Chapter 1: General introduction & background
1 General introduction & background
1.1 GENERAL INTRODUCTION
Industrialisation since the 1940’s resulted in a massive explosion of urban populations, which in turn led to widespread economic expansion fuelled largely by demand for energy resources. The post war years relied heavily upon the manufacturing and utilisation of various chemicals, designed to meet the growing standards of living. Several of these chemicals were tainted with heavy metal(loid)s and inappropriately used and disposed of in urban and agricultural surroundings. The result of industrialisation over the past 60 years has resulted in widespread heavy metal(loid) contamination of soils and groundwater across the globe. Today their legacy poses one of the biggest environmental challenges of the 21 st century: to effectively and economically restore and remediate heavy metal(loid)s contaminated sites (Bhandari 2007). Heavy metal(loid)s may be used to describe both biologically essential (copper; Cu, manganese; Mn and zinc; Zn), often referred to as ‘trace elements’ or ‘micro elements’, and nonessential (cadmium; Cd, lead; Pb, mercury; Hg, arsenic; As) elements. The nonessential elements are of greatest importance as they pose a serious threat to ecosystem, human and animal health. (Raskin et al. 1997). For example, heavy metal(loid) uptake (such as Cd) in agricultural and/or horticultural production systems can potentially threaten food quality, safety and marketability, and may have deleterious effects on plant growth through phytotoxicities (McLaughlin et al. 2000a; Kachenko and Singh 2006).
Once in the environment, heavy metal(loid)s express contrasting behaviours due to different thermodynamic and phisico-chemical properties (Mukherjee 2001). For example, Cr, Hg and Pb are very strongly retained by the solid phase of most soils, hence accumulation of these elements poses negligible risk to plant and soil organisms. Conversely, Cd is readily accumulated by plants at levels that are well below those where phytotoxicity is expressed, potentially threatening animal and human consumers (McLaughlin et al. 2000b).
1 A. G. Kachenko
In recent times, there has been growing awareness regarding the development and implementation of cost-effective strategies to alleviate the threat of heavy metal(loid) contaminates in the soil environment (Mulligan et al. 2001). There are thousands of sites throughout the world where corrective action is required to satisfy the growing need for sustainable management of the environment. In Australia alone, there are an estimated 80,000 contaminated sites, of which many are related to former mining activities (Naidu et al. 2002). For many of these sites, there has been a trend toward enhanced environmental management practices particularly the implementation of green environmentally neutral technologies to rectify the ailing state of the environment (Swindoll and Firth 1998).
The possibility of utilising plants to remediate and potentially decontaminate the environment began during the later 19 th century when Baumann (1885) observed up to 1.7% Zn in violet (Viola caliminara) and mustard (Thalaspi calaminara) plants growing on Zn rich soils. However, almost a century passed before these unique plant species were first considered as a possible green solution to remediate heavy metal(loid)s contaminated soils. It was in 1977 that Brooks and others first described this phenomenon as hyperaccumulation (defined in section 1.3.3) and today, this term has since been associated with > 450 species world wide (Prasad and Freitas 2006). More recently, studies have focused on the ecophysiology of hyperaccumulation (e.g. Fernando et al. 2000a; Whiting et al. 2003; Bhatia et al. 2005a), and many have attempted to identify additional hyperaccumulating species (e.g. Meharg 2003; Wang et al. 2007).
In this introductory chapter, an overview is presented on the mechanisms of heavy metal(loid) tolerance and the occurrence of hyperaccumulator species. The chapter begins with a brief background on heavy metal(loid)s in the soil environment and the process of heavy metal(loid) uptake and (hyper)accumulation. Further, this chapter provides an overview on the tolerance mechanisms employed by hyperaccumulator species in order to neutralise potentially inimical concentrations of accumulated heavy metal(loid)s, and the potential use of these species in the remediation of polluted soils. The chapter concludes with a discussion on heavy metal(loid)s in ferns followed by the rationale, aims and objectives of this thesis, including a thesis outline.
2 Chapter 1: General introduction & background
1.2 HEAVY METAL(LOID)S IN THE SOIL ENVIRONMENT 1.2.1 Definition The term heavy metals is a widely used term in literature to describe a disparate group of potentially toxic elements, however, its definition is arbitrary and somewhat imprecise (Hodson 2004). In an environmental context, it is often used to describe a group of elements associated with pollution and potential toxicity (Hodson 2004). Chemically, heavy metals are defined as a group of elements with an atomic density greater than 6 g cm 3 (Phipps 1981). Although this definition appears precise, great debate has evolved over the correct interpretation and scientific merit of the term heavy metals. A concise summary of current definitions based on six parameters; density (specific gravity), atomic weight (relative atomic mass), atomic number, chemical properties, toxicity or criteria used to define heavy metals before 1936 has been recently reviewed (Duffus 2002).
Phipps (1981) discussed a biologically significant classification of metals based on the four (s-, p-, d- and f-) blocks of the periodic table. The d-block elements are particularly important as their redox behaviours and complex formation properties underlie their catalytic role in enzymes, whilst higher atomic number (p-block) elements bind strongly to sulphur, often resulting in a toxic effect. A weakness in Phipps (1981) proposal is that there is no differentiation between the metal ions in each of the different blocks, as the scheme is based on reactivity (Duffus 2002). An alternative theory suggested by Neiboer and Richardson (1980) classified metal ions in terms of differential Lewis acidity based on their relative affinity for N-, O- and S- containing ligands. This theory, based on chemical properties alone, is not absolute. It is largely empirically based and ion specific, however is relevant in a biological, toxicological and environmental contexts (Duffus 2002). A summary of the Lewis acidity based classification of heavy metals is presented in Table 1.1.
3 A. G. Kachenko
Table 1.1 Classification of heavy metals based on Lewis acidity adapted from Neiboer and Richardson (1980).
Class A (Hard) metals Lewis acids (electron acceptors) of small size, low polarizability (hardness), are easily displaced and mobile, generally forming ionic bonds.
Examples : Li +, Fe 3+, Rb +, Sr 2+, Zr 4+ , Cs +, Hf 4+ , Ti4+ , Co 3+
Class B (Soft) metals Lewis acids (electron acceptors) of large size, high polarizability (softness), show strong affinity to soft ligands such as sulphide or sulphur donors, often forming covalent bonds.
Examples : Cu +, Pd 2+ , Ag +, Cd 2+ , Hg +, Pt2+
Borderline (Intermediate) metals Size and polarizability between class A and class B metals, form relatively stable complexes with both hard and soft donor ligands.
Examples : Pb 4+ , Cu 2+ , Fe 2+ , Co 2+ , Ni 2+ , Zn 2+
In this thesis, class B and borderline metals will be collectively referred to as heavy metals. Although imprecise, this term is still frequently used to describe metals of environmental significance with no unambiguous alternative available (McGrath et al. 1997; Citterio et al. 2003). Further, in this thesis the metalloid As will be referred to as a heavy metalloid when discussed.
1.2.2 Sources The soil environment acts as a large repository of heavy metal(loid)s that enter from numerous sources classified as either natural or anthropogenic. Irrespective of their origin, their persistent nature in soils pose significant environmental hazards to plant, animal and most importantly human health (McIntyre 2003; Krämer 2005). Human exposure to As my result in short term diseases such as hypertension or cardiovascular disease or long term illness such as skin, lung or bladder cancer. The clinical manifestations of As exposure has resulted in United States Environment Protection listing As as the number one toxin of prioritised pollutants (Ng et al. 2003). Heavy metal(loid)s arising from natural sources result
4 Chapter 1: General introduction & background from weathering of parent material, wind borne particles, volcanoes, forest fires and biogenic processes (Ernst 1998). For example, geogenic As is commonly associated with potable groundwater supplies throughout South East Asia, placing millions of people at direct risk of As exposure (Juhasz et al. 2003). Anthropogenic sources of heavy metal(loid)s include mining, smelting and refining of metalliferous ore including by-products such as slag, emissions from industrial manufacturing processes including electroplating, energy and fuel production, and agricultural inputs such as the application of fertilizers, pesticides, fungicides and municipal sludges to land (Singh 2001). For example, at the abandoned Mount Perry Cu mine, South Eastern Queensland, a recent study reported mining-influenced stream sediment contained Cu and As levels that were typically higher than international sediment quality guidelines (Ashley et al. 2003). Historically, anthropogenic activities have been the principle source of heavy metal(loid) contamination in Australia, and still are responsible for widespread contamination in both urban and agricultural regions (Martley et al. 2004; Pietrzak and McPhail 2004).
1.2.3 Bioavailability To understand the availability and solubility of heavy metals, they are often considered in different pools in the soil system. Tessier et al. (1979) showed that the concentration of a given heavy metal in the soil is the sum of the amounts contained within the following five pools: (i) The soil solution – elements as free ions or soluble inorganic and organometallic complexes; (ii) Elements adsorbed to the colloidal phase, composed of clay minerals and humic compounds. This phase is also referred to as the cation exchange complex; (iii) Ions adsorbed and bound to hydrous oxides of manganese, iron and aluminium; (iv) Insoluble organometallic complexes; (v) Insoluble precipitates, such as sulphides, phosphates and carbonates.
Bioavailability is largely determined by the equilibrium between heavy metal(loid)s in soil solution and the solid phase. In general, the equilibrium is influenced through various reactions including adsorption, ion exchange, complexation with organic and inorganic
5 A. G. Kachenko ligands, redox reactions and precipitation-dissolution (Morel 1997). These reactions can potentially affect the free ion concentration of heavy metal(loid)s at the soil-water interface, hence solubility. The fraction available to plants is not the same as the total concentration in the soil, much of which is locked up in the solid phase. Bioavailable metal(loid)s are either in soil solution, weakly absorbed to the solid phase or adsorbed to the solid phase but able to transfer in solution during plant growth. Therefore, of these five pools, the first three are considered most important in plant uptake.
The unavailable fraction (those rendered immobile) is adsorbed firmly to soil particles, or integrated in organic and inorganic compounds (Morel 1997). It is understood that high concentrations of metals may not necessarily imply their release or their availability for plant uptake (Sánchez-Martin et al. 2007). Recently, Doelsch et al. (2008) evaluated fractionation patterns in soils with elevated Cr (106–175 mg kg –1), Cu (34–118 mg kg –1), Ni (89–310 mg kg –1) and Zn (104–242 mg kg –1) concentrations and reported < 5% of the total heavy metal contents were present in the exchangeable fractions. Conversely, the residual fraction was found to contain the largest fraction of the studied heavy metals, and ranged from 39–97.9% of the total Cu, Cr, Ni and Zn concentrations. Similarly, Sánchez-Martin et al. (2007) studied the fractionation patterns of soils amended with heavy metals tainted sewage sludge and found that > 60% of Cd, Cr, Cu, Ni, Pb and Zn were present in the residual fraction. A schematic representation of heavy metal(loid) availability is given in Figure 1.1.
a) Status
Total Solution Solid phase
Free ions Firmly bound
b) Mobility
Soluble Weakly Desorbed Non m obile ions during ions adsorbed during plant growth plant growth
Available Non Available
Figure 1.1 A model for depicting available heavy metal(loid)s in soil adapted from Morel (1997).
6 Chapter 1: General introduction & background
The bioavailability of heavy meatl(loid)s is governed not only by their total concentrations, but also by physical, chemical and biological processes within the soil environment. Both physical and chemical aspects provide the framework in which biological factors can modify heavy metal(loid) availability. Physical processes are largely dependent on soil type and include physical resistance restricting root penetration, soil structure and low water storage capacity (Ernst 1996). Physical properties such as soil texture may also influence the distribution of elements among soil fractions. Kambala and Singh (2001) investigated the fractions of Cu, Pb and Zn in four contrasting soil profiles in the vicinity of a Cu smelter. In one soil profile dominated by sand fraction, the authors suggested that low clay content and negligible sorption by dominant quartz grains resulted in a low percentage of elements associated with the residual fraction.
Chemical aspects such as soil acidity, redox potential (E h) and speciation may also influence the lability and plant uptake of heavy metal(loid)s (Ernst 1996). The role of soil pH is well documented in determining heavy metal mobilisation. Several studies have indicated that a decline in soil pH increases plant heavy metal uptake (Sappin-Didier et al. 2005; Tsadilas et al. 2005). Like soil pH, soil E h is also a well recognised soil parameter that can control the fate of heavy metal(loid)s in soils (Chuan et al. 1996; Carbonell-Barrachina et al. 2000). Generally, under reducing conditions, heavy metal(loid) bioavailability is considered high
(Carbonell-Barrachina et al. 2000). For example, Chuan et al. (1996) studied the effect of E h on Cd, Pb and Zn mobility in contaminated soils. The authors found that reducing conditions were most favourable for metal solubilisation. They hypothesised that these elements may have been originally adsorbed onto Fe-Mn oxyhydroxides and that dissolution of these Fe-Mn oxyhydroxides under reducing conditions may have resulted in the subsequent release of heavy metals. Chemical speciation of certain heavy metal(loid)s may limit their bioavailability for plant uptake. For example, As V is considered less toxic and mobile than As III (Hossain 2006). Similarly, the mobility of Cr in soil depends on its oxidation state with Cr VI considered more toxic and mobile than Cr III (Kumpiene et al. 2008).
