Strategies for Soil Protection and Remediation

CHAPTER 10

Strategies for Soil Protection and Remediation David Fangueiro1, Petra S. Kidd2, Paula Alvarenga1, Luke Beesley3 and Amarilis de Varennes1 1 LEAF, Instituto Superior de Agronomia, Universidade de Lisboa, Tapada da Ajuda, Lisbon, Portugal Instituto de Investigaciones Agrobiolo´gicas de Galicia (IIAG), CSIC, Santiago de Compostela, Spain 3 The James Hutton Institute, Aberdeen, United Kingdom 2

10.1 INTRODUCTION As shown in previous chapters, soil is an extremely complex medium which performs vital functions and generates numerous ecosystem services. As its formation is an extremely slow process, soil can be considered as an essential and nonrenewable resource. Contamination can lead to a decline in soil quality, which in turn reduces its capacity to perform ecosystem functions and provide these essential ecosystem services. Contaminants in soils can be broadly classified as organic or inorganic in origin. Organic contaminants include petroleum hydrocarbons, polychlorinated biphenyls (PCBs), polyaromatic hydrocarbons (PAHs), pharmaceutical substances, pesticides and herbicides, sulfonated aromatics, phenolics, nitroaromatics, explosives, and solvents [1]. Inorganic contaminants include “heavy metals” which are those elements which have an atomic specific mass greater than 6 g cm21. They can be essential elements such as cobalt (Co), copper (Cu), chromium (Cr), manganese (Mn), and zinc (Zn) or nonessential elements such as cadmium (Cd), lead (Pb), and mercury (Hg). Arsenic (As), boron (B), and selenium (Se) fit into this category, although they are metalloids or nonmetals [2]. Throughout the present chapter such elements will be referred to as potentially toxic trace elements (PTEs). PTE can be released into the environment from natural and anthropogenic sources. The former includes weathering of minerals, erosion, and volcanic activity while the latter includes mining, smelting, electroplating, usage in agriculture of pesticides, phosphatic fertilizers and biosolids, sludge dumping, and emission from industry [2,3]. Unlike organic substances, PTEs are not degraded and can accumulate in the environment and in the tissues of living organisms (bioaccumulation), often with increased concentration as they pass from lower to higher trophic levels Soil Pollution DOI: http://dx.doi.org/10.1016/B978-0-12-849873-6.00010-8

r 2018 Elsevier Inc. All rights reserved.

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(biomagnification) [3]. Metal(loid)s are amongst the most frequently occurring soil contaminants at polluted sites across Europe [4], and their presence in elevated concentrations has been identified by the European Commission (EC) as one of the eight major threats to European soils (COM(2002) 179 final). Hence, special attention will be given in the present chapter to remediation of PTE-contaminated soils. There are two principal approaches for soil remediation based on in situ and ex situ methods. Ex situ treatments involve the partial or complete removal of the contaminant from where it occurs [5], e.g., where hydrophobic pollutants (such as PAHs and PCBs) are not effectively removed by in situ treatments due to limited bioavailability. Such conventional remediation techniques are thought to account for about one-third of implemented management measures in Europe [6]. However, during the last three decades, there has been growing interest in the use of “more gentle” soil remediation options, which include in situ contaminant stabilization and plant-based options (phytoremediation). The rising popularity of in situ remediation techniques are, in part due to their relative low cost, capacity to simultaneously treat a wide variety of organic and inorganic pollutants, and potential to restore soil quality and functionality, thus minimizing the risks associated with hazardous waste transportation and disposal ex situ [5]. Two examples of in situ remediation techniques are bioremediation, which relies on microbial metabolism of pollutants, and phytoremediation, which is primarily based on processes harbored or stimulated by plants; it is generally accepted that only through the successful establishment of a vegetation cover will long-term rehabilitation of soils be achieved. As severely contaminated sites are very unfavorable environments for plants, some form of assistance, by applying soil amendments, may be needed to reduce the bioavailability of PTE or promote degradation of organic substances [7]. Such treatments are often referred to as “assisted phytoremediation.” In the present chapter, we will focus on phytoremediation of contaminated soils with special emphasis on remediation techniques applied to mine and industrially polluted soils. Three case studies from different European regions will be presented by way of applied examples. A brief overview of the main environmental policies and sustainable management systems for soil protection will also be presented in order to stress the limitations of existing regulations and highlight the challenges for any future soil policy framework.

10.2 PHYTOREMEDIATION TECHNIQUES AND PRINCIPLES 10.2.1 Fundamentals of phytoremediation Phytoremediaton techniques are those facilitated by the use of plants and they are considered to be more cost-effective and less ecologically invasive than conventional civil engineering techniques [8 10]. Phytoremediation techniques have been developed to target both organic contaminants and PTEs [11 14] and are primarily

Strategies for Soil Protection and Remediation

deployed with the aim of modifying and minimizing the labile (or bioavailable) pool of PTEs, or degrading organic contaminants. Examples of phytoremediation processes that have been developed are as follows: Phytodegradation refers to the use of the metabolic capabilities of plants and rhizosphere microorganisms to uptake, store, and/or degrade organic pollutants. This includes phytovolatilization (rhizosphere and endophytic bacteria) to transform pollutants into volatile compounds that are then released into the atmosphere. An associated process called rhizofiltration removes contaminants from aqueous sources by plant roots and associated microorganisms. Phytostabilization employs metal-tolerant plants to reduce the bioavailability of metals in root systems (in situ stabilization) assisting the growth of a vegetation cover. It is usually aided using soil amendments and referred to as aided or assisted phytostabilization. Phytostabilization does not lead to the actual removal of contaminants but minimizes the potential toxic effects of metals in the environment by reducing pollutant bioavailability and transfer to other environmental compartments or the food chain. Phytostabilization techniques have been suggested to be the most reasonable option for large contaminated areas and with a large metal availability [15]. Phytoexclusion aims to progressively reduce the bioavailability of pollutants by immobilizing or binding them to the soil matrix through the incorporation of organic or inorganic amendments, singly or in combination, to prevent the excessive uptake of essential elements and nonessential contaminants into the food chain. This is often the precursor to the implementation of a stable vegetation cover using trace elementexcluder crop plants which do not accumulate contaminants in their harvestable/ edible tissues. Phytoextraction involves the cultivation of tolerant plants that accumulate and concentrate soil metal contaminants in their aboveground tissues. Phytoextraction can also be aided by use of soil amendments (termed aided or assisted phytoextraction). At the end of the growth period, plant biomass can be harvested and burned to produce metal-enriched ash or “bio-ore” (the process is then known as phytomining). The advantages of the above options, compared to ex situ methods, are that they can restore soil structure, functionality, and quality, and positively influence other ecosystem services (e.g., C sequestration, water filtration and drainage management, restoration of plant, microbial and animal communities, etc.) and local socioenvironmental quality. In addition, these methods could also remediate land to allow the growth of renewable biomass crops, contributing toward achieving EU targets for renewable energy sources (EU directive 2009/28/EC) and avoiding the diversion of EU croplands to biofuel production and other nonfood crops.

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Phytoremediation techniques have been criticized for the length of time necessary for their benefit to be seen. For example, there has been much debate regarding the ineffectuality of phytoextraction to reduce “total” PTE concentrations in soil to below regulatory limits. The main factor determining the duration of phytoextraction is the mass of PTEs removed by the crop per unit of time (years) compared to the mass of PTE in the soil. A decline in plant uptake as soil PTE concentrations deceases can lead to further increases in the time needed for clean-up [16]. Furthermore, its effectiveness can also be reduced by the limited rooting depth of plants and the presence of contaminant mixtures. Nonetheless, the wider benefits mentioned above in terms of ecosystem services can contribute toward increasing the acceptance of these techniques. Benefits may be in the form of direct revenue generating opportunities (e.g., biomass revenues), an increase in natural or cultural capital in an area (e.g., soil and water improvement, provision of green infrastructure, amenity space, etc.), or provision of tangible or intangible economic benefits (e.g., increase in property values, job generation, etc.).

10.2.2 Plant selection Appropriate plant selection is crucial for the successful implementation of phytoremediation strategies and several decades of research have been dedicated to the screening and selection of PTE-tolerant plant species or genotypes. While tolerance to the contaminant in question will always be vital, at other times the selected plant will depend on the remediation option to be used, e.g., PTE-accumulating plants (phytoextraction) or PTE-excluding plants or crop species (phytostabilization/in situ metal immobilization with phytoexclusion). Several studies evaluating the suitability of plant species for phytostabilization have shown that some plants induce significant reductions in plant-available metal fractions in their rhizosphere. These reductions are largely attributed to the root exudation of organic ligands, which chelate metals in the rhizosphere or apoplastic space, preventing their entry into the symplast, and can also vary between plant species, cultivars, and varieties [12]. For phytoextraction, plants must be able to accumulate and tolerate high PTE concentration(s) in their harvestable parts (e.g., shoots) and have a reasonably high biomass production. PTE-hyperaccumulators (such as Noccaea caerulescens, Alyssum murale, and Alyssum corsicum) are able to accumulate extreme concentrations of metal(loid)s (e.g., Cd, Ni, Zn, Se, and As) in their aboveground biomass; such plants are often endemic to metal-enriched substrates, such as ultramafic or calamine soils [11,17,18]. An argument in favor of hyperaccumulators is the possible recovery of PTE from PTE-rich biomass (bio-ores). However, effective recycling of PTE from PTE-loaded plants has only been proven economically feasible in the case of Ni, based on field case studies carried out in the USA [11] and in Albania [19]. The main factor limiting the practical application of

