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Food byproducts as amendments in trace elements contaminated soils
Food byproducts as amendments in trace elements contaminated soils Rafael Clemente, Tania Pardo, Paula Madej´on, Engracia Madej´on, M. Pilar Bernal PII: DOI: Reference:
S0963-9969(15)00145-3 doi: 10.1016/j.foodres.2015.03.040 FRIN 5756
To appear in:
Food Research International
Received date: Revised date: Accepted date:
15 December 2014 19 March 2015 24 March 2015
Please cite this article as: Clemente, R., Pardo, T., Madej´ on, P., Madej´on, E. & Pilar Bernal, M., Food byproducts as amendments in trace elements contaminated soils, Food Research International (2015), doi: 10.1016/j.foodres.2015.03.040
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ACCEPTED MANUSCRIPT Food byproducts as amendments in trace elements contaminated soils
Centro de Edafología y Biología Aplicada del Segura (CEBAS)-CSIC. Campus
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a
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Rafael Clementea,*, Tania Pardoa, Paula Madejónb, Engracia Madejónb, Mª Pilar Bernala
Universitario de Espinardo, 30100 Murcia, Spain. b
Instituto de Recursos Naturales y Agrobiología de Sevilla (IRNAS)-CSIC. Av. Reina
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Mercedes 10, 41080 Seville, Spain.
*Corresponding author: R. Clemente (CEBAS-CSIC).
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Tel.: +34968396385. Fax: +34968396213.
Abstract
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E-mail: [email protected]
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Trace elements stabilization in contaminated soils is frequently the most convenient option for the remediation of large areas with moderate to high levels of contaminants. The use of organic and/or inorganic amendments is often necessary for the amelioration of the soil in order to immobilize the trace metals and metalloids and prevent their transfer to groundwater or to any living organism in the soil, and to allow the establishment of plants when the aim is the regeneration of a vegetation cover. The great upsurge of the agri-food industry in recent decades has led to the generation of huge amounts of waste materials and byproducts throughout the entire cycle of food production and processing. Some of these byproducts, which may give rise to serious environmental concerns regarding their accumulation, handling and transformation,
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ACCEPTED MANUSCRIPT have been satisfactorily used as amendments in the remediation of trace elements contaminated soils. This review offers a general outline of the use of food byproducts
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from the olive oil, sugar beet, wine, and certain other industries for the stabilization of
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trace metals and metalloids in the soil. Both the benefits of using these materials and the
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potential limitations or inconveniences that their utilization may pose, as well as some considerations regarding possible future uses and challenges, are pointed out in the text.
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Keywords: heavy metals; metalloids; soil remediation; sugar beet lime; olive mill
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waste; agri-food waste recycling.
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ACCEPTED MANUSCRIPT 1. Restoration of trace elements contaminated soils and the use of soil amendments. 1.1. Soil contamination with trace elements: sources and general remediation strategies.
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1.2. The use of organic (waste) materials as soil amendments in restoration procedures.
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1.3. Agri-food byproducts generation and use in soil remediation.
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2. The use of olive oil industry byproducts in contaminated soils remediation. 2.1. Olive oil industry byproducts and associated problems.
2.2. Physico-chemical and biological characteristics of wastes from olive processing:
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composting and adding value.
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2.3. The use of fresh and composted olive mill wastes (OMWs) as amendments in soil remediation.
2.3.1. Trace elements solubility and fractionation in OMWs treated soils.
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2.3.2. Nutrients availability and plant establishment. 2.3.3. Soil biological and ecotoxicological properties.
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3. Sugar beet residues as amendments for contaminated soils. 3.1. Sugar beet byproducts production and common uses.
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3.2. Trace elements stabilization in the soil using sugar beet lime. 3.3. Effects on vegetation and soil microbial populations. 3.4. Other sugar beet byproducts used in trace elements contaminated soils. 4. Other types of byproducts used in the remediation of trace elements contaminated soils. 5. Final considerations and aims for the future.
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ACCEPTED MANUSCRIPT 1. Restoration of trace elements contaminated soils and the use of soil amendments.
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1.1. Soil contamination with trace elements: sources and general remediation
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strategies.
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The term ―trace elements‖ is frequently used to refer to heavy metals (like Cd, Pb, and Zn) and metalloids (such as As, Sb, and Se) that can cause environmental and toxicological problems (Alloway, 2013). As the data available regarding the
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concentrations of these elements in soils and plants increase, it is gradually more
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apparent that considerable areas of soil in many parts of the world are contaminated and pose toxicity problems (Alloway, 2013) and that trace elements pollution of land is extremely widespread. Soil pollution exists where, due to human activities, a substance
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is present at a concentration above the natural (background) level and has a negative impact on some or all of the constituents of the environment (Adriano, 2001). The main
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anthropogenic sources of pollution can be divided into two main groups: extensive contamination over large areas (atmospheric, flooding and sediment deposition) and
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localized contamination (e.g., agriculture and horticulture, urban soils and industrial, mining and smelting contamination) (Alloway, 2013). The soil total trace elements concentrations considered to represent pollution vary according to the element, legislative body or country, and soil use. However, total concentrations do not necessarily indicate the toxicity of a soil and phytotoxicity can be absent at elevated total concentrations whilst, conversely, biological processes can be affected at trace elements levels below the designated maximum permissible values. An important distinction in the distribution of trace elements in soils is whether they are able or not to interact with plants (phytoavailable), although toxicity and accumulation will depend upon plant species (Robinson, Bañuelos, Conesa, Evangelou, & Schulin,
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ACCEPTED MANUSCRIPT 2009)Trace elements contamination of soils has received great attention due to the potential risks it poses to human health and because contamination of agricultural soils
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by these pollutants may threaten food safety and disturb the ecosystem (Lee et al.,
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2013). Although some trace elements, such as Cu, Fe, Mn, Ni, and Zn, perform
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biological functions, the majority of them are non-essential. In general, they cannot be degraded by microbial activity or physico-chemical processes and thus they can be considered as ―persistent‖ (Bernal, Clemente, & Walker, 2007). Direct toxicological
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effects of heavy metals on living organisms arise from displacement of other divalent
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metal ions in biomolecules, inhibition of membrane function, enzyme activities, respiration, and protein synthesis, genetic disruptions, oxidative damage, and, in the case of plants, inhibition of photosynthesis and water and nutrient ion uptake (Bernal et
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al., 2007).
The huge volumes of waste generated worldwide by the mining industry have been
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reported to be the main source of heavy metal and metalloids contamination of agricultural soils and crops, and of water, vegetation, and soil invertebrates (Hwang,
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Neculita, & Han, 2012), and therefore form a potential health risk for humans, animals, and plants (Adriano, 2001). Mine spoil can possess elevated levels of pyrites (sulfides, S2-) and the products of their oxidation and hydrolysis, including sulfuric acid (H2SO4), lead to both soil acidification and reaction with heavy metal sulfides to produce soluble metal sulfates. This process can decrease soil pH to below 4 and induce toxicity, due to heavy metals, aluminum, and the high H+ concentration (Bernal et al., 2007). In trace elements contaminated mine soils, the spatial and temporal heterogeneity of factors such as metal solubility and pH can strongly inhibit the establishment of a vegetation cover (de la Fuente et al., 2014).
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ACCEPTED MANUSCRIPT For degraded and/or contaminated soils, physico-chemical problems are aggravated by poor soil structure and a lack of nutrients and organic matter (OM), particularly for sites
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contaminated by mining or other industrial processes. These problems can be
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exacerbated by removal of the topsoil, either during the contaminating activity itself or
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as part of the remediation process (Clemente, Almela & Bernal, 2006a), since this is the most fertile soil horizon, rich in microbial activity, OM and nutrients and with a good physical structure. The lower soil horizons require major improvements in their
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physico-chemical attributes to permit the establishment of a vegetation cover. In
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addition, poor soil structure results in the formation of surface crusts, soil compaction, poor aeration, water retention, and porosity, and physical instability, leading to erosion by air or water and thus to the dissemination of the contaminants (Bernal et al., 2007).
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A multitude of remediation technologies has been developed for the cleanup of trace elements polluted soils. Traditional methods, such as excavation, thermal treatment, and
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chemical soil washing, are typically expensive and destructive. Strategies ―ex situ‖ cause a significant deterioration of soil structure and often come at a high economic
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cost, limiting their use on large contaminated areas. Techniques ―in situ‖ are less expensive and require less soil management, although their implementation can present constraints, like the difficulty of getting extracting/transforming agents in touch with the soil mass (Adriano, Wenzel, Vangronsveld, & Bolan, 2004). The financial aspect cannot be ignored when deciding which soil remediation technique should be used. The cost of each technique is different and this factor is at least as important as the others when deciding which method should be used (Dickinson, Baker, Doronila, Laidlaw, & Reeves, 2009). For example, soils can naturally reduce the mobility and bioavailability of heavy metals; this natural attenuation (natural remediation) relies upon un-enhanced natural processes that provoke the breakdown or immobilization of the pollutant, which
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ACCEPTED MANUSCRIPT is obviously a cheap remedy. Enhanced natural attenuation, which can be considered as one form of bioremediation, includes treatments, such as the addition of inorganic
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and/or organic amendments, which can accelerate the process (Bernal et al., 2007).
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Phytoremediation (the use of green plants to remove pollutants from the environment or
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render them harmless) is appropriate for sites having shallow contamination where the heavy metal levels warrant intervention but are not phytotoxic (Dickinson et al., 2009). The most relevant types of phytoremediation, regarding trace elements contaminated
Phytoextraction: accumulation of the pollutants in the harvestable parts of the
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-
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soils, are:
plants, which can then be removed from the site. -
Phytoimmobilization: reduction of the ―bioavailability‖ of the pollutant, by
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adsorption or absorption, precipitation, or chelation in the root or rhizosphere. Phytostabilization: physical stabilization of the soil, to diminish the risk of
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contaminant transfer via erosion or leaching. Control of the biological and physico-chemical properties of the soil has a fundamental
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role in phytoremediation. Modification of soil conditions will directly affect soil-plant transfer of the contaminants - for example, by changes in the chemical (pH, redox potential, and metal speciation) and microbiological (type and activity of bacteria and mycorrhizae) properties of the rhizosphere. The improvement of the soil physicochemical properties is usually needed when the aim of the remediation of trace elements contaminated soils is the re-establishment of a vegetation cover. This frequently requires the addition of soil amendments, most commonly organic materials (including recycled agricultural or industrial byproducts). But, possible negative effects associated with the use of soil amendments should be considered, with respect to the mobility of the pollutants and their possible assimilation by living organisms (Bernal et al., 2007).
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1.2. The use of organic (waste) materials as soil amendments in restoration
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procedures.
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Application of different types of residual organic materials to the soil as a source of OM
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is a common practice to improve soil properties and might be an environmentally friendly and cost-effective approach to restore extensive areas with a moderate level of contamination (Dickinson et al., 2009). The use of waste materials with this aim would
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mean an additional environmental benefit, as an alternative for their recycling and
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reutilization.
Some materials traditionally used in agriculture have been employed in different experiments on the bioremediation of trace elements contaminated soils. These
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materials improve soil fertility and increase plant productivity, and can also influence trace elements availability. Organic amendments such as composts or peat, which
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contain a high proportion of humified OM, can decrease the bioavailability of metals in soil by adsorption and by the formation of stable complexes with humic substances
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(Clemente & Bernal, 2006). Humic acids have a great capacity to retain or to bind metals, and their molecular size is usually larger than the soil pore size, resulting in low mobility and little leaching through the soil profile (Bernal et al., 2007). A central issue in the application of waste materials to land is balancing potential benefits from waste-derived soil amendments against the risks due to undesired pollutants and pathogenic organisms that may accumulate in the soil system, and which may be transferred to ground and surface water, plants, and animals, including humans. Particularly, the use of organic wastes can lead to problems pertaining to their trace elements content (Karaka, 2004). For example, the uncontrolled addition of animal manures, biosolids, and composts to agricultural soils and landfills has, in some cases,
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ACCEPTED MANUSCRIPT turned them into a source of pollutants for soils and plants, and may contribute excess N and P to the environment (Bernal et al., 2007). Considering the large amounts of low
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grade composts that are produced in an attempt to divert waste from landfill, there is an
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urgent need to find alternative uses for these products; remediation and stabilization of
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trace elements contaminated sites may provide a viable option (Farrell & Jones, 2010). Interestingly, compost quality - in terms of legislative standards (British Standards Institution [BSI], 2005) has shown no correlation with the success of plant
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establishment in soil remediation experiments (Farrell & Jones, 2010).
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Due to recent changes in legislation the main challenge facing waste managers is the development of strategies and techniques as alternatives to landfill for biodegradable waste materials. For example, sewage sludge has been widely used as a soil
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amendment, leading to metal accumulation in soils where it has been added periodically over a long time (McGrath, Chaudri, & Giller, 1995), although improved techniques in
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the treatment of waste waters and stabilization of the sludge gave rise to the so-called ‗extra quality‘ biosolids, suitable for use as soil amendments (Brown, Chaney, Angle, &
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Ryan, 1998). Similarly, municipal solid wastes are produced in extremely high amounts and their composition (specifically, the concentrations of potentially-toxic metals and organic compounds) has usually prevented their use as soil amendments. However, improved source-separation of wastes has allowed the development of high-quality amendments that can be added safely to soils (Bernal et al., 2007). Soil OM has been of particular interest in studies of trace elements retention in soils due to the tendency of transition metals to form stable complexes with organic ligands, one of the most important factors controlling the solubility and bioavailability of metals in the plant-soil system. The addition of OM to the soil can qualitatively (type of ligand) and quantitatively (surface area, etc.) modify its metal retention capacity and alters the
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ACCEPTED MANUSCRIPT soil chemical conditions (pH, redox potential) and the concentrations of chelating and complexing agents in the soil solution and in the solid phase (Bernal et al., 2007).
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Organic amendments do not decrease trace elements concentrations in soils but can
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decrease their bioavailability by shifting them to fractions of low availability associated
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with OM, carbonates, or metal oxides. The direct effect of the OM of the organic materials on trace elements availability cannot be separated from collateral effects on other soil properties. For instance, the presence of phosphates, aluminum compounds,
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and other inorganic minerals in some organic amendments is also believed to be
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responsible for the retention of metals (Brown et al., 1998). The addition of OM to mine waste-contaminated soils has been employed to aid their revegetation, by improving soil fertility and structure and by decreasing the trace
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elements ―bioavailability‖ (chemical forms which can be taken up by different soil organisms and plants; Madejón, Pérez de Mora, Felipe, Burgos, & Cabrera, 2006). For
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instance, Hwang et al. (2012) found mixtures of food waste-based compost and zeolite to be effective in preventing trace element leaching from a mine contaminated soil and
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proposed their addition as one feasible way to treat and/or prevent acid mine drainage. However, the medium- and long-term evolution of low-solubility organo-metallic compounds formed in the soils is of utmost importance, as their solubility could either decrease (favoring trace metal stabilization) or increase (acting as a ‗chemical time bomb‘) with time (Madejón et al., 2010).
1.3. Agri-food byproducts generation and use in soil remediation. The revolution in agriculture has resulted in the fast development of food processing industries all over the world. The term ―food byproducts‖ usually refers to wastes derived from the processing of raw vegetable and animal materials into foodstuffs,
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ACCEPTED MANUSCRIPT which generally consists of the extraction or separation of the nutritious part from the remains having little nutritional value or inedible components (Oreopoulou & Russ,
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2007). Their production covers the whole food life cycle: from agriculture to industrial
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manufacturing and processing and retail and household consumption (Mirabella,
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Castellani, & Sala, 2014). Although they are not classified as hazardous wastes, food industrialization has generated waste in huge quantities, causing environmental pollution (Ajila, Brar, Verma, & Prasada Rao, 2012). Every year, huge amounts of
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biomass in the form of agro byproducts are accumulated; for example, the food waste
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produced by the 27 countries of the EU in 2006 was estimated to be 89 Mt, 39% of this coming from food manufacture and processing (Mirabella et al., 2014). There is a wide variety of agri-food byproducts. Those of a vegetal origin (the objective of the present
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review) are generally organized into different categories according to the origin of the byproducts: crop waste and residues, fruits and vegetables, sugar, starch and
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confectionary industry, oil industry, grain and legumes, distilleries and breweries, food processing wastes, and energy crops and biofuel production wastes (Ajila et al., 2012).
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Crop wastes mainly include materials remaining after harvesting and processing, such as vines, peels and leaves (Ajila et al., 2012), while in the vegetal food-processing industries solid and liquid wastes are generated, including sludge from treatment of their wastewaters. For instance, the major wastes and byproducts of the sugar cane industry are bagasse, molasses, and sugar cane press mud, while the wine making industry produces grape pomace as its main byproduct – this represents an estimated 13% by weight of the grapes (Torres et al., 2002). The management of agro-industrial residues is one of the main issues in agriculture and in agro-industry and has an impact on their economy and day to day operations (Ajila et al., 2012). Agri-food waste materials and byproducts have been traditionally used as
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ACCEPTED MANUSCRIPT bedding for animals and livestock feed, or have been burned in the fields or simply disposed of in landfill and open-dumping sites. Alternative recent roles of these
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byproducts include their use as soil conditioners or green fertilizers, as sources of
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biofuels, thermoplastics, activated charcoal, or components of other composite
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materials, and as a source of food dietary fiber (Mirabella et al., 2014). The different types of utilization of byproducts from the agri-food processing industry can be mainly classified into five categories: as food/feed ingredients; as a carbon source for growing
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useful microorganisms; as fertilizer materials after composting; for energy production
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(biogas); and as a source of high-added-value products (Ajila et al., 2012). The use of these products as soil amendments or conditioners has been limited, due in part to the high application rates required to produce agronomic benefits, a lack of consistency in
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the composition of some products, poor public perception of their utility, and a lack of unbiased scientific research into the agricultural potential of these products (Quilty &
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Cattle, 2011). On some occasions, the cost of the collection, transportation, and processing of the byproducts can exceed the selling price.
