Keywords

1 Introduction

Soil is a key resource for the life of organisms in terrestrial environments; it is the basis of the ecosystem and of our farming system of food production (Chesworth 2008). It is imperative to acknowledge that current and future exploitation of mineral resources will produce even more wastes and impose more threats on Earth. Pollution of soil with Potentially Harmful Elements (PHEs) is a major concern worldwide since it can cause the transfer of contaminants to other environmental compartments, such as surface water, groundwater and biota, with a consequent spread of pollution over a wider area than the initial one (Lal 2006). PHEs in the soil can be partly absorbed by the plants that derive nutrients from that soil, with the risk of accumulation and toxic effects along the food chain and also to humans (Markert et al. 2003).

Mining is only one of the pathways by which metals enter the environment. Mining itself affects relatively small areas, and this could not pose severe environmental problems. The environmental impact arises when ores are mined, milled and smelted, and a certain amount of metals is released in the surrounding areas and to waterways. Depending on the nature of the waste rock and tailings deposits, a wide dispersion of the metals both in solution and in particulate form is possible. A consequence of the high content of toxic heavy metals, in combination with reduced soil thickness, leads to discontinuous vegetation coverage that is composed mainly of crust lichens, mosses, fescue and wild vegetation genetically tolerant to high metal concentrations (Wahsha et al. 2012b).

Plants may be classified into three groups on the basis of their ability to accumulate metals in their aerial parts (Baker and Brooks 1989) Excluders are those plants whose metal concentrations remain unaffected by metal contents in soils up to a critical level, when toxic symptoms appear; Indicator plants are those whose metal concentrations reflect those the related soil; Accumulator plants have the ability to take up and concentrate metals from soils, independently on levels of metal concentrations. Among the species that may tolerate high metal concentrations in their tissues, plants presenting exceptional accumulating ability are referred to as hyperaccumulators.

2 PHEs Contamination in Soil

Soils are presently contaminated with PHEs, which are originated from different sources in particular by heavy metals, which can negatively affect soil’s chemical, physical and biological properties worldwide. In this context, heavy metals can be added to the soil through diverse natural or/and anthropogenic sources. The major natural sources of heavy metals in soils are: weathering (such as soil erosion and deposition of windblown dust and running water); volcanic eruptions and bushfire.

The major anthropogenic input of heavy metals to soils mainly occurs by: Atmospheric deposition, result from energy production, power generation emissions, metal mining, smelting of metalliferous ores and manufacturing, waste and wood burning; sewage sludge, municipal and organic wastes and co-products from agriculture and food industries; Land disposal of industrial waste; Fertilizer and pesticide (insecticides, herbicides, fungicides) and soil heavy metal pollution through chemical and biological warfare (Wahsha et al. 2012d; Alloway 1995).

The environment is rather strongly influenced by manmade activities through heavy mining, especially in areas where low-grade metal ores are being extracted, which produces higher quantities of soil wastes (e.g. Heavy metals).

Heavy metals found in soils are originated from different sources. In some places we can find natural forms of contamination, due to the presence of particular rock types, such as serpentinites (Ghaderian et al. 2007). In other places, anthropogenic metal accumulation is recorded (Zhang et al. 2009). Indeed, human activity is considered to be the major source of pollution by heavy metals in soils. Among the various causes of this type of contamination we can mention mining and processing of metals, which are linked to generate several environmental problems (e.g. acid mine drainage, impacts on the landscape, pollution of water, soil removal and pollution). The ever-increasing demand for metals lead to more mining and, therefore, to larger amounts of waste production. Accessing mined produces is always accompanied by excavation of range volumes of waste rocks, whose disposal is another important problem of mining. However, little attention has been paid to processing and disposal of mining waste, which have been generally dumped on the ground, to become a potential source of pollution (Lottermoser 2010; Lal 2006; Selim and Sparks 2001).

Dispersal pattern of mine spoils can be recognised by mining activity. The effect of these irregular anthropogenic deposits is the development of a new soil material that covers the original one. Immature Entisols (Lithic Spolic Xerorthents and Spolic Xerorthents,) are formed on mine spoils <100 year old. These soils are characterized by a thin solum (<30 cm), little organic matter accumulation (mean 14 g/kg organic carbon, range 1–33 g/kg), dark brown (10YR3/3) to reddish (5YR4/6) colour, coarse texture (sandy loam to loamy), and subalkaline pH (mean 7.4, range 6.9–7.8). Detailed descriptions of soil properties of a selected profile on mine waste are given in Table 5.1.