Strategies for Soil Protection and Remediation

hyperaccumulators is the low biomass production of most of these species and the high number of cropping cycles required to remove/extract reasonable concentrations of metal(loid)s. Thus, high-biomass crops (annuals or perennials) and woody plants have been proposed as viable alternatives to hyperaccumulators for phytoextraction of TEs (particularly Cd, Se, and Zn), if they also show relevant shoot PTE removals [i.e., moderate high bioconcentration factor and high shoot yield]. Over the last two decades, both high-yielding crop species, such as tobacco (Nicotiana tabacum) and sunflower (Helianthus annuus), and specific clones of several members of the Salicaceae family have been assessed in Europe for their suitability in phytoextraction processes. The use of high-biomass crops and woody tree species, and examples of field case studies using these plant types, was reviewed by Kidd et al. [13]. For phytostabilization, numerous grass species have been proven effective in Europe in establishing long-term plant cover on heavily contaminated sites, namely Poa pratensis, Agrostis capillaris, Festuca arundinacea, Festuca rubra, and Festuca ovina [20]. In addition to stabilizing soil metals or extracting them (in the case of phytoextraction), the plant species with economic value can be selected. For example, bioenergy crops to produce biofuel, in the forms of ethanol, biodiesel, and biomass for biogas production [21], or plants for the extraction of essential oils for medicinal purposes [22,23]. A more detailed description of plants used for phytostabilization is provided below. On contaminated and/or remediated agricultural land, the implementation of phytoexclusion and in situ stabilization can contribute toward reducing the entrance of harmful PTEs into the human-food chain [13]. Cultivars within species from major staple crops such as wheat, barley, rice, potato, or maize differ widely in their ability to accumulate metal(loid)s and excluder-phenotype commercial cultivars can be selected [24 26]. Cadmium is one element of concern regarding metal uptake into the food chain [27]. Selection of the most appropriate cultivar for use in contaminated sites can ensure that food and forage production is in compliance with the respective regulations on threshold PTE contents. Friesl-Hanl et al. cultivated PTE excluderand accumulator-phenotypes of barley in a contaminated field in Austria [28]. The excluder-phenotype had about 40% less Cd in the grain. When cultivated in combination with the incorporation of gravel sludge and red mud into the contaminated soil, a further decrease of about 30% in Cd uptake could be obtained. Five years after soil treatment in the field, the amendments were still effective immobilizing agents, confirming that the combination of soil treatment with cultivar selection can be an easy and cost-effective option for farmers to improve their situation.

10.2.3 Improvement of phytoremediation techniques Much research has been dedicated to the screening and selection of trace metaltolerant plant species or genotypes for application in phytoremediation. However, the

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number of field trials implementing phytoremediation techniques continues to be well below the number of studies carried out at greenhouse level. Several authors have highlighted the importance of incorporating agronomic practices into the remediation process when moving from greenhouse to field conditions. Aspects such as crop rotations, intercropping, planting density, fertilization, irrigation schemes, bioaugmentation, weed, pest, and herbivory management can greatly affect the overall outcome [10,13,29 31]. Most plant species for phytoremediation techniques have been studied as monocultures. However, monocultures can lead to a decline in biomass yields due to the depletion of nutrients, occurrence of diseases, pests, and weeds and have a negative effect on soil fertility [8,32,33]. In contrast, alternative cropping patterns, such as rotations or plant intercropping, can improve plant productivity and nutrition, enhance biodiversity, and habitat, or aid pest control, thus improving the success of the remediation process. Intercropping of plants aims to stimulate interspecific below-ground interactions, which may result in improved nutrient availability and increased yield of crops [34]. Intercropping can also alter conditions in the shared rhizosphere and thereby affect the availability of selected PTE to neighboring plants [35]. Fertilization regimes can also be designed to increase plant uptake of PTEs, in addition to supplying nutrients to plants. Inorganic fertilizers affect PTE bioavailability, and the type and amount of fertilizer used and interactions between PTE and major nutrients (N, P, and S) and among PTE themselves (e.g., Zn Cd and Fe Cd) are key players in PTE uptake by crops [36,37]. Furthermore, the fertilizer itself can also be a source of PTE; several P fertilizers have been shown to be rich in Cd [38]. More recently, biotechnological approaches have been developed as a means of improving phytoremediation processes. Several authors have suggested that by manipulating plant microbial soil interactions in the rhizosphere phytoremediation processes can be optimized. Inoculation with mycorrhizal fungi and/or plant-associated bacterial strains (rhizobacteria and endophytes) can not only improve plant growth but also modify soil contaminant behavior and their plant uptake, translocation, and accumulation. Phytoremediation of soils contaminated by organic compounds can greatly benefit from plant bacterial associations [39,40]; the rhizosphere and plant endosphere host plant growth-promoting bacteria and microbes with the capacity to degrade organic contaminants and/or modify their bioavailability [39,41,42]. In contrast, the application of plant-associated microorganisms in phytoremediation strategies targeting PTE-contaminated sites is still under development. For phytostabilization purposes, several authors propose the use of microorganisms which can immobilize PTE through sorption to cell components or exopolymers, transport and intracellular sequestration, or the release of metal-binding compounds or precipitation as insoluble organic or inorganic molecules [42]. The effectivity of these plant-associated bioinoculants depends on numerous factors, including the host plant species, soil type and

Strategies for Soil Protection and Remediation

properties, and nature and type of contamination [40,44 46]. Furthermore, their performance under natural conditions has rarely been investigated and future studies should be carried out at a field scale.

10.3 ASSISTED PHYTOSTABILIZATION IN MINE-CONTAMINATED SOILS 10.3.1 Introduction Mining, smelting, and associated anthropogenic activities can contribute to the entrance and dispersal of relatively large amounts of PTEs in the environment. Intensive mining activity in some areas has produced large amounts of waste tailings, with high concentrations of metal(loid)s, which have been deposited in the surrounding areas without further treatment [47,48]. Surface erosion, by wind and water, and effluent drainage has resulted in the dispersal of these pollutants and in the permanent contamination of nearby terrestrial and aquatic ecosystems, generating large areas where soils are highly enriched in PTE, with low pH, poor nutritional conditions, and scarce vegetation cover [47 50]. Phytostabilization when supported by amendments application has been proposed as a suitable technique to decrease the environmental risks associated with metal(loid)enriched mine tailings. This type of phytomanagement has been shown to reduce PTE mobility by altering speciation, as well as to improve soil physicochemical properties. Furthermore, it can increase microbial diversity in soils and restore functionality of PTE-contaminated sites in the longer term [51,52]. Depending on the characteristics of the amendments, the scale of improvement in the chemical and biological characteristics of the contaminated mine soils varies [53,55]. The amelioration of adverse soil properties allows for the establishment of a plant cover, which stabilizes the soil to help prevent dispersion of metal-contaminated particles by water and wind erosion and reduce metal mobility by rhizosphere-induced adsorption and precipitation processes [53,56].

10.3.2 Fundamentals of phytostabilization 10.3.2.1 Soil amendments Soil amendments can strongly reduce the availability of PTE to plants, thus reducing phytotoxicity and facilitate the revegetation of contaminated sites. Numerous amendments have been evaluated with the specific aim of being used in aided phytostabilization of metal-contaminated soils, from organic amendments, e.g., sewage sludge (SS), composted SS, municipal solid-waste compost, green waste-derived compost, cow and pig slurry, poultry manure, peat, paper mill sludge, sugarbeet lime, leonardite, and biochar, to inorganic amendments, e.g., lime, alkaline fly ash, red mud (bauxite residue), vermiculite, iron oxide, and iron grit, or to a combination of both organic and inorganic amendments. Several case studies have illustrated the successful use of soil

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amendments to support the establishment of a persistent plant cover, reduce bioavailability, and mobility of TEs, and augment organic carbon and nutrients needed to support permanent vegetation cover [57]. In terms of reducing bioavailability and mobility of PTEs, the most important processes involved are the transformation of metals in soils, through precipitation dissolution, adsorption desorption, complexation processes, and ion exchange. The practice of liming [chalk or limestone (CaCO3), quicklime (CaO), or hydrated lime (Ca(OH)2] to increase soil pH has been commonly used and associated effects on metal mobility are well known (reviewed by Vangronsveld et al. [10]). However, in some cases liming can be ineffective or even counterproductive in metal immobilization. For example, McLaughlin et al. [58] reported an increase in Cd accumulation in potato tubers after lime addition. The prolonged effects and sustainability of liming are also questionable [59,60]. For this reason, alternative amendments have been proposed. Soils amended with Fe-(hydr)oxides or by-products, rich in Fe oxides, usually reveal a decrease in the most labile PTE fractions (i.e., soluble and exchangeable) and increase in the reducible fraction (i.e., due to oxide binding) [61]. Application of cyclonic ashes (formerly known as beringite) which are rich in clay minerals were shown to induce a strong decrease in plant-available Pb, Zn, and Cd concentrations, and restored vegetation, in the Zn-smelter affected area of Lommel, Maatheide (Belgium) [62,63]. Many of these amendments are by-products of industrial activities and are therefore inexpensive, abundant, and available in large amounts [64]. Furthermore, their usage can provide an environmentally sustainable means for the recycling of such residues (instead of incineration, landfilling, etc.). Metal immobilization, and, in particular, Pb immobilization, has been studied using a range materials rich in phosphates, such as synthetic and natural apatites and hydroxyapatites, phosphate rock, phosphate-based salts, diammonium phosphate, phosphoric acid, and their combinations [26,65]. In general, these materials are considered more effective for Pb immobilization than for Zn, Cu, or Cd. However, some risks associated with the use of phosphate materials have been identified. For example, in cases of soils cocontaminated with Pb and As, P addition can effectively reduce Pb availability but inadvertently solubilize As. Applying organic residues also provides additional benefits by improving soil physical, chemical, and biological properties by modifying organic matter content, increasing water-holding capacity and modifying PTE mobility [66,67]. The formation of stable organo-metal complexes can reduce the availability of PTE to plants for uptake [68]. One, amongst other, key aspects in the sustainability and (self)-maintenance of phytostabilization techniques is the monitoring of amendment-induced effects and the need for new applications [especially in the case of liming or organic (OM) addition], since soil reacidification after liming or OM mineralization may result in the remobilization of sorbed metals.