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Therefore, the disposal of agri-food byproducts can often be difficult because of their: (i) biological instability and potential growth of pathogens; (ii) water content; (iii) rapid auto-oxidation; and (iv) changes due to enzymatic activities, which may cause the production of undesired odors and environmental problems (Ajila et al., 2012). In addition, most food wastes have low protein content and are therefore not ideal for animal feed, while their high lignin content (as in the cases of olive waste and sugarcane bagasse) also limits their utilization as animal feed since it makes the waste difficult to digest (Laufenberg, Kunz, & Nystroem, 2003). A large part of these wastes, the lignocellulose, remains largely underused and could serve as a feedstock for bioenergy production (bioethanol, biogas), although certain food processing wastes can also
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ACCEPTED MANUSCRIPT contain compounds that may be inhibitory to the fermentation process and this has to be taken into account (Van Dyk, Gama, Morrison, Swart, & Pletschke, 2013). Value-added
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products (e.g., fiber, sugars, enzymes, biofuels or food additives) are also often
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extracted from these materials, but the majority of the waste is currently unused and
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discarded, such as citrus, apple, grape, and sugar beet wastes (Van Dyk et al., 2013). The large amount of waste produced by the food industry, in addition to representing a serious economic and environmental problem, also means a loss of potentially valuable
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materials (Mirabella et al., 2014).Therefore, possible alternative uses should be investigated. The efficient utilization of these byproducts would also contribute to the
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―zero waste economy‖.
A clear example is the fact that agri-food waste materials can be utilized as soil
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conditioners or fertilizers (Arvanitoyannis, Ladas, & Mavromatis, 2006). This can be done through spreading of the untreated food waste on the soil, thereby increasing the
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organic content and microbial biomass of the soil. Tomato waste and olive husks have been utilized in this manner (Van Dyk et al., 2013). In some cases, transformation by
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composting can be undertaken prior to utilization as a soil amendment. Composted fruit and vegetable wastes can provide significant amounts of available K, Ca, and Mg and some P for crop nutrition. Also, although fruit and vegetable wastes are usually acidic, after composting the materials become neutral to alkaline (Grub‘s Up, 2009), facilitating their application to soil. For example, in South Korea, the use of food wastebased compost as organic amendment would be of interest due to the recent prohibition of food waste disposal in landfills, as well as the banning of food wastewater dumping into the sea from 2013 by the London Convention Protocol (Kim & Kim, 2010). However, food waste is considered unsuitable for agricultural purposes because of its high salt content (Hwang et al., 2012). Therefore, alternative markets for food compost,
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ACCEPTED MANUSCRIPT such as its use as an organic amendment for the prevention and/or treatment of acid mine drainage, are urgently required (Kim & Kim, 2010).
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The addition of agricultural wastes to soil, despite its potential benefits, can lead to
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problems pertaining to their trace elements content and their potential accumulation in
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the soil. The latter is a crucial factor leading to restricted agricultural use of compost (Arvanitoyannis et al., 2006). In turn, food byproducts have proved to be good candidates for use in the restoration of trace elements contaminated soils (Clemente,
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Paredes & Bernal, 2007b; de la Fuente, Clemente, Martínez-Alcalá, Tortosa, & Bernal,
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2011; Arvanitoyannis et al., 2006). Various studies have demonstrated that the application of agro-industrial wastes, such as sawdust and rice husk (Singh & Prasad, 2014) or sugar beet residues (Madejón et al., 2006), is able to reduce the toxicity of
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contaminated soils by reducing trace elements availability. Lindsay et al. (2011), in a 3year project, found that spent brewing grain promoted sulfate reduction and metal(oid)s
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removal, whereas municipal biosolids and conifer-derived peat were ineffective. Agricultural materials rich in cellulose have metal adsorption capacity: the cellulose,
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hemicelluloses, lignin, sugars, proteins, and starch contain a variety of functional groups that facilitate metal retention (Hashem, Abdelmonem, & Farrag, 2007). Therefore, many agri-food byproducts show potential as viable options for trace elements adsorption due to their low cost and availability in abundance. Also, the use of agroresidues for metal remediation will not only reduce the heavy metal pollution but also mitigate the biological pollution resulting from the agro-residues (Ajila et al., 2012). In the following sections of the review, a complete discussion about the use of two major agri-food vegetal byproducts (from olive oil and sugar beet production) in soil remediation is presented, together with an overview of some particular examples of other minor byproducts used with this aim.
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2.1. Olive oil industry byproducts and associated problems.
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2. The use of olive oil industry byproducts in contaminated soils remediation.
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The olive oil production industry is widely distributed around the world, and represents a significant socio-economic sector for many countries, especially in the Mediterranean region. According to the Food and Agriculture Organisation of the United Nations
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(FAOSTAT 2012/2013), 3.32 million tons of olive oil are produced annually
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worldwide, of which 72% is produced in Europe - with Spain (41.6%), Italy (17.7%) and Greece (10.5%) standing out as the highest producers - where olive (Olea europaea L.) cultivation is a centuries-old tradition. Olive oil production is not restricted to the
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Mediterranean basin, and the output of other, smaller producers without a tradition of olive cultivation is steadily increasing in Asia, Africa, and America (with 14.5%,
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11.7%, and 1.5% of worldwide production, respectively; FAOSTAT 2012/2013). The environmental problems associated with this agri-food sector are caused by: (i) the
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large amount of solid and liquid wastes and byproducts (OMWs) generated throughout the oil extraction process; (ii) their concentrated generation in a short period of the year (normally from November to February); (iii) their phytotoxic characteristics; and (iv) the lack of optimal and economical solutions to olive wastes management (Azbar et al., 2004; Alburquerque, Gonzálvez, García, & Cegarra, 2004). The olive fruit consists of the pulp (70-90% on a total weight basis), stone (9-27%), and seed (2-3%), and the two main constituents (water and oil) are mainly concentrated in the pulp and seed. Three systems are used for the industrial-scale extraction of the olive oil: the traditional press-cake system (almost obsolete), the three-phase decanter system, and the two-phase centrifugation system, the latter being currently the one most
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ACCEPTED MANUSCRIPT commonly used in Spain (Alburquerque et al., 2004; Morillo, Antizar-Ladislao, Monteoliva-Sánchez, Ramos-Cormenzana, & Russell, 2009).
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In the three-phase decanter system the oil, tissue water, and solid phase of the olive are
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separated in a continuous process. This technology implies the generation of large
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quantities of a liquid phase called olive mill wastewater (OMWW or alpechín) - which is composed of the olive tissue water plus the water added in the different steps of the process (Paredes, Cegarra, Roig, Sánchez-Monedero, & Bernal, 1999) - and of a solid
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byproduct composed of a mixture of olive pulp and olive stones, called olive cake, olive
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pomace, or orujo (Brunetti, Plaza, & Senesi, 2005; Morillo et al., 2009), from which further oil is extracted by chemical methods.
The two-phase centrifugation system, developed during the 1990s, is considered a more
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ecological system because it significantly reduces both the water and energy consumption (by >80 and >20%, respectively, relative to the three-phase system) and
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the volume of wastewater generated (Azbar et al., 2004; Alburquerque et al., 2004). In this system, water is not added during the centrifugation process (or only small
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quantities), which produces a high quality olive oil and a unique semi-solid byproduct called solid olive mill waste, olive wet husk, wet pomace, or alperujo (AL) (Borja, Rincón, & Raposo, 2006). From each ton of olives, 600 kg of solid waste and 1200 kg of OMWW are obtained with the three-phase system, whereas 800 kg of AL are generated by the two-phase method (Azbar et al., 2004).
2.2. Physico-chemical and biological characteristics of wastes from olive processing: composting and adding value. The solid and liquid olive oil industry byproducts (OMWs) are dark-colored wastes and contain high amounts of OM. The characteristics of OMWs vary widely depending on
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ACCEPTED MANUSCRIPT the oil production method, but also on several other variables, such as fruit variety, maturation degree, climatic conditions during cultivation, and storage options.
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The OMWW is a liquid violet to dark brown in color with a strong smell of olives. It is
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characterized by the presence of most of the water-soluble chemical species present in
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the olive fruit, a high load of low degradability OM (COD/BOD5 ratio between 2.5 and 5; Azbar et al., 2004), an acidic pH, and high electrical conductivity (Morillo et al., 2009; Alburquerque et al., 2004; Paredes et al., 1999). The organic fraction of OMWW
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polyphenolic compounds (up to 80 g l−1).
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contains large amounts of proteins, lipids, polysaccharides, and aromatic and
The solid waste (AL) is a dense sludge characterized by its strong odor, slightly acidic character, and elevated content of OM - mainly lignin, hemicelluloses, and cellulose,
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but also proteins, lipids, and a small proportion of phenolic compounds (Morillo et al., 2009; Alburquerque et al., 2004).
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The major problems associated with OMWs pollution are caused by their color and by the high content of phenolic compounds, which are originally synthesized by the olive
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plants and also formed during the oil extraction process (Morillo et al., 2009). Numerous studies consider that these compounds are responsible for the phytotoxicity and inhibition of microbial growth exhibited by OMWs (Hachicha et al., 2009; Pierantozzi et al., 2012), which limit their direct use in soils. However, the benefits of these wastes are linked to their elevated concentration of K, which, after degradation of the phytotoxic compounds in the soil, has been demonstrated to be beneficial for plant growth (Clemente et al., 2007b). Because of the high OM load and the richness of essential nutrients (N, K, Ca, Mg, and Fe) of OMWs, composting seems an appropriate solution for the production of organic fertilizers.
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ACCEPTED MANUSCRIPT Composting of organic wastes is a spontaneous biooxidative process in a predominantly aerobic environment that involves the mineralization and partial humification of the OM
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of the materials being composted. During the process, bacteria, fungi, and other
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microorganisms break down organic materials to produce a stable material, rich in
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humic-like organic substances and free of phytotoxicity and pathogens, called compost. The OMWs have certain properties that are inadequate for composting - such as high levels of compounds that are not easily degradable, or a dense and sticky texture in the
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case of AL- and which usually require them to be mixed with bulking materials to
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provide aerobic conditions and an adequate C/N ratio for the development of the microorganisms involved in the process. There are many examples of successful composting of these byproducts (Alburquerque, Gonzálvez, García, & Cegarra, 2006;
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Hachicha et al., 2009; Paredes, Roig, Bernal, Sánchez-Monedero, & Cegarra, 2000) and the Spanish legislation for fertilizer products even includes “Compost of Alperujo” as
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an organic amendment (Ministerio de la Presidencia, 2013). Some olive oil producers have established composting systems for their byproducts and
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the Government of Andalusia (Junta de Andalucía) - the main olive oil producing region of Spain - has promoted the composting of agricultural byproducts, mainly from oil production. As a result, the production of compost in Andalusia increased from 5000 t in 2007 to 94000 t in 2012. Most of the composts produced at an industrial level come from mixtures of alperujo with olive leaves, animal manures, or straw. The physicochemical characteristics of experimentally and industrially produced solid-OMWs (mainly AL and pomace, but also OMWW sludge) composts are shown in Table 1.
2.3. The use of fresh and composted olive mill wastes as amendments in soil remediation.
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ACCEPTED MANUSCRIPT 2.3.1. Trace elements solubility and fractionation in OMWs treated soils. Several studies have been developed to assess the influence of OMWs on trace elements
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solubility and fractionation at the laboratory (Alburquerque, de la Fuente, & Bernal,
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2011; Pardo, Clemente, & Bernal, 2011; Beesley et al., 2014), greenhouse (de la Fuente
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et al., 2011), and field (Clemente et al., 2007b, 2012; Pardo, Clemente, Epelde, Garbisu, & Bernal, 2014c; Pardo, Martínez-Fernández, Clemente, Bernal, & Walker, 2014d) scale. Different results have been reported for the different types of wastes used, mainly
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related to their application to the soil in a fresh or composted form. An example of the
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difference in behavior of fresh and composted AL was reported by de la Fuente et al. (2011), who compared the effects of composted and fresh AL and their water-soluble fractions on heavy metals availability in the soil. They observed that compost reduced
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the availability and plant uptake of metals, while fresh AL increased their bioavailability (especially Mn; Figure 1), produced phytotoxicity, and reduced plant
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growth.
The slightly acidic nature of fresh OMWs and their high concentration of soluble
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organic compounds, such as polyphenols (Azbar et al., 2004), can increase metals solubility in soils and, therefore, their leaching and availability to plants (Clemente et al., 2007b; de la Fuente, Clemente, & Bernal, 2008; Burgos, Madejón, Cabrera, & Madejón, 2010). Soluble phenols can chelate heavy metals, thereby blocking their sorption and promoting leaching through the formation of soluble metal complexes, but also can highly influence redox processes in soils (de la Fuente et al., 2008). Clemente et al. (2007b) found solubilization of Cu (increased DTPA-extractable Cu concentrations) in a calcareous soil indirectly affected by mining activities, up to a year after the application of AL to field experimental plots, which was associated with the chelation of this element by the soluble phenolic compounds of AL. In addition, in
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ACCEPTED MANUSCRIPT numerous studies, increases have observed in the solubility of Fe, Mn, and other elements associated with Fe and Mn oxides, such as Zn and Pb, after the application of
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fresh OMWs to contaminated soils, which have been attributed to changes in the soil
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redox conditions during the oxidation of phenolic compounds of these wastes (Clemente
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et al., 2007b; de la Fuente et al., 2008). Soluble monomeric phenols of OMWs are oxidized in soil, while soil Mn and Fe oxides are reduced to divalent forms: Mn2+ and Fe2+, which are generally highly soluble and easily extractable (Clemente et al., 2007b).
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Whereas Fe2+ is quickly oxidized to Fe3+, Mn2+ oxidation is quite slow, so soluble Mn
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species persist longer in the soil. This is in agreement with the results found by Piotrowska, Iamarino, Antonietta, and Gianfreda (2006), who reported an increase in the soil extractable concentrations of Fe and Mn immediately after the application of
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different rates of OMWW. Also, de la Fuente et al. (2008) observed that the addition of AL to a calcareous, metal-contaminated soil increased the soluble fraction of Mn and
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the EDTA-extractable Zn and Pb concentrations in the soil after 56 days of incubation, presumably due to the oxidation of water-soluble phenols. The mineralization process,
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in addition to changing soil redox conditions, leads to the formation of inorganic compounds (like phosphates), and in calcareous soils the CO2 produced can alter the carbonate/bicarbonate equilibrium, affecting metal precipitation (Clemente, Escolar, & Bernal, 2006b; de la Fuente et al., 2008). Contrastingly, the use of composted OMWs in contaminated soils generally reduces the solubility and availability of trace elements, mainly due to their alkaline character (that may increase soil pH) and the complexing capacity of their OM (de la Fuente et al., 2011; Pardo, Bernal, & Clemente, 2014a; Moreno-Jiménez, Clemente, Mestrot, & Meharg, 2013). The influence of the buffering capacity of these materials was reported by Fornes, García de la Fuente, Belda, and Abad (2009) and Alburquerque et al. (2011),
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ACCEPTED MANUSCRIPT who reported that the increase in soil pH after the addition of AL compost to contaminated soils reduced the concentrations of available Al and the most available
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(CaCl2-extractable) fraction of Mn and Zn in the soil, while increasing those of the least
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available ones. The main cause of the effects of composted OMWs on trace elements
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availability is thought to be their richness in humic substances able to interact with trace elements by chelation, adsorption, or retention in their exchange complex. Clemente and Bernal (2006) found that the application of humic acids isolated from a compost
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prepared with the solid fraction of OMWW to an acid, contaminated soil caused
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significant Zn and Pb immobilization, increasing the proportion of the least available fractions of these metals in the soil, as a result of the provision of exchange sites by the humic acids. But, also in this study, Cu and Fe were slightly mobilized (increased CaCl2
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and/or NaOH extractable concentrations) as a consequence of their adsorption on particulate OM or the formation of soluble chelates.
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Moreover, in contrast to fresh OMWs, the changes in redox conditions provoked by composts are normally not significant, as their OM mineralization in soil is very low
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due to the high microbial stability of these composts (Clemente et al., 2006b). Alburquerque et al. (2011) and Pardo et al. (2011) found 3.8-10% mineralization of the organic-C of AL compost after 56 days of incubation. In addition, during composting the majority of the phenolic compounds present in OMWs are degraded (Hachicha et al., 2009), although the compost could possess a small, resistant fraction of this type of compound that may provoke mobilization of trace elements when oxidized in soil. Clemente, Walker, Pardo, Martínez-Fernández, and Bernal (2012) and Pardo et al. (2014d) observed, in two field experiments, significant increases in the CaCl2- and DTPA-extractable Mn concentrations, respectively, up to one year after AL compost addition to two soils with differing characteristics (contamination degree and pH),
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ACCEPTED MANUSCRIPT which then decreased with time. However, this rise in soil Mn solubility did not result in Mn-enriched leaves and fruits of the established vegetation (Atriplex halimus and
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Bituminaria bituminosa, respectively).
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In accordance with the well-known affinity of Cu for soil OM, numerous authors have
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observed Cu mobilization in contaminated soils after OMWs compost addition (Burgos et al., 2010; Pardo et al., 2011; Moreno-Jiménez et al., 2013). However, Pardo et al. (2014a) found, in a column experiment, that complexation of Cu by the OM provided
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by AL compost reduced its leaching through the soil profile (the CaCl2-extractable
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concentrations in the deep layers were lower than those in soils treated with lime and in control soils).
Mobilization of As after the addition of OMWs compost to contaminated soils has been
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reported in several studies (Pardo et al., 2011, 2014a, 2014d; Clemente et al., 2012; Moreno-Jiménez et al., 2013; Beesley et al., 2014). The liming effect of compost can
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increase As solubility in the soil solution, and the presence of phosphates and soluble organic compounds able to compete with As for adsorption sites in the soil also
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contribute to this solubilization (Fitz & Wenzel, 2002). This could be one of the main constraints regarding the use of this type of compost in the remediation of trace elements contaminated soils, and implies that the dose applied must be carefully studied to avoid the possible solubilization of As (and/or any other toxic elements) in the soil (Clemente et al., 2012; Moreno-Jiménez et al., 2013). 2.3.2. Nutrients availability and plant establishment. Raw or composted OMWs have a high content of OM, are generally rich in K but poor in P, Ca and Mg, and usually contain a rather low amount of N (Alburquerque et al., 2004). Therefore, the addition of this type of organic waste to contaminated soils normally leads to significant increases in the total and water-soluble organic C (TOC
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ACCEPTED MANUSCRIPT and WSOC), total or dissolved N (TN), and available K in the soils, whether they are used fresh or composted (Romero, Benítez, & Nogales, 2005; Clemente, de la Fuente,
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Moral, & Bernal, 2007a, Clemente et al., 2012; Moreno-Jiménez et al., 2013; Curaqueo,
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Schoebitz, Borie, Caravaca, & Roldán, 2014). This is of great importance for
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stimulation of the nutrient cycling in these typically poor soils (Pardo, Clemente, Alvarenga, & Bernal, 2014b; Pardo et al., 2014c).