Table 5.1 Selected properties of the Spolic Technosols
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Soils developed on old mine dumps or in the proximal parts (<0.5 km) of the dumps, have a >50 cm thick solum, sandy loam to loam texture, blocky structure, slightly acidic pH (mean value 6.3, range 4.9–7.7), humus accumulation (up to 14 % organic matter in the A horizon), moderate to low cation exchange capacity (CEC) (mean 20 cmolc kg−1), with significant desaturation (base saturation <60 %). As a rule, they have distinct A-B-C horizonation and a well formed cambic horizon. Thus, they are Inceptisols Spolic Haploxerepts or Spolic Dystroxerepts, (Table 5.2). Frequently, a discontinuity occurs between the upper and the lower portions of the profile, which developed from the underlying bedrock. Data (not reported) indicate relevant differences and a remarkable polycyclic evolution, due to the superposition of mine spoil over the normal soil. Colour, texture, reaction, and CEC are the most prominent features that present important differences in soils. Soil horizons show dark brown (7.5YR3/2) to dark reddish brown (5YR 3/3) colour, well individualized structure, from crumby to fine blocky beds. Texture is coarse (sandy loam to loam) in surface horizons of mine spoils and loamy to clayey underneath. Values of pH around 6.3 at the surface indicate a soil reaction that is slightly acidic; this is slightly subalkaline (pH 7.4) and base-saturated at the bottom. Cation exchange capacity increases with depth, from 15 to 25 cmol (+)/kg (Bini 2010).

Table 5.2 Selected properties of the Spolic Dystroxerept
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Shrubby vegetation with shallow trees (i.e. holm-oak, strawberry tree, heath, etc.) is the typical flora of these sites, where rock-rose is the dominant plant.

Far from the mine dumps (>0.5 km) the soil samples show little evidence of mine spoil in the profile. Sulphide minerals are found especially at the surface, as revealed by mineralogical and chemical investigation (Bini 2012). An abrupt textural change (Table 5.3) indicates a marked discontinuity between the upper and lower part of the soil profile. The upper part (A and E horizons) has dark brown (10YR2/2) to yellowish brown (7.5YR3/4) colours, loam to sandy loam texture, crumby structure, high organic carbon content (mean 21 g/kg), and subalkaline pH. The lower part (Bt horizon) presents reddish colours (5YR3/4–2.5YR3/4), a strong clay content increase (clay loam to clayey texture), organic carbon decreases; pH is subalkaline with traces of carbonate. Therefore, they are classified as Alfisols. Since there is evidence for mine waste in the profile, these soils should be classified as Spolic Rhodoxeralfs or Spolic Haploxeralfs. However, considering the net discontinuity already mentioned, these soils could be classified as Spolic Xerorthent over Typic Rhodoxeralf (or Haploxeralf) (Bini 2012).

Table 5.3 Selected properties of the Spolic Haploxeralf
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3 Mine Soils Pollution

The availability of resources from mining sites is not an easy task, where some products are existing in almost in 100 % of the excavated sites, such as the production of sand, clay and gravel, while other products such as gold are barely reaching few parts per million of the raw material, therefore generating large volumes of wastes. Copper, zinc, and gold mining goes through several processes; such as grinding, washing, and sizes of the raw material of rocks, which generates some byproducts such as dust and other fine-grained material, the later could be used as natural fertilizers. In general, forma mining is always producing volumes of waste material known as “gangue”. Of the processes with such large volumes of mining wastes, ranked from high waste to low are coal, non-ferrous ores, ferrous ones, and industrial minerals. The type and magnitude of the released contaminants from mining sites is controlled by many factors, such as the geology, topology and climate of the mined sites, and the techniques applied for mining and processing of the mined resources. Understanding all these factors should be deemed critically important for better control of waste discharge from mining areas. Improper operations of mining activities have been facing increasingly strong criticism by environmentalists (Lottermoser 2010).

Both physical and chemical processes work during rock alteration. Physical processes appear to be a leading role during the initial stages. Steep morphology areas, where most mining districts are located, are the most affected by these processes. Loose and coarse-grained material forms as a consequence of rock fragmentation. Rock fragments migration on instable slopes, erosion of fine particles by wind plus runoff, all these processes contribute to land modelling.

The rock transformations also occur by chemical processes such as oxidation potential (Eh >250 mV), acidification (pH <7), hydrolysis, metal leaching, precipitation of oxy-hydroxides and sulphates, argillogenesis. Wet and high temperature climatic conditions increase the role of chemical processes and stimulate mine waste reactivity. Thus, the physical and chemical processes of soils developed from mine soils lead to new soils with a different set of physical and chemical features.