Strategies for Soil Protection and Remediation

Several authors have recommended the use of organic amendments with a low mineralization rate within a neutral pH range to avoid the release of metals and reduce their uptake and bioaccumulation by plants [69]. Although the most widely described effect of OM additions on metal mobility is one of immobilization, some studies have also shown an enhanced metal solubilization [68,70,71]. The effects of organic amendments on metal(loid) availability depend on the nature of the OM, and on the particular soil type and elements concerned [57,72,73]. Other organic amendments, such as municipal wastewater, SS, landfill leachate, or wood ash, have also been applied in phytostabilization processes. Another important soil organic amendment which has received growing attention in recent years is biochar. Biochar consists of charred organic matter that has undergone controlled thermal decomposition [pyrolysis] at reduced oxygen conditions. When used as a soil amendment, some biochars have the potential to modify the soil properties which are beneficial for the establishment of a plant cover, by neutralizing soil acidity, adsorbing, or precipitating toxic elements, and by improving the soil structure and the ability to retain water [74,75], though some pH-related mobilization of As has been found in some studies. 10.3.2.2 Plants As previously mentioned, a very important aspect to be considered in a phytostabilization-based technique is the choice of an appropriate plant. Autochthonous plant species, growing spontaneously in mine-contaminated soils, are often tolerant to the unfavorable characteristics of these soils and can potentially be used in their phytoremediation [76 78]. For example, Cistus ladanifer and Erica andevalensis (Fig. 10.1) were already highlighted in previous studies as potential candidates to be used in the phytostabilization of mine-contaminated soils from the Iberian Pyrite Belt, since they are often found as colonizers of these soils [23,77,79]. Alvarenga et al. [80] demonstrated, as part of a study about the (bio)availability of PTEs in the abandoned mine area of Sa˜o Domingos (Portugal) (Table 10.1, unpublished results), that these two species present low accumulation factors for the PTEs thus reducing their effectuality in these soils (As, Cu, Pb, and Zn). 10.3.2.3 Limitations of aided [assisted] phytostabilization In some cases, there may be a need for repeated applications of the amendment, if a single application is not sufficient to prevent changes in soil characteristics and the maintenance of PTE immobilization. In fact, the longevity of the effects of the amendments in the field, especially in semiarid regions and areas of rapid mineralization of applied soil amendments, is of particular concern and should be previously evaluated and subsequently monitored [81].

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Figure 10.1 Erica andevalensis (top) growing in mine tailings (middle) and near water streams affected by acidic mine drainage (bottom) at the abandoned mine area of São Domingos (Portugal), in the Iberian Pyrite Belt.

Another important limitation to aided phytostabilization are the contradictory effects for metal(loid) mobility as a response to different organic amendments and alkaline materials [54,82]. In fact, the application of organic and/or inorganic amendments to PTE-contaminated soils, which should promote their immobilization, can,

Strategies for Soil Protection and Remediation

Table 10.1 Concentration Ranges for Trace Elements in Soils and in Erica andevalensis and Cistus ladanifer Shoots, Collected at Different Locations of the São Domingos Mine Area (n 5 5 Sampling Sites, n 5 3 Replicates) As Cu Pb Zn (mg kg21 DM)

Soil concentration ranges

746 5598

484 1928

1049 14,041

816 2140

E. andevalensis

2.0 7.9

7.0 20.0

,DL

52 101

0.001 0.006 1.3 10.3

0.004 0.03 7.0 11.0

n.c. ,DL

0.03 0.09 69 205

0.001 0.008

0.004 0.030

n.c.

0.05 0.75

C. ladanifer

Concentration ranges AF Concentration ranges AF

Accumulation factors (AF) were calculated: AF 5 [metal concentration in the shoots]/[total metal concentration in the soil]. Soil concentrations were adapted from Ref. [80]. DM, dry matter; DL, detection limit; DL(Pb) 5 6.7 mg kg21 DM; n.c., not calculated.

in some cases, lead instead to their mobilization; this can be the case where soils are simultaneously contaminated with metals and metalloids, in particular arsenic [54,74,82,83].

10.3.3 Case study 1 10.3.3.1 Context and description of the case study Aljustrel is a mine located in the southwest of Portugal, in the western sector of the Iberian Pyrite Belt (Fig. 10.2). The mineralization is characterized by the dominance of pyrite (FeS2), associated with other ore minerals, the most important of which are chalcopyrite (CuFeS2), spharelite (ZnS), and galena (PbS). Aljustrel mine was extensively exploited from 1850 to 1991. Afterward, the production was discontinued and reestablished in 2008. Currently, the Aljustrel mining project, developed by Almina Company, is focused in the Moinho, Algares, S. Joa˜o and Feitais ore deposits, with high Zn and Cu reserves. This region is characterized by a Mediterranean mesothermic humid climate, with hot and dry summers and with an annual average rainfall of 500 650 mm. Several studies were conducted by Alvarenga et al. [66,67,84 88], with soils from Aljustrel mine. These soils have a sandy loam texture and low nutrient content and are extremely acidic with low salinity. In these studies, the possibility of using organic wastes as soil amendments for phytostabilization was assessed. In Portugal, landfill deposition of wastes is still relatively high [89], so their diversion to other uses could contribute to a reduction of greenhouse gases emissions and to the end-of-waste policy in Europe [90,91]. The soil selected for the presently discussed study was very acidic (pH 3.9), with low organic

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Figure 10.2 Overview of the Aljustrel mine—case study 1.

matter content (3.7 g kg21) and high concentrations of Cu, Pb, and Zn (362; 4350, and 245 mg kg21, dry matter basis, respectively). Previous incubation studies [84,87] preselected SS, municipal solid-waste compost (MSWC), and garden-waste compost (GWC) (Table 10.2) as suitable soil amendments for this trial at equivalent application rates of 25, 50, and 100 t ha21 (dry weight basis).

Strategies for Soil Protection and Remediation

Table 10.2 Characteristics of the Organic Wastes (Mean 6 Standard Deviation, n 5 3) Parameter SS MSWC GWC

pH EC (dS m21) Organic matter (%) NKjeldahl (%) Ptotal (% DW) Ktotal (g kg21) Cd (mg kg21) Cr (mg kg21) Cu (mg kg21) Ni (mg kg21) Pb (mg kg21) Zn (mg kg21)

6.7 6 0.2 2.90 6 0.09 72 6 2 7.2 6 0.3 1.36 6 0.06 1.3 6 0.3 1.46 6 0.04 15.3 6 0.2 98 6 5 10.0 6 0.1 37 6 1 491 6 12

8.2 6 0.2 5.69 6 0.09 37 6 0.4 1.8 6 0.4 0.61 6 0.09 8.1 6 0.9 4.3 6 1.1 56 6 13 357 6 12 56 6 6 269 6 24 583 6 26

7.9 6 0.1 2.47 6 0.08 39 6 2 1.0 6 0.4 0.094 6 0.004 7.6 6 0.3 1.4 6 0.1 13 6 1 14 6 2 16 6 3 34 6 3 35 6 10

SS, sewage sludge; MSWC, municipal solid-waste compost; GWC, green-waste compost; EC, electrical conductivity. All concentrations report to a dry matter basis. Adapted from Alvarenga et al. [66,67,84 87].

10.3.3.2 Greenhouse pot experiment Soil chemical characteristics, perennial ryegrass (Lolium perenne) growth and metal concentrations in ryegrass were measured in this pot test. Furthermore, soil phytotoxicity, enzymatic activities, and the composition and toxicity of the soil leachate toward Vibrio fischeri and Daphnia magna [66,67,85] were also studied. The results showed that the organic wastes were able to neutralize soil acidity, increase the organic matter content, N, P, and K concentrations, while decreasing the mobile fractions of Cu, Pb, and Zn (extracted by 0.01 M CaCl2), in a direct relation to the application doses of the organic wastes [66]. Results also showed that the greatest increase in ryegrass growth, against the nonamended control soil, was obtained in the presence of 50 t MSWC ha21, followed by SS at the same application dose. However, ryegrass growth decreased at the higher application rate, 100 t ha21, both for SS and MSWC, but especially for SS, showing some phytotoxic symptoms (Table 10.3). Contrastingly, GWC did not contribute to a significant increase in ryegrass growth (Table 10.3), due to its reduced effects on soil pH and supply of essential macronutrients (N, P, K) (Table 10.1; [66]). As for the biological indicators of soil quality, the application of SS led to the greatest overall microbial and biochemical activity in amended soils, evaluated by the activities of dehydrogenase, acid phosphatase, β-glucosidase, protease, and urease (Table 10.3). This can be ascribed to the fact that the SS was not stabilized at the source wastewater treatment plant and was thus applied in an active stage of biodegradation. On the contrary, GWC was applied in a very mature stage and did not increase these enzymatic activities, relative to the unamended soil. Cellulase activity increased with increasing application rates of the

263

Soil Treatments (mean, n 5 3, n 5 9 for the control) Urease (mmol Cellulase Protease NH41 N g21 h21) (mmol (mmol Gluc g21 h21) Tyr g21 h21)

Table 10.3 Relative Organic amendment (t ha21)

Growth of Perennial Ryegrass Grown and Enzymatic Activity in the Different Dehydrogenase Relative β-Glucosidase Acid (mg growth (mmol phosphatase TPF g21 h21) (%) PNP g21 h21) (mmol PNP g21 h21)

Control soil

100a

0.03a

1.3a

0.97a

0.038a

0.34a

,LD

197bc 277de 118a 240cd 323e 238cd 138ab 159ab 158ab

0.40ab 1.26c 1.80d 0.17a 0.28ab 0.77bc 0.05a 0.06a 0.12a

2.9a 17.4b 44.8c 0.7a 1.1a 0.7a 1.8a 4.1a 3.4a

1.03ab 1.74c 2.06c 1.04ab 0.80a 0.74a 1.52ab 1.08ab 0.95a

0.074bc 0.117f 0.087cd 0.057b 0.095de 0.114ef 0.080cd 0.110ef 0.121f

9.75b 19.56c 36.63d 0.49a 0.74a 1.29a 0.35a 0.36a 0.29a

0.56a 3.53c 7.76d 0.16a 2.28b 2.08b ,LD 0.06a 0.37a

SS

MSWC

GWC

25 50 100 25 50 100 25 50 100

MSWC, municipal solid-waste compost; GWC, garden-waste compost; SS, sewage sludge; SD, standard deviation. Values marked with the same letter are not significantly different (Tukey test, P . 0.05). Adapted from Ref. [67].