Clemente et al. (2012) and Pardo et al. (2014d) observed high concentrations of TOC,
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TN, NaHCO3-extractable P, and NaNO3-extractable K in two contaminated soils with
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different physico-chemical characteristics two years after AL compost addition to field experimental plots, reporting levels around 10-fold and 5-fold those in control soils for K and P in the amended soils, respectively. Kinetic models of C-mineralization
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developed in contaminated soils after OMWs application show that a high proportion of the easily available OM is provided by raw materials (de la Fuente et al., 2008), while
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composts are also an important source of slowly mineralisable compounds that remain longer in the soil (Pardo et al., 2011). However, slow microbial degradation of the OM
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of fresh AL has been observed by some authors (Romero et al., 2005; de la Fuente et al., 2011) and was assumed to be a consequence of the presence of phenolic compounds and/or the large amounts of highly-resistant ligno-cellulosic compounds present in this material. Nitrogen in OMWs is mainly present in organic forms, and several studies have found a modest release of available N (mainly NO3--N) in the short-term by microbial mineralization in contaminated or agricultural soils after raw or composted OMWs application (Brunetti et al., 2005; Alburquerque et al., 2006; Gómez-Muñoz, Hatch, Bol, Dixon, & García-Ruiz, 2011). However, a risk of nitrate and K leaching in soils amended with composted OMWs has been reported (Pardo et al., 2014a), indicating
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ACCEPTED MANUSCRIPT again that the doses applied must be carefully adjusted. In addition, it must be taken into account that the increase in the soil OM content by OMWs may also change the soil
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aggregation and structure (López-Piñeiro, Albarrán, Rato Nunes, & Barreto, 2008). In
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agreement with this, Mahmoud, Janssen, Haboub, Nassour, and Lennartz (2010) and
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Mahmoud, Janssen, Peth, Horn, and Lennartz (2012) reported that the regular application of OMWW for 5-15 years increased soil aggregates stability and soil hydrophobicity, and reduced the effective diffusion coefficient into aggregates and the
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drainable porosity.
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Vegetation is frequently scarce or absent in trace elements contaminated soils, which leaves them heavily exposed to erosion and leaching processes that lead to the dispersion of the contaminants. The enhancement of plants establishment and their
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improved growth and nutritional status when OMWs are added to this type of soils, due to the improvement of soil conditions, have been reported (Martínez-Fernández &
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Walker, 2012; Curaqueo et al., 2014; Pardo et al., 2014a, 2014b, 2014c, 2014d). In this regard, raw materials are generally less effective than compost (de la Fuente et al.,
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2011), and various studies have found negative effects of OMWW and AL on plant growth, this being attributed to the high concentration of phenolic compounds in these materials (Piotrowska et al., 2006; Pierantozzi et al., 2012). For example, Clemente et al. (2007b) and de la Fuente et al. (2011) observed, in field and pot experiments, inadequate growth of Beta maritima and low yields of plant biomass in AL-amended soils, due to the phytotoxicity of AL, which enhanced the solubility of metals in soil and their accumulation in leaves. Contrastingly, high percentages of plant cover of Atriplex halimus and spontaneous species were reported by Clemente et al. (2012) and Pardo et al. (2014d) in mine soils treated with AL compost, compared with unamended soils. In the latter study, a higher
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ACCEPTED MANUSCRIPT richness of native species was also observed in compost treated plots, including several species that were unique to that treatment. A significant increase in the shoot biomass of
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A. halimus in a mine waste contaminated soil due to AL compost addition was also
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observed in a greenhouse assay by Martínez-Fernández and Walker (2012); this was
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associated with improvements in the leaf N and K contents. Curaqueo et al. (2014) found a significant increase in the biomass production of Tetraclinis articulata in a mine soil with the use of AL compost in combination with arbuscular mycorrhizal fungi
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(AMF) inoculation, mainly due to the improvement of nutrients uptake. Pardo et al.
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(2014a, 2014b) concluded that reduction of trace elements mobility, together with the increase of soluble nutrients concentrations induced by the application of AL compost in a highly contaminated mine soil (Figure 2), allowed the growth of Lolium perenne
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(null in untreated soil) and improved its nutritional status (high leaf K concentration). 2.3.3. Soil biological and ecotoxicological properties.
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The soils affected by this type of contamination often have low biological activity, and therefore, limited soil functioning (Pérez de Mora, Ortega-Calvo, Cabrera, & Madejón,
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2005; Clemente et al., 2007a; Pardo et al., 2011), as high concentrations of trace elements provoke adverse effects on the development and survival of soil microorganisms (Brookes, 1995). The addition of OMWs (especially composted) has been proven to be able to stimulate soil microbial communities, reducing their stress and increasing their growth, activity and functional diversity (Romero et al., 2005; Clemente et al., 2007a, 2007b; de la Fuente et al. 2011; Pardo et al., 2014b, 2014c, 2014d). This, again, is associated with the improvement of soil conditions, the reduction of trace elements availability, and the supply of essential nutrients. Several studies have shown the stimulation of soil microorganisms after the addition of composted OMWs (Fornes et al., 2009; Clemente et al., 2012; Pardo et al., 2014b,
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ACCEPTED MANUSCRIPT 2014c, 2014d). Elevated microbial biomass C and N (BC and BN) values, hydrolase activities, and respiration rates, and the reduction of stress indicators have been reported
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by Romero et al. (2005), Clemente et al. (2006b), Alburquerque et al. (2011), and Pardo
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et al. (2011, 2014b). Fornes et al. (2009) observed that the reduction of Al availability
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and the supply of nutrients in two trace elements contaminated soils after AL compost addition allowed increases in the populations of heterotrophic bacteria, actinomycetes, fungi, and yeast, and in their activity (FDA activity) in the soil. Pardo et al. (2014c),
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showed that two and half years after AL compost addition to a mine soil, soil hydrolase
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activity (ß-glucosidase, urease, and arylsulfatase activities), the overall functional activity (FDA activity and AWCD values from Biolog EcoPlates), and the status of soil microbial communities (dehydrogenase activity, basal respiration, and the metabolic
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quotient) were significantly improved. Increases in the WSOC, WSN (water soluble N), and available-K in the soil were crucial for such improvements.
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Contrastingly, a short-term negative response of the soil microbial community after the application of fresh OMWs has been reported. Romero et al. (2005) found that the main
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soil enzymatic activities were scarcely affected four months after the incorporation of AL into a contaminated soil, whereas the soluble and AB-DTPA-extractable Pb and Zn concentrations had significantly increased. Nevertheless, a medium- to long-term positive effect of raw materials in the soil has been reported (Clemente et al., 2007a; Burgos et al., 2010). Low values of BC and BN after AL addition to a soil indirectly affected by mine activity were initially found by Clemente et al. (2007a), but these were significantly increased throughout the experiment, while values of CO2-C/BC, which indicates stress conditions for soil microorganisms, declined. This positive evolution of microbial activity was associated with the gradual degradation of toxic compounds (such as polyphenols) and the re-oxidation of Mn (II) to Mn (IV). In agreement with
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ACCEPTED MANUSCRIPT this, important decreases in the concentrations of water soluble phenolic compounds in soils treated with AL or its water extracts were observed by de la Fuente et al. (2008 and
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2011), two months after their addition to a contaminated soil. Burgos et al. (2010)
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reported significant increases of BC concentrations and of dehydrogenase and
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arylsulfatase activities in two soils amended with fresh AL after 40 weeks of incubation, compared to a non-amended control, which were related to the rise of water soluble organic C.
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Besides the reactivation of biogeochemical cycles, the application of OMWs is able to
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improve the soil habitat function and minimize the environmental impact of trace elements contaminated soils in other ecosystems (Pardo et al., 2014c). Ecotoxicological tests have been applied in soils treated with compost (Pardo et al., 2011, 2014b, 2014c)
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and, in general, they reflected a reduction in the potential ecological risks associated with the different exposure routes of these soils. Pardo et al. (2014b, 2014c)
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demonstrated, in greenhouse and field experiments, the potential of AL compost for reducing the risk of contamination of surface water and groundwater (decrease of
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toxicity for Thamnocephalus platyurus and Vibrio fischeri), while also decreasing the direct toxicity for plants (Lactuca sativa, Lepidium sativum, Lolium perenne, and Zea mays) in the short- (after 4.5 months) and long-term (2.5 years). The efficiency of AL compost with regard to minimizing the potential risks posed by trace elements was also reported by Beesley et al. (2014), who observed that reduction of their availability in a contaminated soil due to the use of AL compost, alone or in combination with biochar, produced a significant decrease of the soil toxicity towards L. sativum (seed germination) and V. fischeri (a luminescent bacterium).
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ACCEPTED MANUSCRIPT 3. Sugar beet residues as amendments for contaminated soils. 3.1. Sugar beet byproducts production and common uses.
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Sugar production from sugar beet results in about 85% (by weight) of wastes and/or
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byproducts. Leaves, weeds and beet tails are often used for the production of composts
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and thus as soil amendments. At the industrial level, about 60% of the processing material is a dry substance that is turned into the product and the remaining 40% are byproducts. The European Union, the USA, and Russia are the world‘s three largest
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sugar beet producers, generating about 177×106 million tons per year (FAOSTAT,
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2012/2013).
Valuable byproducts of sugar production are the pulp, molasses, and carbonation lime residue (sugar beet lime) (Mosen, 2007). Pulp, the material remaining after the
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extractable sugar is taken out (exhausted beet cossettes), is usually dried and sold as animal feed due to its high nutritional value. Modern sugar beet factories produce up to
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70 kg of dried pulp per ton of sliced and extracted beets (Broughton, Dalton, Jones, & Williams, 1995). Beet molasses are the runoff syrup from the final stage of
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crystallization, at which further separation is not possible with conventional equipment. They are generated in amounts from 4 to 5% of the weight of the beets, contain about 50% sugars, and are mainly used as an animal feed component (mixed cattle feed), for yeast fermentation, and in the production of pharmaceutics and alcohol (Grub‘s Up, 2009; Mosen, 2007). Carbonation lime or sugar beet lime (SL) is generated in the raw juice purification stage, resulting from the purification-flocculation of colloidal matter in the beet extract using limestone and carbon dioxide. It is frequently used on agricultural land for de-acidifying and balancing the pH of the soil (Grub‘s Up, 2009) as it is very rich in active lime. Apart from its calcium carbonate and hydroxide content (80-90%), it contains also a considerable amount of OM (0.85%) and other nutrients
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ACCEPTED MANUSCRIPT such as P, Mn, and Fe (Mose, 2007). As an example, the Amalgamated Sugar Company LLC (USA) generates approximately 250,000 Mg of SL per year in southern Idaho and
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western Oregon from the processing of sugar beet for sugar recovery (Ippolito, Strawn,
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& Scheckel, 2013).
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The first applications of SL to correct soil pH date back to the 1950s, when positive effects on soil physico-chemical properties after SL addition were first reported. Since then, this material has been widely used in agricultural acidic soils (De Sutter &
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Godsey, 2010) with very positive responses, like significantly increasing crop yields
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(González-Fernández, Ordóñez-Fernández, Mariscal-Sancho, & Espejo-Serrano, 2012). Gómez-Paccard et al. (2013) showed, in a field experiment, that SL significantly increased the soil OM content and pH and reduced exchangeable Al concentrations in
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the surface layer of agricultural acidic soils. When used in combination with biosolids, SL has been shown to be more effective than other liming agents for crop production
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(Shaheen & Tsadilas, 2013), due to its elevated P fertilizer potential. García-Zamareño et al. (2013), in experiments under field conditions using different crops, found that SL
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had a very positive effect, increasing P concentrations in the soils. López, Vidal, Blázquez, & Urbano (2001) found that SL was more effective than lime alone, not only in increasing the soil cation-exchange capacity (CEC), but also in increasing forage production and the Ca, Mg, K, Cu, and Zn concentrations in the plants. The benefits of SL application to soils also include long-term control of Aphanomyces (common rootrot disease; Ag Gold Standard, 2008). Although this product has been mainly applied to soil, SL has been also used for waste water treatment (Ippolito et al., 2013), reducing Cu concentrations in water effluents from dairy industries, which use Cu in hoof bath solutions to prevent hoof diseases.
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ACCEPTED MANUSCRIPT 3.2. Trace elements stabilization in the soil using sugar beet lime. Industrial wastes or byproducts such as SL could be used to reduce of the mobility of
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the contaminants or to increase their retention in the soil. González-Núñez, Alba, Orta,
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Vidal, and Rigol (2012a) proved the suitability of this byproduct for the remediation of
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metal-contaminated soils, which showed simultaneous increases in pH and metal sorption after SL addition.
Most of the studies that used SL as a soil amendment reported data obtained under
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laboratory or controlled greenhouse conditions - in pots or containers or in column
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experiments. Many authors coincided in affirming that this product decreased trace elements availability (Madejón et al, 2006; Madejón, Burgos, Cabrera, & Madejón 2009b; Pérez-de-Mora, Madejón, Burgos, & Cabrera, 2006c). Furthermore, in field
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studies, SL also proved to be an excellent ameliorating agent for trace elements contaminated acid soils. For example, it was used successfully in the restoration of the
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Green Corridor of the Guadiamar river (SW Spain), created after the Aznalcóllar mine spill in 1998. Different studies carried out in this area (Clemente, Walker, and Bernal,
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2005; Madejón et al., 2006; Burgos, Pérez-de-Mora, Madejón, Cabrera, & Madejón, 2008; Madejón, Madejón, Burgos, Pérez de Mora, & Cabrera, 2009a) have demonstrated, at different scales, the effectiveness of this byproduct in the remediation/stabilization of trace elements contaminated soils. In this area, SL was able to raise soil pH to neutral values and its effect was sufficient to maintain soil pH close to neutrality over a 5-year period after the last application (Madejón et al., 2010). This result demonstrated that SL was a more time-efficient alkalinizing soil amendment than other organic materials. When the alkalinity of soils is increased, insoluble trace elements may precipitate, forming insoluble compounds and complexes and secondary minerals. Therefore,
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ACCEPTED MANUSCRIPT increases in soil pH are theoretically linked to a decline in the availability of metals in the soil. However, some studies have shown a clear reduction on trace elements
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availability after SL addition to soil (Burgos et al., 2010; Pérez de Mora et al., 2006c,
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Madejón et al., 2009a), while others have reported less clear or even contrasting results
(Jiménez Moraza, Iglesias, & Palencia, 2006).
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regarding the behavior of certain elements like Zn, Mn, and Cd after SL application
Different studies on the leachability and solubility of trace elements in the soil after SL
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addition in columns (Garrido, Illera, Campbell, & García-González, 2006; Aguilar-
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Carrillo, Garrido, Barrios, & García-González, 2009), pots (González-Núñez, Alba, Orta, Vidal, & Rigol, 2012b), and semi-field experiments (Pérez-de-Mora, Burgos, Cabrera, & Madejón, 2007), have found that this amendment is able to reduce the
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concentrations of these elements in leachates and, consequently, the risk of groundwater contamination. In agreement with this, the results from a columns experiment using an
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acidic contaminated soil showed that SL decreased significantly the mobility of Cd, Cu, and Pb in the soil (Campbell, Garrido, Illera, & García-González, 2006). Similarly,
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Aguilar-Carrillo et al. (2009), in another columns experiment, found that SL reduced the potential leachability of As, Cd, and Tl through the profile of an acidic soil. In the latter study, results from a sequential extraction revealed that the amendment of contaminated soils with SL provoked redistribution of the toxic elements from soluble and exchangeable (highly available) pools to a less available fraction (associated with Al, Fe, and Mn (hydro)oxides). Direct remediation of subsoil acidity in mine wastes has been tested in different studies using SL in combination with other organic amendments. Svendson, Henry, and Brown, (2007) studied the use of different mixtures in mine tailings rehabilitation and concluded that the most successful, among different types of liming materials, in
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ACCEPTED MANUSCRIPT reducing 0.01 M Ca(NO3)2-extractable Cd and Zn concentration, was biosolid sludge in combination with SL. These authors remarked that this material is an industrial
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byproduct and therefore offers a low-cost alternative to commercial limestone. Shaheen
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et al. (2014) found similar results using mixtures of SL and sewage sludge and
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demonstrated the effectiveness of these mixtures in reducing the availability of Cu and Zn to plants. Alvarenga et al. (2008) showed that SL in combination with other organic amendments decreased the level of the mobile/effectively bioavailable fraction of Cu,
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Pb, and Zn in mining soils in Portugal. Similarly, Brown, Svendsen and Henry (2009)
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reported a reduction in the Cd and Zn concentrations extracted with Ca(NO3)2 from fluvial mine tailings following the application of mixtures of SL and biosolid, in a longterm field experiment in Arkansas River, Connecticut (USA).
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Regarding the recommended doses of SL in soils, Jiménez-Moraza et al. (2006) calculated the most adequate amount of SL for pyrite neutralization in mine tailings
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restoration. They concluded that this product was highly effective for soil alkalization and that the stoichiometric dose was enough to prevent or delay the leaching of Al, Cu,
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and Fe, whereas the release of Cd, Mn, and Zn was influenced by the mass ratio SL/sulfide sulfur of the soil. As commented on previously, SL has non-negligible nutrient and OM contents. This fact is of special interest in soil remediation procedures, as many polluted soils are also characterized by negative properties such as poor nutrient availability, lack of soil structure, low OM content, high salinity, and/or acid pH (Adriano, 2001). In this sense, SL acts as a powerful fertilizer, mainly improving K and P availability (Mosen, 2007; García-Zamareño et al., 2013; Shaheen & Tsadilas, 2013) and also increasing the OM content and soluble concentrations (Pérez de Mora et al., 2005, Gómez–Paccard et al., 2013) in the soil.
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3.3. Effects on vegetation and soil microbial populations.