The formation of a biologically active substrate may occur as the parent material is finely subdivided and weathered, and this permits the development of a pioneer vegetation (lichens, mosses). Litter accumulation (OL horizon) is the process that characterizes the early stage of soil formation. Subsequently, organic matter decomposition (OF horizon), humus formation (OH horizon), mineralisation (A horizon) constitute a first pedogenetic phase (Bini 2012, 2010). According to Jabiol et al. (2007), this phase may bring to the differentiation of several types of humus as a function of litter composition, microflora and microfauna activity, pH and climate conditions.

A second pedogenetic phase is determined by in situ mineral transformations (e.g. acid hydrolysis), oxyhydroxides and clay formation (stage of cambic horizon formation). In this phase, colour varies from very dark brown (10YR 3/3) to dark brown (7,5YR 3/3), reddish brown (2,5YR 3/2), dark yellowish brown (10YR 4/5), or blackish (5YR 2, 5/5), in relation to the nature of the bedrock, and/or to the amount of mine waste.

A third pedogenetic phase is consistent with solute leaching and particles migration towards bottom (stage of argillic horizon formation); precipitation of new minerals (e.g. carbonate, sulphate) is also likely to occur. However, this third phase is difficult to assess in mine waste materials. Since the time elapsed from mining operations in most cases is not sufficient for Bt formation, considering that the landscape morphology is generally undulated, with slopes ranging from 15 to 45 %, which makes erosion a dominant process. However, soils developed from waste dumps are generally shallow (20–100 cm), skeletal, coarse-textured (sandy loam to loamy sand), little developed, with limited horizonation (Bini 2007, 2012).

Mine waste production varies between countries, depending on intensity of mining activities in each country, and on methods of mining and processing (Table 5.4).

Table 5.4 Mining waste production compared to municipal waste in different countries of the world
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Earth’s crust is noticeably enriched with different metallic and mineral compounds accumulated over time due to various geological processes. Geological processes leads to the formation of different deposits according to the amount and type of mineral enrichment in each deposit. Mineral deposits formation and differentiation is an important aspect in understanding the environment. Deposits geology (included rock composition, geomorphology etc.) may influence the characteristics of soil chemistry, surface and ground water properties leading to the natural enrichment of soils, waters, and sediments with metals and metalloids (Lottermoser 2010).

Natural enrichment of metals and metalloids varies between sites, leading to the adaptation of the covering vegetation to specific soil contaminants. For example, serpentine flora is the type of vegetation highly adapted to lands with overlying soils highly enriched with iron, nickel, chromium and magnesium. Soils enriched with boron favor the growth of cereal and legume crops. Soils enriched in selenium have tolerant vegetation covering, although animals grazing on such plants will suffers acute toxicity.

Accordingly, elemental enrichment in ores and rocks could cause unfavorable effects on local and regional ecosystems due to the signature of such elemental enrichment to surrounding water bodies, soils and sediments. Such adverse effects could result from geological processes or artificially imposed due to the improper mining practices and mine wastes disposal and management (Lottermoser 2010).

4 Environmental Impacts of Abandoned Mines, Scientific Issues and Case Studies

The impact of PHEs from abandoned mine sites on the environment will be discussed through case studies from Europe (NE Italy). The environmental effects recognised for these specific sites could be valid other mining sites worldwide.

The area we are considering is an abandoned mine site in North-East Italy called “Imperina Valley”, at an altitude ranging between 543 and 990 m above sea level. The geological substrate consists of rocks of the metamorphic basement (Pre-Permian), in tectonic contact with dolomite rocks (Dolomia Principale, Upper Triassic).

The Imperina stream crosses the valley; even if no settlements can be found in this area, many buildings and tunnel outlets still witness the past mining activity. Part of the area (right side and a portion of the bottom) lies within the National Park of the Belluno Dolomites. The mined area, located along the tectonic contact between basement and dolomites, consists of a deposit of mixed sulfides, composed primarily of cupriferous pyrite, pyrite and chalcopyrite, with minor amounts of other metallic minerals (Frizzo and Ferrara 1994). Copper and sulphur were the main products extracted. Until the beginning of the twentieth century, copper was extracted and processed directly in situ through roasting, a method with a severe impact on the area due to acid rains formation and intensive wood cutting. Today, the vegetation cover is mainly constituted of mixed forests (Abies alba Mill., Picea abies (L.) H. Karst., 1881, Fagus sylvatica L. and Ostrya carpinifolia Scop.), with clearances where herbaceous and shrubby vegetation prevails over the arboreal one (Dissegna et al. 1997).

The first certain historical records indicate that mining in the Valle Imperina dates back to the first years of the fourteenth century, thanks to research funded by the Republic of Venice into the mining of copper. Very probably, the extraction of minerals had already commenced in pre-Roman times, considering the particular condition and position of the rocks and the proximity of the Agordo valley; this theory is also backed up by the large amount of items made from bronze and copper found in archaeological excavations in the whole of the Belluno valley.