Strategies for Soil Protection and Remediation

amendments tested, but decreased at the highest SS application rate (Table 10.3). The organic amendments could suppress soil toxicity to levels that did not affect D. magna, when applied at 50 and 100 t ha21, but SS, at the same application rates, increased the soil leachate toxicity toward V. fischeri. [67]). The results obtained in this study suggested that ryegrass may be viable for aided-phytostabilization for this type of minecontaminated soil and that MSWC, and to a lesser extent SS, applied at 50 t ha21, were effective in the in situ immobilization of metals. The results also demonstrated the importance of using biological and ecotoxicological indicators of soil quality to monitor the remediation processes. Although SS immobilized trace elements and increased soil pH, when used at higher application rates its application led to toxicity of soil leachate toward V. fischeri, decreased soil cellulase activity, and impaired ryegrass growth. These effects would be impossible to assess only with physicochemical indicators of soil quality [67]. 10.3.3.3 Field experiment The previous experiments showed that MSWC had the potential to be successfully deployed in the remediation of a highly acidic metal-contaminated soil further than GWC. A field experiment was conducted to evaluate the use of MSWC and of GWC, applied at 50 t ha21, as immobilizing agents in aided-phytostabilization of the same type of highly acidic soil, contaminated with trace elements, but, in this case, using Agrostis tenuis, and supplementing the GWC-amended soil with lime (4.6 t CaO ha21) and mineral fertilization with N, P, and K (100 kg N ha21, 79 kg P ha21, and 150 kg K ha21, supplied as NH4NO3,P2O5, and K2O, respectively) [88]. The results showed that both treatments raised soil organic matter, pH, and soil nutrient concentrations, whilst decreasing Cu and Zn extractable concentrations (Table 10.4). In contrast, extractable As concentration increased greater than 2-fold (Table 10.4). However, the risk of As uptake to plants was still judged to be very low, because it remained as a small fraction of its pseudototal content [88]. From this study, and using soil enzymatic activities as indicators of soil quality, it was concluded that the amended soil had higher enzymatic activities, especially in the presence of plants, which is a good indicator of their beneficial role in the remediation process. The soil amendments allowed the establishment of a healthy plant cover, but the phytoxicity of the mine soil ultimately resulted in their mortality (Table 10.5). Accumulation factors for As, Cu, Pb, and Zn by A. tenuis were low, and their concentrations in the plant were lower than the maximum tolerable levels for grazing cattle. As a consequence, the use of A. tenuis can be recommended for assisted phytostabilization of this type of soil, in combination with one of the compost treatments evaluated (Table 10.5) [88].

265

Table 10.4 Physicochemical Characteristics and CaCl2 0.01 M Extractable Trace Elements Concentrations in Soil with the Different Soil Treatments (mean, n 5 4) Kavailable Asextractable pH OM (% NKjeldahl Pavailable Cuextractable Pbextractable Znextractable w/w) (% w/w) (mg (mg (μg kg21) (mg kg21) (mg kg21) (mg kg21) 21 21 P2O5 kg ) K2O kg )

Control MSWC GWC 1 NPK 1 CaO

NS S NS S NS S

3.9a 3.9a 5.8bc 5.5b 7.0c 6.5bc

0.49ab 0.40a 1.68c 1.69c 1.57c 1.29bc

0.05a 0.07c 0.23d 0.23d 0.13bc 0.12b

17.1a 18.4a 110.4c 77.1bc 95.3c 50.1ab

7.7a 9.8a 157.9b 151.5b 164.7b 102.0b

33a 35a 79a 67a 145b 71a

32b 32b 1.2a 1.4a ,QL ,QL

2.5a 2.6a 2.3a 2.7a 2.6a 2.6a

171bc 240c 92ab 72ab 0.6a 5a

MSWC, municipal solid-waste compost; GWC 1 NPK 1 CaO, green-waste compost 1 mineral fertilization (NPK) 1 liming with CaO; NS, not seeded; S, seeded; DM, dry matter; QL, Quantification limit; QL(Cu) 5 0.5 mg kg21 DM. All concentrations report to a dry matter basis. Values in each column marked with the same letter are not significantly different (Tukey test, P . 0.05). Adapted from Ref. [88].

Strategies for Soil Protection and Remediation

Table 10.5 Plant Biomass and Trace Elements Concentrations in A. tenuis shoots with the different soil treatments (mean, n 5 4) Plant biomass (t ha21)

As

Cu

Shoot (mg kg21 DM)

AF

Control soil MSWC GWC 1 NPK 1 CaO Plant concentrations excessive or toxica Maximum tolerable level for cattleb

Pb

Shoot (mg kg21 DM)

AF

Zn

Shoot (mg kg21 DM)

AF

Shoot (mg kg21 DM)

8 9

0.03 0.03

287 189

AF

Plants have died 1.7 3.0

1.4 1.3

0.007 0.007

12 10

0.04 0.04

5 20

2 20

30 100

30

40

100

0.27 0.26 100 400 500

MSWC, municipal solid-waste compost; GWC 1 NPK 1 CaO, green-waste compost 1 mineral fertilization (NPK) 1 liming with CaO; AF, accumulation factor 5 [metal concentration in the shoots]/[total metal concentration in the soil]; Concentrations refer to a dry matter basis.

Adapted from Alvarenga et al. [88].

10.3.4 Case study 2 10.3.4.1 Context and description of the case study A field trial of aided phytostabilization was implemented in the Cu-rich mine tailings of the Touro mine (A Corun˜a, NW Spain) [92]. The mine was active from 1974 until 1988 and the area is now used only for extraction of materials for road construction. The mine tailings cover an area of around 550 ha over the geological substrate of amphibolite, with significant quantities of metal sulfides (pyrite, pyrrhotite, and chalcopyrite). The mine soils are shallow and extremely gravelly, and are classified as Spolic Technosols (Episkeletic). The uncontrolled oxidation of sulfides generates hyperacidic (pH 2 3) and hyperoxidizing (.500 mvol) soils and waters, with a high electrical conductivity (EC) and high concentrations of sulfates, Fe, Al, and potentially toxic trace metal(loid)s. Furthermore, the tailings soil is deficient in nutrients such as N, P, and K, and poor in OM. The area selected for the experimental field plots was barren, and microbial activity was almost nonexistent (Fig. 10.3). The climate of the region is Atlantic (oceanic), with a mean annual precipitation of 1900 mm and mean annual temperature of 12.6 C [92]. The mine tailings were amended with composted sewage sludges and planted with Salix spp., Populus nigra or A. capillaris cv. Highland. Plant growth, nutritional status and metal accumulation, and soil physico- and biochemical properties were monitored over 3 years (4 years for plant growth). The total soil bacterial community and specific bacterial groups were studied by denaturing gradient gel electrophoresis of 16s rDNA fragments. The principal objective of the study was to establish a healthy plant cover (either a short rotation coppice (SRC) system or a grass cover) by reducing Cu mobility and phytotoxicity and improving soil fertility and health using assisted phytostabilization.

267

Figure 10.3 (A) Overview of the mine tailings where assisted phytostabilization was implemented and the incorporation of composted SS (B). Plant growth 6 months after planting (C) and after 4 years (Salix viminalis (D), Populus nigra (E), and Agrostis capillaris (F)).

Strategies for Soil Protection and Remediation

Table 10.6 Evolution of Soil Physicochemical Properties in the Mine Tailings Soil after Amendment with Compost (Considering the Whole Experimental Plot). Mean values 1 / 2 standard error (in italics) T50 T 5 1 year T 5 2 years T 5 3 years

pHH2 O pHKCl

3.18 6 0.02c 3.06 6 0.02c

5.37 6 0.15b 5.16 6 0.15b

6.22 6 0.11a 5.71 6 0.10a

6.48 6 0.11a 6.04 6 0.09a

467 6 22a 1.2 6 0.4b 33.2 6 3.9a 0.16 6 0.11b 0.44 6 0.28b 29.3 6 3.7a 0.85 6 0.13a 2.43 6 0.42a 243.0 6 29.8a 6.67 6 0.72a 0.45 6 0.07b

529 6 41a 0.3 6 0.1b 28.5 6 1.6a ,loq 0.20 6 0.09b 24.8 6 1.4ab 0.95 6 0.06a 2.68 6 0.15a 194.3 6 10.1ab 8.14 6 0.72a 0.70 6 0.06a

481 6 29a 0.4 6 0.1b 17.4 6 0.8b ,loq 0.18 6 0.09b 14.8 6 0.7b 0.64 6 0.06a 1.73 6 0.09ab 133.3 6 10.6b 7.91 6 0.94a 0.65 6 0.05a

Cu concentration (mg kg21) Pseudototal NaNO3-extractable CEC (cmolc kg21) H1 Al31 Ca21 K1 Mg21 Polsen (mg kg21) %C %N

523 6 21a 33.0 6 3.1a 19.5 6 3.5b 7.96 6 1.22a 7.78 6 1.28a 2.3 6 0.6c ,loq 1.45 6 0.50b 2.1 6 0.3c 0.63 6 0.08b 0.11 6 0.00c

Different letters indicate significant differences with time (P , 0.05). loq, limit of quantification: 0.01 cmolc kg21. Adapted from [92].

10.3.4.2 Resulting findings The implementation of (aided) phytostabilization effectively reduced PTE phytotoxicity while improving soil quality and microbial activity [92]. Compost addition (independent of plant cover) had a strong effect on all physicochemical properties, trace metal availability and biological activity (Table 10.6). One year after compost addition, soil pH was significantly increased and continued to increase over the following two years (P , 0.05). A progressive increase in total C and N content was also observed through time. Compost addition also induced significant changes in the effective CEC; before amendment exchange sites were dominated by the acidic cations H1 and Al31 (with a base saturation of ,20%), while after 3 years, the mean base saturation was .90% (predominantly dominated by Ca21). Available P (P olsen) also increased after compost addition compared to the concentrations in nonamended tailings soil. Compost addition did not significantly alter the pseudototal concentration of Cu but significantly reduced Cu availability. The mean NaNO3extractable Cu concentration in the mine tailings soil before compost addition was 61 mg kg21, and it decreased to 2.9 mg kg21 after 1 year and to 0.5 mg kg21 after 3 years (P , 0.05). Compost addition also induced a significant increase in all assayed enzymes (catalase, dehydrogenase, invertase, β-glucosidase, urease, acid phosphomonoesterase, alkaline phosphomonoesterase, and arylsulfatase). In nonamended soil, the