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As a result of the soil condition improvement, several case studies using SL have
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shown, under different environments, enhanced natural remediation processes that have
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resulted in substantially improved vegetation growth. The beneficial effects of SL use have been demonstrated also when an induced vegetation cover was established under semi-controlled conditions (Pérez-de-Mora, Madejón, Burgos, & Cabrera, 2006b;
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Shaheen et al., 2014; De Sutter & Godsey, 2010) or when native vegetation colonized a
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contaminated soil under field conditions (Clemente et al., 2006a; Madejón et al., 2006, Pérez de Mora et al., 2011). Some authors have also shown that not only do plant cover and biomass increase in SL treated soils, but also the plant biodiversity of the area is
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greatly improved (Pérez de Mora et al., 2011). Another important finding related to the creation or restoration of a vegetation cover is a
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general reduction of trace elements translocation to the plant aerial parts (Madejón et al., 2006; Burgos et al., 2008; Pérez de Mora et al., 2011; Shaheen et al., 2014). This
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represents an important goal for phytoremediation, especially when phytostabilization strategies are chosen. A well-established vegetation cover with low concentrations of contaminants in the aerial parts maintains trace elements stabilized in the soil, reduces losses by erosion processes, and avoids transfer to the food chain (Bolan, Park, Robinson, Naidu, & Huh, 2011). Another important aspect to be considered is the microbiological status of the contaminated soils. Trace elements pollution exerts a negative impact on soil microbial activity, which greatly alters nutrient cycles (Domínguez, Marañón, Murillo, Schulin, & Robinson, 2010). The detrimental effects of trace elements on soil quality are largely related to both soil fertility status and biological activity (Gil-Sotres, Trasar-Cepeda,
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ACCEPTED MANUSCRIPT Leiros, & Seoane, 2005), hindering soil restoration programs. In this regard, as a result of the improvement of soil physico-chemical conditions and fertility and of the
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establishment of plants in the soil, the addition of SL has a very positive effect on the
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soil microbiological status. For example, in an incubation experiment, Burgos et al.
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(2010) found that, in neutral soil with low trace elements availability, although the addition of SL did not alter soil trace elements extractability, it significantly enhanced soil microbial activity. Also, Pérez de Mora et al. (2005, 2006a) reported the
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biochemical restoration of a contaminated soil treated with SL under semi-field
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conditions, by the increase of microbial biomass C, enzymatic activities and heterotrophic potential. These authors also demonstrated, by ARDRA (amplified ribosomal DNA restriction analysis), changes in the structural diversity of both the
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bacterial and fungal community under the SL treatment. Fingerprinting patterns of the 16S rDNA obtained with HinfI and of the 18S rDNA with HpaII revealed higher
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similarity percentages among samples from the same treatment with respect to samples from other inorganic and organic treatments. They concluded that the use of SL could
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improve soil chemical properties, as well as the microbial community function and structure (Pérez de Mora et al., 2006a). In agreement with this, Hinojosa, Carreira, Rodríguez-Maroto, and García-Ruíz (2008), who evaluated the recovery of enzyme activities in polluted soil after addition of SL in a pot experiment, reported that although soil recovery was still incomplete in terms of bioavailable trace elements, the recovery of certain soil enzyme activities was fully successful.
3.4. Other sugar beet byproducts used in trace elements contaminated soils. Although the main byproduct from the sugar beet industry used in trace elements contaminated soils is SL, other sugar beet wastes have been also used for this purpose.
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ACCEPTED MANUSCRIPT Alguacil et al. (2011) used fermented sugar beet pulp in a trace elements contaminated soil. These authors found that the addition of this amendment increased the population
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of AMF and decreased trace elements accumulation in the shoots of the host plants,
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concluding that this waste can be used for the remediation and/or phytostabilization of
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mine tailings. Medina, Vassilev, Barea, and Azcón (2005) studied the use of the same byproduct in a soil contaminated artificially (spiked) with Cd and found that the addition of this waste together with the colonization by AM endophytes was paramount
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for the revegetation of the contaminated soil, as both factors (mycorrhizal and organic
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sugar beet waste addition) increased plant growth and improved soil quality. Similar results were reported by Medina, Vassilev, Barea, and Azcón (2006) for a soil
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contaminated artificially with Zn.
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4. Other types of byproducts used in the remediation of trace elements contaminated soils.
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Other agri-food byproducts have been tested as soil amendments, with different results, and just a few of these have been used in contaminated and degraded land remediation. Byproducts from the wine industry are one of the most significant examples. Grapes are one of the world‘s largest fruit crops, with an annual production of more than 60 million tons - of which about 80% are used for wine making. The winery and distillery industries generate an increasing amount of solid and liquid byproducts and wastes such as grape stalk, grape pomace (marc), wine lees, exhausted grape marc, and vinasse (Rubio et al., 2013). Grape pomace is the most important solid residue remaining after juice extraction in the winemaking process and represents 20-30% of the weight of processed grapes; it includes skins, seeds, and stalks, and has an elevated moisture
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ACCEPTED MANUSCRIPT content (about 60%). Only in Europe, 4 million tons of solid wastes are produced annually by the wine sector in wine producing countries; these are mainly used as cattle
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feed or dumped in disposal sites (Grub‘s Up, 2009). There is, therefore, high interest in
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finding alternative uses for these byproducts.
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For instance, wine byproducts have been successfully composted (Bustamante, Paredes, Morales, Mayoral, & Moral, 2009) and their composts used effectively as fertilizers (Rubio et al., 2013) or in growing media (García-Martínez et al., 2009), with positive
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results regarding crop yield and quality. Even the direct application of winery
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wastewater to soil has been demonstrated to be a suitable alternative for its disposal (Mosse, Patti, Smernik, Christen, & Cavagnaro, 2012). Díaz, Madejón, López, López, and Cabrera (2002) reported that grape marc could be recycled as a soil conditioner in
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view of its organic and nutrient contents. Wine waste is characterized by the presence of natural antioxidants and can be potentially used as a soil conditioner, adsorbent for
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heavy metals, or fertilizer.
Composts derived from winery wastes have been found to be of good quality with good
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physico-chemical characteristics and high concentrations of nutrients – properties which confer on them an elevated agronomic value and make them adequate for use as soil conditioners (Arvanitoyannis et al., 2006; Bustamante, Said-Pullicino, Paredes, Cecilia, & Moral, 2010). The use of compost derived from wine byproducts in vineyards is of growing interest because of the general low levels of OM in these soils and their exposure to erosion (). Combinations of wine byproducts with other materials are frequently used to ameliorate the negative properties of single materials, such as heterogeneity, high salinity, low content of OM, low CEC or the presence of contaminants (Ingelmo, Canet, Ibañez, Pomares, & García, 1998).
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ACCEPTED MANUSCRIPT Regarding trace elements remediation, winery waste sludge has been found to be an effective adsorbent for the removal of Cr, Ni and Cu from aqueous solutions
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(Villaescusa et al., 2004). The sorption of metals by winery wastes might be attributable
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to their proteins, carbohydrates, and phenolic compounds - that have carboxyl,
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hydroxyl, sulfate, phosphate, and amino groups that can bind metal ions (Villaescusa et al., 2004).
In an incubation experiment, Karaka (2004) studied the effects of composted grape
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marc and mushroom compost on the extractability of metals (Cd, Cu, Ni, and Zn) in an
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alkaline soil. The amendments increased the OM content of the soil, slightly decreased soil pH, and caused significant decreases in the extractable (DTPA) Cd, Cu, and Ni concentrations; these decreases became greater with increasing rates of both organic
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amendments. A significant, negative correlation was found between the extractable Cu concentrations and the OM added to the soil, a consequence of the ability of Cu to form
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stable complexes with soil OM. In contrast, the DTPA-extractable Zn concentrations increased with increasing rates of both organic amendments, and this was assumed to be
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a consequence of the pH decrease. Mushroom compost is thus another by-product that potentially can be used for trace elements remediation. The most significant edible fungi cropped around the world are Agaricus bisporus and Pleurotus ostreatus - which, therefore, generate the substrate wastes with the greatest environmental implications (García-Delgado, Jiménez-Ayuso, Frutos, Gárate, & Eymar, 2013). A large amount of spent A. bisporus substrate is generated; just in Europe, more than 3.5×106 t are produced every year. The accumulation of spent substrate causes important environmental pollution in mushroom production areas due to salts and soluble organic carbon leaching during its storage (Guo, Chorover, & Fox, 2001).
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ACCEPTED MANUSCRIPT In the search for new or alternative uses for spent mushroom substrates, García-Delgado et al. (2013) studied the metal adsorption capacity of two of them (P. ostreatus and A.
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bisporus), which were originally formed of wheat straw and wheat straw in combination
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with poultry litter and grape marc, respectively. According to adsorption isotherms,
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both materials showed high Cd and Pb adsorption potential, which was related to the presence of carboxyl acidic groups that increase the CEC of these materials and induced the formation of heavy metal complexes. Spent substrate used as an A. bisporus
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inoculum carrier was described as being useful for soil bioremediation, producing the
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immobilization of heavy metals and the enhancement of PAH biodegradation (GarcíaDelgado et al., 2013). In agreement with this, Shuman (1998) reported that spent mushroom compost was able to lower Cd and Pb availability in an artificially
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contaminated (spiked) soil, by redistribution of the metals from exchangeable and/or OM related fractions to less available fractions in the soil. In the same experiment,
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cotton gin litter (short fibers and plant trash combed from the cotton before processing for spinning) caused much lower immobilization of Cd and Pb.
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Another example of a byproduct with potential use as an amendment in trace elements contaminated soils is rice husk, whose addition to a Cd contaminated soil provoked a significant reduction in the concentrations of Cd accumulated in the leaves of Amaranthus caudatus and improved plant yield to levels comparable with those obtained in non-contaminated sites (Singh & Prasad, 2014). Finally, rapeseed residue, in addition to being a source of biofuel, can be used as an amendment to improve soil fertility (Tejada, Hernández, & García, 2009). Lee et al. (2013) treated soils from a rice paddy area adjacent to a mine site in Korea, contaminated by Cd and Pb (11.3 and 1233 mg kg−1, respectively), with rapeseed residue and eggshells. This increased soil pH values and decreased significantly the
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ACCEPTED MANUSCRIPT extractable (acetic acid-TCLP and diluted HCl) Cd and Pb concentrations, that declined with increasing rates of eggshell addition (Lee et al., 2013). Eggshell has been used in
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soil remediation experiments as a lime-based waste material, as it is very readily
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obtained and contains >95% CaCO3 (Ahmad et al., 2012). It can be a good alternative to
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commercial limes, with many advantages for immobilizing heavy metals in soils (Ok et al., 2010).Application of rapeseed residue and eggshell waste was thus shown to be a cost-effective way to remediate the soil contaminated with these metals, mitigating soil
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acidity and immobilizing the metals in the soil (Lee et al., 2013).
5. Final considerations and aims for the future.
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Agricultural land is being significantly reduced (through degradation, erosion, salinization and contamination processes) or diverted to non-agricultural uses, while
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total food production is notably increasing, this generating an environmental footprint which represents a major issue that must be urgently addressed (Lal, 2013). Agri-food
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industry is one of the largest production sectors worldwide, and is thus of primary importance for the global economy. One of the main issues challenging this sector is the huge generation of byproducts and wastes of many different kinds. Only in the EU the total amount of food byproducts produced has been estimated at roughly 222 million tons per year (AWARENET, 2004) and approximately 30% of global agricultural products end up as residues and refuses (Ajila et al., 2012). In the last few years, new methods and policies for waste treatment have been introduced, and the global perception of agricultural wastes is changing rapidly in response to the need for environmental preservation and sustainability, and for global food security (Ajila et al., 2012). In addition, the efficiency of food production can be
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ACCEPTED MANUSCRIPT enhanced by reusing and recycling the byproducts generated, turning them from wastes into resources (Lal, 2013).
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In this sense, the use of food byproducts, fresh or transformed, alone or in combination
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with other (residual) materials, in the remediation of trace elements contaminated soils
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(summarized in Table 2) has been clearly shown in this review to represent an attractive alternative to conventional landfill or incineration treatments, especially when further acquisition of valuable products is not feasible (e.g., after biogas production or
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extraction of valuable chemical compounds from waste materials). Land disposal of
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organic and/or inorganic food byproducts may alter trace elements solubility and bioavailability in the soil, likely provoking their immobilization in the soil and improving plant establishment and growth by improving the nutrient status and the
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physico-chemical characteristics of the soil. This may offer a low-cost, sustainable solution for the remediation of trace elements contaminated soils and the establishment
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of a healthy and self-sustaining vegetation cover, which would protect the soil against further degradation and prevent the dispersion of the contaminants in the ecosystem
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(Clemente et al., 2012; van Herwijnen et al., 2007). Nevertheless, some aspects may have to be carefully considered in order to elude any possible negative effects in the environment; namely, the type (stability) of organic matter and the amount (dose) of byproduct added may have to be adjusted to control the effects on soil pH and redox conditions that could cause mobilization of certain elements like As or Mn. In any case, it has also been evidenced here that byproducts recycling in trace elements contaminated soils would restore or improve soil health and quality, contributing significantly to the stimulation of the growth and activity of microorganisms, as well as to the completion of the cycles of the principal nutrients, which are usually inhibited in these soils. This would also positively influence the overall ecosystem functioning and
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ACCEPTED MANUSCRIPT help to minimize the risks associated with the presence of trace elements in the soil and
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the potential environmental impact they may pose.
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Acknowledgements
The authors thank Dr. David J. Walker for the English corrections. Part of R.
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Clemente‘s salary is paid by the European Social Fund of the EU.
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ACCEPTED MANUSCRIPT Madejón, P., Burgos, P., Cabrera, F., & Madejón, E. (2009b). Phytostabilization of amended soils polluted with trace elements using the Mediterranean Shrub:
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McGrath, S. P., Chaudri, A. M., & Giller, K. E. (1995). Long-term effects of metals in sewage sludge on soils, microorganisms and plants. Journal of Industrial Microbiology, 14, 94-104. Medina, A., Vassilev, N., Barea, J. M., & Azcón, R. (2005). Application of Aspergillus niger-treated agrowaste residue and Glomus mosseae for improving growth and nutrition of Trifolium repens in a Cd-contaminated soil. Journal of Biotechnology, 116, 369–378. Medina, A., Vassilev, N., Barea, J. M., & Azcón, R. (2006). The growth-enhancement of clover by Aspergillus-treated sugar beet waste and Glomus mosseae inoculation in Zn contaminated soil. Applied Soil Ecology, 33, 87–98.
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Mirabella, N., Castellani, V., & Sala, S. (2014). Current options for the valorization of
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ACCEPTED MANUSCRIPT Pardo, T., Clemente, R., & Bernal, M. P. (2011). Effects of compost, pig slurry and lime on trace element solubility and toxicity in two soils differently affected by mining
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activities. Chemosphere, 84, 642-650.
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the phytostabilisation efficiency in a trace elements contaminated soil using soil
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health indicators. Journal of Hazardous Materials, 268, 68-76. Pardo, T., Martínez-Fernández, D., Clemente, R., Bernal, M. P., & Walker, D. J. (2014d). The use of olive-mill waste compost to promote the plant vegetation cover
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in a trace element-contaminated soil. Environmental Science and Pollution Research,
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Paredes, C., Cegarra, J., Roig, A., Sánchez-Monedero, M. A., & Bernal, M. P. (1999). Characterization of olive mill wastewater (alpechín) and its sludge for agricultural
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purpose. Bioresource Technology, 67, 111-116. Paredes, C., Roig, A., Bernal, M.P., Sánchez-Monedero, M.A., Cegarra, J. (2000). Evolution of organic matter and nitrogen during co-composting of olive mill wastewater with solid organic wastes. Biology and Fertility of Soils, 32, 222-227. Pérez de Mora, A., Ortega-Calvo, J. J., Cabrera, F., & Madejón, E. (2005). Changes in enzyme activities and microbial biomass after ―in situ‖ remediation of a heavy metalcontaminated soil. Applied Soil Ecology, 28, 125-137. Pérez-de-Mora, A., Burgos, P., Cabrera, F., & Madejón, E. (2007). ―In situ‖ amendments and revegetation reduce trace element leaching in a contaminated soil. Water, Air and Soil Pollution, 185, 209–222.
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ACCEPTED MANUSCRIPT Pérez-de-Mora, A., Burgos, P., Madejón, E, Cabrera, F., Jaeckel, P., & Schloter, M. (2006a). Microbial community structure and function in a soil contaminated by
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heavy metals: effects of plant growth and different amendments. Soil Biology &
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Biochemistry, 38, 327–341.
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Pérez-de-Mora, A., Madejón, E, Burgos, P., & Cabrera, F. (2006b). Trace element availability and plant growth in a mine-spill-contaminated soil under assisted natural remediation: II. Plants. Science of the Total Environment, 363, 38-45.
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Pérez-de-Mora, A., Madejón, E., Burgos, P., & Cabrera, F. (2006c). Trace element
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availability and plant growth in a mine-spill contaminated soil under assisted natural remediation I. Soils. Science of the Total Environment, 363, 28–37. Pérez-de-Mora, A., Madejón, P., Burgos, P., Cabrera, F., Lepp, N. W., & Madejón, E.
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(2011). Phytostabilization of semiarid soils residually contaminated with trace elements using by-products: Sustainability and risks. Environmental Pollution, 159,
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Maestri, D. (2012). Physico-chemical and toxicological assessment of liquid wastes from olive processing-related industries. Journal of the Science of Food and Agriculture, 92, 216–223. Piotrowska, A., Iamarino, G., Antonietta, M., & Gianfreda, L. (2006). Short-term effects of olive mill wastewater (OMW) on chemical and biochemical properties of a semiarid Mediterranean soil. Soil Biology & Biochemistry, 38, 600–610. Quilty, J. R., & Cattle, S.R. (2011). Use and understanding of organic amendments in Australian agriculture: a review. Soil Research, 49, 1–26.
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ACCEPTED MANUSCRIPT Romero, E., Benítez, E., & Nogales, R. (2005). Suitability of wastes from olive-oil industry for initial reclamation of a Pb/Zn mine tailing. Water, Air, & Soil Pollution,
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165, 153–165.
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Rubio, R., Perez-Murcia, M. D., Agullo, E., Bustamante, M. A., Sanchez, C., Paredes,
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C., & Moral, R. 2013. Recycling of agro-food wastes into vineyards by composting: agronomic validation in field conditions. Communications in Soil Science and Plant Analysis, 44, 502-516.