A notable increase in the production of copper derives from the passage of the property rights of the mines from private hands to the Republic of Venice; in fact, initially the mine was subdivided in tunnels and “mints” owned by private individuals with obvious management problems and lack of efficiency of the mining. This passage started gradually at the end of 1,600, but it was only in 1,835 that these mines formally become public property, in the hands of the Austro-Hungarian Empire.

During the first years of the last century, the whole process was electrified, and in 1925 the new standard-gauge electric railway line allowed for the loading of the materials which had been extracted from the mine and taken to the situation by cableway.

After the First World War, production stabilised at around 50,000 tonnes a year up to the period 1940–1944. Following this period, after a first attempt to modernise the mine in the early 1950s, the condition of the mineral deposit and its exhaustion led initially to staff reductions and, finally, to the complete closure of the mine on 8 September 1962 (Wahsha et al. 2011).

4.1 Case Study 1

4.1.1 PHEs Contamination in Soils and Plants of an Abandoned Mine

Over the past decades, human activities such as metal mining and milling operations have been recognized as one of the most important sources of contamination in the environment, along with mine and mill waste water (Jung 2001; Navarro et al. 2008). Heavy metal contamination has been a serious problem in the vicinity of abandoned mine sites due to the discharge and dispersion of mine-waste materials into the ecosystem (Jung and Thornton 1996). These heavy metals have a potential to contaminate soil and water (Haque et al. 2008; Lim et al. 2008). Extraction of metals from sulphide minerals usually results in large amounts of waste materials that often contain elevated concentrations of potentially harmful metals such as Cu, Zn, Cd, and Pb (Jian-Min et al. 2007; Lee et al. 2001). The degree of heavy metal contamination around mines varies depending upon geochemical characteristics of elements and degree of mineralization of the tailing (Navarro et al. 2008). Yun-Guo et al. (2006) reported that abandoned mining sites represent significant sources of metal pollutants in water and soils and a threat to the ecosystem. These metals can be transported, dispersed and accumulated in plants and then passed through the food chain to human beings as the final consumer.

The restoration of metal-contaminated sites is one of the most important environmental issues. Soil pollution by chemicals poses serious hazards to surface and ground waters, plants and humans, and presents relevant social, sanitary and economic costs (only in the U.S. up to 250$ m−3 soil; Adriano et al. 1995; Bini 2010). Metal accumulation in soil diminishes soil fertility, microbial activity and plant growth (Lehoczky et al. 1996). Moreover, trace elements are very persistent, can interact with plant roots by adsorption or release from the soil particles, and therefore increase the risk of long-term soil pollution and of toxic effects on organisms (Rosselli et al. 2006).

The assessment of soil contamination by metals has been extensively carried out through plant analysis (Blaylock et al. 2003; Brooks 1998; Ernst 1996; Wenzel et al. 1993); both wild and cultivated plant species have been frequently used as (passive accumulative) bioindicators for large scale and local soil contamination (Baker 1981; Baker and Brooks 1989; Bargagli 1993; Zupan et al. 1995; Zupan et al. 2003).

In the last decades, attention has been deserved to plants as tools to clean up metal-contaminated soils by the low cost and environmental friendly technique of phytoremediation (Adriano et al. 1995; Baker et al. 2000). This technology is focused on the ability of plants to accumulate high heavy metal concentrations (up to 100 times the normal concentration) in their aerial parts (i.e. they are hyperaccumulator plants as defined by Baker 1981). The plant ability to uptake metals was firstly applied in phytomining projects (Brooks and Robinson 1998; Ernst 1993; Helios-Rybicka 1996; Mc Grath 1998; Vergnano Gambi 1992), and only successively, when environmental contamination became a global concern, it was recognized as an useful tool for remediation projects (Adriano et al. 1995; Bini 2010, 2005; Bini et al. 2000b; Mc Grath 1998; Salt et al. 1995). Indeed, tolerant or accumulator populations of higher plants may colonize naturally or even anthropogenic metal-enriched areas, accompanying the disappearance of sensitive plants. Therefore, they may be utilized in restoration of such areas. The choice of plants is a crucial aspect for the remediation techniques. Up to now, more than 400 plants that accumulate metals are reported, Brassicaceae being the family with the largest number of accumulator species (Bini 2010; Marchiol et al. 2004; Mc Grath 1998).

Heavy metal accumulation is known to produce significant physiological and biochemical responses in vascular plants (Mangabeira et al. 2001). As stated by Preeti and Tripathi (2011), there is a direct relationship between chemical characteristics of soil, heavy metals concentration and morphological and biochemical responses of plants. Yet, metabolic and physiological responses of plants to heavy metal concentration can be viewed as potentially adaptive changes of the plants during stress.