Biological factors including soil bacterial and fungal rhizosphere associations and higher plants may significantly modify the chemical and physical conditions which determine heavy metal(loid) bioavailability (Ernst 1996). Wu et al. (2006) studied the influence of bacterial inoculation on the metal release from Pb/Zn tailings. In the case of Zn, bacteria colonies
7 A. G. Kachenko decreased pH and enhanced metal mobility and bioavailability. Conversely, Pb was absorbed by bacterial cell walls and resulted in a great reduction of water-soluble Pb. Mycorrhizal association of roots may result in an increase in root surface area for heavy metal(liod) acquisition. For example, Agely et al. (2005) reported that arbuscular mycorrhizal (AM) fungi could increase aboveground biomass, As accumulation, translocation, and bioconcentration in Pteris vittata . Lie et al. (2005) also observed that AM fungi increased P. vittata biomass and consequently increased the amount of As removed from the soil. The role of mycorrhizal association is described in Section 1.4.1.1. Plant uptake of heavy metal(loids) continuously alters the concentration and speciation in contaminated soils through release of root exudates and rhizosphere acidification (Ernst 1996). These aspects are discussed in Section 1.4.1.2.
1.3 PLANT RESPONSE TO HEAVY METAL(LOID)S
In plants, accumulation of a given heavy metal(loid) is a function of uptake capacity and intracellular binding sites (e.g. cation transporters). Intracellular binding of heavy metal(loid) ions is described in Section 1.4.1.4. Accumulation is further complicated by tissue and cell specific differences and intercellular transport (Clemens et al. 2002). It is unlikely that plants have evolved specific heavy metal(loid) uptake mechanisms. Instead, it is more plausible that these unwanted ions are transported across the root membrane and substituted for nutrients through the same transport mechanisms (Ross and Kaye 1994). A classical example is that of arsenate and phosphate where several authors have suggested that As competes with phosphate as a substrate for the phosphate uptake system (Jacobs and Keeney 1970; Asher and Reay 1979). Following entry into plant tissues, heavy metal(loid)s can affect various physiological and biochemical processes resulting in a reduction of plant growth, inhibition of photosynthesis and respiration, and degeneration of main cell organelles (Vangronsveld and Clijsters 1994). At a cellular level, heavy metal(loid)s may further block functional groups of biologically important molecules, displace and/or substitute essential metal ions, denature and inactivate enzymes and disrupt cell or membrane integrity (Ross and Kaye 1994).
Generally, for essential elements such as Ca, there is a zone of deficiency followed by a zone of tolerance where soil concentrations are adequate for plant needs. If external concentrations exceed toxicity threshold, a breakdown in metabolic control followed by passive uptake and eventual death occurs. For non-essential elements such as As, there is no deficiency zone and tissue concentrations generally increase until the external concentration is toxic and results in
8 Chapter 1: General introduction & background eventual death. These relationships are limited to plants that do not exhibit heavy metal(loid) tolerance strategies. The response of plants to essential and non essential metal(loid)s is illustrated in Figure 1.2.
Tolerance Toxicity
(Optimum level) Non harmful Toxicity threshold
Deficiency Le thal Growth Response Growth Response Growth Growth Response Growth Response Growth
Soil Heavy Metal(loid) Concentration
Figure 1.2 A dose response curve for essential (solid line) and non-essential (broken line) elements in plants. Modified after Morel (1997).
1.3.1 Heavy metal(loid) tolerance The phenomenon of heavy metal(loid) tolerance is ubiquitous, a characteristic found in all organisms in order to maintain the concentration of essential metals within physiological limits and to minimise the detrimental effects of non-essential metals (Clemens 2001). In plants, heavy metal(loid) tolerance is thought to be a constitutive property of all cells, tissues and organs (Ernst et al. 1992). Macnair et al. (2000) defined tolerance as the ability of a plant to survive in soils toxic to other plants of the same or different species, evident by an interaction between a genotype and its environment. According to Baker (1987) tolerance is the ability of a plant to survive internal stress. He suggested that tolerance was conferred by the possession of physiological mechanisms enabling growth on heavy metal(loid) rich soils, and implied a genetic basis, hence heritable tolerance. Macnair (1993) implied that a knowledge of a simple genetic basis might be useful in the exploitation of tolerance, by using genetic engineering techniques to clone tolerant genes.
1.3.2 Plant tolerance strategies Plants have developed unique strategies to protect themselves against heavy metal(oid) stress. Baker (1981) proposed two contrasting strategies of plant tolerance, namely exclusion and accumulation . In exclusion, the heavy metal(loid) concentration in the aboveground biomass
9 A. G. Kachenko is maintained at a low concentration with roots acting as barriers. For example, several studies that have examined accumulation and partitioning of heavy metals in mangroves indicate greater heavy metals concentrations in roots (Peters et al. 1997; MacFarlane et al. 2007). Exclusion is effective up to a certain threshold concentration with a leaf:root heavy metal(loid) concentration ratio of < 1 after which the mechanism breaks down resulting in unrestricted uptake and possible death (Baker 1981; Baker et al. 2000). Physiological mechanisms enable these species to survive in the presence of high concentrations of heavy metal(loid)s. Mechanisms may include a change in the metal binding capacity of cell walls, exudation of chelating substances in the rhizosphere or altered membrane permeability (Ghosh and Singh 2005). Colonisation of plant roots may also provide a barrier to heavy metal(loid) uptake (Baker and Walker 1990). These mechanisms are further discussed in Section 1.4.1.
Conversely, in the accumulation strategy, plants can actively accumulate high levels of heavy metal(loid)s in the above-ground biomass without adverse affects on plant growth until soil conditions become toxic and plant growth suppressed. These species are characterized by a leaf:root heavy metal(loid) concentration ratio of > 1 (Baker et al. 2000). Plants that fall into this category are termed heavy metal(loid) hyperaccumulators and are described in Section 1.3.3.
In addition to these tolerance strategies, it is not uncommon for some species to be referred to as indicators , defined as plants exhibiting a proportional relationship between soil and plant heavy metal(loid) concentrations. Indicators accumulate heavy metal(loid)s into their above- ground tissues at similar levels to those in the soil and often results in a reduction of plant growth as soil concentrations increase (Baker 1981). Indicator species may signify areas of soil heavy metal contamination and can be utilised in prospecting for potential ore bodies (Brooks 1976; Brummer and Woodward 1999). In Figure 1.3, the three conceptual soil-plant relationships and their corresponding soil-plant growth relationships are presented.
10 Chapter 1: General introduction & background
Excluder
(a) (b )
Accumulator
(c ) (d )
Plant growth Plant
Indicator
Heavy metal(loid) concentration in plant in concentration metal(loid) Heavy (e ) (f )
Heavy metal(loid) concentration in soil
Figure 1.3 Plant tolerance strategies in response to increasing soil heavy metal(loid) concentrations, modified from Baker (1981; a, c, e). Corresponding plant growth characteristics in response to increasing soil characteristics, are also presented (Bhatia, 2003; b, d, f).
Excluders (a) restrict the movement of heavy metal(loid)s into their above-ground biomass up to a certain threshold concentrations where this strategy breaks down resulting in a sudden increase in heavy metal(loid) concentration and the onset of severe toxicity (b). Accumulators (c) actively concentrate heavy metal(loid)s into their above-ground tissues until soil conditions become toxic and plant growth suppressed (d). Indicators (e) accumulate heavy metal(loid)s into their above-ground tissues at similar levels to those in the soil and often results in a reduction of plant growth as soil concentrations increase (f).
11 A. G. Kachenko
1.3.3 Heavy metal(loid) hyperaccumulation The term hyperaccumulator was first used by Brooks et al. (1977), to describe a plant species that could accumulate substantial amounts of a given heavy metal(loid) in aboveground tissue without deleterious effects to the plant. This definition was developed with a focus on Ni hyperaccumulating plants, however, several hyperaccumulators for As, Cd, Cu, Co, Mn, Pb, Se, Ti and Zn have also been described. Arbitrary threshold guideline concentrations have been set for hyperaccumulating species for different elements. For example, Cd concentrations > 100 mg kg –1 (0.01% dry weight; DW), As, Cu, Cr , Ni and Pb concentrations > 1,000 mg kg–1 (0.1% DW) and Zn concentrations > 10,000 mg kg –1 (1% DW) (Baker and Brooks 1989; Cai and Ma 2003).
The first documented evidence of hyperaccumulation occurred in the 19 th Century with the identification of two species growing on Zn and Cd rich soils (Baumann 1885). Violet ( Viola caliminara) and mustard Thalaspi calaminara (later called Thalaspi caerulescens ) accumulated approximately 1% and 1.7% Zn DW in leaves on DW basis, respectively. Some fifty years passed before the accumulation of Se (0.6% DW basis) was discovered by Byers (1935) in plants of the genus Astragalus . Soon after, plants capable of hyperaccumulating Ni were identified by Minguzzi and Vergnano (1948), who observed 1% Ni DW in leaves of Alyssum bertolonii growing on Ni-enriched serpentenitic soils.
Since these early findings, over 450 species from 45 families of metal hyperaccumulators have been identified (Prasad and Freitas 2006). A large number of those identified belong to the Brassicaseae and Cruciferae families, and when combined, contain more than 150 hyperaccumulating species (Palmer et al. 2001). A summary of hyperaccumulating taxa identified to date is presented in Table 1.2. Since the initial discovery of As hyperaccumulating Pteris vittata (Ma et al. 2001), additional As hyperaccumulating species have been reported (e.g. Wang et al. 2007), and it is likely that additional heavy metal(loid) hyperaccumulating species will be discovered in future field and screening investigations. Reeves and Baker (2000) suggested some important considerations when assessing the validity of hyperaccumulation data, such as possible contamination of samples from soil and aerial fallout, and errors with analytical methodologies. For example, Wild (1974b) reported high Cr concentrations in leaves of Dicoma niccolifera ( ca. 1,500 mg Cr kg –1 DW) and Sutera fodina ( ca. 2,400 mg Cr kg –1 DW) near a chrome mine in Zimbabwe, however, in a later
12 Chapter 1: General introduction & background
study by Brooks and Yang (1984), Cr concentrations were up to 20 times lower in the same species. Reeves and Baker (2000) suggested that the samples in the previous study may have been contaminated with windblown ore dust resulting in spurious concentrations.
Table 1.2 Heavy metal(loid) hyperaccumulating plants. Normal concentrations Element Threshold value (mg kg –1)a Number of reported taxa range in plants (mg kg –1) Arsenic 0.3–6 1,000 20 Cadmium 0.03–20 100 3 Cobalt 0.03–2 1,000 26 Copper 5–25 1,000 37 Lead 0–10 1,000 14 Manganese 2–2000 10,000 11 Nickel 0.2–100 1,000 330 Zinc 5–2,000 10,000 18 a Concentration in above-ground biomass required to be classified a hyperaccumulator. Data summarised from: (Brooks et al. 1998; Baker et al. 2000; Reeves and Baker 2000; Kabata-Pendias 2001; Ma et al. 2001; Francesconi et al. 2002; McGrath et al. 2002; Zhao et al. 2002; Ashley et al. 2003; Chen et al. 2003; Meharg 2003; Reeves 2003; Xue et al. 2004; Du et al. 2005; Sridokchan et al. 2005; Srivastava et al. 2006; Wang et al. 2006; Wang et al. 2007).
1.3.4 Serpentine soil: A hyperaccumulator host environment Serpentine soils are a noteworthy example of naturally occurring metalliferous environments and are considered one of the most important soil types to host hyperaccumulating species (Reeves and Baker 2000). The term serpentine refers to three main mineral polymorphs
formed by the hydration of ferromagnesium minerals, (MgFe) 2SiO 4, and include antigorite, chrysotile and lizardite (Malpus 1992). Although serpentine may also be present, it is considered a misnomer to classify soils as serpentine with ultramafic a more appropriate term emphasising their high magnesium (Mg) and iron (Fe) concentrations (Proctor 1999). Nevertheless, serpentine is still frequently used by ecologists and botanists to describe soils derived from ultramafic (ultra-magnesium-ferric) rocks containing minerals such as olivine, pyroxene and augite (Brooks 1987; Wenzel et al. 2003; Doherty et al. 2008). Serpentine soils derived from ultramafic rocks occupy < 1% of the Earths surface and occur across several continents. In Australia, serpentine soils occur mainly in Queensland, New South Wales, Victoria and Western Australia (Davie and Benson 1997).
13 A. G. Kachenko
Serpentine soils are often enriched with heavy metal(loid)s such as Ni, Cr and Co; and are characterised by a low calcium to magnesium ratio. These soils are typically deficient in essential plant nutrients such as nitrogen (N), potassium (K), and phosphorus (P). In addition, serpentine soils typically present unfavourable physical conditions. For example, serpentine environments are often steep, rocky and are consequently exposed to increased erosion risk resulting in shallow soils (Figure 1.4). Moreover, serpentine environments are often low in silt and clay fractions and combined with chemical limitations, yield a hostile environment with little moisture and depressed nutrient levels (Brady et al. 2005).