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Shaheen, S. M, Shams, M. S., Ibrahim, S. M., Elbehiry, F. A., Antoniadis, V., & Hooda,
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P. S. (2014). Stabilization of sewage sludge by using various by-products: effects on soil properties, biomass production, and bioavailability of copper and zinc. Water, Air and Soil Pollution, 225:2014.
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Shaheen, S. M., & Tsadilas C. (2013). Phosphorus sorption and availability to canola grown in an alfisol amended with various soil amendments. Communications in Soil
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Science and Plant Analysis, 44, 89–103. Shuman, L. M. (1998). Effect of organic waste amendments on cadmium and lead in
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soil fractions of two soils. Communications in Soil Science and Plant Analysis, 29, 2939-2952.
Singh, A., & Prasad, S. M. 2014. Effect of agro-industrial waste amendment on Cd uptake in Amaranthus caudatus grown under contaminated soil: An oxidative biomarker response. Ecotoxicology and Environmental Safety, 100, 105-113. Svendson, A., Henry, C., & Brown, S. (2007). Revegetation of high zinc and lead tailings with municipal biosolids and lime: greenhouse study. Journal of Environmental Quality, 36, 1609–1617. Tejada, M., Hernández, M. T., & García, C. (2009). Soil restoration using composted plant residues: effect on soil properties. Soil & Tillage Research, 102, 109–117.
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ACCEPTED MANUSCRIPT Torres, J. L., Varela, B., García, M. T., Carilla, J., Matito, C., Centelles, J.J., Cascante, M., Sort, X., & Bobet, R. (2002). Valorization of grape (Vitis vinifera) byproducts.
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Chemistry, 50, 7548-7555.
Van Dyk, J. S., Gama, R., Morrison, D., Swart, S., & Pletschke, B. I. (2013). Food processing waste: Problems, current management and prospects for utilization of the
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lignocellulose component through enzyme synergistic degradation. Renewable and
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Sustainable Energy Reviews, 26, 521–531.
van Herwijnen, R., Hutchings, T. R., Al-Tabbaa, A., Moffat, A. J., Johns, M. L., & Ouki, S. K. (2007). Remediation of metal contaminated soil with mineral-amended
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composts. Environmental Pollution, 150, 347-354. Villaescusa, I., Fiol, N., Martínez, M., Mirrales, N., Poch, J., & Seralocs, J. (2004).
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Removal of copper and nickel ions from aqueous solutions by grape stalks wastes. Water Research, 38, 992–1002.
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Walker, D. J., Clemente, R., Roig, A., & Bernal, M. P. (2003). The effects of soil amendments on heavy metal bioavailability in two contaminated Mediterranean soils. Environmental Pollution, 122, 303–312.
Figure legends
Figure 1. Effects of the addition of AL, AL-compost (CM), and their corresponding water extracts (WE-AL and WE-COM) on Mn and Zn availability (sequential
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treatment.
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Figure 2. Principal component analysis (PCA) of the physico-chemical (nutrients and
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trace elements availability) and biochemical (enzymatic activities) parameters of a contaminated soil treated with AL compost (CM) and other amendments (CT: control;
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HL: hydrated lime; PS: pig slurry) in a glasshouse column experiment.
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ACCEPTED MANUSCRIPT Table 1. Chemical composition of olive mill wastewater (OMWW, fresh weight), and alperujo (AL) and experimental and industrial solid-OMWs composts (dry weight).
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Adapted from Alburquerque et al. (2004), Morillo et al. (2009), Chowdhury, Akratos,
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Vayenas, & Pavlou (2013) and Álvarez de la Puente (2006).
Dry matter (%) pH -1
EC (dS m )
Industrial
Composts
Composts
AL
6.33-7.19
49.6-71.4
n.d.
24.8-45.3
4.2-5.2
4.9-6.8
5.4-9.5
7.8-9.1
0.98-5.2
1.2-7.3
1.7-5.3
848-976
260-900
n.d.
5.5-12.0 -1
Experimental
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OMWW
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Parameters
46.5-62.1
-1
34.2-39.8
495-539
110-580
n.d.
TN (g kg )
0.61-2.10
7.0-18.5
11-54
16-29
C/N
52.3-54.3
28.2-72.9
9-36
14-20
0.16-0.31
0.7-2.2
1-30
0.8-2.2
1.97-8.97
7.7-29.7
6-44
6.5-24
0.11-0.42
0.5-1.6
2-41
6-49
0.20-0.64
1.7-9.2
7-72
19-72
0.04-0.22
0.7-3.8
0.9-57
1.6-12.7
18.3-120
78-1462
100-4100
991-1188
1.5-6
12-29
1.4-79
14-126
1.1-12
5-39
13-131
56-199
a
39-321
b
n.d. n.d.
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Organic Matter (g kg ) TOC (g kg )
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-1
P (g kg ) -1
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K (g kg ) -1
-1
Ca (g kg ) -1
Mg (g kg ) -1
Fe (mg kg ) -1
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Cu (mg kg )
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Na (g kg )
-1
Mn (mg kg ) -1
Zn (mg kg )
2.4-12
10-37
38-138
Phenols (%)
0.98-10.7
0.5-2.4
0.1-3.8
Lipids (%)
1.64-12.2
3.7-18
0.2-8.1 b
Carbohydrates (%) 1.4-16.1 9.6-19.3 0.1-1.2 n.d. n.d.: not determined; a Alburquerque, personal communication; b water soluble fractions.
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Table 2. The main agri-food byproducts used in the remediation of trace elements contaminated soils (OMWW: olive mill wastewater; AL: solid
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olive mill waste (alperujo); SL: sugar beet lime). Byproduct
Soils
Effects
Olive oil
OMWW
Artificially contaminated soils,
Acidification of soil. Increased trace elements solubility (especially Fe and Mn) and leaching.
agricultural soils
Inhibition of plant growth.
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Source
References Madrid & Díaz-Barrientos, 1998; Mahmoud et al., 2010, 2012; Nogales et al., 1997; Piotrowska et al., 2006
Increased soil aggregates stability and soil hydrophobicity. Acid and calcareous contaminated soils
In acid soil, Pb and Zn immobilization and slight Cu and Fe mobilization. In calcareous soil, slight increase in Cu, Pb and Zn availability.
AL
Mine soils, contaminated calcareous soils
Acidification of soil. Increased Cu and Mn solubility. Immobilization of some metals (like Pb and Zn) after alteration of carbonate/bicarbonate equilibrium.
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Humic acids from OMWWcompost
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Inhibition of plant growth and phytotoxicity.
Clemente & Bernal, 2006
Burgos et al., 2010; Clemente et al., 2007a, 2007b; de la Fuente et al., 2008, 2011; Pierantozzi et al., 2012; Romero et al., 2005
AL-compost
Mine soils, soils indirectly contaminated by mine activities, and soils affected by pyritic sludge spill
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Short-term negative effects on soil microbial communities. Longterm stimulation of soil microbial biomass and activity. Slow mineralisation of OM. Increase in soil pH and content of OM and essential nutrients (N, K, P). Risk of nitrate and K leaching. Decreased trace elements solubility and leaching. Slight solubilization of As and Cu. Facilitation of plant establishment and growth, and reduction of trace elements uptake by plants. Stimulation of the development, activity and diversity of soil microbial communities. Reduction of soil direct and indirect toxicity to microorganisms, aquatic and terrestrial invertebrates, and plants.
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Alburquerque et al., 2011; Beesley et al., 2014; Clemente et al., 2006b, 2012; Curaqueo et al., 2014; de la Fuente et al., 2011; Fornes et al., 2009; MartínezFernández & Walker, 2012; MorenoJiménez et al., 2013; Pardo et al., 2011, 2014a, 2014b, 2014c, 2014d
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Mine soils, soils affected by pyritic sludge spill, agricultural acidic soils
Increase in soil pH and content of OM and essential nutrients (mainly K, P).
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SL
Decrease in trace elements availability and leaching. Solubilization of trace metals (Cd, Mn and Zn).
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Sugar beet
Increased plan establishment and reduced trace elements translocation in plants.
Mine soils, artificially contaminated soils
Decreased trace elements uptake and accumulation in plants. Increased plant growth.
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Fermented sugar beet pulp
US
Improved growth and activity of the soil microbial communities.
Aguilar-Carrillo et al., 2009; Alvarenga et al., 2008; Brown et al., 2009; Burgos et al., 2010; Clemente et al., 2005, 2006a; de Sutter & Godsey, 2010; Garrido et al., 2006; Gómez-Paccard et al., 2013; González-Nuñez et al., 2012b; Hinojosa et al., 2008; Jiménez-Moraza et al., 2006; Madejón et al., 2006, 2009a; Pérez de Mora et al., 2005, 2007, 2011; Svendson et al., 2007; Shaheen et al, 2014 Alguacil et al., 2011; Medina et al., 2005, 2006
Stimulation of arbuscular mycorrhizal fungi populations. Aqueous contaminated solutions, agricultural soils
Decreased trace elements solubility (high adsorption capacity).
Winery wastescompost
Agricultural soils
Increase in soil nutrients content. Immobilization of trace elements. Increased crop production biomass and quality
Arvanitoyannis et al., 2006; Bustamante et al., 2010; García-Martínez et al., 2009; Karaka, 2004; Rubio et al., 2013
Mushroom cultivation
Spent substrate /compost
Alkaline soils, artificially contaminated soils
Slight decrease in soil pH. Increase in OM content.
García-Delgado et al., 2013; Karaka, 2004; Shuman ,1998
Eggs
Eggshell
Increase in soil pH. Immobilization of trace elements.
Rice
Rice husk
Contaminated soils, paddy soils Contaminated soils
Rapeseed
Rapeseed waste
Contaminated soils, paddy soils
Increase in soil pH. Decreased trace elements solubility. Enhancement of soil fertility
Lee et al., 2013; Tejada et al., 2009
Brewing industry
Spent brewing grain
Contaminated soils
Sulfate reduction. Removal of TEs.
Lindsay et al., 2011
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Winery waste sludge
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Enhancement of soil fertility vs. increase in soil electrical conductivity.
Reduction in trace elements solubility (high adsorption capacity). Degradation of PAH.
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Wine industry
Reduction of trace elements uptake by plants, increased plant growth
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Díaz et al., 2002; Ingelmo et al., 1998; Villaescusa et al., 2004
Ahmad et al., 2012; Lee et al., 2013; Ok et al., 2010 Singh & Prasad, 2014
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Huge amounts of byproducts are generated every year and cause environmental
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-
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Highlights
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concern
Agro-wastes can be used as amendments in trace elements contaminated soils
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Food byproducts offer a low-cost and sustainable solution for soils remediation
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Type of byproduct and dose have to be tested to avoid negative effects
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This would help protecting the ecosystem and minimizing environmental risks
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S0963-9969(15)00145-3 doi: 10.1016/j.foodres.2015.03.040 FRIN 5756
To appear in:
Food Research International
Received date: Revised date: Accepted date:
15 December 2014 19 March 2015 24 March 2015
Please cite this article as: Clemente, R., Pardo, T., Madej´ on, P., Madej´on, E. & Pilar Bernal, M., Food byproducts as amendments in trace elements contaminated soils, Food Research International (2015), doi: 10.1016/j.foodres.2015.03.040
This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
ACCEPTED MANUSCRIPT Food byproducts as amendments in trace elements contaminated soils
Centro de Edafología y Biología Aplicada del Segura (CEBAS)-CSIC. Campus
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Rafael Clementea,*, Tania Pardoa, Paula Madejónb, Engracia Madejónb, Mª Pilar Bernala
Universitario de Espinardo, 30100 Murcia, Spain. b
Instituto de Recursos Naturales y Agrobiología de Sevilla (IRNAS)-CSIC. Av. Reina
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Mercedes 10, 41080 Seville, Spain.
*Corresponding author: R. Clemente (CEBAS-CSIC).
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Tel.: +34968396385. Fax: +34968396213.
Abstract
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E-mail: [email protected]
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Trace elements stabilization in contaminated soils is frequently the most convenient option for the remediation of large areas with moderate to high levels of contaminants. The use of organic and/or inorganic amendments is often necessary for the amelioration of the soil in order to immobilize the trace metals and metalloids and prevent their transfer to groundwater or to any living organism in the soil, and to allow the establishment of plants when the aim is the regeneration of a vegetation cover. The great upsurge of the agri-food industry in recent decades has led to the generation of huge amounts of waste materials and byproducts throughout the entire cycle of food production and processing. Some of these byproducts, which may give rise to serious environmental concerns regarding their accumulation, handling and transformation,
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ACCEPTED MANUSCRIPT have been satisfactorily used as amendments in the remediation of trace elements contaminated soils. This review offers a general outline of the use of food byproducts
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from the olive oil, sugar beet, wine, and certain other industries for the stabilization of
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trace metals and metalloids in the soil. Both the benefits of using these materials and the
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potential limitations or inconveniences that their utilization may pose, as well as some considerations regarding possible future uses and challenges, are pointed out in the text.
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Keywords: heavy metals; metalloids; soil remediation; sugar beet lime; olive mill
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waste; agri-food waste recycling.
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ACCEPTED MANUSCRIPT 1. Restoration of trace elements contaminated soils and the use of soil amendments. 1.1. Soil contamination with trace elements: sources and general remediation strategies.
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1.2. The use of organic (waste) materials as soil amendments in restoration procedures.
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1.3. Agri-food byproducts generation and use in soil remediation.
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2. The use of olive oil industry byproducts in contaminated soils remediation. 2.1. Olive oil industry byproducts and associated problems.
2.2. Physico-chemical and biological characteristics of wastes from olive processing:
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composting and adding value.
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2.3. The use of fresh and composted olive mill wastes (OMWs) as amendments in soil remediation.
2.3.1. Trace elements solubility and fractionation in OMWs treated soils.
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2.3.2. Nutrients availability and plant establishment. 2.3.3. Soil biological and ecotoxicological properties.
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3. Sugar beet residues as amendments for contaminated soils. 3.1. Sugar beet byproducts production and common uses.
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3.2. Trace elements stabilization in the soil using sugar beet lime. 3.3. Effects on vegetation and soil microbial populations. 3.4. Other sugar beet byproducts used in trace elements contaminated soils. 4. Other types of byproducts used in the remediation of trace elements contaminated soils. 5. Final considerations and aims for the future.
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ACCEPTED MANUSCRIPT 1. Restoration of trace elements contaminated soils and the use of soil amendments.
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1.1. Soil contamination with trace elements: sources and general remediation
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strategies.
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The term ―trace elements‖ is frequently used to refer to heavy metals (like Cd, Pb, and Zn) and metalloids (such as As, Sb, and Se) that can cause environmental and toxicological problems (Alloway, 2013). As the data available regarding the
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concentrations of these elements in soils and plants increase, it is gradually more
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apparent that considerable areas of soil in many parts of the world are contaminated and pose toxicity problems (Alloway, 2013) and that trace elements pollution of land is extremely widespread. Soil pollution exists where, due to human activities, a substance
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is present at a concentration above the natural (background) level and has a negative impact on some or all of the constituents of the environment (Adriano, 2001). The main
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anthropogenic sources of pollution can be divided into two main groups: extensive contamination over large areas (atmospheric, flooding and sediment deposition) and
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localized contamination (e.g., agriculture and horticulture, urban soils and industrial, mining and smelting contamination) (Alloway, 2013). The soil total trace elements concentrations considered to represent pollution vary according to the element, legislative body or country, and soil use. However, total concentrations do not necessarily indicate the toxicity of a soil and phytotoxicity can be absent at elevated total concentrations whilst, conversely, biological processes can be affected at trace elements levels below the designated maximum permissible values. An important distinction in the distribution of trace elements in soils is whether they are able or not to interact with plants (phytoavailable), although toxicity and accumulation will depend upon plant species (Robinson, Bañuelos, Conesa, Evangelou, & Schulin,
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ACCEPTED MANUSCRIPT 2009)Trace elements contamination of soils has received great attention due to the potential risks it poses to human health and because contamination of agricultural soils
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by these pollutants may threaten food safety and disturb the ecosystem (Lee et al.,
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2013). Although some trace elements, such as Cu, Fe, Mn, Ni, and Zn, perform
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biological functions, the majority of them are non-essential. In general, they cannot be degraded by microbial activity or physico-chemical processes and thus they can be considered as ―persistent‖ (Bernal, Clemente, & Walker, 2007). Direct toxicological
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effects of heavy metals on living organisms arise from displacement of other divalent
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metal ions in biomolecules, inhibition of membrane function, enzyme activities, respiration, and protein synthesis, genetic disruptions, oxidative damage, and, in the case of plants, inhibition of photosynthesis and water and nutrient ion uptake (Bernal et
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al., 2007).
The huge volumes of waste generated worldwide by the mining industry have been
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reported to be the main source of heavy metal and metalloids contamination of agricultural soils and crops, and of water, vegetation, and soil invertebrates (Hwang,
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Neculita, & Han, 2012), and therefore form a potential health risk for humans, animals, and plants (Adriano, 2001). Mine spoil can possess elevated levels of pyrites (sulfides, S2-) and the products of their oxidation and hydrolysis, including sulfuric acid (H2SO4), lead to both soil acidification and reaction with heavy metal sulfides to produce soluble metal sulfates. This process can decrease soil pH to below 4 and induce toxicity, due to heavy metals, aluminum, and the high H+ concentration (Bernal et al., 2007). In trace elements contaminated mine soils, the spatial and temporal heterogeneity of factors such as metal solubility and pH can strongly inhibit the establishment of a vegetation cover (de la Fuente et al., 2014).
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ACCEPTED MANUSCRIPT For degraded and/or contaminated soils, physico-chemical problems are aggravated by poor soil structure and a lack of nutrients and organic matter (OM), particularly for sites
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contaminated by mining or other industrial processes. These problems can be
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exacerbated by removal of the topsoil, either during the contaminating activity itself or
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as part of the remediation process (Clemente, Almela & Bernal, 2006a), since this is the most fertile soil horizon, rich in microbial activity, OM and nutrients and with a good physical structure. The lower soil horizons require major improvements in their
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physico-chemical attributes to permit the establishment of a vegetation cover. In
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addition, poor soil structure results in the formation of surface crusts, soil compaction, poor aeration, water retention, and porosity, and physical instability, leading to erosion by air or water and thus to the dissemination of the contaminants (Bernal et al., 2007).