Plants growing on abandoned mine sites and naturally metal-enriched soils (e.g. serpentine soils) are of particular interest in this perspective, since they are genetically tolerant to high metal concentrations, as reported by several authors (Bini 2005; Giuliani et al. 2008; Maleci et al. 1999; Brooks 1998; Pandolfini et al. 1997; Vergnano Gambi 1992), who studied endemic serpentine flora (Alyssum bertoloni, A. murale, Silene paradoxa, Stachys serpentini, Thymus ophioliticus) at various sites in the world. All these authors agree that morphological, physiological and phytochemical characters of serpentine plants are strongly dependent on the substrate composition (what Jenny, in 1989, called “the serpentine syndrome”), and that they are likely metal accumulator or tolerant ecotypes.

Understanding the mechanisms of metal bioaccumulation by plants species and of metal bioreduction by microorganisms is a clue to the efficiency of phytoremediation techniques. The localization and the chemical form of metals in cells are key information for this purpose (Kidd et al. 2009; Sarret et al. 2001). After their assimilation by plants, heavy metals could interfere with metabolic processes and are potentially toxic (Lopareva-Pohu et al. 2011); phytotoxicity results in chlorosis, weak plant growth, yield depression, and may be accompanied by disorders in plant metabolism such as reduction of the meristematic zone (Maleci et al. 2001), plasmolysis and reduced chlorophyll and carotenoids production (Corradi et al. 1993). Mangabeira et al. (2001) studied the ultrastructure of different organs of tomato plants (root, stem, leaf) which showed visible symptoms of Cr toxicity, and argued that CrVI induces changes in the ultrastructure of these organs. Similar findings were reported by Vasquez et al. (1991) for Cd in vacuoles and nuclei of bean roots. Since both these metals are known to be inessential to plant nutrition, it is suggested that they are likely confined in roots by a barrier-effect as defense strategy during stress. Conversely, essential metals such as Zn and Cu are easily conveyed to the aerial parts, as reported by Fontana et al. (2010).

Among wild plants, the common dandelion (Taraxacum officinale Web) has received attention (Bini et al. 2000a; Królak 2003; Zupan et al. 2003; Simon et al. 1996) as bioindicator plant, and has been also suggested in remediation projects (Turuga et al. 2008), given its ability to uptake and store heavy metals in the aerial tissues. T. officinale is a very common species, widely diffused in Central and Southern Europe, easy to identify and greatly adaptable to every substrate (Keane et al. 2001; Malawska and Wilkomirski 2001). Moreover, this species is commonly collected to be used in cooking as fresh salad or boiled vegetable, and is used also in ethnobotany and traditional pharmacopoeia (Rosselli et al. 2006). Therefore, when grown on heavily contaminated soils, it may be potentially harmful if introduced in dietary food, as it occurs in many countries.

Previous studies of our research group (Bini et al. 2000a; Fontana et al. 2010) investigated the heavy metal concentration of soils developed from mine waste material, and the wild plants (Plantago major, Silene dioica, Stachys alopecuros, Stellaria nemorum, Taraxacum officinale, Vaccinium myrtillus, Gymnocarpium dryopteris, Gymnocarpium robertianum, Salix caprea, Salix eleagnos, Salix purpurea) growing on those contaminated soils, in order to determine the extent of heavy metal dispersion, and the uptake by both known and unreported metal-tolerant plant species.

In the last few years many studies have focused on the potential use of trees as a suitable vegetation cover for phytoremediation (French et al. 2006; Jensen et al. 2009). A very suitable tree for use in phytoremediation is willow (Landberg and Greger 2002). Pulford and Watson (2003) detailed the phytoremediation potential of willow in heavy metal contaminated areas. Willows have not been included in the group of hyperaccumulators of heavy metals, but on the other hand they provide potential bioindicator of pollution (Mleczek et al. 2009). However, metal concentrations in willows depend on species, growth performance, root density, distribution within the soil profile and sampling period (Chehregani et al. 2009). Moreover, willow has been recently recognized as a good accumulator of heavy metals (Meers et al. 2007).

Our aim was to assess total concentration of six potentially toxic metals (Cd, Cr, Cu, Pb, Zn and Fe) in the soil and plant samples of T. officinale and three dominant willow species (Salix purpurea L., Salix caprea L. and Salix eleagnos Scop.) collected from abandoned mixed sulphide mine dumps of Imperina Valley, in order to propose these plants for phytoremediation plans (Table 5.5).

