and green fertilizer in strawberry production. Hortic Sci 45:1567–1572Google Scholar
  7. Bañuelos GS, Stushnoff C, Walse SS, Zuber T, Yang SI, Pickering IJ, Freeman JL (2012) Biofortified, selenium enriched, fruit and cladode from three Opuntia cactus pear cultivars grown on agricultural drainage sediment for use in nutraceutical foods. Food Chem 135:9–16CrossRefGoogle Scholar
  8. Boyd RS (2007) The defense hypothesis of elemental hyperaccumulation: status, challenges and new directions. Plant Soil 293:153–176CrossRefGoogle Scholar
  9. Broadhurst CL, Chaney RL, Angle JS, Erbe EF, Maugel TK (2004) Nickel localization and response to increasing Ni soil levels in leaves of the Ni hyperaccumulator Alyssum murale. Plant Soil 265:225–242CrossRefGoogle Scholar
  10. Broadhurst CL, Tappero RV, Maugel TK, Erbe EF, Sparks DL, Chaney RL (2009) Interaction of nickel and manganese in accumulation and localization in leaves of the Ni hyperaccumulators Alyssum murale and Alyssum corsicum. Plant Soil 314:35–48CrossRefGoogle Scholar
  11. Brooks RR (1987) Serpentine and its vegetation. A multidisciplinary approach. Dioscorides Press, PortlandGoogle Scholar
  12. Brooks RR, Lee J, Reeves RD, Jaffré T (1977) Detection of nickeliferous rocks by analysis of herbarium specimens of indicator plants. J Geochem Explor 7:49–57CrossRefGoogle Scholar
  13. Centofanti T, Siebecker MG, Chaney RL, Davis AP, Sparks DL (2012) Hyperaccumulation of nickel by Alyssum corsicum is related to solubility of Ni mineral species. Plant Soil 359:71–83CrossRefGoogle Scholar
  14. Chaney RL (1983) Plant uptake of inorganic waste constituents. In: Parr JF, Marsh JMK (eds) Land treatment of hazardous wastes. Noyes Data Corp, Park RidgeGoogle Scholar
  15. Chaney RL, Angle JS, Broadhurst CL, Peters CA, Tappero RV, Sparks DL (2007) Improved understanding of hyperaccumulation yields commercial phytoextraction and phytomining technologies. J Environ Qual 36:1429–1433CrossRefGoogle Scholar
  16. Chaney RL, Centofanti T, Broadhurst CL (2010) Phytoremediation of soil trace elements. In: Hooda PS (ed) Trace elements in soils. Wiley, Chichester, p 352Google Scholar
  17. Cherian S, Oliveira MM (2005) Transgenic plants in phytoremediation: recent advances and new possibilities. Environ Sci Technol 39:9377–9390CrossRefGoogle Scholar
  18. Clemens S, Palmgren MG, Krämer U (2002) A long way ahead: understanding and engineering plant metal accumulation. Trends Plant Sci 7:309–315CrossRefGoogle Scholar
  19. Conesa HM, Evangelou MWH, Robinson BH, Schulin R (2012) A critical view of current state of phytotechnologies to remediate soils: still a promising tool. Sci World J. doi: 10.1100/2012/173829, Article ID 173829Google Scholar
  20. Dhankher OP, Pilon-Smits EAH, Meagher RB, Doty S (2012) Biotechnological approaches for phytoremediation. In: Altman A, Hasegawa PM (eds) Plant biotechnology and agriculture prospects for the 21st century. Academic, MA, pp 309–328CrossRefGoogle Scholar
  21. Dickinson NM, Baker AJM, Doronila A, Laidlaw S, Reeves RD (2009) Phytoremediation of inorganics: realism and synergies. Int J Phytoremediat 11:97–114CrossRefGoogle Scholar
  22. GACGC (1994) World in transition: the threat to soils. German Advisory Council on Global Change, Annual report, Economica Verlag GmbH, Bonn, GermanyGoogle Scholar
  23. Gomes HI (2012) Phytoremediation for bioenergy: challenges and opportunities. Environ Technol Rev 1:59–66CrossRefGoogle Scholar
  24. Granel T, Robinson B, Mills T, Clothier B, Green S, Fung L (2002) Cadmium accumulation by willow clones used for soil conservation, stock fodder, and phytoremediation. Aust J Soil Res 40:1331–1337CrossRefGoogle Scholar