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A multitude of remediation technologies has been developed for the cleanup of trace elements polluted soils. Traditional methods, such as excavation, thermal treatment, and
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chemical soil washing, are typically expensive and destructive. Strategies ―ex situ‖ cause a significant deterioration of soil structure and often come at a high economic
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cost, limiting their use on large contaminated areas. Techniques ―in situ‖ are less expensive and require less soil management, although their implementation can present constraints, like the difficulty of getting extracting/transforming agents in touch with the soil mass (Adriano, Wenzel, Vangronsveld, & Bolan, 2004). The financial aspect cannot be ignored when deciding which soil remediation technique should be used. The cost of each technique is different and this factor is at least as important as the others when deciding which method should be used (Dickinson, Baker, Doronila, Laidlaw, & Reeves, 2009). For example, soils can naturally reduce the mobility and bioavailability of heavy metals; this natural attenuation (natural remediation) relies upon un-enhanced natural processes that provoke the breakdown or immobilization of the pollutant, which
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ACCEPTED MANUSCRIPT is obviously a cheap remedy. Enhanced natural attenuation, which can be considered as one form of bioremediation, includes treatments, such as the addition of inorganic
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and/or organic amendments, which can accelerate the process (Bernal et al., 2007).
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Phytoremediation (the use of green plants to remove pollutants from the environment or
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render them harmless) is appropriate for sites having shallow contamination where the heavy metal levels warrant intervention but are not phytotoxic (Dickinson et al., 2009). The most relevant types of phytoremediation, regarding trace elements contaminated
Phytoextraction: accumulation of the pollutants in the harvestable parts of the
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-
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soils, are:
plants, which can then be removed from the site. -
Phytoimmobilization: reduction of the ―bioavailability‖ of the pollutant, by
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adsorption or absorption, precipitation, or chelation in the root or rhizosphere. Phytostabilization: physical stabilization of the soil, to diminish the risk of
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contaminant transfer via erosion or leaching. Control of the biological and physico-chemical properties of the soil has a fundamental
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role in phytoremediation. Modification of soil conditions will directly affect soil-plant transfer of the contaminants - for example, by changes in the chemical (pH, redox potential, and metal speciation) and microbiological (type and activity of bacteria and mycorrhizae) properties of the rhizosphere. The improvement of the soil physicochemical properties is usually needed when the aim of the remediation of trace elements contaminated soils is the re-establishment of a vegetation cover. This frequently requires the addition of soil amendments, most commonly organic materials (including recycled agricultural or industrial byproducts). But, possible negative effects associated with the use of soil amendments should be considered, with respect to the mobility of the pollutants and their possible assimilation by living organisms (Bernal et al., 2007).
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1.2. The use of organic (waste) materials as soil amendments in restoration
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procedures.
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Application of different types of residual organic materials to the soil as a source of OM
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is a common practice to improve soil properties and might be an environmentally friendly and cost-effective approach to restore extensive areas with a moderate level of contamination (Dickinson et al., 2009). The use of waste materials with this aim would
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mean an additional environmental benefit, as an alternative for their recycling and
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reutilization.
Some materials traditionally used in agriculture have been employed in different experiments on the bioremediation of trace elements contaminated soils. These
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materials improve soil fertility and increase plant productivity, and can also influence trace elements availability. Organic amendments such as composts or peat, which
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contain a high proportion of humified OM, can decrease the bioavailability of metals in soil by adsorption and by the formation of stable complexes with humic substances
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(Clemente & Bernal, 2006). Humic acids have a great capacity to retain or to bind metals, and their molecular size is usually larger than the soil pore size, resulting in low mobility and little leaching through the soil profile (Bernal et al., 2007). A central issue in the application of waste materials to land is balancing potential benefits from waste-derived soil amendments against the risks due to undesired pollutants and pathogenic organisms that may accumulate in the soil system, and which may be transferred to ground and surface water, plants, and animals, including humans. Particularly, the use of organic wastes can lead to problems pertaining to their trace elements content (Karaka, 2004). For example, the uncontrolled addition of animal manures, biosolids, and composts to agricultural soils and landfills has, in some cases,
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ACCEPTED MANUSCRIPT turned them into a source of pollutants for soils and plants, and may contribute excess N and P to the environment (Bernal et al., 2007). Considering the large amounts of low
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grade composts that are produced in an attempt to divert waste from landfill, there is an
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urgent need to find alternative uses for these products; remediation and stabilization of
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trace elements contaminated sites may provide a viable option (Farrell & Jones, 2010). Interestingly, compost quality - in terms of legislative standards (British Standards Institution [BSI], 2005) has shown no correlation with the success of plant
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establishment in soil remediation experiments (Farrell & Jones, 2010).
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Due to recent changes in legislation the main challenge facing waste managers is the development of strategies and techniques as alternatives to landfill for biodegradable waste materials. For example, sewage sludge has been widely used as a soil
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amendment, leading to metal accumulation in soils where it has been added periodically over a long time (McGrath, Chaudri, & Giller, 1995), although improved techniques in
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the treatment of waste waters and stabilization of the sludge gave rise to the so-called ‗extra quality‘ biosolids, suitable for use as soil amendments (Brown, Chaney, Angle, &
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Ryan, 1998). Similarly, municipal solid wastes are produced in extremely high amounts and their composition (specifically, the concentrations of potentially-toxic metals and organic compounds) has usually prevented their use as soil amendments. However, improved source-separation of wastes has allowed the development of high-quality amendments that can be added safely to soils (Bernal et al., 2007). Soil OM has been of particular interest in studies of trace elements retention in soils due to the tendency of transition metals to form stable complexes with organic ligands, one of the most important factors controlling the solubility and bioavailability of metals in the plant-soil system. The addition of OM to the soil can qualitatively (type of ligand) and quantitatively (surface area, etc.) modify its metal retention capacity and alters the
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ACCEPTED MANUSCRIPT soil chemical conditions (pH, redox potential) and the concentrations of chelating and complexing agents in the soil solution and in the solid phase (Bernal et al., 2007).
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Organic amendments do not decrease trace elements concentrations in soils but can
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decrease their bioavailability by shifting them to fractions of low availability associated
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with OM, carbonates, or metal oxides. The direct effect of the OM of the organic materials on trace elements availability cannot be separated from collateral effects on other soil properties. For instance, the presence of phosphates, aluminum compounds,
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and other inorganic minerals in some organic amendments is also believed to be
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responsible for the retention of metals (Brown et al., 1998). The addition of OM to mine waste-contaminated soils has been employed to aid their revegetation, by improving soil fertility and structure and by decreasing the trace
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elements ―bioavailability‖ (chemical forms which can be taken up by different soil organisms and plants; Madejón, Pérez de Mora, Felipe, Burgos, & Cabrera, 2006). For
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instance, Hwang et al. (2012) found mixtures of food waste-based compost and zeolite to be effective in preventing trace element leaching from a mine contaminated soil and
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proposed their addition as one feasible way to treat and/or prevent acid mine drainage. However, the medium- and long-term evolution of low-solubility organo-metallic compounds formed in the soils is of utmost importance, as their solubility could either decrease (favoring trace metal stabilization) or increase (acting as a ‗chemical time bomb‘) with time (Madejón et al., 2010).
1.3. Agri-food byproducts generation and use in soil remediation. The revolution in agriculture has resulted in the fast development of food processing industries all over the world. The term ―food byproducts‖ usually refers to wastes derived from the processing of raw vegetable and animal materials into foodstuffs,
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ACCEPTED MANUSCRIPT which generally consists of the extraction or separation of the nutritious part from the remains having little nutritional value or inedible components (Oreopoulou & Russ,
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2007). Their production covers the whole food life cycle: from agriculture to industrial
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manufacturing and processing and retail and household consumption (Mirabella,
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Castellani, & Sala, 2014). Although they are not classified as hazardous wastes, food industrialization has generated waste in huge quantities, causing environmental pollution (Ajila, Brar, Verma, & Prasada Rao, 2012). Every year, huge amounts of
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biomass in the form of agro byproducts are accumulated; for example, the food waste
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produced by the 27 countries of the EU in 2006 was estimated to be 89 Mt, 39% of this coming from food manufacture and processing (Mirabella et al., 2014). There is a wide variety of agri-food byproducts. Those of a vegetal origin (the objective of the present
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review) are generally organized into different categories according to the origin of the byproducts: crop waste and residues, fruits and vegetables, sugar, starch and
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confectionary industry, oil industry, grain and legumes, distilleries and breweries, food processing wastes, and energy crops and biofuel production wastes (Ajila et al., 2012).
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Crop wastes mainly include materials remaining after harvesting and processing, such as vines, peels and leaves (Ajila et al., 2012), while in the vegetal food-processing industries solid and liquid wastes are generated, including sludge from treatment of their wastewaters. For instance, the major wastes and byproducts of the sugar cane industry are bagasse, molasses, and sugar cane press mud, while the wine making industry produces grape pomace as its main byproduct – this represents an estimated 13% by weight of the grapes (Torres et al., 2002). The management of agro-industrial residues is one of the main issues in agriculture and in agro-industry and has an impact on their economy and day to day operations (Ajila et al., 2012). Agri-food waste materials and byproducts have been traditionally used as
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ACCEPTED MANUSCRIPT bedding for animals and livestock feed, or have been burned in the fields or simply disposed of in landfill and open-dumping sites. Alternative recent roles of these
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byproducts include their use as soil conditioners or green fertilizers, as sources of
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biofuels, thermoplastics, activated charcoal, or components of other composite
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materials, and as a source of food dietary fiber (Mirabella et al., 2014). The different types of utilization of byproducts from the agri-food processing industry can be mainly classified into five categories: as food/feed ingredients; as a carbon source for growing
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useful microorganisms; as fertilizer materials after composting; for energy production
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(biogas); and as a source of high-added-value products (Ajila et al., 2012). The use of these products as soil amendments or conditioners has been limited, due in part to the high application rates required to produce agronomic benefits, a lack of consistency in
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the composition of some products, poor public perception of their utility, and a lack of unbiased scientific research into the agricultural potential of these products (Quilty &
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Cattle, 2011). On some occasions, the cost of the collection, transportation, and processing of the byproducts can exceed the selling price.
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Therefore, the disposal of agri-food byproducts can often be difficult because of their: (i) biological instability and potential growth of pathogens; (ii) water content; (iii) rapid auto-oxidation; and (iv) changes due to enzymatic activities, which may cause the production of undesired odors and environmental problems (Ajila et al., 2012). In addition, most food wastes have low protein content and are therefore not ideal for animal feed, while their high lignin content (as in the cases of olive waste and sugarcane bagasse) also limits their utilization as animal feed since it makes the waste difficult to digest (Laufenberg, Kunz, & Nystroem, 2003). A large part of these wastes, the lignocellulose, remains largely underused and could serve as a feedstock for bioenergy production (bioethanol, biogas), although certain food processing wastes can also
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ACCEPTED MANUSCRIPT contain compounds that may be inhibitory to the fermentation process and this has to be taken into account (Van Dyk, Gama, Morrison, Swart, & Pletschke, 2013). Value-added
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products (e.g., fiber, sugars, enzymes, biofuels or food additives) are also often
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extracted from these materials, but the majority of the waste is currently unused and
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discarded, such as citrus, apple, grape, and sugar beet wastes (Van Dyk et al., 2013). The large amount of waste produced by the food industry, in addition to representing a serious economic and environmental problem, also means a loss of potentially valuable
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materials (Mirabella et al., 2014).Therefore, possible alternative uses should be investigated. The efficient utilization of these byproducts would also contribute to the
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―zero waste economy‖.
A clear example is the fact that agri-food waste materials can be utilized as soil
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conditioners or fertilizers (Arvanitoyannis, Ladas, & Mavromatis, 2006). This can be done through spreading of the untreated food waste on the soil, thereby increasing the
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organic content and microbial biomass of the soil. Tomato waste and olive husks have been utilized in this manner (Van Dyk et al., 2013). In some cases, transformation by
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composting can be undertaken prior to utilization as a soil amendment. Composted fruit and vegetable wastes can provide significant amounts of available K, Ca, and Mg and some P for crop nutrition. Also, although fruit and vegetable wastes are usually acidic, after composting the materials become neutral to alkaline (Grub‘s Up, 2009), facilitating their application to soil. For example, in South Korea, the use of food wastebased compost as organic amendment would be of interest due to the recent prohibition of food waste disposal in landfills, as well as the banning of food wastewater dumping into the sea from 2013 by the London Convention Protocol (Kim & Kim, 2010). However, food waste is considered unsuitable for agricultural purposes because of its high salt content (Hwang et al., 2012). Therefore, alternative markets for food compost,
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The addition of agricultural wastes to soil, despite its potential benefits, can lead to
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problems pertaining to their trace elements content and their potential accumulation in
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the soil. The latter is a crucial factor leading to restricted agricultural use of compost (Arvanitoyannis et al., 2006). In turn, food byproducts have proved to be good candidates for use in the restoration of trace elements contaminated soils (Clemente,
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Paredes & Bernal, 2007b; de la Fuente, Clemente, Martínez-Alcalá, Tortosa, & Bernal,
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2011; Arvanitoyannis et al., 2006). Various studies have demonstrated that the application of agro-industrial wastes, such as sawdust and rice husk (Singh & Prasad, 2014) or sugar beet residues (Madejón et al., 2006), is able to reduce the toxicity of
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contaminated soils by reducing trace elements availability. Lindsay et al. (2011), in a 3year project, found that spent brewing grain promoted sulfate reduction and metal(oid)s
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removal, whereas municipal biosolids and conifer-derived peat were ineffective. Agricultural materials rich in cellulose have metal adsorption capacity: the cellulose,
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hemicelluloses, lignin, sugars, proteins, and starch contain a variety of functional groups that facilitate metal retention (Hashem, Abdelmonem, & Farrag, 2007). Therefore, many agri-food byproducts show potential as viable options for trace elements adsorption due to their low cost and availability in abundance. Also, the use of agroresidues for metal remediation will not only reduce the heavy metal pollution but also mitigate the biological pollution resulting from the agro-residues (Ajila et al., 2012). In the following sections of the review, a complete discussion about the use of two major agri-food vegetal byproducts (from olive oil and sugar beet production) in soil remediation is presented, together with an overview of some particular examples of other minor byproducts used with this aim.
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2.1. Olive oil industry byproducts and associated problems.
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2. The use of olive oil industry byproducts in contaminated soils remediation.
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The olive oil production industry is widely distributed around the world, and represents a significant socio-economic sector for many countries, especially in the Mediterranean region. According to the Food and Agriculture Organisation of the United Nations
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(FAOSTAT 2012/2013), 3.32 million tons of olive oil are produced annually
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worldwide, of which 72% is produced in Europe - with Spain (41.6%), Italy (17.7%) and Greece (10.5%) standing out as the highest producers - where olive (Olea europaea L.) cultivation is a centuries-old tradition. Olive oil production is not restricted to the
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Mediterranean basin, and the output of other, smaller producers without a tradition of olive cultivation is steadily increasing in Asia, Africa, and America (with 14.5%,
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11.7%, and 1.5% of worldwide production, respectively; FAOSTAT 2012/2013). The environmental problems associated with this agri-food sector are caused by: (i) the
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large amount of solid and liquid wastes and byproducts (OMWs) generated throughout the oil extraction process; (ii) their concentrated generation in a short period of the year (normally from November to February); (iii) their phytotoxic characteristics; and (iv) the lack of optimal and economical solutions to olive wastes management (Azbar et al., 2004; Alburquerque, Gonzálvez, García, & Cegarra, 2004). The olive fruit consists of the pulp (70-90% on a total weight basis), stone (9-27%), and seed (2-3%), and the two main constituents (water and oil) are mainly concentrated in the pulp and seed. Three systems are used for the industrial-scale extraction of the olive oil: the traditional press-cake system (almost obsolete), the three-phase decanter system, and the two-phase centrifugation system, the latter being currently the one most
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In the three-phase decanter system the oil, tissue water, and solid phase of the olive are
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separated in a continuous process. This technology implies the generation of large
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quantities of a liquid phase called olive mill wastewater (OMWW or alpechín) - which is composed of the olive tissue water plus the water added in the different steps of the process (Paredes, Cegarra, Roig, Sánchez-Monedero, & Bernal, 1999) - and of a solid
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byproduct composed of a mixture of olive pulp and olive stones, called olive cake, olive
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pomace, or orujo (Brunetti, Plaza, & Senesi, 2005; Morillo et al., 2009), from which further oil is extracted by chemical methods.
The two-phase centrifugation system, developed during the 1990s, is considered a more
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ecological system because it significantly reduces both the water and energy consumption (by >80 and >20%, respectively, relative to the three-phase system) and
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the volume of wastewater generated (Azbar et al., 2004; Alburquerque et al., 2004). In this system, water is not added during the centrifugation process (or only small
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quantities), which produces a high quality olive oil and a unique semi-solid byproduct called solid olive mill waste, olive wet husk, wet pomace, or alperujo (AL) (Borja, Rincón, & Raposo, 2006). From each ton of olives, 600 kg of solid waste and 1200 kg of OMWW are obtained with the three-phase system, whereas 800 kg of AL are generated by the two-phase method (Azbar et al., 2004).
2.2. Physico-chemical and biological characteristics of wastes from olive processing: composting and adding value. The solid and liquid olive oil industry byproducts (OMWs) are dark-colored wastes and contain high amounts of OM. The characteristics of OMWs vary widely depending on
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ACCEPTED MANUSCRIPT the oil production method, but also on several other variables, such as fruit variety, maturation degree, climatic conditions during cultivation, and storage options.
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The OMWW is a liquid violet to dark brown in color with a strong smell of olives. It is
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characterized by the presence of most of the water-soluble chemical species present in
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the olive fruit, a high load of low degradability OM (COD/BOD5 ratio between 2.5 and 5; Azbar et al., 2004), an acidic pH, and high electrical conductivity (Morillo et al., 2009; Alburquerque et al., 2004; Paredes et al., 1999). The organic fraction of OMWW
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polyphenolic compounds (up to 80 g l−1).