Table 5.5 Average concentration of PHEs in soils and plants of interest expressed as mg kg−1
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Comparing the values found with those of control levels of Angelone and Bini (1992), the total concentrations of most of the investigated metals (Cd, Cu, Pb, Zn and Fe) in the soil samples were significantly higher (p <0.05), and almost above the toxicity threshold according to the Italian legislation (D.L. 152/2006).

The area is almost not contaminated by Cr, whereas there are a contamination by Zn, Cu, Pb and Fe that show high concentrations, particularly at sites affected by mining activities and ore processing.

There is a linear positive correlation between Pb, Cu, Zn and Fe (Cu/Pb 0.867; Pb/Zn 0.616; Cu/Zn 0.688; Cu/Fe 0.933). This is consistent with their calcophile geochemical behaviour, since these metals tend to form compounds with sulfur, as chalcopyrite (CuFeS2), sphalerite (ZnS) and galena (PbS), commonly found in the Imperina Valley ore deposits (Frizzo and Ferrara 1994). Cr is negatively correlated with Cu (−0.847), Pb (−0.816), Zn (−0.604) and Fe (−.0754). Conversely, Fe indicates a significant positive correlation with Pb (Fe/Pb 0.734). Furthermore, Fe it is not significantly correlated with Cd. Therefore, it is likely that most of iron in soils of the study area derives from the alteration of pyrite and chalcopyrite mineralization. For all elements, it might be that the same type of elements combination occurring in the mineralization of Imperina Valley is found in soil. This means that no element of the mineralization has been removed in a special way, since the factors of pedogenesis have acted for a some decades in the areas affected by ore processing, and thus the soil chemical characteristics still seem to be close to those of the parent material, as it was found by Bini (2005) in mine soils of Tuscany.

It is noteworthy to point out, however, that willows ability to accumulate heavy metals in different parts is independent of the species; rather, it depends on local factors as soil and pedoclimate (particularly temperature, aeration and water content) and on plant physiology and aging (Wahsha et al. 2012b). Moreover, a counteracting behavior of essential and toxic heavy metals is likely to occur as a barrier effect of the roots (Fontana et al. 2011).

The calculation of translocation factors (TF) highlights that willows translocate and accumulate metals in the aerial parts, in particular Cu (S. purpurea TFCu = 4.72), Pb (S. purpurea TFPb = 3.42), Zn (S. caprea TFZn = 3.48), Fe (S. purpurea TFFe = 1.44), and Cr (S. purpurea TFCr = 1.24). Most of the plant species had BCF less than one and TF more than one, although the concentration of heavy metals remained below 1,000 mg kg−1. In general, metal concentrations in plants vary with plant species; plant uptake of heavy metals from soil may occur either passively with the mass flow of water into the roots, or through active transport from root cells (Kabata-Pendias 2004; Mun et al. 2008).

In this case study, the metal translocation ability, combined with rapid growth and a higher biomass than herbaceous plants, qualifies willows as good candidates for phytoremediation of polluted soils (Bini 2007). Since most of the studied willows were capable to uptake and translocate more than one metal from roots to shoots and, based on high TF values, they can be used for phytoextraction. On the other hand, Cd shows very low translocation factors in all investigated plants, and proved to be blocked in the roots, since it is known to be unessential to plants, thereby suggesting some exclusion strategy by plants (Vandecasteele et al. 2002).

Anthropic influence related to mining activity in soils of the studied area is evident. Soils in the mining site are highly contaminated by heavy metals, mainly Cu, Zn, Pb and Fe. The metal content in willows show relatively high concentrations of these elements. The results of this study indicated that there is an increasing need for further research on the mechanisms whereby such plants are able to survive in contaminated soils. Furthermore, studies should aim to determine the growth performance, biomass production and metal accumulation of these species in metal contaminated soils for their better management and conservation.

Concerning Taraxacum (plants), data show that this species is tolerant to high metal concentrations, which supports the use as a bioindicator plant. Metals accumulated preferentially in roots, but also leaves proved accumulator organs, being able to store up to 200 mg kg−1 Pb and 160 mg kg−1 Zn, with only little damages (e.g. reduced foliar surface, reduced plant development).

Soil analysis of the studied area (data not reported) showed low pH, low cation exchange capacity, percentage of sand higher than 50 %, absence of structure and low capacity of the soil to retain water and metals. High heavy metal (Cd, Cr, Cu, Pb, Zn, Fe) concentrations were recorded in both soils and selected plants (T. officinale) growing on mine tailings. There is a relationship between metal content in soils and plants, which qualifies T. officinale as an indicator plant, rather than accumulator.

The ability of T. officinale to uptake and translocate heavy metals, particularly the essential micronutrients Zn and Fe, from soil to plant was ascertained.