269

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Soil Pollution

activity of the oxido-reductase enzymes was null, while 1 year after compost amendment mean values of catalase and dehydrogenase activities of 1.17 mmol H2O2 consumed g21 h21 and 0.28 μmol INTF g21 h21 were recorded. For the hydrolase enzymes, enzymes activities were increased by 4- to 23-fold (depending on the enzyme) after 1 year. The improved soil properties led to the establishment of a healthy vegetation cover (both woody and grass species) during the 4 years of the experiment (Fig. 10.4). Mean plant height after 4 years of growth was 219 cm for S. viminalis, 189 cm for S. caprea, and 240 cm for P. nigra (Fig. 10.4). Shoot DW yields after 4 years ranged from 3.7 6 3.0 t ha21 in S. caprea to 4.3 6 2.3 t ha21 in S. viminalis, while for P. nigra a mean biomass production of 4.3 6 0.7 t ha21 was obtained. A. capillaris presented a good establishment and coverage within the first year of seeding and shoot DW yield of close to 5 t ha21 were obtained. Biomass production obtained in the A. capillaris plots was similar to that reported by Alvarenga et al. [88] for A. tenuis (up to 3 t ha21) cultivated in pyrite mine soils amended with green waste-derived compost. The main changes observed in soil physicochemical properties were the result of the compost addition rather than due to plant growth and activity (Fig. 10.4), although soil pH and CEC in compost-amended tailings was higher under plant cover than in unplanted plots. Plant root activity also stimulated soil-enzyme activities and induced important shifts in the bacterial community structure over time. Mean values of enzyme activities were always higher in planted soils than in unplanted soils and were also influenced by the plant species. Mean values in soils under S. viminalis were generally higher than those recorded under A. capillaris. The bacterial community associated with tree or grass species was clearly separated, particularly for α-Proteobacteria, and a higher diversity was observed in soils under S. viminalis. These results suggest that the analysis of specific bacterial groups may be of interest to monitor the management of PTE-contaminated soils and study the role that different members of the bacterial community play in these processes. The beneficial effects of the phytostabilization process were maintained for at least 3 years after treatment. Furthermore, heavily contaminated mine tailings such as those studied here are generally considered unsuitable for the cultivation of food or fodder crops but the production of energy crops, such as SRC of willow and poplar, offer an alternative method for the management of this type of land. The use of waste-derived amendments, such as the compost used in this study, represent a practical solution for the valorization of waste products and is in accordance with sustainable management practices within current European policies (EU’s Zero Waste Policy 2013).

Strategies for Soil Protection and Remediation

7.00

Unplanted S. vinunalis S. caprea P. nigra A. capillaris

a ab ab b

a b

a

a

a a

a

b o

pH value

6.00 5.00

o o o

o 4.00

o o

o

o

S. caprea P. nigra

Unplanted S. vinunalis

aaa

60 CEC (cmolc kg–1)

8.00

a

A. capillaris

a a

ab

40 ab ab ab

b

a

b

20

3.00 0

2.00 0

1

2

3

0

2

140

S. caprea P. nigra

Unplanted S. vinunalis

A. capillaris

400

25

120

20

100 15

80

10

60

5

40

0

20

ab ab ab b

0 0

1

b a

ab

2

ab

2

Time (years)

0 3

B

B 0.5

B

B

B 1

A

A A

aa a a

A

A A

a b b b b

a

b

a

b a

3

a

ab a

a a

300

200

a

a

S. vinunalis S. caprea P. nigra

100 a 1

3

Time (years)

Tree height (cm)

NaNO3-extractable Cu (mg kg–1)

Time (years)

A

A

A

B 2

3

4

Time (years)

Figure 10.4 Soil pH, CEC, and NaNO3-extractable Cu (median, range, and standard deviation) in unplanted and planted compost-amended mine tailings, and growth of tree species and Agrostis capillaris, over time. Different capital letters denote significant differences between years within each plant species and different lower case letters indicate significant differences between plant species within the same year (P , 0.05). Reprinted from [92]. With permission from Elsevier.

10.4 INDUSTRIALLY CONTAMINATED SITE REMEDIATION 10.4.1 Introduction Thousands of synthetic organic compounds have been produced for many uses since industrialization began. Many end up in soils and sediments, mainly as a consequence

271

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Soil Pollution

of improper disposal, but also as a result of accidents such as leakages. Microorganisms are often unable to degrade substances with chemical links other than those present in natural molecules, and these xenobiotic substances will accumulate in the environment and be eventually absorbed by plants and animals thereby contaminating the food chain. Other organic substances are by-products or result from incineration or disposal of the original compounds. Persistent organic pollutants are of particular concern as they tend to bioaccumulate, i.e., to be found at greater concentrations at higher levels of the food chain. Soil contamination with oil usually results from local spills or leakage from containers or pipes. As oil is extracted in large quantities, it is not surprising that contamination from oil or its derivatives is very serious and widespread. Oil is a mixture of different hydrocarbons that can be refined to produce materials such as plastics, paints, and solvents. Solvents extracted or released from oil include benzene, toluene, ethylbenzene, and xylene, usually designated together as BTEX compounds. Most plastics end up in landfills or in oceans as they are recalcitrant materials presenting a high risk to marine organisms [93]. Eventually, they may be broken down into very small particles that can be taken up and concentrate in the aquatic food chain [94]. When plastics are incinerated, PAHs are produced. Dioxins can also be emitted when plastics contain chlorine, as for example in poly(vinyl chloride). PAHs are naturally formed during volcanic eruptions and whenever organics substances are burned [95]. They result in a diffuse contamination of the atmosphere that may travel far and reach soils through deposition. PAHs have generally low solubility in water (which decreases with molecular mass) and tend to be adsorbed to soil particles or sediments, a process that increases their resistance to degradation. A more complete discussion relative to organic compounds contamination can be found in Chapter 5, Organic Pollutants in Soils.

10.4.2 Context and description of the case study Significantly elevated levels of inorganic and organic pollutants are often in coexistence in soils from previously heavily industrialized areas. The Northwest and Midlands region of England (UK) was the hub of vast chemical and metallurgic manufacturing and processing sites, many such sites being located in the vicinity of the Widnes area. These sites produced wastes contaminated with heavy metals, arsenic, antimony, and various organic contaminants such as PAHs. Following the decline of heavy manufacturing, many industrial sites have been abandoned and remain so without remedial works to their contaminated and often phytotoxic substrates. As previously discussed in other parts of the chapter, organic materials can be used as soil amendments to reduce the bioavailability of organic and inorganic contaminants.

Strategies for Soil Protection and Remediation

To evaluate the ability of green-waste compost (GWC) to reduce the mobility and toxicity of heavy metals and PAHs a 60-day pot trial was conducted on a sedimentderived soil from a canal bank site at Kidsgrove in the Staffordshire North Midlands region of England (UK). This soil was thought to have been contaminated by industrial effluent discharged into a neighboring canal as well as by a local gasworks, and subsequent dredging of canal bottom sediments, resulting in the formation of a spontaneously vegetated sediment-derived soil in the last 50 70 years. The soil profile consisted of highly elevated total concentrations of As, Cd, and Zn (100 300 mg kg21) variously distributed. GWC was sources from local municipal green [garden] waste collections and was mixed with soil before being potted, watered, and maintained under controlled experimental conditions. Soil pore water analysis was used to determine the mobility of metal(loid)s following the application and mixing of GWC. Bioavailable PAHs were measured by cyclodextrin extraction of soils from the pots following the 60-day experiment, while shoot emergence phytotoxicity testing using perennial ryegrass (L. perenne var Cadix) indicated the relative improvement in soil conditions relevant to plant growth, resulting from the compost applications. Compost application reduced Cd and Zn concentrations in soil pore water, whilst also reducing bioavailable 2 and 3 ringed PAHs and also reducing 4 and 5 ringed PAHs (Fig. 10.5). This resulted in the promotion of ryegrass shoot emergence from 61% (control soil) to ,75% (compost treated soil). In common with most experiments involving soils, large differences between replicates were observed. However, these results, and those from other experiments, indicate that the amendment of soils contaminated simultaneously with inorganic and organic contaminants may achieve useful reductions in metal(loid) and PAH bioavailability and result in more fruitful revegetation attempts.

10.5 SUSTAINABLE MANAGEMENT SYSTEMS AND ENVIRONMENTAL POLICIES FOR SOIL PROTECTION Soil has to be managed in order to preserve its functions for the next generations [96]. In this sense, the European Union (EU) prepared, in 2006, a draft for a Soil Framework Directive (EC, 2006) which presented a Thematic Strategy toward soil protection considering eight large-scale threats to European soils: (1) erosion, (2) loss of organic matter, (3) contamination, (4) compaction and other physical soil degradation, (5) salinization, (6) decline of biodiversity, (7) soil sealing by infrastructure, and (8) floods and landslides [97]. This proposal was very important and promoted a discussion on how to translate soil science into environmental policies [98]. However, the proposal was withdrawn by the Commission in May 2014 and, to date, soil protection is not a specific objective of EU legislation [99].

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Soil Pollution

35 Bioavailable PAHs Concentration in soil extraction (mg kg–1)

274

30 Soil Soil+Compost 25

20

15

10

5

0 s

ing

2-r

s

ing

3-r

s

ing

4-r

s

H

ing

5-r

um

of

PA

S

Figure 10.5 Bioavailable PAH concentrations in a contaminated soil amended with green-waste compost.