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contains large amounts of proteins, lipids, polysaccharides, and aromatic and
The solid waste (AL) is a dense sludge characterized by its strong odor, slightly acidic character, and elevated content of OM - mainly lignin, hemicelluloses, and cellulose,
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but also proteins, lipids, and a small proportion of phenolic compounds (Morillo et al., 2009; Alburquerque et al., 2004).
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The major problems associated with OMWs pollution are caused by their color and by the high content of phenolic compounds, which are originally synthesized by the olive
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plants and also formed during the oil extraction process (Morillo et al., 2009). Numerous studies consider that these compounds are responsible for the phytotoxicity and inhibition of microbial growth exhibited by OMWs (Hachicha et al., 2009; Pierantozzi et al., 2012), which limit their direct use in soils. However, the benefits of these wastes are linked to their elevated concentration of K, which, after degradation of the phytotoxic compounds in the soil, has been demonstrated to be beneficial for plant growth (Clemente et al., 2007b). Because of the high OM load and the richness of essential nutrients (N, K, Ca, Mg, and Fe) of OMWs, composting seems an appropriate solution for the production of organic fertilizers.
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ACCEPTED MANUSCRIPT Composting of organic wastes is a spontaneous biooxidative process in a predominantly aerobic environment that involves the mineralization and partial humification of the OM
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of the materials being composted. During the process, bacteria, fungi, and other
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microorganisms break down organic materials to produce a stable material, rich in
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humic-like organic substances and free of phytotoxicity and pathogens, called compost. The OMWs have certain properties that are inadequate for composting - such as high levels of compounds that are not easily degradable, or a dense and sticky texture in the
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case of AL- and which usually require them to be mixed with bulking materials to
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provide aerobic conditions and an adequate C/N ratio for the development of the microorganisms involved in the process. There are many examples of successful composting of these byproducts (Alburquerque, Gonzálvez, García, & Cegarra, 2006;
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Hachicha et al., 2009; Paredes, Roig, Bernal, Sánchez-Monedero, & Cegarra, 2000) and the Spanish legislation for fertilizer products even includes “Compost of Alperujo” as
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an organic amendment (Ministerio de la Presidencia, 2013). Some olive oil producers have established composting systems for their byproducts and
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the Government of Andalusia (Junta de Andalucía) - the main olive oil producing region of Spain - has promoted the composting of agricultural byproducts, mainly from oil production. As a result, the production of compost in Andalusia increased from 5000 t in 2007 to 94000 t in 2012. Most of the composts produced at an industrial level come from mixtures of alperujo with olive leaves, animal manures, or straw. The physicochemical characteristics of experimentally and industrially produced solid-OMWs (mainly AL and pomace, but also OMWW sludge) composts are shown in Table 1.
2.3. The use of fresh and composted olive mill wastes as amendments in soil remediation.
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ACCEPTED MANUSCRIPT 2.3.1. Trace elements solubility and fractionation in OMWs treated soils. Several studies have been developed to assess the influence of OMWs on trace elements
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solubility and fractionation at the laboratory (Alburquerque, de la Fuente, & Bernal,
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2011; Pardo, Clemente, & Bernal, 2011; Beesley et al., 2014), greenhouse (de la Fuente
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et al., 2011), and field (Clemente et al., 2007b, 2012; Pardo, Clemente, Epelde, Garbisu, & Bernal, 2014c; Pardo, Martínez-Fernández, Clemente, Bernal, & Walker, 2014d) scale. Different results have been reported for the different types of wastes used, mainly
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related to their application to the soil in a fresh or composted form. An example of the
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difference in behavior of fresh and composted AL was reported by de la Fuente et al. (2011), who compared the effects of composted and fresh AL and their water-soluble fractions on heavy metals availability in the soil. They observed that compost reduced
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the availability and plant uptake of metals, while fresh AL increased their bioavailability (especially Mn; Figure 1), produced phytotoxicity, and reduced plant
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growth.
The slightly acidic nature of fresh OMWs and their high concentration of soluble
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organic compounds, such as polyphenols (Azbar et al., 2004), can increase metals solubility in soils and, therefore, their leaching and availability to plants (Clemente et al., 2007b; de la Fuente, Clemente, & Bernal, 2008; Burgos, Madejón, Cabrera, & Madejón, 2010). Soluble phenols can chelate heavy metals, thereby blocking their sorption and promoting leaching through the formation of soluble metal complexes, but also can highly influence redox processes in soils (de la Fuente et al., 2008). Clemente et al. (2007b) found solubilization of Cu (increased DTPA-extractable Cu concentrations) in a calcareous soil indirectly affected by mining activities, up to a year after the application of AL to field experimental plots, which was associated with the chelation of this element by the soluble phenolic compounds of AL. In addition, in
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fresh OMWs to contaminated soils, which have been attributed to changes in the soil
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redox conditions during the oxidation of phenolic compounds of these wastes (Clemente
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et al., 2007b; de la Fuente et al., 2008). Soluble monomeric phenols of OMWs are oxidized in soil, while soil Mn and Fe oxides are reduced to divalent forms: Mn2+ and Fe2+, which are generally highly soluble and easily extractable (Clemente et al., 2007b).
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Whereas Fe2+ is quickly oxidized to Fe3+, Mn2+ oxidation is quite slow, so soluble Mn
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species persist longer in the soil. This is in agreement with the results found by Piotrowska, Iamarino, Antonietta, and Gianfreda (2006), who reported an increase in the soil extractable concentrations of Fe and Mn immediately after the application of
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different rates of OMWW. Also, de la Fuente et al. (2008) observed that the addition of AL to a calcareous, metal-contaminated soil increased the soluble fraction of Mn and
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the EDTA-extractable Zn and Pb concentrations in the soil after 56 days of incubation, presumably due to the oxidation of water-soluble phenols. The mineralization process,
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in addition to changing soil redox conditions, leads to the formation of inorganic compounds (like phosphates), and in calcareous soils the CO2 produced can alter the carbonate/bicarbonate equilibrium, affecting metal precipitation (Clemente, Escolar, & Bernal, 2006b; de la Fuente et al., 2008). Contrastingly, the use of composted OMWs in contaminated soils generally reduces the solubility and availability of trace elements, mainly due to their alkaline character (that may increase soil pH) and the complexing capacity of their OM (de la Fuente et al., 2011; Pardo, Bernal, & Clemente, 2014a; Moreno-Jiménez, Clemente, Mestrot, & Meharg, 2013). The influence of the buffering capacity of these materials was reported by Fornes, García de la Fuente, Belda, and Abad (2009) and Alburquerque et al. (2011),
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ACCEPTED MANUSCRIPT who reported that the increase in soil pH after the addition of AL compost to contaminated soils reduced the concentrations of available Al and the most available
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(CaCl2-extractable) fraction of Mn and Zn in the soil, while increasing those of the least
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available ones. The main cause of the effects of composted OMWs on trace elements
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availability is thought to be their richness in humic substances able to interact with trace elements by chelation, adsorption, or retention in their exchange complex. Clemente and Bernal (2006) found that the application of humic acids isolated from a compost
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prepared with the solid fraction of OMWW to an acid, contaminated soil caused
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significant Zn and Pb immobilization, increasing the proportion of the least available fractions of these metals in the soil, as a result of the provision of exchange sites by the humic acids. But, also in this study, Cu and Fe were slightly mobilized (increased CaCl2
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and/or NaOH extractable concentrations) as a consequence of their adsorption on particulate OM or the formation of soluble chelates.
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Moreover, in contrast to fresh OMWs, the changes in redox conditions provoked by composts are normally not significant, as their OM mineralization in soil is very low
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due to the high microbial stability of these composts (Clemente et al., 2006b). Alburquerque et al. (2011) and Pardo et al. (2011) found 3.8-10% mineralization of the organic-C of AL compost after 56 days of incubation. In addition, during composting the majority of the phenolic compounds present in OMWs are degraded (Hachicha et al., 2009), although the compost could possess a small, resistant fraction of this type of compound that may provoke mobilization of trace elements when oxidized in soil. Clemente, Walker, Pardo, Martínez-Fernández, and Bernal (2012) and Pardo et al. (2014d) observed, in two field experiments, significant increases in the CaCl2- and DTPA-extractable Mn concentrations, respectively, up to one year after AL compost addition to two soils with differing characteristics (contamination degree and pH),
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ACCEPTED MANUSCRIPT which then decreased with time. However, this rise in soil Mn solubility did not result in Mn-enriched leaves and fruits of the established vegetation (Atriplex halimus and
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Bituminaria bituminosa, respectively).
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In accordance with the well-known affinity of Cu for soil OM, numerous authors have
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observed Cu mobilization in contaminated soils after OMWs compost addition (Burgos et al., 2010; Pardo et al., 2011; Moreno-Jiménez et al., 2013). However, Pardo et al. (2014a) found, in a column experiment, that complexation of Cu by the OM provided
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by AL compost reduced its leaching through the soil profile (the CaCl2-extractable
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concentrations in the deep layers were lower than those in soils treated with lime and in control soils).
Mobilization of As after the addition of OMWs compost to contaminated soils has been
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reported in several studies (Pardo et al., 2011, 2014a, 2014d; Clemente et al., 2012; Moreno-Jiménez et al., 2013; Beesley et al., 2014). The liming effect of compost can
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increase As solubility in the soil solution, and the presence of phosphates and soluble organic compounds able to compete with As for adsorption sites in the soil also
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contribute to this solubilization (Fitz & Wenzel, 2002). This could be one of the main constraints regarding the use of this type of compost in the remediation of trace elements contaminated soils, and implies that the dose applied must be carefully studied to avoid the possible solubilization of As (and/or any other toxic elements) in the soil (Clemente et al., 2012; Moreno-Jiménez et al., 2013). 2.3.2. Nutrients availability and plant establishment. Raw or composted OMWs have a high content of OM, are generally rich in K but poor in P, Ca and Mg, and usually contain a rather low amount of N (Alburquerque et al., 2004). Therefore, the addition of this type of organic waste to contaminated soils normally leads to significant increases in the total and water-soluble organic C (TOC
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ACCEPTED MANUSCRIPT and WSOC), total or dissolved N (TN), and available K in the soils, whether they are used fresh or composted (Romero, Benítez, & Nogales, 2005; Clemente, de la Fuente,
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Moral, & Bernal, 2007a, Clemente et al., 2012; Moreno-Jiménez et al., 2013; Curaqueo,
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Schoebitz, Borie, Caravaca, & Roldán, 2014). This is of great importance for
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stimulation of the nutrient cycling in these typically poor soils (Pardo, Clemente, Alvarenga, & Bernal, 2014b; Pardo et al., 2014c).
Clemente et al. (2012) and Pardo et al. (2014d) observed high concentrations of TOC,
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TN, NaHCO3-extractable P, and NaNO3-extractable K in two contaminated soils with
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different physico-chemical characteristics two years after AL compost addition to field experimental plots, reporting levels around 10-fold and 5-fold those in control soils for K and P in the amended soils, respectively. Kinetic models of C-mineralization
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developed in contaminated soils after OMWs application show that a high proportion of the easily available OM is provided by raw materials (de la Fuente et al., 2008), while
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composts are also an important source of slowly mineralisable compounds that remain longer in the soil (Pardo et al., 2011). However, slow microbial degradation of the OM
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of fresh AL has been observed by some authors (Romero et al., 2005; de la Fuente et al., 2011) and was assumed to be a consequence of the presence of phenolic compounds and/or the large amounts of highly-resistant ligno-cellulosic compounds present in this material. Nitrogen in OMWs is mainly present in organic forms, and several studies have found a modest release of available N (mainly NO3--N) in the short-term by microbial mineralization in contaminated or agricultural soils after raw or composted OMWs application (Brunetti et al., 2005; Alburquerque et al., 2006; Gómez-Muñoz, Hatch, Bol, Dixon, & García-Ruiz, 2011). However, a risk of nitrate and K leaching in soils amended with composted OMWs has been reported (Pardo et al., 2014a), indicating
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ACCEPTED MANUSCRIPT again that the doses applied must be carefully adjusted. In addition, it must be taken into account that the increase in the soil OM content by OMWs may also change the soil
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aggregation and structure (López-Piñeiro, Albarrán, Rato Nunes, & Barreto, 2008). In
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agreement with this, Mahmoud, Janssen, Haboub, Nassour, and Lennartz (2010) and
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Mahmoud, Janssen, Peth, Horn, and Lennartz (2012) reported that the regular application of OMWW for 5-15 years increased soil aggregates stability and soil hydrophobicity, and reduced the effective diffusion coefficient into aggregates and the
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drainable porosity.
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Vegetation is frequently scarce or absent in trace elements contaminated soils, which leaves them heavily exposed to erosion and leaching processes that lead to the dispersion of the contaminants. The enhancement of plants establishment and their
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improved growth and nutritional status when OMWs are added to this type of soils, due to the improvement of soil conditions, have been reported (Martínez-Fernández &
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Walker, 2012; Curaqueo et al., 2014; Pardo et al., 2014a, 2014b, 2014c, 2014d). In this regard, raw materials are generally less effective than compost (de la Fuente et al.,
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2011), and various studies have found negative effects of OMWW and AL on plant growth, this being attributed to the high concentration of phenolic compounds in these materials (Piotrowska et al., 2006; Pierantozzi et al., 2012). For example, Clemente et al. (2007b) and de la Fuente et al. (2011) observed, in field and pot experiments, inadequate growth of Beta maritima and low yields of plant biomass in AL-amended soils, due to the phytotoxicity of AL, which enhanced the solubility of metals in soil and their accumulation in leaves. Contrastingly, high percentages of plant cover of Atriplex halimus and spontaneous species were reported by Clemente et al. (2012) and Pardo et al. (2014d) in mine soils treated with AL compost, compared with unamended soils. In the latter study, a higher
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ACCEPTED MANUSCRIPT richness of native species was also observed in compost treated plots, including several species that were unique to that treatment. A significant increase in the shoot biomass of
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A. halimus in a mine waste contaminated soil due to AL compost addition was also
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observed in a greenhouse assay by Martínez-Fernández and Walker (2012); this was
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associated with improvements in the leaf N and K contents. Curaqueo et al. (2014) found a significant increase in the biomass production of Tetraclinis articulata in a mine soil with the use of AL compost in combination with arbuscular mycorrhizal fungi
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(AMF) inoculation, mainly due to the improvement of nutrients uptake. Pardo et al.
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(2014a, 2014b) concluded that reduction of trace elements mobility, together with the increase of soluble nutrients concentrations induced by the application of AL compost in a highly contaminated mine soil (Figure 2), allowed the growth of Lolium perenne
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(null in untreated soil) and improved its nutritional status (high leaf K concentration). 2.3.3. Soil biological and ecotoxicological properties.
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The soils affected by this type of contamination often have low biological activity, and therefore, limited soil functioning (Pérez de Mora, Ortega-Calvo, Cabrera, & Madejón,
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2005; Clemente et al., 2007a; Pardo et al., 2011), as high concentrations of trace elements provoke adverse effects on the development and survival of soil microorganisms (Brookes, 1995). The addition of OMWs (especially composted) has been proven to be able to stimulate soil microbial communities, reducing their stress and increasing their growth, activity and functional diversity (Romero et al., 2005; Clemente et al., 2007a, 2007b; de la Fuente et al. 2011; Pardo et al., 2014b, 2014c, 2014d). This, again, is associated with the improvement of soil conditions, the reduction of trace elements availability, and the supply of essential nutrients. Several studies have shown the stimulation of soil microorganisms after the addition of composted OMWs (Fornes et al., 2009; Clemente et al., 2012; Pardo et al., 2014b,
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ACCEPTED MANUSCRIPT 2014c, 2014d). Elevated microbial biomass C and N (BC and BN) values, hydrolase activities, and respiration rates, and the reduction of stress indicators have been reported
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by Romero et al. (2005), Clemente et al. (2006b), Alburquerque et al. (2011), and Pardo
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et al. (2011, 2014b). Fornes et al. (2009) observed that the reduction of Al availability
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and the supply of nutrients in two trace elements contaminated soils after AL compost addition allowed increases in the populations of heterotrophic bacteria, actinomycetes, fungi, and yeast, and in their activity (FDA activity) in the soil. Pardo et al. (2014c),
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showed that two and half years after AL compost addition to a mine soil, soil hydrolase
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activity (ß-glucosidase, urease, and arylsulfatase activities), the overall functional activity (FDA activity and AWCD values from Biolog EcoPlates), and the status of soil microbial communities (dehydrogenase activity, basal respiration, and the metabolic
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quotient) were significantly improved. Increases in the WSOC, WSN (water soluble N), and available-K in the soil were crucial for such improvements.
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Contrastingly, a short-term negative response of the soil microbial community after the application of fresh OMWs has been reported. Romero et al. (2005) found that the main
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soil enzymatic activities were scarcely affected four months after the incorporation of AL into a contaminated soil, whereas the soluble and AB-DTPA-extractable Pb and Zn concentrations had significantly increased. Nevertheless, a medium- to long-term positive effect of raw materials in the soil has been reported (Clemente et al., 2007a; Burgos et al., 2010). Low values of BC and BN after AL addition to a soil indirectly affected by mine activity were initially found by Clemente et al. (2007a), but these were significantly increased throughout the experiment, while values of CO2-C/BC, which indicates stress conditions for soil microorganisms, declined. This positive evolution of microbial activity was associated with the gradual degradation of toxic compounds (such as polyphenols) and the re-oxidation of Mn (II) to Mn (IV). In agreement with
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ACCEPTED MANUSCRIPT this, important decreases in the concentrations of water soluble phenolic compounds in soils treated with AL or its water extracts were observed by de la Fuente et al. (2008 and
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2011), two months after their addition to a contaminated soil. Burgos et al. (2010)
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reported significant increases of BC concentrations and of dehydrogenase and
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arylsulfatase activities in two soils amended with fresh AL after 40 weeks of incubation, compared to a non-amended control, which were related to the rise of water soluble organic C.