Figure 1.4 A typical serpentine site (Nature Reserve of Monterufoli) in Tuscany, Italy showing lack of vegetation cover and steep, rocky slopes (May, 2006).
Despite adverse growth conditions, serpentine soils can sustain an abundance of unique serpentine endemic flora species and communities (Brady et al. 2005). Many of the identified Ni hyperaccumulating species were discovered on serpentine soils and it has been recently suggested that studies of serpentine systems may result in further discoveries of hyperaccumulating species (Reeves and Baker, 2000). Serpentine species typically exhibit several physiological and morphological modifications that induce survival under such hostile growing conditions. Some of these modifications include xerophytism (adaptation to drought), nanism (dwarfism), glaucescence (epicuticular waxiness), plagiotropism (oblique or horizontal growth) and erythrism (colour changes; anthocyanic, chlorotic) (Kruckeberg 1992).
14 Chapter 1: General introduction & background
1.3.5 Heavy metal(loid) hyperaccumulators in Australia Several Australian ecological studies have reported heavy metal(loid) accumulating or tolerant plant species (e.g. Khan et al. 1998; Archer and Caldwell 2004), however, few have identified Australian hyperaccumulating species. A study by McCray and Hurwood (1963) in north-western Queensland reported two Se hyperaccumulators in the Leguminosae (Fabaceae) family, Acacia cana (1,121 mg Se kg –1 DW) and Neptunia amlpexicaulis (4,334 mg Se kg –1 DW). In the Bulman-Waimuna Springs area, Northern Territory, Cole et al. (1968) reported up to 1,044 mg Pb kg –1 DW in Polycarpaea synandra (Family Caryophyllacea). In a field survey of industrially contaminated soils, Tam and Singh (2004) identified Leptospermum juniperinum as a possible Pb hyperaccumulator and Nephrolepis cordifolia as a possible Cu hyperaccumulator, with concentration of 1,029 mg kg–1 Pb DW and 2,324 mg kg –1 Cu DW, respectively. Controlled glasshouse experiments are required to confirm the hyperaccumulation status of these species. Bidwell et al. (2002) identified Austromyrtus bidwillii (nomenclatural synonymous with Gossia bidwillii) as Australia’s first Mn hyperaccumulator with Mn concentrations of up to 19,200 mg kg –1 DW and 26,500 mg kg –1 DW in leaves and young bark, respectively. Aside from these species, there are three known Australian Ni hyperaccumulators: Hybanthus floribundus (3 subspecies; 13,500 mg Ni kg –1 DW) in the Violaceae family (Severne and Brooks 1972), Stackhousia tryonii Bailey (41,300 mg Ni kg –1) in the Stackhousiaceae family (Batianoff et al. 1990) and Pimelea leptospermoides (1,620 mg Ni kg –1 DW) in the Thymelaeaceae family (Batianoff and Specht 1992). The latter two are rare species confined to serpentine soils of central Queensland. Conversely, Hybanthus species are different in that some taxa are not restricted to serpentine soils and are scattered throughout southern Australia in New South Wales, Victoria, South Australia and Western Australia (Elliot and Jones 1990). Moreover, it was recently shown that Hybanthus floribundus subsp. floribundus ecotypes from non-serpentine soils were capable of hyperaccumulating Ni, suggesting that hyperaccumulation in this species may be a constitutive trait (Bidwell 2001). These unique characteristics make H. floribundus subsp. floribundus an ideal species for further investigation in particular non-serpentine endemic populations. In this thesis, ecophysiological aspects of H. floribundus subsp. floribundus hyperaccumulation have been investigated in populations sourced from South Mandurang, Victoria, Australia.
15 A. G. Kachenko
Several studies have reported hyperaccumulating species in Australia, though in all cases, there has been a lot of variation among replicates in heavy metal(loid) uptake. For example, in a survey of Queensland serpentine soils, both Commelina ensifolia and Rostellularia adscendens var. hispida species were reported to exhibit Ni concentrations > 1,000 mg kg –1 DW, however, in both cases, only a single specimen contained abnormal Ni levels (Reeves 2003). It is possible that there are several other studies where hyperaccumulators have been falsely reported and warrants reinvestigation of the species in question for confirmation of their hyperaccumulation status.
1.3.5.1 Background on Hybanthus floribundus There are some 150 species of Hybanthus (native violets) throughout the world with the main populations in Central and South America. In Australia, there are eleven species of which three have been reported as Ni hyperaccumulators (Bennett 1972; Severne 1974; Bidwell 2001). Two of the species, H. floribundus subsp. curvifolius and subsp. adpressus are restricted to serpentine soils in the Ravensthorpe and Norseman regions of Western Australia, respectively, whereas the third species, subsp. floribundus is found throughout southern parts of Australia (Bennett 1972; Severne and Brooks 1972; Severne 1974).
Hybanthus floribundus species are typical under story shrubs that generally occur on sandy and gravelly soils amongst other low shrubs or in Mallee country across southern Australia (Elliot and Jones 1990). These species are perennial woody shrubs, up to 1.5 m in height and subspecies of this genus are distinguished by distinct morphological features such as leaf orientation. For example, subspecies adpressus has narrow conduplicate leaves which are appressed upwards against the stems and pale blue petaloid sepals, whereas, subspecies curvifolius has distinctly curved conduplicate leaves with an uncinate apex and dark blue to green sepals. The greatest morphological variation exists in the subspecies floribundus (Figure 1.5) which has pale blue petaloid sepals or dark green or blue non-petaloid sepals, however, leaves are typically broad and flat (Bennett 1978). Hybanthus floribundus species can be characterized by their poor germination rate (Roche et al. 1997) which may be overcome by using tissue culture micropropagation techniques as described by Bidwell et al . (2001).
16 Chapter 1: General introduction & background
(a) (b)
(c)(c)(c) (d)
Figure 1.5 A typical Hybanthus floribundus subsp. floribundus in natural habitat in Bendigo Regional Park, South Mandurang, Victoria.
(a) Typical habitat (April 2006); (b) A young H. floribundus subsp. floribundus growing naturally in this habitat (March 2006); (c) Filter paper impregnated with a 1% dimethylglyoxime solution returns a magenta colour upon crushing a sample of H. floribundus subsp. floribundus suggesting Ni accumulation (February 2007); (d) inflorescence (June 2007).
17 A. G. Kachenko
1.3.6 Ecological significance of hyperaccumulation It has been proposed that heavy metal(loid) hyperaccumulation may have evolved as an adaptive trait, possibly as an extension of the metal tolerance trait, however, the current view is that the trait itself evolved independently in various genera (Assunção et al. 2003). Boyd and Martens (1992) proposed the following five hypothetical selective forces that could potentially explain the evolutionary/ecological significance of the heavy metal(loid) hyperaccumulating trait: (i) Chemical defence against herbivores or pathogens; (ii) Xeromorphic adaptation (drought resistance); (iii) Tolerance or disposal of soil metal ions; (iv) Interference with neighbouring plants; (v) Inadvertent uptake.
Boyd and Martens (1998) indicated that hypotheses (i) to (iv) suggest an ecological benefit accrues to heavy metal(loid) hyperaccumulating species whereas inadvertent uptake (v) attributes no selective value to heavy metal(loid) hyperaccumulation. These authors suggested that this trait is a by-product of other physiological processes in hyperaccumulating species.
To date, the defence mechanism against herbivory (e.g Pollard and Baker 1997; Boyd and Moar 1999) as well as fungal and bacterial pathogens (Boyd et al. 1994; Ghaderian et al. 2000) have been tested in great detail and less attention has been given to the remaining hypotheses (Boyd and Martens 1998). For example, Rathinasabapathi et al. (2007) exposed nymphs of the grasshopper Schistocerca americana to As-laden (46.4 mg kg –1 As) and untreated (2.9 mg kg –1 As) Pteris vittata fronds for a period of two days. Significantly less frond material was consumed in S. americana exposed to As-laden fronds compared to untreated fronds, and significantly more faecal pellets and increased body weight were observed in S. americana fed untreated ferns. The authors suggested that hyperaccumulation in P. vittata was an elemental defence against S. americana herbivory.
Studies investigating the remaining hypotheses are limited, speculative and contradictory in their nature. For example, the reduction of cuticular transpiration (Severne 1974) and osmotic adjustment (Baker and Walker 1990) have been proposed as possible mechanisms arising from hyperaccumulation in order to provide tolerance to drought stress. Severne (1974)
18 Chapter 1: General introduction & background suggested that Ni accumulation in epidermal cells of H. floribundus reduced cuticular transpiration and proposed Ni hyperaccumulation was a xeromorphic adaptation, thereby suggesting plants were resistant to drought stress. Recently, Bidwell et al. (2004) reported low K concentrations in the epidermal vacuoles of H. floribundus subsp. curvifolius leaves rich in Ni, and speculated that this may be attributed to osmotic or electrical adjustment within cells. Bhatia et al. (2005a) suggested that Ni hyperaccumulation in Stackhousia tryonii was involved in osmotic adjustment and protection of plants against drought. In contrast, Whiting et al. (2003) demonstrated that Ni hyperaccumulation in A. murale and Zn hyperaccumulation in T. caerulescens did not enhance their resistance to water stress as metal hyperaccumulation did not ameliorate growth inhibition in response to drought treatments applied.
In summary, various aspects pertaining to the ecological significance of hyperaccumulation remains unclear in the literature, particularly in relation to the physiological role of heavy metal(loid)s in osmotic adjustment. Furthermore, the ecological significance of hyperaccumulation has been experimentally explored with few hyperaccumulating species and it is possible that the ecological significance of hyperaccumulation is genotype specific.
1.4 MECHANISMS OF HEAVY METAL(LOID) TOLERANCE
Tolerant plants have evolved a complex network of homeostatic mechanisms that serve to control heavy metal(loid) uptake, accumulation, trafficking and detoxification (Clemens 2001). Heavy metal(loid) homeostasis is the fundamental concept underlying tolerance in plants and involves both cellular and extracellular mechanism. For example, Tong et al. (2004) classified cellular mechanisms for heavy metal(loid) tolerance into two basic strategies and suggested that in the first strategy, plants aimed to keep a low concentration of the heavy metal(loid) ions in the cytoplasm by preventing the heavy metal(loid)s from being transported across the plasma membrane. They suggested that this was achieved by increased binding of heavy metal(loid) ions to the cell wall, by reducing uptake through modified ion channels or by pumping the heavy metal(loid) out of the cell with active efflux pumps. The authors proposed that the second strategy involved the detoxification of heavy metal(loid) ions entering the cytoplasm through inactivation via chelation, conversion into a less toxic form or compartmentalization. These two strategies are an extension on the principles of plant metal tolerance, outlined by Baker (1981) and can be defined as either exclusion/avoidance or accumulation and sequestration tolerance mechanisms (Figure 1.6).
19 A. G. Kachenko
PC -M PC -M-S
Figure 1.6 An overview of possible mechanisms involved in heavy metal(loid) tolerance. Modified after Marschner (1997).
Mechanisms may include: 1. Restriction into roots by mycorrhizas; 2. Binding to cell wall and root exudates; 3. Reduced efflux across the plasma membrane; 4. Active efflux into the apoplast; 5. Chelation in cytosol by various ligands (PCs: phytochelatins; MTs: metallothioneins) 6. Repair and protection of plasma membrane under stress conditions (HSP: heat shock proteins); 7. Transport of phytochelatin-metal(loid) complex into the vacuole and 8. Transport and storage in vacuoles.
1.4.1 Avoidance tolerance mechanisms 1.4.1.1 Mycorrhizal associations Mycorrhizas may impact on the availability of heavy metal(loid)s for plant uptake, however, their exact role has not been established (Whiting et al. 2001). The presence of mycorrhizal associations between fungi and roots of host plants growing on heavy metal(loid) contaminated soils, indicates an important relationship in plant tolerance and accumulation (Killham and Firestone 1983; Marschner 1997; Jentschke and Godbols 2000). Ectomycorrhizas (ECO) and arbuscular mycorrhizas (AM; previously referred to as vesicular arbuscular mychorriza or VAM) are the two most common types of mycorrhizal associations in plants growing on heavy metal(loid) contaminated soils (Leyval et al. 1997). Several studies indicate that AM reduce the influx of heavy metal(loid)s into host plants and essentially provide an effective exclusion barrier to heavy metal(loid) uptake. For example, Joner and Leyval (1997) investigated the uptake of Cd in Trifolium subterraneaum in symbiosis with AM Glomus mosseae and reported that extraradical hypae of AM could transport Cd from soil to roots, however, the transfer from fungus to plant was restricted as
20 Chapter 1: General introduction & background
Cd was sequestered in hyphae. Several mechanisms to explain how mycorrhizas reduce the influx of heavy metal(loid)s have been proposed and include absorption by the hyphal sheath, reduced access to the apoplast due to the hydrophobicity of the fungal sheath, chelation by fungal exudates and adsorption onto the external mycelium (Hall 2002).