Nevertheless, requirements for soil protection are often included in other EU policies, such as the Nitrates Directive and the Water Framework Directive, and in the national legislations of various European countries, specifically addressing water, waste, and mining regulations. Although these policies consider soil contamination and contribute indirectly to soil protection, they only feature soil as a secondary objective [99]. However, some countries have specific legislation and regulations concerning the management of contaminated sites. Rodrigues et al. [100,101] exhaustively reviewed soil policies from different European countries dealing with soil contamination aspects. Most of these soil protection policies started to be implemented in the 1990s, and they mainly focused on soil contamination as the major issue. They mostly include, on one hand, preventive regulations intended to minimize the introduction of new contaminants into some environmental compartments (namely soil), and on the other hand, investigation and clean-up of contaminated land. However, these policies covered different soil types and contaminants that were mostly relevant at a national level [100,101]. Environmental policies for soil protection need to change so as to consider soil management in an integrated perspective, and some studies have demonstrated the importance of this approach [102,103]. Monitoring soil changes, namely as a consequence of the impact of human activities, has been the main action for soil protection

Strategies for Soil Protection and Remediation

in several countries in recent years [104] but, according to the ITPS [105], it is also important to understand the spatial and temporal variation in soil functions. Nevertheless, due to the complexity of soil processes, such information is available only through soil quality models. Recent studies recommended an integrated approach to define land-use criteria taking into consideration environmental and social aspects [98,100,102,103]. This might imply a multisector and multiactors approach since some alteration in policies and legislation of other sectors, such as waters or waste disposal, will probably be needed [102]. An integrated and common approach is therefore required and several actions were taken over the last years at both EU and international levels toward that aim. Another aspect that is important to consider is the fact that contaminants do not respect borders. So, any actions to prevent or remediate soil contamination cannot be limited to a regional or national scale [103]. Even if the geographic proximity of countries is the main driver for common soil policies, other aspects such as soil characteristics, nature and source of contamination, or soil use can easily be used to build some integrated soil management regulations. Some international Environmental Agreements such as the Kyoto Protocol (relative to carbon storage in soil) or the United Nation Convention to combat desertification also have an influence on soil management and protection. All agriculture policies generally have a strong relationship with soils. In fact, six of the soil degradation processes recognized at the EU level [97] are closely linked to agricultural practices: erosion, organic carbon depletion, soil biodiversity decline, compaction, contamination, and salinization and sodification [99]. Therefore, the Common Agricultural Policy of the EU is increasingly designed to meet a wide range of needs, including soil protection. However, all these isolated measures cannot be considered as a full and strong soil protection policy. The increase in food demand in some regions has also led to more concerns relative to soil quality, since it is directly linked to soil productivity and agriculture sustainability [106]. Integrated soil management policies have to face a major challenge in meeting the demands of an increasing world population, while maintaining sustainable agro-ecosystems, which should be considered from long-term soil fertility, environmental and socioeconomic perspectives [96,107]. There is, therefore, the need for a policy framework that recognizes the environmental importance of soil, takes into account problems arising from the competition between its concurrent uses, both ecological and socioeconomic, and that aims at maintaining soil’s multiple functions [108]. Changes to policy, governance, and funding worldwide will be needed to conserve and manage the soil resource and to restore already degraded systems. Such changes must engage all land-use stakeholders, must involve educational, training and extension programs and must embrace the multidisciplinarity required for effective soil conservation and management [107].

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Soil Pollution

ACKNOWLEDGMENTS The authors acknowledge the support of the LIFE programme (project NoWaste LIFE14 ENV/PT/ 000369), the Interreg-Sudoe programme (PhytoSUDOE SOE1/P5/E0189), the Scottish Government’s Rural and Environmental Sciences and Analytical Services (RESAS), and the FCT—Portuguese Foundation for Science and Technology (Project support UID/AGR/04129/2013 attributed to LEAF).

REFERENCES [1] L. Wu, Z. Li, C. Han, L. Liu, Y. Teng, X. Sun, et al., Phytoremediation of soil contaminated with cadmium, copper and polychlorinated biphenils, Int. J. Phytoremed 14 (2012) 570 584. [2] A.P. Pinto, A. Varennes, R. Fonseca, D. Martins Teixeira, Phytoremediation of soils contaminated with heavy metals: techniques and strategies, in: A.A. Ansari, S.S. Gill, R. Gill, G.R. Lanza, L. Newman (Eds.), Management of Environmental Contaminants, 1, Springer, Wien, 2015, pp. 133 155. [3] H. Ali, E. Khan, M.A. Sajad, Phytoremediation of heavy metals—concepts and applications, Chemosphere 91 (2013) 869 881. [4] European Environment Agency (EEA), Progress in management of Contaminated Sites (LSI 003), European Environmental Agency, Copenhagen, Denmark, 2014 ,http://www.eea.europa.eu/dataandmaps/indicators/progress-in-management-of-contaminated-sites3/assessment#toc-2. (accessed: 7.10.2014). [5] P.J.J. Alvarez, W.A. Illman, Bioremediation technologies, in: J.L. Schnoor, H. Zehnder (Eds.), Bioremediation and Natural Attenuation. Process Fundamentals and Mathematical Models. An Environmental Science and Technology, Wiley, Hoboken, NJ, 2006, pp. 351 455. [6] M. van Liedekerke, G. Prokop, S. Rabl-Berger, M. Kibblewhite, G. Louwagie, Progress in the Management of Contaminated Sites in Europe, Institute for Environment and Sustainability, European Commission, Luxembourg, 2014. [7] S.L. Brown, R.L. Chaney, J. Hallfrisch, Q. Xue, Effect of biosolids processing on lead bioavailability in an urban soil, J. Environ. Qual 32 (2003) 100 108. [8] M. Mench, N. Lepp, V. Bert, J.P. Schwitzgue´bel, S. Gawronski, P. Schro¨der, et al., Successes and limitations of phytotechnologies at field scale: outcomes, assessment and outlook from COST Action 859, J. Soils Sediments 10 (2010) 1039 1070. [9] M. Mench, J.P. Schwitzgue´bel, P. Schroeder, V. Bert, S. Gawronski, S. Gupta, Assessment of successful experiments and limitations of phytotechnologies: contaminant uptake, detoxification and sequestration, and consequences for food safety, Environ. Sci. Pollut. Res. 16 (2009) 876 900. [10] J. Vangronsveld, R. Herzig, N. Weyens, J. Boulet, K. Adriaensen, A. Ruttens, et al., Phytoremediation of contaminated soils and groundwater: lessons from the field, Environ. Sci. Pollut. Res. Int 16 (2009) 765 794. [11] R.L. Chaney, J.S. Angle, C.L. Broadhurst, C.A. Peters, R.V. Tappero, D.L. Sparks, Improved understanding of hyperaccumulation yields commercial phytoextraction and phytomining technologies, J. Environ. Qual 36 (2007) 1429 1443. [12] P. Kidd, J. Barcelo´, M.P. Bernal, F. Navari-Izzo, C. Poschenrieder, S. Shilev, et al., Trace element behavior at the root soil interface: implications in phytoremediation, Environ. Exp. Bot. 67 (2009) 243 259. [13] P. Kidd, M. Mench, V. A´lvarez-Lo´pez, V. Bert, I. Dimitriou, W. Friesl-Hanl, et al., Agronomic practices for improving gentle remediation of trace element-contaminated soils, Int. J. Phytorem. 17 (2015) 1005 1037. [14] E. Pilon-Smits, Phytoremediation, Annu. Rev. Plant Biol. 56 (2005) 15 39. [15] J. Schwitzgue´bel, Phytoremediation of polluted soils: hype, hope and facts, in: Paper Presented at the Invited Lecture Presented at the International Congress on Phytoremediation of Polluted Soils, E´cole Polytecnique Fe´de´rale de Lausanne, Switzerland, 2014.

Strategies for Soil Protection and Remediation

[16] B.H. Robinson, G. Ban˜uelos, H.M. Conesa, M.W.H. Evangelou, R. Schulin, The phytomanagement of trace elements in soil, Crit. Rev. Plant Sci. 28 (2009) 240 266. [17] A. van der Ent, A.J.M. Baker, R.D. Reeves, A.J. Pollard, H. Schat, Hyperaccumulators of metal and metalloid trace elements: Facts and fiction, Plant Soil 362 (2013) 319 334. [18] A. van der Ent, A.J.M. Baker, M.M.J. van Balgooy, A. Tjoa, Ultramafic nickel laterites in Indonesia (Sulawesi, Halmahera): mining, nickel hyperaccumulators and opportunities for phytomining, J. Geochem. Explor. 128 (2013) 72 79. [19] A. Bani, G. Echevarria, S. Sulc¸e, J.L. Morel, A. Mullai, In-situ phytoextraction of Ni by a native population of Alyssum murale on an ultramafic site (Albania), Plant Soil 293 (2007) 79 89. [20] T. Stuczynski, G. Siebielec, W.L. Daniels, G. McCarty, R.L. Chaney, Biological aspects of metal waste reclamation with biosolids, J. Environ. Qual. 36 (2007) 1154 1162. [21] A.G. Babu, J. Shim, K.S. Bang, P.J. Shea, B.T. Oh, Trichoderma virens PDR-28: a heavy metaltolerant and plant growth-promoting fungus for remediation and bioenergy crop production on mine tailing soil, J. Environ. Manage. 132 (2014) 129 134. [22] M.M. Abreu, E.S. Santos, M. Ferreira, M.C.F. Magalha˜es, Cistus salviifolius a promising species for mine wastes remediation, J. Geochem. Explor. 113 (2012) 86 93. [23] E.S. Santos, M.M. Abreu, C. Nabais, M.C.F. Magalha˜es, Trace element distribution in soils developed on gossan mine and Cistus ladanifer L. tolerance and bioaccumulation, J. Geochem. Explor 123 (2012) 45 51. [24] P. Stolt, H. Asp, S. Hultin, Genetic variation in wheat cadmium accumulation on soils with different cadmium concentrations, J Agron. Crop Sci. 192 (2006) 201 208. [25] H. Yu, J.L. Wang, W. Fang, J.G. Yuan, Z.Y. Yang, Cadmium accumulation in different rice cultivars and screening for pollution-safe cultivars of rice, Sci. Total Environ. 370 (2006) 302 309. [26] F. Chen, J. Dong, F. Wang, F.B. Wu, G.P. Zhang, G.M. Li, et al., Identification of barley genotypes with low grain Cd accumulation and its interaction with four microelements, Chemosphere 67 (2007) 2082 2088. [27] C.A. Grant, J.M. Clarke, S. Duguid, R.L. Chaney, Selection and breeding of plant cultivars to minimize cadmium accumulation, Sci. Total Environ. 390 (2008) 301 310. [28] W. Friesl-Hanl, K. Platzer, O. Horak, M.H. Gerzabek, Immobilising of Cd, Pb, and Zn contaminated arable soils close to a former Pb/Zn smelter: a field study in Austria over 5 years, Environ. Geochem. Health 31 (2009) 581 594. [29] M.J. McLaughlin, D. Parker, J. Clarke, Metals and micronutrients—food safety issues, Field Crops Res 60 (1999) 143 163. [30] D. Claus, H. Dietze, A. Gerth, W. Grosser, A. Hebner, Application of agronomic practice improves phytoextraction on a multipolluted site, J. Environ. Eng. Landscape Manage 15 (2007) 208 212. [31] T. Vamerali, L. Marchiol, M. Bandiera, G. Fellet, N.M. Dickinson, P. Lucchini, et al., Advances in agronomic management of phytoremediation: methods and results from a 10-year study of metalpolluted soils, Italian J. Agron. 7 (2012) 323 330. [32] S. Facknath, B. Lalljee, Allelopathic strategies for eco-friendly crop protection, in: S.S. Narwal, R. E. Hoagland, R.H. Dilday, M.J. Reigosa (Eds.), Allelopathy in Ecological Agriculture and Forestry: Proceedings of the III International Congress on Allelopathy in Ecological Agriculture and Forestry, Dharwad, India, 18 21 August 1998, Springer Netherlands, Dordrecht, 2000, pp. 33 46. [33] M.M. Lasat, Phytoextraction of metals from contaminated soil: a review of plant/soil/metal interaction and assessment of pertinent agronomic issues, J. Hazard Substan. Res. 2 (2000) 1 25. [34] G. Wieshammer, R. Unterbrunner, T.B. Garcı´a, M.F. Zivkovic, M. Puschenreiter, W.W. Wenzel, Phytoextraction of Cd and Zn from agricultural soils by Salix ssp. and intercropping of Salix caprea and Arabidopsis halleri, Plant Soil 298 (2007) 255 264. [35] Y.T. Tang, T.H.B. Deng, Q.H. Wu, S.Z. Wang, R.L. Qiu, Z.B. Wei, et al., Designing cropping systems for metal-contaminated sites: a review, Pedosphere 22 (2012) 470 488. [36] M.J. Mench, Cadmium availability to plants in relation to major long-term changes in agronomy systems, Agric. Ecosys. Environ 67 (1998) 175 187. [37] K. Tiller, D. Oliver, M. McLaughlin, R. Merry, R. Naidu, I. Iskander, et al., Managing cadmium contamination of agricultural land, Remediation of Soils Contaminated with Metals. Proceedings of a Conference on the Biogeochemistry of Trace Elements, 5 10 September 1993, Science Reviews Ltd, Taipei, 1997, pp. 225 255.