  25. Harris AT, Naidoo K, Nokes J, Walker T, Orton F (2009) Indicative assessment of the feasibility of Ni and Au phytomining in Australia. J Clean Prod 17:194–200CrossRefGoogle Scholar
  26. Heller MC, Keoleian GA, Volk TA (2003) Life cycle assessment of a willow bioenergy cropping system. Biomass Bioenergy 25:147–165CrossRefGoogle Scholar
  27. Hu Y, Nan Z, Su J, Wang N (2013) Heavy metal accumulation by poplar in calcareous soil with various degrees of multi-metal contamination: implications for phytoextraction and phytostabilization. Environ Sci Pollut Res 20:7194–7203CrossRefGoogle Scholar
  28. Keller C, Hammer D (2004) Metal availability and soil toxicity after repeated croppings of Thlaspi caerulescens in metal contaminated soils. Environ Pollut 131:243–254CrossRefGoogle Scholar
  29. Koh LP, Ghazoul J (2008) Biofuels, biodiversity, and people: understanding the conflicts and finding opportunities. Biol Conserv 141:2450–2460CrossRefGoogle Scholar
  30. Krämer U (2010) Metal hyperaccumulation in plants. Annu Rev Plant Biol 61:517–534CrossRefGoogle Scholar
  31. Leblanc M, Petit D, Deram A, Robinson BH, Brooks RR (1999) The phytomining and environmental significance of hyperaccumulation of thallium by Iberis intermedia from southern France. Econ Geol 94:109–113CrossRefGoogle Scholar
  32. Letey J, Williams CF, Alemi M (2002) Salinity, drainage and selenium problems in the Western San Joaquin Valley of California. Irrig Drain Syst 16:253–259CrossRefGoogle Scholar
  33. Lewandowski I, Schmidt U, Londo M, Faaij A (2006) The economic value of the phytoremediation function – assessed by the example of cadmium remediation by willow (Salix ssp). Agric Syst 89:68–89CrossRefGoogle Scholar
  34. Li Y-M, Chaney R, Brewer E, Roseberg R, Angle JS, Baker A, Reeves R, Nelkin J (2003a) Development of a technology for commercial phytoextraction of nickel: economic and technical considerations. Plant Soil 249:107–115CrossRefGoogle Scholar
  35. Li Y-M, Chaney RL, Brewer EP, Angle JS, Nelkin J (2003b) Phytoextraction of nickel and cobalt by hyperaccumulator Alyssum species grown on nickel-contaminated soils. Environ Sci Technol 37:1463–1468CrossRefGoogle Scholar
  36. Maxted AP, Black CR, West HM, Crout NMJ, McGrath SP, Young SD (2007) Phytoextraction of cadmium and zinc by Salix from soil historically amended with sewage sludge. Plant Soil 290:157–172CrossRefGoogle Scholar
  37. Meers E, Vandecasteele B, Ruttens A, Vangronsveld J, Tack FMG (2007) Potential of five willow species (Salix spp.) for phytoextraction of heavy metals. Environ Exp Bot 60:57–68CrossRefGoogle Scholar
  38. Milner MJ, Kochian LV (2008) Investigating heavy-metal hyperaccumulation using Thlaspi caerulescens as a model system. Ann Bot 102:3–13CrossRefGoogle Scholar
  39. Mirck J, Isebrands JG, Verwijst T, Ledin S (2005) Development of short-rotation willow coppice systems for environmental purposes in Sweden. Biomass Bioenergy 28:219–228CrossRefGoogle Scholar
  40. Morvan X, Saby NPA, Arrouays D, Le Bas C, Jones RJA, Verheijen FGA, Bellamy PH, Stephens M, Kibblewhite MG (2008) Soil monitoring in Europe: a review of existing systems and requirements for harmonisation. Sci Total Environ 391:1–12CrossRefGoogle Scholar
  41. Nicks LJ, Chambers MF (1998) A pioneering study of the potential of phytomining for nickel. In: Brooks RR (ed) Plants that hyperaccumulate heavy metals. CABI, Wallingford, p 380Google Scholar
  42. Ohlendorf HM, Hoffman DJ, Saiki MK, Aldrich TW (1986) Embryonic mortality and abnormalities of aquatic birds: apparent impacts of selenium from irrigation drainwater. Sci Total Environ 52:49–63CrossRefGoogle Scholar