Several studies have suggested that AM colonisation can enhance the uptake of heavy metal(loid)s in host plants at higher soil concentrations. Agely et al. (2005) reported that AM increased frond dry mass and As uptake in Pteris vittata treated with 100 mg kg –1 As, however, no effect was observed when ferns were treated with 50 mg kg –1 As. Similarly, Turnau and Mesjasz-Przybyłowicz (2003) noted that Berkheya coddii Roessler, B. zeyheri , Senecio coronatus and S. anomalochrous growing on ultramafic soils in South Africa were consistently colonised by AM fungi, with the abundant formation of arbuscules.
1.4.1.2 Rhizosphere effects 1.4.1.2.1 Root exudates Root exudates consist of a combination of diffusates, excretions and secretions released into the rhizosphere zone by plants. Diffusates include; simple and complex sugars, organic acids, amino acids, water and inorganic ions. Excretions include; carbon dioxide, bicarbonate ions, protons and ethylene whereas secretions include; mucilage, protons, siderophores and allelopathic compounds (Uren and Raisenauer 1988; Marschner 1997). There are few studies that address the role of root exudates in heavy metal(loid) chelation. For example, Zhao et al. (2001) studied the influence of root exudates (total soluble organic carbon) on Cd and Zn mobilization in ecotypes of the hyperaccumulator T. caerulescens , and two non-accumulators, wheat and canola. The results indicated that the root exudates from the hyperaccumulator plants did not significantly enhance mobilization of Cd and Zn and the authors suggested that they were not involved in Zn or Cd hyperaccumulation. This finding is consistent with a study by Salt et al. (2000) who found no evidence of the presence of any high-affinity Ni-chelating compounds in the root exudates of the Ni hyperaccumulator Thlaspi goesingense . However, elevated histidine and citrate (possible Ni-chelators) were identified in the root exudates of the non hyperaccumulator Thlaspi arvense , suggesting that root exudates may reduce Ni uptake and thus play a role in Ni tolerance in this species.
21 A. G. Kachenko
1.4.1.2.2 Rhizosphere acidification Several studies have examined rhizosphere acidification as a result of H + extrusion from roots and suggest no direct correlation between rhizosphere acidification and heavy metal(loid) uptake in hyperaccumulating and non hyperaccumulating plants (McGrath et al. 2001; Jauert et al. 2002). McGrath et al. (1997) investigated the effect of pH on Zn mobilisation in Thlaspi caerulescens and non hyperaccumulating Thlaspi ochroleucum in a pot experiment using a rhizobag technique. The authors found the degree of rhizosphere acidification between the two species was insignificant and the authors suggested that Zn accumulation by T. caerulescens was not related to rhizosphere acidification. Instead, the authors suggested that T. caerulescens was effective in removing Zn from less soluble fractions. These findings are similar to those of Bernal et al. (1994) who observed that the acidification and reducing activity of Ni hyperaccumulating A. murale roots were smaller than that of non- hyperaccumulating Raphanus sativus , and suggested that metal solubilisation in A. murale did not involve rhizosphere acidification or the release of reductants from roots.
1.4.1.3 Cell wall Plant cell walls act as a continuous matrix consisting of a network of cellulose, hemicellulose (including pectins) and glycoproteins, the latter representing up to 10% of the dry weight of the cell walls (Marschner 1997). In this network, a variable proportion of the pectins consists of polugalacturonic acid acting as a cation exchanger, holding variable quantities of heavy metal(loid)s and providing for some exclusion (Ernst et al. 1992). The cell wall associated heavy metal(loid)s are bound to carboxylic groups (R −COO –) of polygaacturonic acids, to which < 10% of the total cellular heavy metal(loid)s are bound (Ernst et al. 1992).
The involvement of cell walls in plant tolerance to heavy metal(loid)s has been controversial and frequently disputed, with limited available literature (Hall 2002). Bringezu et al . (1999) investigated the intracellular distribution of heavy metals (Cu, Ni and Zn) in the leaves of Silene vulgaris using energy dispersive X-ray analysis (EDXA). The results showed that heavy metal concentrations were highest in the outer and anticlinal cell walls of epidermal cells, whereas in the parenchyma and bundle cell walls, trace concentrations were found. The binding state of Cu and Zn were also reported using electron energy loss spectroscopy (EELS). Copper was tightly bound to a protein with oxalate oxidase activity, suggesting a high homology to germin, whereas Zn was accumulated in the cell wall as silicate. In a
22 Chapter 1: General introduction & background similar study, MacFarlane and Burchett (2000) examined the distribution of Cu, Pb and Zn in the root and leaf tissues of the grey mangrove, Avicennia marina using EDXA. In roots, EDXA revealed accumulation of these metals predominantly in epidermal, endodermal, ground tissue, xylem and phloem cell walls. In leaf tissues, Cu and Zn concentrations were higher in cell walls than intracellular concentrations, however, Pb concentrations were not detected. The authors suggested that the root epidermis provided a barrier in the transport of Pb from root to shoot.
Localisation of heavy metals in the cell walls of hyperaccumulating plants has also been observed. Frey et al. (2000) using EDXA studied subcellular localisation of Zn in leaves and roots of hyperaccumulator Thalaspi caerulesces . In leaves, the authors found highest Zn concentrations in leaf epidermal (177 ± 36 mmol kg –1) and mesophyll (143 ± 32 mmol kg –1) cell walls and hypothesised that cationic Zn was bound to the anionic sites in the cell wall. Similarly, in roots Zn was detected in the cortical cell walls (9 ± 4 mmol kg –1). The authors hypothesised that Zn bound to the negatively charged sites of cell walls may have contributed substantially to its accumulation in roots; however, this observation was seen as less significant in Zn tolerance as concentrations were five times higher in shoots than in roots. In Ni hyperaccumulating Hybanthus floribundus , the pectic components of the cell walls have been suggested to be involved in Ni storage (Severne 1974; Farago et al. 1975). In a recent study, Bidwell et al. (2004) using EDXA revealed that Ni in leaves of Hybanthus floribundus subsp. curvifolius was predominantly localised in the vacuoles of epidermal cells (128 ± 12.7 mmol kg –1) as well as the outside of epidermal cell walls (46 ± 13.1 mmol kg –1). In Ni hyperaccumulating T. goesingense 68 ± 6% of total leaf Ni was associated with the cell wall, with the remaining associated with citrate 28 ± 7% or histidine 4 ± 3% (Krämer et al. 2000).
1.4.1.4 Plasma membrane Following mobilization in the rhizosphere, heavy metal(loid) ions in soil solution need to be captured by the root cells and transported across the root cell plasma membrane. This membrane is considered the first functional site of contact between root cells and heavy metal(loid) ions (Astolfi et al. 2005). Transport across the plant plasma membrane is driven against a prevailing electrochemical gradient of protons generated by plasma membrane H +- ATPases and the presumably very low activity of cytosolic metals (Welch and Norvell 1999). These primary transporters pump protons out of the cell, resulting in pH and electrical
23 A. G. Kachenko potential differences across the plasma membrane. As a result, a large negative membrane potential provides enough energy to drive uptake of metal ions (Guerinot 2000). Secondary transporters such as channel proteins and/or H +-coupled carrier proteins then utilize these gradients to facilitate uptake of metal ions (Mäser et al. 2001). Studies have shown that heavy metal sensitive H+-ATPases enzyme systems and the lipid and protein constituents of plasma membranes, render plasma membranes susceptible to heavy metal toxicity (Quartacci et al. 2001; Astolfi et al. 2003).
Phylogenetic studies have been used to identify cation transporters that may be involved in mediating transport across plant membranes (Mäser et al. 2001), which act as lipophilic barriers restricting the free movement of charged ions (Lasat 1999). Transport proteins play important roles in several steps of nutrition and plant signalling, and mediate uptake in root cells, transfer between cells and organs and thus, are considered essential in maintaining intracellular ion homeostatis (Thomine et al. 2000; Mäser et al. 2001). Several transporters have been identified by supplementing Arabidopsis genetics and genomics with the power of yeast and bacterial genetics (Clemens 2001), however, many plant transporters remain to be identified at a molecular level. The transporters that are thought to assist in the uptake of micronutrients are in the zipper integrating (ZIP; Zn regulated transporters [ZRT], Fe- regulated transporters [IRT] – like protein), natural resistance-associated macrophage protein (NRAMP) and cation diffusion facilitator (CDF) families (Mäser et al. 2001). Studies have indicated that transporters have a broad substrate range. For example, Korshunova et al. (1999) studied the substrate range of IRT 1 when expressed in yeast and showed that it could transport Fe, Mn, Zn and possibly Cd, enabling heavy metals to enter cells. Further studies are warranted to understand the exact nature of transport proteins and their involvement in uptake of heavy metal(loid)s, particularly following the recent discoveries of As hyperaccumulating species.
1.4.1.5 Metal efflux P-type ATPases constitute a large ubiquitous group of proteins involved in the transport of heavy metals across biological membranes. P-type ATPases can be divided into many subfamilies on the basis of both sequence and functional similarities, and include H +-ATPases
(type 3 A) and heavy metal transporting ATPases (type 1 B) (Axelsen and Palmgren 1998). P- type ATPases that mediate the efflux of heavy metals have been identified in certain bacteria
24 Chapter 1: General introduction & background strains (Ghosh and Rosen 2002). Various studies have indicated their importance in heavy metal homeostasis, transport and tolerance (Williams et al. 2000; Axelsen and Palmgren
2001; Hall and Williams 2003). The first P 1B -type ATPase to be cloned and characterised was AtHMA4 from Arabidopsis of the Zn/Co/Cd/Pb group (Mills et al. 2003). It was found that this protein conferred Cd resistance when heterologously expressed in a wild strain of Saccharomyces cerevisiae . Moreover, it was reported that AtHMA4 appeared to reduce the Zn sensitivity of the Escherichia coli zntA mutant and may have a role in Cd and Zn transport. Recently, Papoyan and Kochian (2004) identified ATPase, TcHMA4 , in T. caerulescens and demonstrated that it could mediate yeast metal tolerance via active efflux of Cd, Cu, Pb and Zn. However, when TcHMA4 was compared with its homologue in Arabidopsis thaliana (AtHMA4 ), tissue-specific and metal-responsive expression was inconsistent suggesting TcHMA4 was not responsible for tolerance in T. caerulescens . Studies on the molecular aspects of heavy metal(loid) tolerance are increasing and it is likely that numerous plant transporters remain to be identified (Yang et al. 2005). Characterisation of heavy metal(loid) transport and efflux genes will shed light on the molecular aspects of heavy metal(loid) tolerance and may be used in the future to enhance heavy metal(loid) uptake in hyperaccumulating species.
1.4.2 Accumulation tolerance mechanisms 1.4.2.1 Plant Ligands Several metal binding ligands have been reported to sequester, transport or store the accumulated heavy metal(loid). Ligands that have been identified include peptides, organic acids and amino acids. Peptides include cysteine-rich phytochelatins (PCs) and metallothioneins (MTs) that contain a sulphur functional group (–SH), which is thought to complex with thiol-reactive metals. In addition to these peptides, a large number of low molecular weight (LMW) oxygen and nitrogen donor ligands, namely organic (carboxylic) and amino acids have also been identified. A selection of these is provided in Figure 1.7.
25 A. G. Kachenko
(a)
glutamic acid-cysteine glycine
(b) (c)
Figure 1.7 Chemical structure of (a) phytochelatins, (b) histidine and (c) citric acid adapted from Callahan et al. 2006.
26 Chapter 1: General introduction & background
1.4.2.1.1 Phytochelatins Phytochelatins (PCs) are low molecular weight, metal(loid) binding peptides composed of three amino acids; glutamate (Glu), cysteine (Cys) and glycine (Gly), with the general structure ( γ-Glu-Cys) n-X, where n = 2-11 and X is commonly Gly (Zenk 1996; Welch and Norvell 1999). Iso-PC molecules are homologues to the chemical structure of PCs, where β- alanine, serine and Glu are substituted for the carboxyl terminal Gly (Cobbett 2000). In the structure of these compounds, Glu occupies the N-terminal position and Cys is bound to its γ- COOH group, forming a γ-peptide bond instead of a α-bond found in all proteins (Figure 1.7a). The primary structure of PCs and iso-PCs suggests that they are structurally related to glutathione. The reduced form of glutathione, GSH, is a tripeptide ( γ-Glu-Cys-Gly) and an important precursor in PC biosynthesis (Foyer et al. 2001; Tomaszewska 2002).
Phytochelatin synthesis in plant cells and PC synthase activity can be induced by the presence of heavy metal(loid) ions listed in order of their decreasing induction ability: Cd 2+ > Pb 2+ > Zn 2+ > Bi 3+ > Ag + > Hg 2+ > As 5- > Cu + > Sn 2+ > Au 3+ > Bi 3+ (Tomaszewska 2002). However, the only heavy metal(loid)-PC complexes that have been isolated from plants contain ions of Cd, Ag, Pb and Hg (Cobbett and Goldsbrough 2002). The PCs bind heavy metal(loid)s in vitro with organic sulphur (R-SH- thiolate bonds) on the cysteine residues of these peptides. Two complexes are formed as a result of these associations, either low or high molecular weight in nature (Zenk 1996). The resulting metal(loid)-PC complex is actively transported from the cytosol across the tonoplast into the vacuole. Vogeli-Lange and Wagner (1990) studied mesophyll protoplasts derived from tobacco plants (Nicotiana rustica var Pavonii) exposed to Cd and reported that Cd is transported into the vacuole by complexation with phytochelatins (PCs) followed by active transport of the PC–Cd complex across the tonoplast. Zenk (1996) reported that once in the vacuole, the metal(loid) can be bound to organic acids after which the PC-peptide is degraded and the original amino acids can re-enter the cytosol.