Plants proved to accumulate heavy metals in their shoots more than in roots. This is an effective and cheap option for phytoremediation of contaminated areas and also to decrease erosion risk. The accumulation of metals in plants, however, affects the normal processes of plant metabolism. The study shows that there is a relationship between high metal contents in plants and their modified morphology: strong reduction of leaf thickness, modified parenchyma structure, and decreased mitochondria organization were ascertained, although toxic symptoms were apparently absent.

The evaluation of metal uptake by plants, combined with geobotanical observations, proved a useful tool to find tolerant plant populations to be used in revegetation programs aimed at reducing the environmental impact of contaminated areas. The selection of new genotypes from metal-tolerant species will bring large advances in phytoremediation of contaminated sites. Further investigations may help understanding if dandelion could be a metal-tolerant plant to grow on slightly metal-contaminated soils for restoration purposes (Maleci et al. 2013; Wahsha et al. 2011).

4.2 Case Study 2

4.2.1 Environmental Impact of PHEs on Native Flora Growing on Mine Dumps

Bioavailable heavy metals can enter the food chain through primary producers, reducing growth cycle and altering some biochemical pathways in plants (Loureiro et al. 2006). Moreover, heavy metals induce oxidative stress by generation of hydrogen peroxide, superoxide radical, hydroxyl radical and singlet oxygen, collectively termed reactive oxygen species (ROS) (Wahsha et al. 2012c; Verma and Dubey 2003). Many organic molecules are exposed to severe damage by free radicals after high accumulation of heavy metals in plants (Alfonso and Puppo 2009; Joshi et al. 2005). Formation of ROS in cells is associated with the development of many pathological states (e. g. reduced root elongation, seed germination, signaling imbalance) (Bini et al. 2008; Wahsha and Al-Jassabi 2009). This has contributed to the creation of the oxidative stress concept; in this view, ROS are unavoidable toxic products of O2 metabolism, and aerobic organisms have evolved antioxidant defenses to protect against this toxicity (Alfonso and Puppo 2009). Oxidative stress can increase sharply in cells either due to the decrease in the activity of the antioxidant defense systems or to the overproduction of ROS (Wahsha et al. 2012d; Mukherjee et al. 2007; Soffler 2007). The most harmful effect induced by ROS in plants is the oxidative degradation of lipids, especially polyunsaturated fatty acids (PUFA) in cell membranes known as lipid peroxidation, which can directly cause biomembrane disorganization (Gobert et al. 2010; Wahsha et al. 2010; Timbrell 2009). Several studies reported that ROS can initiate lipid peroxidation through the action of hydroxyl radicals (Armstrong 2008; Katoch and Begum 2003). Lipid peroxidation reactions are usually free radical-driven chain reactions in which one radical can induce the oxidation of PUFA (Abuja and Albertini 2001). The lipid peroxide Malondialdehyde (MDA) is one of the major end-product of lipid peroxidation process (Yadav 2010). In this case, membrane destabilization and fusion are directly correlated with MDA production (Wahsha and Al-Jassabi 2009; Wahsha et al. 2010). The determination of MDA content is widely used as a reliable tool to detect the oxidative stress hazard by estimating the formation of lipid peroxides in biological material (Loureiro et al. 2006; Taulavuori et al. 2001; Zielinska et al. 2001). Furthermore, the formation of ROS and an increased MDA production were observed in plants exposed to different heavy metals as Cr, Pb, Cu and Zn under laboratory conditions (Aravind and Prasad 2003; Baryla et al. 2000; Sinha et al. 2005; Verma and Dubey 2003).

The LPO levels (expressed as MDA contents) in the common dandelion (Taraxacum officinale Weber ex F.H. Wigg. 1780), and different willows (Salix purpurea L., Salix caprea L., and Salix elaeagnos Scop.) vary proportionally with the level of heavy metals in soils of the corresponding site indicating a close relationship between MDA and metals, thus confirming the LPO test to be effective in environmental contamination assessment (see Table 5.6).

Table 5.6 The contents of MDA of dandelion and willows
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The control plants of T. officinale exhibited normal levels of LPO, and it was 0.2063 μM in leaves and 0.1450 μM in roots. There was a dramatic increase in MDA level in leaves and root homogenate from T. officinale collected from Imperina Valley. In agreement with previous results by Savinov et al. (2007), the increase of MDA production in T. officinale was expected because when heavy metal levels increase in soil their absorption by roots will increase, and the lipid peroxidation through the possible excessive generation of free radicals will be incremented. T. officinale responds to the increased heavy metal contents by intensification of LPO processes, which are related to the concentrations of Cu, Zn, Pb and Fe in the soil, as a result of an imbalance in the homeostasis of the antioxidant defence system (Alfonso and Puppo 2009).