  43. Pilon-smits E (2005) Phytoremediation. Annu Rev Plant Biol 56:15–39CrossRefGoogle Scholar
  44. Pulford ID, Watson C (2003) Phytoremediation of heavy metal-contaminated land by trees—a review. Environ Int 29:529–540CrossRefGoogle Scholar
  45. Robinson BH, Chiarucci A, Brooks RR, Petit D, Kirkman JH, Gregg PEH, De Dominicis V (1997a) The nickel hyperaccumulator plant Alyssum bertolonii as a potential agent for phytoremediation and phytomining of nickel. J Geochem Explor 59:75–86CrossRefGoogle Scholar
  46. Robinson BH, Brooks RR, Howes AW, Kirkman JH, Gregg PEH (1997b) The potential of the high-biomass nickel hyperaccumulator Berkheya coddii for phytoremediation and phytomining. J Geochem Explor 60:115–126CrossRefGoogle Scholar
  47. Robinson B, Green S, Mills T, Clothier B, van der Velde M, Laplane R, Fung L, Deurer M, Hurst S, Thayalakumaran T, van den Dijssel C (2003a) Phytoremediation: using plants as biopumps to improve degraded environments. Aust J Soil Res 41:599–611CrossRefGoogle Scholar
  48. Robinson B, Fernández JE, Madejón P, Marañón T, Murillo JM, Green S, Clothier B (2003b) Phytoextraction: an assessment of biogeochemical and economic viability. Plant Soil 249:117–125CrossRefGoogle Scholar
  49. Robinson BH, Green SR, Chancerel B, Mills TM, Clothier BE (2007) Poplar for the phytomanagement of boron contaminated sites. Environ Pollut 150:225–233CrossRefGoogle Scholar
  50. Robinson BH, Bañuelos G, Conesa HM, Evangelou MWH, Schulin R (2009) The phytomanagement of trace elements in soil. Crit Rev Plant Sci 28:240–266CrossRefGoogle Scholar
  51. Rowe RL, Street NR, Taylor G (2009) Identifying potential environmental impacts of large-scale deployment of dedicated bioenergy crops in the UK. Renew Sustain Energy Rev 13:271–290CrossRefGoogle Scholar
  52. Rugh CL, Gragson GM, Meagher RB, Merkle SA (1998) Toxic mercury reduction and remediation using transgenic plants with a modified bacterial gene. Hortic Sci 33:618–621Google Scholar
  53. Sheoran V, Sheoran AS, Poonia P (2009) Phytomining: a review. Miner Eng 22:1007–1019CrossRefGoogle Scholar
  54. Šyc M, Pohořelý M, Kameníková P, Habart J, Svoboda K, Punčochář M (2012) Willow trees from heavy metals phytoextraction as energy crops. Biomass Bioenergy 37:106–113CrossRefGoogle Scholar
  55. Tappero R, Peltier E, Gräfe M, Heidel K, Ginder-Vogel M, Livi KJT, Rivers ML, Marcus ML, Chaney RL, Sparks DL (2007) Hyperaccumulator Alyssum murale relies on a different metal storage mechanism for cobalt than for nickel. New Phytol 175:641–654CrossRefGoogle Scholar
  56. van der Ent A, Baker AJM, Reeves RD, Joseph Pollard A, Schat H (2013) Hyperaccumulators of metal and metalloid trace elements: facts and fiction. Plant Soil 362:319–334CrossRefGoogle Scholar
  57. Vervaeke P, Luyssaert S, Mertens J, Meers E, Tack FMG, Lust N (2003) Phytoremediation prospects of willow stands on contaminated sediment: a field trial. Environ Pollut 126:275–282CrossRefGoogle Scholar
  58. Volk TA, Abrahamson LP, Nowak CA, Smart LB, Tharakan PJ, White EH (2006) The development of short-rotation willow in the northeastern United States for bioenergy and bioproducts, agroforestry and phytoremediation. Biomass Bioenergy 30:715–727CrossRefGoogle Scholar
  59. Zhao FJ, Lombi E, Breedon T, McGrath SP (2000) Zinc hyperaccumulation and cellular distribution in Arabidopsis halleri. Plant Cell Environ 23:507–514CrossRefGoogle Scholar

Copyright information

© Springer India 2015

Authors and Affiliations

  1. 1.Department of Plant ScienceCalifornia State University-FresnoFresnoUSA

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