Several studies have failed to identify PCs in Ni hyperaccumulating species (eg. Sagner et al. 1998) or the coordination of Ni with S (Krämer et al. 1996; Krämer et al. 2000). Similarly, the role of PCs in As hyperaccumulating species is inconclusive. For example, Zhao et al . (2003) reported PC synthesis upon exposure to As in roots and shoots of Pteris vittata . Only
PC 2 (short-chain glycine C-terminal PC) was observed whereas studies involving other As tolerant plant species found PCs of longer length (e.g. Schmöger et al. 2000). Moreover, it
27 A. G. Kachenko was noted that concentrations of PCs were considerably lower than the values reported in other As tolerant plants under comparable periods of exposure, indicting that PCs play a limited role in the detoxification of As in P. vittata . These findings were confirmed by Zhang et al . (2004) who also reported the presence of PCs in P. vittata . They suggested that PCs were partly responsible for the detoxification of As and indicated that PC-independent sequestration of As in vacuoles may also be involved.
1.4.2.1.2 Metallothioneins Metallothioneins (MTs) are groups of Cys-rich cytoplasmic metal(loid) binding peptides similar to PCs. They typically have low molecular weights (Cobbett and Goldsbrough 2002) and in their reduced state, provide thiols for chelation (Rauser 1999). Metallothioneins are grouped into four classes based on the alignment of Cys or amino acid sequence, with the majority of plant MT genes categorized as either Type 1 or Type 2 (Cobbett and Goldsbrough 2002; Hall 2002). Metallothioneins have been identified in a range of plants including barley, maize, wheat, rice, cotton, kiwifruit, tobacco, tomato and coffee (Rauser 1999).
Metallothioneins act in the detoxification of several heavy metals in fungi and animals (Cai and Ma 2003). In plants, Cd, Cu and Zn associations with MTs have been studied (Goldsbrough 2000), however, their role in detoxification has not been established (Schat et al. 2000; Hall 2002). Murphy and Taiz (1995) studied the expression of MT genes in 10 select ecotypes of Arabidopsis seedlings and identified 2 ecotypes that displayed Cu tolerance, and further, demonstrated a positive correlation between Cu tolerance and MT2 gene expression. Van Hoof et al. (2001) screened a complimentary deoxyribonucleic acid (cDNA) library derived from a highly Cu tolerant population of Silene vulgaris using Arabidopsis -based MT probes and identified a MT2b-like gene. It was suggested that MTs were responsible for raising the level of Cu tolerance in highly Cu tolerant Silene vulgaris populations , however, when these populations were crossed with Cu sensitive populations, SvMT2b expression did not cosegregate with Cu tolerance. They concluded that overexpression of SvMT2b conferred Cu tolerance that was confined within the genetic background of a Cu tolerant plant.
28 Chapter 1: General introduction & background
1.4.2.1.3 Organic acids The association between organic acids and heavy metals is well documented (e.g. Kersten et al. 1980; Gabrielli et al. 1997; Bidwell et al. 2002; Bhatia et al. 2005b; Sun et al. 2006) however, the precise role of organic acids in tolerance remains unclear. Correlations between organic acids and the degree of heavy metal exposure have not been observed (Clemens 2001). For example, Tolrá et al. (1996) investigated the influence of Zn concentrations on organic acid levels in roots and shoots of T. caerulescens , and reported high concentration of organic acids, notably malate, irrespective of applied Zn concentrations. It was suggested that the variation in organic acid concentrations was associated with cation-anion balance rather than Zn tolerance and, furthermore, indicated that the high concentrations of organic acids observed in shoots was a constitutive property of this species. Shen et al. (1997) observed constitutively high concentrations of malate in leaves of Zn hyperaccumulator T.i caerulescens and non-hyperaccumulating T. ochroleucum , noting that they were not affected by applied Zn. It was suggested that high concentrations of malate may play a role in Zn chelation, however, the levels present were not sufficient to explain the species specificity of Zn tolerance and hyperaccumulation.
It is doubtful, whether the concentrations of organic acids identified in other hyperaccumulators are sufficient to contribute significantly to heavy metal sequestration and possible detoxification. For example, Lee et al. (1978) reported a strong correlation between Ni and citric acid concentrations in hyperaccumulators from New Caledonia, however, the molar ratio of Ni to citric acid in nearly half of the investigated samples was above 2, suggesting that there was insufficient citrate available in the system to complex Ni ions. Sagner et al . (1998) identified citric acid as the dominant organic acid in the latex of Ni hyperaccumulator Sebertia acuminata, although only 37% of Ni was bound to citrate and a further 50% was available as free ions. It was suggested that additional counter ions, in particular nitrates, may stabilise the Ni citrate complex, however, quantification was not achieved. Rengasamy and Doran (2003) studied heavy metal uptake and organic acid response in hairy roots of Cd hyperaccumulator T. caerulescens , and the Ni- hyperaccumulator, A. bertolonii . They observed high constitutive levels of citric, malic and malonic acids in the hairy roots of both species and reported only 13% of the total Cd in Thalaspi caerulescens hairy roots and 28% of the total Ni in A. bertolonii hairy roots were associated with organic acids.
29 A. G. Kachenko
Powerful non-invasive analytical tools such as X-ray absorption spectroscopy (XAS) have recently been applied to elucidate the nature of organic acid and metal complexes. Salt et al. (1999) employed XAS to study Zn hyperaccumulation in T. caerulescens tissue and reported that 21% of Zn in xylem sap was bound to citrate and speculated that the remaining Zn was likely to be present as free Zn 2+ ions. In shoots, 38% of total Zn was associated with citrate and 26% as aqueous ions. Recently, Perrier et al . (2004) using extended X-ray absorption fine structure (EXAFS) suggested Ni was bound to citrate in leaves of the Ni-hyperaccumulator S. acuminate , findings that support earlier observations (Sagner et al. 1998). X-ray absorption spectroscopy provides pertinent information of the ligand environment of the accumulated heavy metal(loid). These techniques have become more widely used in order to understand and unravel the biochemistry of heavy metal(loid) hyperaccumulation (Krämer et al. 2000; Pickering et al. 2006).
There have been few studies that have successfully quantified the involvement of organic acids in Hybanthus floribundus species. Severne (1974) first observed a small (molecular weight < 250) water-soluble cationic compound in Hybanthus floribundus leaf tissue, but its identity and function was not determined. In 1977, Lee et al. indicated that most of the hyperaccumulated Ni in New Caledonian Hybanthus caledonicus and H. austrocaledonicus leaf extracts was in the form of a negatively charged citronickelate complex, however, this complex was not quantified. The authors further investigated the relationship between citrate and Ni in leaves of the same Hybanthus species and reported a molar ratio (Ni to citric acid) between 2.04 and 2.62 (Lee et al. 1978). In both studies, the authors failed to determine if there were other dominant organic acids complexed with Ni or if citrate was a constitutive component of Hybanthus species. Similarly, Farago and Mahmoud (1983) suggested either soluble galacturonates or organic acids such as citrate were bound to Ni in Hybanthus floribundus leaves, however, could not identify complexes and whether they existed under physiological conditions. Bidwell (2001) attempted to quantify the organic acid content in Hybanthus floribundus subsp curvifolius sourced from serpentine soils using an array of techniques including high precision liquid chromatography (HPLC), gas chromatography- mass spectrometry (GC-MS) and enzymatic assays. The results showed only citric acid with an average molar ratio of 5:1. Bidwell (2001) suggested that there were insufficient concentrations of citric acid to complex accumulated Ni.
30 Chapter 1: General introduction & background
1.4.2.1.4 Amino acids Nitrogen-donor ligands, particularly free amino acids, are thought to play a role in heavy metal tolerance. For example, Krämer et al . (1996) using HPLC reported a 36-fold increase in the concentration of histidine in xylem sap of Ni-hyperaccumulating Alyssum lesbiacum after exposure to 0.3 mM Ni compared to control plants. Histidine is regarded as the most important amino acid involved in Ni hyperaccumulation, acting as a tridentate ligand through its carboxylate, amine and imadazole functions (Callahan et al. 2006). A linear relationship between histidine and Ni was observed in Ni hyperaccumulating A. murale and A. bertolonii species, and EXAFS analysis demonstrated that Ni was directly complexed with histidine in A. lesbiacum (Krämer et al . 1996). In addition, EXAFS analysis revealed no evidence of Ni coordination with S, inferring that phytochelatins were not involved. The question of whether or not histidine was a constitutive trait responsible for Ni transport and accumulation in the Alyssum genus was considered by exposing a non-hyperaccumulator, Alyssum montanum to histidine. There was a rise in histidine levels proportional to higher Ni tolerance and shoot Ni content, suggesting that histidine may indeed enhance Ni tolerance and transport in non- hyperaccumulating species. It was concluded that histidine may also be responsible for the Ni hyperaccumulating phenotype in Alyssum . In a later study, Krämer (2000) using µ-X-ray absorption near-edge structure (µ-XANES) spectroscopy demonstrated that 36 ± 4% of the total leaf Ni in non-hyperaccumulating T. arvense was associated with histidine as compared to 4 ± 3% in Ni hyperaccumulating T. geosingense . It was suggested that in T. arvense , Ni accumulated in the cytoplasm as Ni-organic complexes whereas in T. geosingense, Ni could effectively compartmentalise in the vacuole as a Ni-organic acid complex.
The involvement of histidine in hyperaccumulation and possible detoxification has been questioned. Persans et al. (1999) using molecular and biochemical techniques suggested that the histidine response was not a universal tolerance mechanism in Ni hyperaccumulating species. They reported no significant differences in histidine concentrations in xylem sap, roots and shoots of Ni hyperaccumulating T. goesingense and non-hyperaccumulating T. arvense following Ni exposure. Furthermore, the study found no changes in expression of three cDNA encoding enzymes involved in histidine biosynthesis. Salt (2001) concluded that histidine in T. geosingense was not involved in Ni hyperaccumulation and suggested that other mechanism such as metal transport proteins may be involved.
31 A. G. Kachenko
Amino acids other than histidine have also been identified as possible ligands in hyperaccumulating species. For example, Farago et al. (1980) reported serine + amide, aspartic acid and isoleucine as dominant amino acids in Ni hyperaccumulating Hybanthus floribundus , however, the study was inconclusive as the complexes dissociated under normal chromatographic conditions. In a subsequent study, Farago and Mahmoud (1983) indicated that serine, aspartic acid and glycine were the dominate amino acids in H. floribundus stem material, however concluded that the relationship between Ni and amino acids was obscure. Homer (1997) reported proline as the most abundant amino acid in aqueous leaf extracts of Ni hyperaccumulators Walsura monophylla , Phyllanthus balgooyi and Dichapetalum gelonioides subsp . tuberculatum. In addition, they observed high concentrations of glutamine in W. monophylla and suggested that future investigations should examine specific samples such as xylem sap to minimise spurious observations. Bhatia et al . (2005b) analysed xylem sap of Ni hyperaccumulating Stackhousia tryonii for changes in amino acid concentrations in plants exposed to low or high Ni levels. They recorded a 4 mM decrease in total amino acid concentrations in plants treated with high Ni and a > 60% decrease in glutamine concentrations with a corresponding increase in alanine, aspartic acid, and glutamic acid upon high Ni exposure. From this study, it was concluded that the role of amino acids in Ni complexation and transport in S. tryonii was unclear and required further investigation using synchrotron based techniques.
1.4.2.2 Localisation of heavy metal(loid)s in plant tissues 1.4.2.3.1 Cellular localisation Over the last two decades, advances in electron microscopy, such as energy dispersive X-ray analysis (EDXA), micro-proton induced X-ray emission (µ–PIXE) and micro-X-ray fluorescence (µ-XRF) spectroscopies, have been utilised to examine elemental localisation within hyperaccumulating plant tissues (e.g. Vázquez et al. 1992; Mesjasz-Przybyłowicz et al. 1994; Lombi et al. 2002; Bhatia et al. 2003). The majority of these studies have employed EDXA (Table 1.3) which lacks sensitivity and provides a semi-quantitative estimation of elemental distribution at low spatial resolution (Stelzer and Lehmann 1993; Ma et al. 2005). These limitations can be overcome using µ-PIXE, as higher analytical sensitivities and signal- to-back ratios are obtainable (Mesjasz-Przybyłowicz and Przybyłowicz 2002).