277

278

Soil Pollution

[38] M.J. McLaughlin, B.R. Singh, Cadmium in Soils and Plants, 85, Kluwer Academic Publishers, Amsterdam, 1999. [39] N. Weyens, D. van der Lelie, S. Taghavi, L. Newman, J. Vangronsveld, Exploiting plant microbe partnerships to improve biomass production and remediation, Trends Biotechnol. 27 (2009) 591 598. [40] N. Weyens, S. Truyens, J. Dupae, L. Newman, S. Taghavi, D. Van Der Lelie, et al., Potential of the TCE-degrading endophyte Pseudomonas putida W619-TCE to improve plant growth and reduce TCE phytotoxicity and evapotranspiration in poplar cuttings, Environ. Pollut. 158 (2010) 2915 2919. [41] V. Andria, T.G. Reichenauer, A. Sessitsch, Expression of alkane monooxygenase (alkB) genes by plant-associated bacteria in the rhizosphere and endosphere of Italian ryegrass (Lolium multiflorum L.) grown in diesel contaminated soil, Environ. Pollut. 157 (2009) 3347 3350. [42] S.D. Siciliano, N. Fortin, A. Mihoc, G. Wisse, S. Labelle, D. Beaumier, et al., Selection of specific endophytic bacterial genotypes by plants in response to soil contamination, Appl. Environ. Microbiol. 67 (2001) 2469 2475. [43] G.M. Gadd, Microbial influence on metal mobility and application for bioremediation, Geoderma 122 (2004) 109 119. [44] C. Becerra-Castro, C. Monterroso, A´. Prieto-Ferna´ndez, L. Rodrı´guez-Lamas, M. Loureiro-Vin˜as, M.J. Acea, et al., Pseudometallophytes colonising Pb/Zn mine tailings: a description of the plantmicroorganism-rhizosphere soil system and isolation of metal-tolerant bacteria, J. Hazard. Mater. 217 218 (2012) 350 359. [45] M.I. Cabello-Conejo, C. Becerra-Castro, A´. Prieto-Ferna´ndez, C. Monterroso, A. Saavedra-Ferro, M. Mench, et al., Rhizobacterial inoculants can improve nickel phytoextraction by the hyperaccumulator Alyssum pintodasilvae, Plant Soil 379 (2014) 35 50. [46] A. Sessitsch, M. Kuffner, P. Kidd, J. Vangronsveld, W.W. Wenzel, K. Fallmann, et al., The role of plant-associated bacteria in the mobilization and phytoextraction of trace elements in contaminated soils, Soil Biol. Biochem. 60 (2013) 182 194. [47] J.M.S. Oliveira, J. Farinha, J.X. Matos, P. A´vila, C. Rosa, M.J.C. Machado, et al., Diagno´stico Ambiental das Principais A´reas Mineiras Degradadas do Paı´s, Bol. Minas—Inst. Geol. Min. (IGM) 39 (2) (2002) 67 85 in Portuguese. [48] J.X. Matos, L.P. Martins, Reabilitac¸a˜o ambiental de a´reas mineiras do sector portugueˆs da Faixa Piritosa Ibe´rica: estado da arte e perspectivas futuras, Bol. Geol. Min. 117 (2) (2006) 289 304 in Portuguese. [49] A.M. A´lvarez-Valero, R. Pe´rez-Lo´pez, J. Matos, M.A. Capita´n, J.M. Nieto, R. Sa´ez, et al., Potential environmental impact at Sa˜o Domingos mining district (Iberian Pyrite Belt, SW Iberian Peninsula): evidence from a chemical and mineralogical characterization, Environ. Geol. 55 (2008) 1797 1809. [50] P. Alvarenga, P. Palma, A. de Varennes, A.C. Cunha-Queda, A contribution towards the risk assessment of soils from the Sa˜o Domingos Mine (Portugal): chemical, microbial and ecotoxicological indicators, Environ. Pollut. 161 (2012) 50 56. [51] T. Pardo, R. Clemente, M.P. Bernal, Effects of compost, pig slurry and lime on trace element solubility and toxicity in two soils differently affected by mining activities, Chemosphere 84 (2011) 642 650. [52] R. Zornoza, J.A. Acosta, S. Martı´nez-Martı´nez, A´. Faz, E. Baath, Main factors controlling microbial community structure and function after reclamation of a tailing pond with aided phytostabilization, Geoderma 245 246 (2015) 1 10. [53] M. Arienzo, P. Adamo, V. Cozzolino, The potential of Lolium perenne for revegetation of contaminated soils from a metallurgical site, Sci. Tot. Environ 319 (2004) 13 25. [54] M. Mench, S. Bussie`re, J. Boisson, E. Castaing, J. Vangronsveld, A. Ruttens, et al., Progress in remediation and revegetation of the barren Jales gold mine spoil after in situ treatments, Plant Soil 249 (2003) 187 202. [55] S.X. Yang, B. Liao, Z.H. Yang, L.Y. Chai, J.T. Li, Revegetation of extremely acid mine soils based on aided phytostabilization: a case study from southern China, Sci. Tot. Environ 562 (2016) 427 434.

Strategies for Soil Protection and Remediation

[56] G.M. Tordoff, A.J.M. Baker, A.J. Willis, Current approaches to the revegetation and reclamation of metalliferous mine wastes, Chemosphere 41 (2000) 219 228. [57] R. Clemente, D.J. Walker, M.P. Bernal, Uptake of heavy metals and As by Brassica juncea grown in a contaminated soil in Aznalco´llar (Spain): the effect of soil amendments, Environ. Pollut. 138 (2005) 46 58. [58] M. McLaughlin, D. Nardecchia, N. Maier, M. Smart, G. Cozens, Remediation of cadmiumcontaminated soils, in: Paper Presented at the Proc. Int. Conf. Soil Remediation Soil Rem2000, Hangzhou, 2000. [59] C.W. Gray, S.J. Dunham, P.G. Dennis, F.J. Zhao, S.P. McGrath, Field evaluation of in situ remediation of a heavy metal contaminated soil using lime and red-mud, Environ. Pollut. 142 (2006) 530 539. [60] A. Ruttens, K. Adriaensen, E. Meers, A. De Vocht, W. Geebelen, R. Carleer, et al., Long-term sustainability of metal immobilization by soil amendments: cyclonic ashes versus lime addition, Environ. Pollut. 158 (2010) 1428 1434. [61] M. Koma´rek, A. Vanˇek, V. Ettler, Chemical stabilization of metals and arsenic in contaminated soils using oxides—a review, Environ. Pollut. 172 (2013) 9 22. [62] J. Vangronsveld, A. Ruttens, J. Colpaert, D. Van der Lelie, In situ fixation and phytostabilization of metals in polluted soils, in: Proc. Int. Conf. Soil Remediation SoilRem2000, Hangzhou, 2000. [63] J. Vangronsveld, A. Ruttens, M. Mench, J. Boisson, N. Lepp, R. Edwards, et al., In situ inactivation and phytoremediation of metal/metalloid contaminated soils: field experiments, in: D.L. Wise, E.J. Cichon, H.I. Inyang, U. Stottmeister (Eds.), Bioremediation of Contaminated Soils, Marcel Dekker, New York, 2000, pp. 859 884. [64] M. Mench, J. Vangronsveld, N.W. Lepp, R. Edwards, Physicochemical aspects and efficiency of trace element immobilization by soil amendments, in: J. Vangronsveld, S.D. Cunningham (Eds.), Metal-Contaminated Soils: In-Situ Inactivation and Phytorestoration, Springer Verlag, Berlin, 1998, pp. 151 182. [65] J. Kumpiene, A. Lagerkvist, C. Maurice, Stabilization of As, Cr, Cu, Pb and Zn in soil using amendments—a review, Waste Manage. 28 (2008) 215 225. [66] P. Alvarenga, A.P. Gonc¸alves, R.M. Fernandes, A. de Varennes, G. Vallini, E. Duarte, et al., Organic residues as immobilizing agents in aided phytostabilization: (I) Effects on soil chemical characteristics, Chemosphere 74 (2009) 1292 1300. [67] P. Alvarenga, P. Palma, A.P. Gonc¸alves, R.M. Fernandes, A. de Varennes, G. Vallini, et al., Organic residues as immobilizing agents in aided phytostabilization: (II) Effects on soil biochemical and ecotoxicological characteristics, Chemosphere 74 (2009) 1301 1308. [68] R. Clemente, M.P. Bernal, Fractionation of heavy metals and distribution of organic carbon in two contaminated soils amended with humic acids, Chemosphere 64 (2006) 1264 1273. [69] L. Blake, K.W.T. Goulding, Effects of atmospheric deposition, soil pH and acidification on heavy metal contents in soils and vegetation of semi-natural ecosystems at Rothamsted Experimental Station, UK, Plant Soil 240 (2002) 235 251. [70] A. Alma˚s, B.R. Singh, B. Salbu, Mobility of cadmium-109 and zinc-65 in soil influenced by equilibration time, temperature, and organic matter, J. Environ. Qual 28 (1999) 1742 1750. [71] S. Tandy, J.R. Healey, M.A. Nason, J.C. Williamson, D.L. Jones, Remediation of metal polluted mine soil with compost: co-composting versus incorporation, Environ. Pollut. 157 (2009) 690 697. [72] J. Kumpiene, M. Mench, C.M. Bes, J.P. Fitts, Assessment of aided phytostabilization of coppercontaminated soil by X-ray absorption spectroscopy and chemical extractions, Environ. Pollut. 159 (2011) 1536 1542. [73] A. Lagomarsino, M. Mench, R. Marabottini, A. Pignataro, S. Grego, G. Renella, et al., Copper distribution and hydrolase activities in a contaminated soil amended with dolomitic limestone and compost, Ecotoxicol. Environ 74 (2011) 2013 2019. [74] L. Beesley, N. Dickinson, Carbon and trace element fluxes in the pore water of an urban soil following greenwaste compost, woody and biochar amendments, inoculated with the earthworm Lumbricus terrestris, Soil Biol. Biochem. 43 (2011) 188 196.