Lipid peroxidation in leaves and roots of willows, measured as MDA content, are given in Table 5.6. Compared to control, heavy metals induced oxidative stress in willows was evident from the increased lipid peroxidation in roots, stems and leaves, indicating an enhanced MDA production, with MDA increasing in leaves in comparison to roots and stems. This is in agreement with data reported by Kuzovkina et al. (2004) and Ali et al. (2003). Generally, in both parts of the plant, the MDA contents were found to be positively correlated with metal accumulation (p <0.05). The high level of MDA observed in investigated plants under metal stress might be attributed to the peroxidation of membrane lipids caused by ROS due to metal stress indicating a concentration-dependent free radical generation (Bini 2010; Ali et al. 2003).

The soils in the mining area are highly contaminated by trace elements, mainly Cu, Zn, Pb and Fe. The observed ability of Salix species and T. officinale to continue growth in the presence of heavy metals and to accumulate metals in their tissues, and particularly in leaves, demonstrated their tolerance to moderate to high levels of metals (has this already been states?). Therefore, they have good potential to be used in phytoremediation projects. Our results show that T. officinale, S. purpurea, S. caprea and S. elaeagnos exposed to great metal concentrations in soils result in an increment in LPO in their tissues, suggesting an important role of oxidative stress in the pathogenesis of heavy metal-induced cellular toxicity, and they can be a promising bioindicator for such research. The LPO process proved to be a useful tool for health assessment of wild-growing plant species, as it reflects the anthropic heavy metal pollution in ecosystems.

4.3 Case Study 3

4.3.1 Biological Soil Quality Evaluation of Mine Dumps

Heavy metals have been reported to disturb the ecosystem structure and functioning for long time, and the results of this study largely agree with published data (Wahsha et al. 2012a). Soil health is the continued capacity of the soil to function as a vital living system, providing essential ecosystem services. Within soils, all bio-geo-chemical processes of the different ecosystem components are combined. These processes are able to sustain biological productivity of soil, to maintain the quality of surrounding air and water environments, as well as to promote plant, animal, and human health (Karlen et al. 2001). A common criterion to evaluate long term sustainability of ecosystems is to assess the quality of soil. Recently, several bioindicators of soil quality and health have been reviewed (Chauvat et al. 2003; Parisi et al. 2005). Among them, microarthropods, due to their high sensitivity to respond to environmental changes, play a fundamental role in the dynamics of organic matter and in the fragmentation of soils, at different scales of time and space (Loranger-Merciris et al. 2007). Thus, they can also contribute to metal translocation through the ecosystem in polluted environments. The Soil Biological Quality index (QBS-ar), which is based on microarthropod groups present in the soil (Parisi et al. 2005), may be applied to assess its biological quality: the higher is the number of microarthropod groups adapted to soil habitats, the higher is soil quality.

In soil samples collected from abandoned mixed sulphide mine dumps of Imperina Valley, the QBS-ar values appeared to decrease significantly (p <0.05) with respect to soil pollution by heavy metals. QBS-ar values appeared to decrease significantly (p <0.05) with respect to soil pollution by heavy metals. The correlation matrix (R2) between heavy metals in soil and QBS-ar values showed that QBS-ar values were negatively correlated with Fe (−0.102), Pb (−0.384), Cu (−0.405) and Zn (−0.702). Conversely, our QBS-ar values indicate a significant positive correlation with Cr (0.298) and Cd (0.55). The presence of Acarina, Symphyla, Protura and Collembola is important, being considered metal-tolerant (Migliorini et al. 2004): for example, Symphyla seem to be quite affected by high lead concentrations and our results show a decrease in their abundance in areas with high concentrations of Pb, Zn and Cu.

Former activities proved to affect the microarthropods community altering both quantity and quality of litter and the chemical-physical structure of the microhabitats. We found in the study area a moderate soil health status of the surface horizons due to the ecological success of secondary recolonization after abandonment, although affected by heavy metal contamination. Even if we could not find a statistical difference between QBS-ar and humus forms/ecosystem type, there seem to be different structures of microarthropods communities in terms of richness and evenness.

In this case we hypothesize that 50 years of biological restoration of the mine site could have improved the microarthropods biodiversity, driving humus development towards a better ecosystem functional stability. QBS-ar index proved a useful tool to evaluate soil biological health. However, there is an increasing need for further research focusing on soil health restoration assessment, combining QBS-ar index with soil bio-physical-chemical indicators.

5 Conclusion

It is imperative to acknowledge that current and future exploitation of mineral resources will produce even more wastes and impose more threats on Earth. Therefore, management and utilization of once unwanted mine wastes should be seriously considered through implementation of proper waste disposal management plans which focus on the utilization of wastes rather than considering it as unwanted.