32 Chapter 1: General introduction & background
Irrespective of the technique employed, studies to date suggest that heavy metal(loid)s taken up by plants tend to accumulate in epidermal and subepidermal tissues, including leaf trichomes (Table 1.3), however, results from these studies are not consistent, nor do they indicate a universal pattern of metal(loid)s localisation. For example, Robinson et al. (2003) suggested that adaxial epidermal cuticle cells were the principle site of Ni enrichment in leaves of Ni hyperaccumulating Berkheya coddii , whereas Budka et al . (2005) indicated that mesophyll cells were the principle site of Ni enrichment in the same species. Mesophyll cells have also been reported as the preferential site of As enrichment in Pteris vittata (Chen et al . 2005) and Pteris cretica var. nervosa (Chen et al . 2003), and Zn enrichment in Arabidopsis halleri (Zhao et al. 2000) and T. caerulescens (Frey et al. 2000). Although many studies associate Ni enrichment with sites of maximum accumulation, the majority do not consider the total amount of Ni storage per unit biomass (e.g. Bhatia et al. 2005). Ma et al. (2005) studied dual hyperaccumulating T. caerulescens and reported 2-folds higher concentrations of Cd and Zn in epidermal tisues than mesophyll tissues. However, they reported that 65–70% of the total leaf Cd and Zn were located in the mesophyll tissues and considering the larger biomass of mesophyll tissues, it was concluded that mesophyll was the major storage site of Cd and Zn. Overall, quantitative localisation studies of metal(loid)s in hyperaccumulating plants are scarce, and the results from various studies suggest a species specific localisation pattern of the accumulated heavy metal(loid)s. Furthermore, there are few studies that address localisation of the accumulated heavy metal(loid) in tissues other than leaves.
33 A. G. Kachenko
Table 1.3 Review of heavy metal(loid)s localisation studies in hyperaccumulating species.
Plant part Analytical Plant species Element Preferential localisation Reference examined technique a Alyssum bertolonii leaves, Ni EDXA epidermis Küpper et al . (2001) stems
Alyssum bertolonii leaves, Ni EDXA stem – epidermis; Marmiroli et al. (2004) stems leaf – epidermis surface; trichome base
Alyssum bracteatum leaves Ni EDXA epidermal cell Asemanch et al. (2006) lumen/walls
Alyssum euboeum leaves Ni EDXA epidermis Psaras et al . (2000)
Alyssum heldreichii leaves Ni EDXA epidermis Psaras et al. (2000)
Alyssum lesbiacum leaves Ni nano-SIMS Peripheral region of leaf Smart et al. (2007) trichomes and epidermis
Alyssum lesbiacum leaves, Ni EDXA epidermis Küpper et al. (2001) stems
Alyssum lesbiacum leaves Ni EDXA epidermis Psaras et al. (2000)
Alyssum lesbiacum leaves Ni EDXA trichomes Krämer et al. (1997)
Alyssum murale leaves Ni EDXA trichome pedicles, Broadhurst et al. (2004) epidermis
Alyssum murale leaves Ni EDXA epidermal cell Asemaneh et al . (2006) lumen/walls
Alyssum murale leaves Ni and µ-XRF and leaves - trichome base McNear Jr et al. (2005) CMT stems Co stem/roots – vascular
roots tissue
Alyssum murale leaves Ni and µ-XRF and Ni – epidermis Tappero et al. (2007) Co CMT Co – ground tissue/apoplasm
Arabidopsis halleri leaves Cd and EDXA base of trichomes Küpper et al. (2000) Zn
34 Chapter 1: General introduction & background
Table 1.3 continued... Plant species Plant part Analytical Element Preferential localisation Reference examined technique a Arabidopsis halleri leaves Zn EDXA base of trichomes Zhao, et al. (2000)
Arabidopsis thaliana leaves Cd µ-PIXE trichomes Ager et al (2002)
Berkheya coddii leaves Ni µ-PIXE leaf mesophyll Mesjasz-Przyby łowicz et al. (1998)
Berkheya coddii leaves Ni µ-PIXE leaf mesophyll Mesjasz-Przyby łowicz et al. (2001)
Berkheya coddii leaves Ni µ-PIXE leaf mesophyll Budka et al . (2005)
Berkheya coddii leaves Ni EDXA adaxial epidermal cuticle Robinson et al. (2003)
Berkheya zeyheri leaves Ni µ-PIXE epidermis and spongy Mesjasz-Przybyłowicz subsp. rehmannii var. parenchyma et al. (1996) rogersiana
Gossia bidwillii leaves Mn µ-PIXE and palisade mesophyll Fernando et al. (2006a) EDXA Gossia bidwillii leaves Mn EDXA palisade mesophyll Fernando et al. (2006b)
Hybanthus floribundus leaves Ni STEM epidermal vacuole Bidwell et al. (2004) subsp . curvifolius
Leptoplax emarginata leaves Ni EDXA epidermis Psaras et al. (2000)
Pteris cretica var. pinnae As µ-XRF mesophyll Chen et al. (2003) nervosa
Pteris vittata pinnae As EDXA epidermis Lombi et al. (2002)
Pteris vittata pinnae As µ-XRF mesophyll Chen et al. (2005)
Sebertia acuminata stems Ni EDXA laticifers of phloem Sagner et al. (1998)
Senecio coronatus leaves Ni µ-PIXE leaf epidermis, Mesjasz-Przybyłowicz stems stem epidermis and cortex et al. (1997)
Senecio coronatus leaves Ni µ-PIXE epidermis Mesjasz-Przybyłowicz et al. (1994)
35 A. G. Kachenko
Table 1.3 continued... Plant part Analytical Plant species Element Preferential localisation Reference examined technique a Senecio coronatus leaves Ni µ-PIXE epidermis Przybyłowicz et al. (1995) Senecio coronatus leaves Ni µ-PIXE epidermis Mesjasz-Przybyłowicz et al. (1996)
Senecio coronatus leaves Ni µ-PIXE epidermis Mesjasz-Przybyłowicz et al. (1997)
Stackhousia tryonii leaves, Ni µ-PIXE epidermis Bhatia et al. (2004) Bailey stems
Thlaspi caerulescens leaves Cd and Physical epidermis Ma et al. (2005) Zn separation
Thlaspi caerulescens leaves Zn EDXA epidermal vacuoles Vázquez et al. (1992)
Thlaspi caerulescens leaves Zn EDXA epidermal vacuoles Küpper et al (1999)(1999)
Thlaspi caerulescens leaves Zn EDXA epidermal vacuoles Frey et al. (2000)
Thlaspi caerulescens leaves Cd/Zn EDXA Cd – apoplast Vázquez et al. (1994) Zn – epidermal and sub epidermal vacuoles
Thlaspi goesingense leaves, Ni EDXA epidermal vacuoles, Küpper et al. (2001) stems mesophyll and epidermal cell walls
Thlaspi montanum leaves Ni EDXA epidermal subsidiary cells Heath et al. (1997) siskiyouense
Thlaspi pindicum leaves Ni EDXA epidermis Psaras et al. (2000)
Virotia neurophylla leaves Mn µ-PIXE and palisade mesophyll Fernando et al. (2006a) EDXA a Where EDXA, energy dispersive X-ray analysis; nano-SIMS, nano-secondary ion mass spectrometry; µ-XRF, micro-X- ray fluorescence; CMT, computed-microtomography; µ-PIXE, micro-proton induced X-ray emission; STEM, scanning transmission electron microscopy.
36 Chapter 1: General introduction & background
1.4.2.3.2 Subcellular localisation Subcellular enrichment of heavy metal(loid) ions occurs presumably in the vacuoles, the main storage organ of toxic compounds in plant cells (Ernst et al. 1992). The size of a cell may dictate the extent of vacuolation and hence accumulation. Küpper et al . (1999) observed a strong correlation between the length of leaf epidermal cells and relative Zn concentrations from leaves excised from Thalspi caerulescens . The importance of vacuoles in sequestration of Ni was reported by Krämer et al. (2000) (2000) who showed that the hyperaccumulator Thalaspi goesingense accumulated approximately 2-folds more Ni in the vacuole than the non accumulator Thalaspi avense . The study also indicated that a small but significant fraction of the total Ni present in leaf tissue was bound to histidine, and suggested that free histidine or some other bound histidine-like ligands may be involved in shuttling Ni across the cytoplasm for loading into the vacuole. Similarly, in As hyperaccumulating Pteris vittata , Lombi et al. (2002) using EDXA suggested accumulated As was localised in vacuoles. Recently, apoplastic compartmentation was observed in Ni hyperaccumulators Hybanthus floribundus subsp. curvifolius (Bidwell et al. 2004) and Thalaspi goesingense (Frey et al. 2000) and may contribute to possible Ni detoxification in these species.
1.4.2.3.3 Localisation within seeds There are limited studies of elemental localisation within seeds (fruits) of tolerant and hyperaccumulating species (Bhatia et al. 2003; Vogel-Mikus et al. 2007). Bhatia et al . (2003) employed µ-PIXE analysis on seeds of Ni hyperaccumulator S. tryonii and reported the highest Ni in the seed wall (pericarp; 4,433 mg kg 1 DW) and least in endospermic and cotyledonary tissues with 309 and 182 mg kg 1 DW, respectively. The authors indicated that movement of Ni within the seed was via symplastic pathways and suggested that localisation in the pericarp might deter herbivores and/or provide protection against pathogen attack. Conversely, Sagner et al . (1998) reported the highest Ni concentration in the endosperm and pulp tissues (mosocarp) of Ni hyperaccumulator Sebertis acuminate , with 14,000 and 8,000 mg kg –1 DW, respectively. Psaras and Manetas (2001) studied Ni localisation in Thlaspi pindicum seed using EDXA and reported preferential accumulation in the micropylar area opposite the radicle and in the epidermis of cotyledons, relative to the mesophyll. A similar pattern of localisation was observed in seeds of Cd hyperaccumulator Thlaspi praecox using µ -PIXE spectroscopy (Vogel-Mikus et al. 2007). It was observed that Cd was preferentially localised along the embryonic axis, with Cd concentrations in the seed tissue following the
37 A. G. Kachenko order epidermis (1,797 mg kg 1) > mesophyll (784 mg kg 1) > aleurone (444 mg kg 1) > testa (47.8 mg kg 1). Preferential localisation within inner tissues indicates that the accumulated metal was able to move apoplastically within the seed. In the Ni hyperaccumulator, Sececio coronatus , Przybyłowicz et al. (1995) found higher concentrations of Ni in the upper part of the seed, particularly in the region covering the radicle and the radicle itself with less Ni observed in the receptacle. Irrespective of the localisation pattern of the accumulated metal, these studies suggest that hyperaccumulators in the early stages of development will already contain some accumulated metal, however, the majority of metal uptake will occur post-emergence.
Similar to hyperaccumulating species, inconsistencies appear with regard to metal distribution in seeds of tolerant species. For example, in seeds of Zn tolerant Silene vulgaris , µ-PIXE analysis revealed a relatively homogeneous Zn distribution with limited accumulation around the testa, hilum and in the endosperm adjacent to the embryo, suggesting Zn was excluded from the embryo (Mesjasz-Przybyłowicz et al. 1999). Conversely, in Zn tolerant Biscutella laevigata L., higher Zn concentrations were observed within endospermic tissues (414 mg kg 1) as compared to the hilum area, pericarp, hypocotyls, cotyledons and testa, with concentrations of 134, 130, 87, 61 and 44 mg kg 1 Zn, respectively (Mesjasz-Przybyłowicz et al. 2001).
Overall, the localisation of elements in hyperaccumulating species is pertinent to understand the physiology of heavy metal(loid) hyperaccumulation. To date, there is no consensus as to the exact physiological explanation of heavy metal(loid) localisation in plant tissues. The disparity between species suggests that cellular and subcellular localisation is genotype specific.
1.4.3 Heavy metal(loid) uptake and translocation in hyperaccumulators As discussed earlier in Section 1.2.3, only a small fraction of heavy metal(loid)s is available for plant acquisition. The remaining fraction is strongly bound to soil particles and/or precipitation render a significant fraction unavailable for plant uptake. As a consequence, hyperaccumulators possess a number of strategies to enhance the uptake of metal(loid) ions. Whiting et al. (2000) outlined three possible mechanisms involved in uptake by hyperaccumulators:
38 Chapter 1: General introduction & background
(i) Enhanced absorption into roots coupled with high rates of translocation from roots to shoots; (ii) Rhizosphere effects such as the excretion of protons, organic acids and/or chelators; (iii) Foraging for metal(loid)s by the roots.
Lasat et al . (1996) using radiotracer techniques studied 65 Zn influx into the root symplasm and translocation to the shoot in Zn hyperaccumulating T. caerule scens and non hyperaccumulating T. arvense . They noted a 4.5-folds higher influx of Zn in T. caerulescens root cells compared to T. arvense , suggesting enhanced absorption into the root may be involved in Zn hyperaccumulation. They further observed a 10-folds increase in Zn translocated to the shoot of T. caerulescens compared with T. arvense , and concluded that transport sites other than entry into the root symplasm are also stimulated in T. caerulescens . It has been suggested that higher Zn influx into the roots of T. caerulescens was due to an increased abundance of Zn transporters in T. caerulescens root cells (Lasat et al. 2000). In non hyperaccumulating T. arvense , it was found that Zn was sequestered in the root vacuoles, hence restricting Zn translocation to the aboveground tissues.