279

280

Soil Pollution

[75] L. Beesley, N.M. Dickinson, Carbon and trace element mobility in an urban soil amended with green waste compost, J. Soils Sediment 10 (2010) 215 222. [76] E. Testiati, J. Parinet, C. Massiani, I. Laffont-Schwob, J. Rabier, H.R. Pfeifer, et al., Trace metal and metalloid contamination levels in soils and in two native plant species of a former industrial site: evaluation of the phytostabilization potential, J. Hazard. Mater. 248 249 (2013) 131 141. [77] R. Pe´rez-Lo´pez, B. Ma´rquez-Garcı´a, M.M. Abreu, J.M. Nieto, F. Co´rdova, Erica andevalensis and Erica australis growing in the same extreme environments: phytostabilization potential of minig areas, Geoderma 230 231 (2014) 194 203. [78] A. Heckenroth, J. Rabier, T. Dutoit, F. Torre, P. Prudent, I. Laffont-Schwob, Selection of native plants with phytoremediation potential for highly contaminated Mediterranean soil restoration: tools for a non-destructive and integrative approach, J. Environ. Manage. 183 (2016) 850 863. [79] P.M. Alvarenga, M.F. Arau´jo, A.L. Silva, Elemental uptake and root-leaves transfer in Cistus ladanifer L. growing in a pyrite contaminated mining area (Aljustrel-Portugal), Water, Air, Soil Pollut 152 (2004) 81 96. [80] P.M. Alvarenga, C. Laneiro, P. Palma, A. Varennes, C. Cunha-Queda, A study on As, Cu, Pb and Zn (bio)availability in an abandoned mine area (Sa˜o Domingos, Portugal) using chemical and ecotoxicological tools, Environ. Sci. Pollut. Res. 20 (2013) 6539 6550. [81] A. Pe´rez-de-Mora, P. Madejo´n, P. Burgos, F. Cabrera, N.W. Lepp, E. Madejo´n, Phytostabilization of semiarid soils residually contaminated with trace elements using by-products: sustainability and risks, Environ. Pollut. 159 (2011) 3018 3027. [82] R. Clemente, W. Hartley, P. Riby, N.M. Dickinson, N.W. Lepp, Trace element mobility in a contaminated soil two years after field-amendment with a greenwaste compost mulch, Environ. Pollut. 158 (2010) 1644 1651. [83] W. Hartley, N.M. Dickinson, P. Riby, E. Leese, J. Morton, N.W. Lepp, Arsenic mobility and speciation in a contaminated urban soil are affected by different methods of green waste compost application, Environ. Pollut. 158 (2010) 3560 3570. [84] P. Alvarenga, A.P. Gonc¸alves, R.M. Fernandes, A. de Varennes, E. Duarte, G. Vallini, et al., Effect of organic residues and liming materials on metal extraction from a mining-contaminated soil, Bioremed. J 12 (2) (2008) 58 69. [85] P. Alvarenga, A.P. Gonc¸alves, R.M. Fernandes, A. de Varennes, G. Vallini, E. Duarte, et al., Evaluation of composts and liming materials in the phytostabilization of a mine soil using perennial ryegrass, Sci. Tot. Environ 406 (2008) 43 56. [86] P. Alvarenga, P. Palma, A.P. Gonc¸alves, N. Baia˜o, R.M. Fernandes, A. de Varennes, et al., Assessment of chemical, biochemical and ecotoxicological aspects in a mine soil amended with sludge of either urban or industrial origin, Chemosphere. 72 (2008) 1774 1781. [87] P. Alvarenga, A.P. Gonc¸alves, R.M. Fernandes, A. de Varennes, G. Vallini, E. Duarte, et al., Reclamation of a mine contaminated soil using biologically reactive organic matrices, Waste Manage Res 27 (2009) 101 111. [88] P. Alvarenga, A. de Varennes, A.C. Cunha-Queda, The effect of compost treatments and a plant cover with Agrostis tenuis on the immobilization/mobilization of trace elements in a minecontaminated soil, Int. J. Phytoremed 16 (2) (2014) 138 154. [89] EEA Report, No. 7/2009, Diverting Waste from Landfill. Effectiveness of Waste-Management Policies in the European Union. European Environment Agency, 2009. ,http://www.eea.europa. eu/publications/diverting-waste-from-landfill-effectiveness-of-wastemanagement-policies-in-theeuropean-union. (accessed: 10.02.2015). [90] Council Directive 2008/98/EC, On Waste and Repealing Certain Directives, November 2008, pp. 3 30, OJ L 312, 22.11.2008. [91] H. Saveyn, P. Eder, End-of-waste criteria for biodegradable waste subjected to biological treatment (compost & digestate): technical proposals, JR Scientific and Policy Reports. European Comission—Joint Research Centre, Seville, 2014. [92] M. Touceda-Gonza´lez, V. A´lvarez-Lo´pez, A. Prieto-Ferna´ndez, B. Rodrı´guez-Garrido, C. Trasar-Cepeda, M. Mench, et al., Aided phytostabilisation reduces metal toxicity, improves soil fertility and enhances microbial activity in Cu-rich mine tailings, J. Environ. Manag. 186 (2017) 301 313.

Strategies for Soil Protection and Remediation

[93] H.K. Webb, J. Arnott, R.J. Crawford, E.P. Ivanova, Plastic degradation and its environmental implications with special reference to poly(ethylene terephthalate), Polymers 5 (2013) 1 18. [94] A.L. Andrady, Microplastics in the marine environment, Mar. Pollut. Bull. 62 (2011) 1596 1605. [95] R. Dabestani, I.N. Ivanov, A compilation of physical, spectroscopic and photophysical properties of polycyclic aromatic hydrocarbons, Photochem. Photobiol. 70 (1999) 10 34. [96] K. Valujeva, L. OSullivan, C. Gutzler, R. Fealy, R.P.O. Schulte, The challenge of managing soil functions at multiple scales: an optimisation study of the synergistic and antagonistic trade-offs between soil functions in Ireland, Land Use Policy 58 (2016) 335 347. [97] European Commission, Communication from the Commission to the Council, the European Parliament, the European Economic and Social Committee, and the Committee of the Regions: Thematic Strategy for Soil Protection, COM (2006) 231 final, Brussels, 2006. [98] J. Bouma, P. Droogers, Translating soil science into environmental policy: a case study on implementing the EU soil protection strategy in The Netherlands, Environ. Sci. Policy 10 (2007) 454 463. [99] G. Louwagie, S.H. Gay, F. Sammeth, T. Ratinger, The potential of European Union policies to address soil degradation in agriculture, Land Degrad. Dev. 22 (2011) 5 17. [100] S.M. Rodrigues, M.E. Pereira, E. Ferreira da Silva, A.S. Hursthouse, A.C. Duarte, A review of regulatory decisions for environmental protection: Part I—challenges in the implementation of natural soil policies, Environ. Int. 35 (2009) 202 213. [101] S.M. Rodrigues, M.E. Pereira, E. Ferreira da Silva, A.S. Hursthouse, A.C. Duarte, A review of regulatory decisions for environmental protection: Part II—the case-study of contaminated land management in Portugal, Environ. Int. 35 (2009) 214 225. [102] H. Posthumus, L.K. Deeks, I. Fenn, R.J. Rickson, Soil conservation in two English catchments: linking soil management with policies, Land Degrad. Dev. 22 (2011) 97 110. [103] A.P. Ferna´ndez-Getino, A.C. Duarte, Soil management guidelines in Spain and Portugal related to EU Soil Protection Strategy based on analysis of soil databases, Catena 126 (2015) 146 154. [104] G.J. Niemi, L.B. Johnson, R.W. Howe, Environmental indicators of land cover, land use, and landscape change, in: R. Armon (Ed.), Environmental Indicators, Springer, Netherlands, 2015, pp. 265 276. [105] Intergovernmental Technical Panel on Soils (ITPS). Plan of Action for Pillar One of the Global Soil Partnership—Promote Sustainable Management of Soil Resources for Soil Protection, Conservation, and Sustainable Productivity, FAO, 2012. [106] V.P. Obade, R. Lal, Assessing land cover and soil quality by remote sensing and geographical information systems (GIS), Catena 104 (2013) 77 92. [107] K. Killham, Integrated soil management—moving towards globally sustainable agriculture, J. Agric. Sci. 149 (2011) 29 36. [108] EEA, Down to earth: soil degradation and sustainable development in Europe, Environmental Issues Series, No. 16, European Environment Agency, Copenhagen, 2000.

281