The suggestion that rhizosphere effects play a role in hyperaccumulation was addressed by McGrath et al. (1997). The authors studied Zn uptake in T. caerulescens plants and reported that this species was effective in mobilising non-mobile Zn fractions in the studied soils. However, the authors did not observe any relationship between rhizosphere acidification and Zn uptake and suggested that changes in rhizosphere pH were a result of excess uptake of cations over anions. A similar reasoning was provided by Bernal et al. (1994) to explain Ni uptake in hyperaccumulator Alyssum murale . The authors also noted that the release of root exudates did not enhance Ni uptake.
Schwartz et al. (1999) demonstrated preferential foraging of T. caerulesens roots for Zn using rhizoboxes filled with soils containing patches of Zn contaminated soils. They noted that roots exhibited a high affinity for the Zn contaminated patches within soils. Short roots in clusters were observed when Zn was present at high concentrations, and finer and longer roots were present in uncontaminated soil. This study suggested that the formation and organisation of T. caerulesens roots is a specific response to Zn concentration and localisation in the soil profile.
39 A. G. Kachenko
They suggested that in soils with a heterogeneous pattern of contamination, a characteristic of most industrial and urban polluted soils, the main part of the root system would colonize the areas of highest concentration (hot-spots) whilst the remainder would develop in lower contamination areas ensuring a higher rate of phytoextraction. Zincophilic root foraging were also observed by Whiting et al. (2000) and Haines (2002) in T. caerulesens , however, the exact mechanisms that enables this species to discriminate between patches of soil with different Zn concentrations remains unclear.
In hyperaccumulating species, it is possible that a physiological requirement for the hyperaccumulated element may account for elevated uptake and translocation. For example, a 19% increase in dry matter yield was observed in T. caerulesens plants when grown in a 10 –3 –3 mmol m Zn solution as compared to a 1 mmol m Zn solution (Shen et al. 1997) . It was suggested that this species expressed a constitutive sequestration mechanism that appeared to detoxify Zn concentrations in shoot tissues, hence decreasing its physiological availability in the cytosol. Hyperaccumulation of Zn and Cd in dual hyperaccumulators Arabidopsis halleri , and T. caerulescens is considered a constitutive trait, however, an evolutionary explanation has not been proposed (Baker and Whiting 2002).
It is apparent that our comprehension of the processes by which plants hyperaccumulate heavy metal(loid)s is limited. Moreover, several disparities exist between studies with several studies focussed on two genera of the Brassicaceae family: Thlaspi and Alyssum. Indeed, more attention is required to investigate other heavy metal(loid) hyperaccumulating species to gain a broader understanding of the hyperaccumulation phenomenon.
1.5 REMEDIATION OF POLLUTED SOILS
Since the onset of industrialisation, the health of the environment has suffered as a consequence. Heavy metal(loid) polluted soils are now a common occurrence impacting on the sustainability, productivity and health of the soil environment and rendering large expanses of land uninhabitable. As the global population continues to grow (UN 2006), there is likely to be an increased demand for uncontaminated land and therefore, it is crucial that efforts concentrate on economically and environmentally viable techniques to remediate contaminated land. Current physico-chemical methods employed to remediate heavy metal(loid)s contaminated soils are costly and are often restricted to small scale applications
40 Chapter 1: General introduction & background
(Glass 1999). During the past 30 years, there has been intense interest surrounding the growing of plants to remove heavy metal(loid)s from contaminated soils, and these novel approaches will be briefly discussed in the following section.
1.5.1 Phytoremediation ~ Phyto (L): plant Remediation: remedium (L): restoring balance; cure. ~
The notion of phytoremediation as an emerging remediation technology is promising, and although not new, is seen as a favourable approach in the remediation of toxic heavy metal(loid)s from soils. Phytoremediation can be defined as a process in which green plants remove, sequester or stabilise heavy metal(loid)s to render them harmless (Salt et al. 1998). The application of phytoremediation as a strategy for soil decontamination was proposed by Chaney (1983) and since then, the ongoing development of phytoremediation has been largely driven by the spiralling costs associated with conventional soil remediation techniques and the desire to use a ‘green’, sustainable process. Glass (1999) indicated that the costs associated with phytoremediation were in some cases 15-folds less expensive than conventional physico- chemical remediation strategies; and hence, suggested that phytoremediation was an economically viable technology. Furthermore, existing physico-chemical technologies are intended primarily for intensive in situ or ex situ treatment of highly contaminated sites, and thus are not appropriate for vast, diffusely polluted areas where contaminates occur at low concentrations (Mulligan et al. 2001). A summary of the advantages and possible disadvantages of phytoremediation is provided in Table 1.4. The table indicates phytoremediation is a clean technology powered by renewable solar energy, a method to generate recyclable heavy metal(loid)-rich plant residues and hence reduce landfill. Further, phytoremediation is a passive in situ technology applicable to a range of toxic heavy metal(loid)s with minimal environmental disturbance and high public acceptance (Flathman and Lanza 1998).
41 A. G. Kachenko
Table 1.4 Advantages and disadvantages of phytoremediation adapted from Glass (1999) and Ghosh and Singh (2005).
Advantages Disadvantages
Cost Time - Does not require expensive equipment or highly - May take up to several years to remediate a specialised personnel contaminated site - Metal(loid) recycling provides further economic - Many hyperaccumulators are slow growers gain
Performance Performance - In Situ applications decrease the amount of soil - Restricted to sites with shallow contamination within disturbance compared to conventional methods rooting zone of remediative plants - Amendable to a variety of organic and inorganic - Not capable of 100% reduction compounds - In Situ /Ex Situ application possible with effluent or - Restricted to sites with low contaminant soil concentrations - In Situ applications decrease spread of contaminant - Consumption/utilisation of contaminated plant via air and water biomass is a cause of concern - Reduces the amount of waste to be land-filled (up - Harvested plant biomass from phytoextraction may be to 95%), can be further utilized as bio-ore of classified as a hazardous waste hence disposal should heavy metal(loid)s be proper - Capable of remediating bioavailable fraction - Climatic conditions are a limiting factor
Other Other - Public acceptance; aesthetically pleasing - Lack of recognized economic performance data - Compatible with risk-based remediation, - Need to displace existing facilities (e.g. wastewater brownfields treatment) - Can be used during site investigation or after - Introduction of non-native species may affect closure biodiversity - In large scale applications the potential energy - Regulators may be unfamiliar with the technology stored can be utilised to generate thermal energy and its capabilities
1.5.2 Subgroups of phytoremediation Phytoextraction , phytomining, and phytostabilisation are three important subgroups of phytoremediation. Phytoextraction is an approach whereby metal(loid) (hyper)accumulating plants remove heavy metal(loid)s from soils and concentrate them in the aboveground harvestable biomass. Dried, ashed, or composted plant biomass highly enriched in heavy metal(loid)s may then be isolated as hazardous waste (Kumar et al. 1995). Ideal phytoextraction requires plants to translocate high amounts of one or more of the target heavy
42 Chapter 1: General introduction & background metal(loid)s from soil into the aboveground biomass. Several studies have investigated the role of chelating agents such as ethylenediaminetetraacetic acid (EDTA) and diethylenetriamine pentaacetate (DTPA) in enhancing metal(loid) uptake, and results have shown mixed success. Huang et al . (1996) indicated that chelates could enhance Pb desorption from soil to soil solution, facilitate Pb transport into the xylem, and increase Pb translocation from roots to shoots in Zea mays and Pisum sativum . Similarly, Blaylock et al. (1996) showed that EDTA application enhanced Cd, Cu, Pb, Ni and Zn uptake in Brassica juncea shoots. Recently, Hsiao et al. (2007) investigated the role of EDTA and DTPA for enhancing Cr and Ni phytoextraction by Brassica juncea on serpentine-mine tailings. They reported that EDTA and DTPA could efficiently mobilise Cr and Ni in the soil solution and subsequently increase Cr and Ni uptake in biomass, however, the two synthetic chelators reduced plant biomass and invariably increased the risk of Cr and Ni leaching into the environment. Similar observations were reported in a study by Kos and Leštan (2004) who used a soil column experiment to demonstrate that EDTA and DTPA enhanced Cu mobility in soil solution, however, reduced Cu uptake in Brassica rapa var. pekinensis .
Phytomining is similar to phytoextraction except that the heavy metal(loid)s extracted from the ashed biomass are recycled as bio-ores (Anderson et al. 1999). This technology is relatively new with the majority of studies centred around laboratory and glasshouse investigations (e.g.(Robinson et al. 1997; Robinson et al. 1999). Recently, a pioneering field study was established at a gold (Au) mine in Brazil to investigate the feasibility of Au phytomining using B. juncea and Zea mays (Anderson et al. 2005). It was reported that B. juncea was more efficient in accumulating Au, however, an average assay of only 39 mg Au kg –1 of biomass was achieved that was probably a result of low Au concentration in the ore. The term phytomining has recently been referred to as phyto -reclamation as this technology is not considered an alternative to conventional mining as it is more closely associated with ‘farming’ than ‘mining’ (Anderson et al. 2005).
Several plant species that exhibit tolerance can be used as biogeochemical indicators of hidden mineral ore deposits (Reeves and Baker 2000). Tolerant species such as Haumaniastrum spp ., also termed ‘copper flowers’ are reputed to be universal indicators and have been identified for potential application in the Shaban Copper Arc of Central Africa (Paton and Brooks 1996). The ‘Zambian copper flower’ Becium centraliafricanum ( B.
43 A. G. Kachenko
homblei ), has been extensively used in locating Cu in the Shaba Province of Zaire and the Zambian Cu belt (Brummer and Woodward 1999).
An alternative remediation technique that does not extract heavy metal(loid)s is referred to as phytostabilisation , also known as in place inactivation or phytorestoration . This strategy utilises tolerant plants to inhabit heavy metal(loid) contaminated environments where they prevent soil erosion, reduce heavy metal(loid) leaching and minimise uptake, hence effectively containing and minimising the spread of heavy metal(loid)s (Berti and Cunningham 2000). These tolerant plant species are often grown in contaminated soils where there is a lack of established vegetation due to toxic effects of the heavy metal(loid)s on endemic vegetation and/or recent physical disturbance. In doing so, these plants help maintain soil integrity and health by minimising erosion and leaching of heavy metal(loid)s (Salt et al. 1995). For example, in a recent survey of a relinquished silver mine at Yerranderie in New South Wales, Australia, Archer and Caldwell (2004) identified Cynodon dactylon (couch), Juncus usitatus (common rush) and Lomandra longifolia ( spiny-headed mat rush) as of potential use in phytostabilisation programs as they showed tolerance and accumulation of Pb and Cd. In a study of industrially-contaminated sites, Tam and Singh (2004) reported high accumulation of Cr, Pb and Zn in kangaroo grass ( Themeda australis ), kikuyu grass (Pennisetum clandestinum ) and ryegrass ( Lolium perenne ), respectively, and suggested that these species could be used for phytoremediation purposes. Such species could be suitable for the remediation of several Pb and Cd contaminated sites such as shooting ranges, smelting sites etc. where alternative remediation techniques may be impractical due to the size of the area or funds available for remediation. Furthermore, phytostabilisation has become an attractive technique at sites where an interim strategy is required to contain the spread of heavy metal(loid)s until final remediation of a site can be completed (Berti and Cunningham 2000).
1.6 HEAVY METAL(LOID)S IN FERNS 1.6.1 Heavy metal(loid)s tolerance in ferns Ferns are a ubiquitous group of plants, which belong to the Pteridophyta division in the plant Kingdom (Jones 1987). Ferns are the second largest group of vascular plants, with more than 10,000 living species (Jones and Clemesha 1976; Schneider et al. 2004), and first appeared on the fossil record around 380 million years ago in the early Carboniferous Period. Recent
44 Chapter 1: General introduction & background reassessment of the fossil record based on molecular data has shown that > 80% of living fern species diversified in the Cretaceous Period, around 140 million years ago (Schneider et al. 2004). Unique to ferns and other more primitive plants is their independent haploid gametophyte phase, which has been preserved during millions of years of evolution (Banks 1999). Although similar to angiosperms in that they possess true leaves and vascular tissue (Jones 1987), it is clear that their evolutionary makeup is quite distinct.
Until recently, studies have largely focused on hyperaccumulating angiosperms despite several studies identifying fern species associated with heavy metal(loid) rich substrates. For example, Cody and Britton (1989) identified Adiantum pedatum subsp. calderi on serpentines in eastern Canada and USA, and three rare ferns of serpentine habitats in British Columbia: Polystichum lemmonii, P. scopulinum and P. kruckebergii . Other serpentine endemic fern species include Adiantum viridimontanum (Paris 1991), Pellea calomelanos and Cheilanthes hirta , the later two also occurring on Cu rich soils (Wild 1968). Ferns from the genus Asplenium are the most typical on soils from serpentine and other ultramafic soils. Species include A. presolanense , A. Cuneifolium , A. adiantum-nigrum , A. viride and several Asplenium hybrids such as A × contrei ( A. septiontrionale × A. adiantum-nigrum ), A. × alternifolia (A . septiontrionale × A. tricchomanes ) and A. × murbeckii ( A. × septentrionale × A. ruta-muraria ) (Page 1988). Brooks and Malaisse (1985) studied pteridophytes from the Shaba Province, Zaire and reported a number of Co and Cu tolerant species attributing their survival to restricted metal uptake. It was noted that Pteris vittata , Dryopteris athamantica , Cheilanthes inaequalis, Mohria lepigera and Ophioglossum lancifolium could accumulate relatively high levels of Cu in comparison to other species surveyed.
In addition to those ferns reported on soils with inherently high heavy metal(loid)s concentrations, several studies have reported a number of ferns on heavy metal(loid) enriched contaminated soils. Nishizono et al. (1987) identified Athyrium yokoscense growing on Cu and Zn/Cd metalliferous habitats in Japan. The fern was shown to accumulate high levels of Cu (up to 5,989 mg kg –1 DW) and Zn (up to 6,384 mg kg –1 DW) while high concentrations of Cd were found in the leaves (mean of 165 mg kg –1 DW). Numerous species of the Asplenium family have also been reported on old mining sites in the United Kingdom and include, A. ruta-muraria , A. trichomanes , and A. viride in alkaline soils, and A. Adiantum-nigrum and A. septionale in more acidic soils (Page 1988). Nephrolepis cordifolia growing at a former
45 A. G. Kachenko steelworks site in Sydney, Australia, was identified as a possible Cu hyperaccumulator, with concentrations of up to 2,324 mg kg –1 DW in fronds (Tam and Singh 2004).
Aquatic ferns have also been investigated for their accumulating ability. The prolific water fern Azolla filiculoides Lamark was found to accumulate 9,612 mg Cd kg –1 DW, 1,448 mg Cr kg –1 DW, 6,088 mg Cu kg –1 DW and 7,142 mg Ni kg –1 DW in its shoots (Sela et al. 1989). Gupta and Devi (1995) studied Cd uptake in the aquatic ferns Salvinia moletsa, Azolla pinnata and Marsilea minuta , and concluded that Salvinia could be considered as an indicator of Cd, Azolla as an ideal bioassay for Cd and Marsilea as a Cd resistant plant. Two further aquatic ferns Salvinia molesta Mitchell and Azolla rubra R.Br have also been identified as possible Cr accumulators (Shiny et al. 2004). The transfer of tolerance between ‘parent’ and ‘daughter’ aquatic ferns was observed by Outridge and Hutchinson (1991). The authors investigated Cd tolerance of ‘daughter’ ramets of Salvinia minima Baker from prior acclimation of Cd by ‘parent’ plants. This study provided what was the first evidence of induced Cd tolerance in a vascular plant other than grasses, and increased Cd tolerance of daughter ramets due to acclimation of their parents.
1.6.2 Hyperaccumulating ferns The ladder brake fern ( Pteris vittata ) has long been associated with arsenical mine dumps (Wild 1974a) and copper/cobalt rich substrates (Malaisse and Grégoire 1978; Brooks and Malaisse 1985). However, it was not until 2001 that its ability to hyperaccumulate As was discovered when Ma et al. (2001) observed that it accumulates up to 22,630 mg As kg –1 DW in the frond. This fern was highly tolerant to As in soil containing up to 500 mg As kg –1 and soils spiked with 50 mg As kg –1 were best for fern growth resulting in biomass production of 3.9 g plant –1 (Tu and Ma 2002). Similarly, the highest bioconcentration (63) and translation factors (25) were observed in soils spiked with 50 mg As kg –1 (Tu and Ma 2002).
Since the discovery of P. vittata , several other fern species have been identified as potential As hyperaccumulating species (Table 1.5) such as the silver back fern, Pityrogramma calomelanos (Visoottiviseth et al. 2002). Zhao et al . (2002) assessed As accumulation in three different accessions of Pteris vittata , two cultivars of Pteris cretica and, Pteris longifolia and Pteris umbrosa . Arsenic concentrations among all accessions and species ranged from 6,200– 7,600 mg kg –1 DW and these authors indicated that As hyperaccumulation is a constitutive
46 Chapter 1: General introduction & background
property of the Pteris genus. It has been shown, however, that Pteris species such as P. straminea , P. tremula (Meharg 2003) and P. semipinnata (Wang et al. 2006) do not hyperaccumulate As.
Table 1.5 Review of confirmed As hyperaccumulating species.
Maximum concentration in Species Family Reference frond (mg As kg –1)
Pteris vittata Pteridaceae 22,630 Ma et al . (2001)
Pityrogramma calomelanos Pteridaceae 8,350 Francesconi et al . (2002)
Pteris cretica var. albo-lineata Pteridaceae 7,600 Zhao et al . (2002)
Pteris cretica var. alexandrae Pteridaceae 7,600 Zhao et al . (2002)
Pteris longifolia Pteridaceae 7,600 Zhao et al . (2002)
Pteris umbrosa Pteridaceae 7,600 Zhao et al . (2002)
Pteris cretica var. nervosa Pteridaceae 2,594 Chen et al . (2003)
Pteris cretica var. chilsii Pteridaceae 1,358 Meharg (2003)
Pteris cretica var. crista Pteridaceae 1,506 Meharg (2003)
Pteris cretica var. mayii Pteridaceae 1,239 Meharg (2003)
Pteris cretica var. parkerii Pteridaceae 2,493 Meharg (2003)
Pteris cretica var. rowerii Pteridaceae 1,425 Meharg (2003)
Pityrogramma calomelanos var. Pteridaceae 3,330 Ashley et al. (2003) calomelanos
Asplenium australasicum a Aspleniaceae 1,240 Sridochan et al. (2005)
Asplenium bulbiferum a Aspleniaceae 2,630 Sridochan et al. (2005)
Pteris multifida Poir. Pteridaceae 1,145 Du et al. (2005)
Pteris oshimensis Pteridaceae 2,142 Wang et al. (2006)
Pteris biaurita L. Pteridaceae 3,650 Srivastava et al. (2006)
Pteris quadriaurita Retz Pteridaceae 3,650 Srivastava et al . (2006)
Pteris ryuensis Pteridaceae 3,650 Srivastava et al . (2006)
Pteris faurier Pteridaceae 1,362 Wang et al. (2007)
Pteris aspericaulis Pteridaceae 2,410 Wang et al . (2007)
a –1 Showed As toxicity symptoms when exposed to concentrations > 50 mg L .
47 A. G. Kachenko
Meharg (2003) noted that ferns exhibiting As hyperaccumulation arrived relatively late in terms of fern evolution and indicated a genetic perspective is needed to understand As tolerance and accumulation of Pteris species. It has been reported that hyperaccumulated As may protect ferns from damage by insect pests, diseases or herbivores (Rathinasabapathi et al. 2006), however, comprehensive studies supporting these hypotheses are lacking.
Information currently available depicting an association between ferns and heavy metal(loid)s highlights their potential use in phytoremediation of contaminated environments. However, there is tremendous scope for subsequent investigations to identify new hyperaccumulating species and the role, if any, of ferns in phytoremediation.
1.6.2.1 Background on Pityrogramma calomelanos var. austroamericana Pityrogramma calomelanos (L.) Link var. austroamericana (Domin) Farw. (Pteridaceae), or gold dust fern, is native to Southern America and is naturalized across the paleotropics including Australia (Chaffey 2002). This species is a terrestrial, rhizomatous fern and is characterized by a yellow waxy indumentum on the abaxial frond surface (Figure 1.8d). It is largely confined to the coastal regions of south eastern Queensland and north eastern New South Wales in Australia. The fern is often found thriving in disturbed areas such as road cuttings, mine overburden and tailings and has also been reported as a weed in banana and pineapple plantations (McCarthy 1998; Ashley et al. 2003). In a survey of Mt Perry Cu mine, Australia, it was reported as a possible As hyperaccumulator with concentrations in fronds ranging from 249–3,330 mg As kg –1 DW (Ashley et al. 2003).
48 Chapter 1: General introduction & background
(a) (((bbb)))
(((ccc))) (((ddd)))
Figure 1.8 Naturally occurring populations of Pityrogramma calomelanos var. austroamericana growing along the Dee River, Mount Morgan, Queensland (a-d).
(a) Populations are indicated by red circles. (b) Note the formation of white sulfate efflorescences (possibly pickeringite [MgAl 2(SO 4)4·22(H 2O)] and hexahydrite [Mg(SO 4)·6(H 2O)]; Edraki et al. (2005)) along the river bed as indicated by red arrows. (c) Illustrates a population growing naturally in moist crevices in rubble adjacent to the Dee River. (d) Illustrates the golden indumentums on the underside of fronds (April 2005).
49 A. G. Kachenko
1.7 AIMS AND OBJECTIVES 1.7.1 Rationale Contaminated environments, in particularly those resulting from human intervention, have propelled researchers into developing strategies to rectify this significant global problem. The discovery of plants growing in heavy metal(loid)s enriched substrates has provided a possible green solution to remediate heavy metal(loid)s contaminated sites. This unique group of plants has the innate ability to tolerate and in some cases hyperaccumulate heavy metal(loid)s in environments which would otherwise present hostile growing conditions. Hyperaccumulating species in particular, may be utilised in the extraction ( phytoextraction ) of heavy metal(loid)s and can be further exploited for economic extraction of certain elements through phytomining .
It has only been seven years since the first hyperaccumulating Pteridophyte ( Pteris vittata ) was reported and since then, the majority of studies have addressed the physiological aspects of As hyperaccumulation in this species. To date, there have been few studies of lesser known As hyperaccumulating Pteridophytes and further, little emphasis has been given to study the influence of heavy metal, other than As, in Pteridophyte species. Clearly, the role of Pteridophytes in phytoremediation is in its infancy with tremendous scope to advance our understanding of the mechanisms by which these species translocate and detoxify accumulated As. Since the discovery of the As hyperaccumulation trait in Pityrogramma calomelanos var. austroamericana , there have been no further studies of this species. Therefore, this lesser-known As hyperaccumulator species is an ideal candidate to explore further to better understand physiological mechanisms leading to its hyperaccumulation.
In addition to Pteridophytes, research on physiological aspects of other hyperaccumulation species particularly, translocation, tolerance and detoxification mechanisms has only just begun. For example, the localisation of elements within plant parts have been investigated in only a few taxa, ca. 6% of the confirmed hyperaccumulators, the majority of which are Ni hyperaccumulating species from ultramafic outcrops. Several of these studies suggest more than one site of preferential localisation within a particular species and indicate a degree of physiological variation among different plant species in response to different heavy metal(loid)s. Further, there have been few studies that address the ecological significance of the hyperaccumulating trait.
50 Chapter 1: General introduction & background
Hybanthus floribundus subsp. floribundus is one of three H. floribundus subspecies and the only subspecies native to serpentine and non-serpentine environments. To date, the majority of studies have investigated serpentine endemic Hybanthus populations with few studies that have examined ecophysiological aspects in this subspecies. Together with P. calomelanos var. austroamericana , this species was chosen as a suitable candidate for further detailed ecophysiological investigation.
1.7.2 Aims The work presented in this thesis is based on three principal aims: i. To investigate heavy metals tolerance in common Australian ferns with the aim of identifying new hyperaccumulating species ii. To characterise As hyperaccumulation and physiological detoxification mechanisms in P. calomelanos var. austroamericana iii. To examine the ecophysiology of Ni hyperaccumulation and detoxification in H. floribundus subsp. floribundus
1.7.3 Objectives The specific objectives of this thesis were to: 1. Determine the degree of heavy metal accumulation and the biomass produced when nine of the most common Australian fern species in addition to P. vittata were exposed to elevated levels of heavy metals (Chapter 2); 2. Compare sample preparation techniques for quantitative µ-PIXE localisation of elements in hyperaccumulating P. calomelanos var. austroamericana and H. floribundus subsp. floribundus tissues (Chapter 3); 3. Compare the level of As hyperaccumulation in P. calomelanos var. austroamericana with that of well known As hyperaccumulating fern P. vittata (Chapter 4); 4. Determine the cellular localisation of As in frond and stipe tissues of P. calomelanos var. austroamericana after exposure to As using µ-PIXE spectroscopy (Chapter 4); 5. Determine the cellular localisation of Ni in leaf, stem and seed tissues of H. floribundus subsp. floribundus after exposure to Ni using µ-PIXE spectroscopy (Chapter 5);
51 A. G. Kachenko
6. Examine localisation and speciation of As in P. calomelanos var. austroamericana using micro-focused X-ray absorption (µ-XAS) and X-ray fluorescence (µ-XRF) spectroscopies (Chapter 6); 7. Identify and quantify major low molecular weight plant ligands (amino and organic acids) involved in possible detoxification of Ni in shoot tissues of H. floribundus subsp. floribundus (Chapter 7); 8. Examine the hypothesis that Ni hyperaccumulation confers drought tolerance in H. floribundus subsp. floribundus (Chapter 8).
This thesis concludes with Chapter 9, which provides a summary of conclusions arising from Chapters 2-8 and directions for future research.
52