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Besides the reactivation of biogeochemical cycles, the application of OMWs is able to
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improve the soil habitat function and minimize the environmental impact of trace elements contaminated soils in other ecosystems (Pardo et al., 2014c). Ecotoxicological tests have been applied in soils treated with compost (Pardo et al., 2011, 2014b, 2014c)
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and, in general, they reflected a reduction in the potential ecological risks associated with the different exposure routes of these soils. Pardo et al. (2014b, 2014c)
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demonstrated, in greenhouse and field experiments, the potential of AL compost for reducing the risk of contamination of surface water and groundwater (decrease of
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toxicity for Thamnocephalus platyurus and Vibrio fischeri), while also decreasing the direct toxicity for plants (Lactuca sativa, Lepidium sativum, Lolium perenne, and Zea mays) in the short- (after 4.5 months) and long-term (2.5 years). The efficiency of AL compost with regard to minimizing the potential risks posed by trace elements was also reported by Beesley et al. (2014), who observed that reduction of their availability in a contaminated soil due to the use of AL compost, alone or in combination with biochar, produced a significant decrease of the soil toxicity towards L. sativum (seed germination) and V. fischeri (a luminescent bacterium).
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ACCEPTED MANUSCRIPT 3. Sugar beet residues as amendments for contaminated soils. 3.1. Sugar beet byproducts production and common uses.
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Sugar production from sugar beet results in about 85% (by weight) of wastes and/or
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byproducts. Leaves, weeds and beet tails are often used for the production of composts
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and thus as soil amendments. At the industrial level, about 60% of the processing material is a dry substance that is turned into the product and the remaining 40% are byproducts. The European Union, the USA, and Russia are the world‘s three largest
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sugar beet producers, generating about 177×106 million tons per year (FAOSTAT,
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2012/2013).
Valuable byproducts of sugar production are the pulp, molasses, and carbonation lime residue (sugar beet lime) (Mosen, 2007). Pulp, the material remaining after the
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extractable sugar is taken out (exhausted beet cossettes), is usually dried and sold as animal feed due to its high nutritional value. Modern sugar beet factories produce up to
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70 kg of dried pulp per ton of sliced and extracted beets (Broughton, Dalton, Jones, & Williams, 1995). Beet molasses are the runoff syrup from the final stage of
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crystallization, at which further separation is not possible with conventional equipment. They are generated in amounts from 4 to 5% of the weight of the beets, contain about 50% sugars, and are mainly used as an animal feed component (mixed cattle feed), for yeast fermentation, and in the production of pharmaceutics and alcohol (Grub‘s Up, 2009; Mosen, 2007). Carbonation lime or sugar beet lime (SL) is generated in the raw juice purification stage, resulting from the purification-flocculation of colloidal matter in the beet extract using limestone and carbon dioxide. It is frequently used on agricultural land for de-acidifying and balancing the pH of the soil (Grub‘s Up, 2009) as it is very rich in active lime. Apart from its calcium carbonate and hydroxide content (80-90%), it contains also a considerable amount of OM (0.85%) and other nutrients
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ACCEPTED MANUSCRIPT such as P, Mn, and Fe (Mose, 2007). As an example, the Amalgamated Sugar Company LLC (USA) generates approximately 250,000 Mg of SL per year in southern Idaho and
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western Oregon from the processing of sugar beet for sugar recovery (Ippolito, Strawn,
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& Scheckel, 2013).
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The first applications of SL to correct soil pH date back to the 1950s, when positive effects on soil physico-chemical properties after SL addition were first reported. Since then, this material has been widely used in agricultural acidic soils (De Sutter &
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Godsey, 2010) with very positive responses, like significantly increasing crop yields
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(González-Fernández, Ordóñez-Fernández, Mariscal-Sancho, & Espejo-Serrano, 2012). Gómez-Paccard et al. (2013) showed, in a field experiment, that SL significantly increased the soil OM content and pH and reduced exchangeable Al concentrations in
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the surface layer of agricultural acidic soils. When used in combination with biosolids, SL has been shown to be more effective than other liming agents for crop production
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(Shaheen & Tsadilas, 2013), due to its elevated P fertilizer potential. García-Zamareño et al. (2013), in experiments under field conditions using different crops, found that SL
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had a very positive effect, increasing P concentrations in the soils. López, Vidal, Blázquez, & Urbano (2001) found that SL was more effective than lime alone, not only in increasing the soil cation-exchange capacity (CEC), but also in increasing forage production and the Ca, Mg, K, Cu, and Zn concentrations in the plants. The benefits of SL application to soils also include long-term control of Aphanomyces (common rootrot disease; Ag Gold Standard, 2008). Although this product has been mainly applied to soil, SL has been also used for waste water treatment (Ippolito et al., 2013), reducing Cu concentrations in water effluents from dairy industries, which use Cu in hoof bath solutions to prevent hoof diseases.
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ACCEPTED MANUSCRIPT 3.2. Trace elements stabilization in the soil using sugar beet lime. Industrial wastes or byproducts such as SL could be used to reduce of the mobility of
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the contaminants or to increase their retention in the soil. González-Núñez, Alba, Orta,
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Vidal, and Rigol (2012a) proved the suitability of this byproduct for the remediation of
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metal-contaminated soils, which showed simultaneous increases in pH and metal sorption after SL addition.
Most of the studies that used SL as a soil amendment reported data obtained under
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laboratory or controlled greenhouse conditions - in pots or containers or in column
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experiments. Many authors coincided in affirming that this product decreased trace elements availability (Madejón et al, 2006; Madejón, Burgos, Cabrera, & Madejón 2009b; Pérez-de-Mora, Madejón, Burgos, & Cabrera, 2006c). Furthermore, in field
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studies, SL also proved to be an excellent ameliorating agent for trace elements contaminated acid soils. For example, it was used successfully in the restoration of the
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Green Corridor of the Guadiamar river (SW Spain), created after the Aznalcóllar mine spill in 1998. Different studies carried out in this area (Clemente, Walker, and Bernal,
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2005; Madejón et al., 2006; Burgos, Pérez-de-Mora, Madejón, Cabrera, & Madejón, 2008; Madejón, Madejón, Burgos, Pérez de Mora, & Cabrera, 2009a) have demonstrated, at different scales, the effectiveness of this byproduct in the remediation/stabilization of trace elements contaminated soils. In this area, SL was able to raise soil pH to neutral values and its effect was sufficient to maintain soil pH close to neutrality over a 5-year period after the last application (Madejón et al., 2010). This result demonstrated that SL was a more time-efficient alkalinizing soil amendment than other organic materials. When the alkalinity of soils is increased, insoluble trace elements may precipitate, forming insoluble compounds and complexes and secondary minerals. Therefore,
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ACCEPTED MANUSCRIPT increases in soil pH are theoretically linked to a decline in the availability of metals in the soil. However, some studies have shown a clear reduction on trace elements
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availability after SL addition to soil (Burgos et al., 2010; Pérez de Mora et al., 2006c,
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Madejón et al., 2009a), while others have reported less clear or even contrasting results
(Jiménez Moraza, Iglesias, & Palencia, 2006).
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regarding the behavior of certain elements like Zn, Mn, and Cd after SL application
Different studies on the leachability and solubility of trace elements in the soil after SL
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addition in columns (Garrido, Illera, Campbell, & García-González, 2006; Aguilar-
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Carrillo, Garrido, Barrios, & García-González, 2009), pots (González-Núñez, Alba, Orta, Vidal, & Rigol, 2012b), and semi-field experiments (Pérez-de-Mora, Burgos, Cabrera, & Madejón, 2007), have found that this amendment is able to reduce the
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concentrations of these elements in leachates and, consequently, the risk of groundwater contamination. In agreement with this, the results from a columns experiment using an
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acidic contaminated soil showed that SL decreased significantly the mobility of Cd, Cu, and Pb in the soil (Campbell, Garrido, Illera, & García-González, 2006). Similarly,
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Aguilar-Carrillo et al. (2009), in another columns experiment, found that SL reduced the potential leachability of As, Cd, and Tl through the profile of an acidic soil. In the latter study, results from a sequential extraction revealed that the amendment of contaminated soils with SL provoked redistribution of the toxic elements from soluble and exchangeable (highly available) pools to a less available fraction (associated with Al, Fe, and Mn (hydro)oxides). Direct remediation of subsoil acidity in mine wastes has been tested in different studies using SL in combination with other organic amendments. Svendson, Henry, and Brown, (2007) studied the use of different mixtures in mine tailings rehabilitation and concluded that the most successful, among different types of liming materials, in
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ACCEPTED MANUSCRIPT reducing 0.01 M Ca(NO3)2-extractable Cd and Zn concentration, was biosolid sludge in combination with SL. These authors remarked that this material is an industrial
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byproduct and therefore offers a low-cost alternative to commercial limestone. Shaheen
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et al. (2014) found similar results using mixtures of SL and sewage sludge and
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demonstrated the effectiveness of these mixtures in reducing the availability of Cu and Zn to plants. Alvarenga et al. (2008) showed that SL in combination with other organic amendments decreased the level of the mobile/effectively bioavailable fraction of Cu,
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Pb, and Zn in mining soils in Portugal. Similarly, Brown, Svendsen and Henry (2009)
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reported a reduction in the Cd and Zn concentrations extracted with Ca(NO3)2 from fluvial mine tailings following the application of mixtures of SL and biosolid, in a longterm field experiment in Arkansas River, Connecticut (USA).
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Regarding the recommended doses of SL in soils, Jiménez-Moraza et al. (2006) calculated the most adequate amount of SL for pyrite neutralization in mine tailings
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restoration. They concluded that this product was highly effective for soil alkalization and that the stoichiometric dose was enough to prevent or delay the leaching of Al, Cu,
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and Fe, whereas the release of Cd, Mn, and Zn was influenced by the mass ratio SL/sulfide sulfur of the soil. As commented on previously, SL has non-negligible nutrient and OM contents. This fact is of special interest in soil remediation procedures, as many polluted soils are also characterized by negative properties such as poor nutrient availability, lack of soil structure, low OM content, high salinity, and/or acid pH (Adriano, 2001). In this sense, SL acts as a powerful fertilizer, mainly improving K and P availability (Mosen, 2007; García-Zamareño et al., 2013; Shaheen & Tsadilas, 2013) and also increasing the OM content and soluble concentrations (Pérez de Mora et al., 2005, Gómez–Paccard et al., 2013) in the soil.
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3.3. Effects on vegetation and soil microbial populations.
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As a result of the soil condition improvement, several case studies using SL have
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shown, under different environments, enhanced natural remediation processes that have
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resulted in substantially improved vegetation growth. The beneficial effects of SL use have been demonstrated also when an induced vegetation cover was established under semi-controlled conditions (Pérez-de-Mora, Madejón, Burgos, & Cabrera, 2006b;
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Shaheen et al., 2014; De Sutter & Godsey, 2010) or when native vegetation colonized a
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contaminated soil under field conditions (Clemente et al., 2006a; Madejón et al., 2006, Pérez de Mora et al., 2011). Some authors have also shown that not only do plant cover and biomass increase in SL treated soils, but also the plant biodiversity of the area is
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greatly improved (Pérez de Mora et al., 2011). Another important finding related to the creation or restoration of a vegetation cover is a
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general reduction of trace elements translocation to the plant aerial parts (Madejón et al., 2006; Burgos et al., 2008; Pérez de Mora et al., 2011; Shaheen et al., 2014). This
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represents an important goal for phytoremediation, especially when phytostabilization strategies are chosen. A well-established vegetation cover with low concentrations of contaminants in the aerial parts maintains trace elements stabilized in the soil, reduces losses by erosion processes, and avoids transfer to the food chain (Bolan, Park, Robinson, Naidu, & Huh, 2011). Another important aspect to be considered is the microbiological status of the contaminated soils. Trace elements pollution exerts a negative impact on soil microbial activity, which greatly alters nutrient cycles (Domínguez, Marañón, Murillo, Schulin, & Robinson, 2010). The detrimental effects of trace elements on soil quality are largely related to both soil fertility status and biological activity (Gil-Sotres, Trasar-Cepeda,
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ACCEPTED MANUSCRIPT Leiros, & Seoane, 2005), hindering soil restoration programs. In this regard, as a result of the improvement of soil physico-chemical conditions and fertility and of the
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establishment of plants in the soil, the addition of SL has a very positive effect on the
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soil microbiological status. For example, in an incubation experiment, Burgos et al.
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(2010) found that, in neutral soil with low trace elements availability, although the addition of SL did not alter soil trace elements extractability, it significantly enhanced soil microbial activity. Also, Pérez de Mora et al. (2005, 2006a) reported the
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biochemical restoration of a contaminated soil treated with SL under semi-field
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conditions, by the increase of microbial biomass C, enzymatic activities and heterotrophic potential. These authors also demonstrated, by ARDRA (amplified ribosomal DNA restriction analysis), changes in the structural diversity of both the
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bacterial and fungal community under the SL treatment. Fingerprinting patterns of the 16S rDNA obtained with HinfI and of the 18S rDNA with HpaII revealed higher
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similarity percentages among samples from the same treatment with respect to samples from other inorganic and organic treatments. They concluded that the use of SL could
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improve soil chemical properties, as well as the microbial community function and structure (Pérez de Mora et al., 2006a). In agreement with this, Hinojosa, Carreira, Rodríguez-Maroto, and García-Ruíz (2008), who evaluated the recovery of enzyme activities in polluted soil after addition of SL in a pot experiment, reported that although soil recovery was still incomplete in terms of bioavailable trace elements, the recovery of certain soil enzyme activities was fully successful.
3.4. Other sugar beet byproducts used in trace elements contaminated soils. Although the main byproduct from the sugar beet industry used in trace elements contaminated soils is SL, other sugar beet wastes have been also used for this purpose.
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ACCEPTED MANUSCRIPT Alguacil et al. (2011) used fermented sugar beet pulp in a trace elements contaminated soil. These authors found that the addition of this amendment increased the population
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of AMF and decreased trace elements accumulation in the shoots of the host plants,
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concluding that this waste can be used for the remediation and/or phytostabilization of
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mine tailings. Medina, Vassilev, Barea, and Azcón (2005) studied the use of the same byproduct in a soil contaminated artificially (spiked) with Cd and found that the addition of this waste together with the colonization by AM endophytes was paramount
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for the revegetation of the contaminated soil, as both factors (mycorrhizal and organic
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sugar beet waste addition) increased plant growth and improved soil quality. Similar results were reported by Medina, Vassilev, Barea, and Azcón (2006) for a soil
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contaminated artificially with Zn.
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4. Other types of byproducts used in the remediation of trace elements contaminated soils.
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Other agri-food byproducts have been tested as soil amendments, with different results, and just a few of these have been used in contaminated and degraded land remediation. Byproducts from the wine industry are one of the most significant examples. Grapes are one of the world‘s largest fruit crops, with an annual production of more than 60 million tons - of which about 80% are used for wine making. The winery and distillery industries generate an increasing amount of solid and liquid byproducts and wastes such as grape stalk, grape pomace (marc), wine lees, exhausted grape marc, and vinasse (Rubio et al., 2013). Grape pomace is the most important solid residue remaining after juice extraction in the winemaking process and represents 20-30% of the weight of processed grapes; it includes skins, seeds, and stalks, and has an elevated moisture
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ACCEPTED MANUSCRIPT content (about 60%). Only in Europe, 4 million tons of solid wastes are produced annually by the wine sector in wine producing countries; these are mainly used as cattle
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feed or dumped in disposal sites (Grub‘s Up, 2009). There is, therefore, high interest in
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finding alternative uses for these byproducts.
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For instance, wine byproducts have been successfully composted (Bustamante, Paredes, Morales, Mayoral, & Moral, 2009) and their composts used effectively as fertilizers (Rubio et al., 2013) or in growing media (García-Martínez et al., 2009), with positive
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results regarding crop yield and quality. Even the direct application of winery
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wastewater to soil has been demonstrated to be a suitable alternative for its disposal (Mosse, Patti, Smernik, Christen, & Cavagnaro, 2012). Díaz, Madejón, López, López, and Cabrera (2002) reported that grape marc could be recycled as a soil conditioner in
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view of its organic and nutrient contents. Wine waste is characterized by the presence of natural antioxidants and can be potentially used as a soil conditioner, adsorbent for
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heavy metals, or fertilizer.
Composts derived from winery wastes have been found to be of good quality with good
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physico-chemical characteristics and high concentrations of nutrients – properties which confer on them an elevated agronomic value and make them adequate for use as soil conditioners (Arvanitoyannis et al., 2006; Bustamante, Said-Pullicino, Paredes, Cecilia, & Moral, 2010). The use of compost derived from wine byproducts in vineyards is of growing interest because of the general low levels of OM in these soils and their exposure to erosion (). Combinations of wine byproducts with other materials are frequently used to ameliorate the negative properties of single materials, such as heterogeneity, high salinity, low content of OM, low CEC or the presence of contaminants (Ingelmo, Canet, Ibañez, Pomares, & García, 1998).
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ACCEPTED MANUSCRIPT Regarding trace elements remediation, winery waste sludge has been found to be an effective adsorbent for the removal of Cr, Ni and Cu from aqueous solutions
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(Villaescusa et al., 2004). The sorption of metals by winery wastes might be attributable
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to their proteins, carbohydrates, and phenolic compounds - that have carboxyl,
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hydroxyl, sulfate, phosphate, and amino groups that can bind metal ions (Villaescusa et al., 2004).
In an incubation experiment, Karaka (2004) studied the effects of composted grape
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marc and mushroom compost on the extractability of metals (Cd, Cu, Ni, and Zn) in an
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alkaline soil. The amendments increased the OM content of the soil, slightly decreased soil pH, and caused significant decreases in the extractable (DTPA) Cd, Cu, and Ni concentrations; these decreases became greater with increasing rates of both organic
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amendments. A significant, negative correlation was found between the extractable Cu concentrations and the OM added to the soil, a consequence of the ability of Cu to form
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stable complexes with soil OM. In contrast, the DTPA-extractable Zn concentrations increased with increasing rates of both organic amendments, and this was assumed to be
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a consequence of the pH decrease. Mushroom compost is thus another by-product that potentially can be used for trace elements remediation. The most significant edible fungi cropped around the world are Agaricus bisporus and Pleurotus ostreatus - which, therefore, generate the substrate wastes with the greatest environmental implications (García-Delgado, Jiménez-Ayuso, Frutos, Gárate, & Eymar, 2013). A large amount of spent A. bisporus substrate is generated; just in Europe, more than 3.5×106 t are produced every year. The accumulation of spent substrate causes important environmental pollution in mushroom production areas due to salts and soluble organic carbon leaching during its storage (Guo, Chorover, & Fox, 2001).
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ACCEPTED MANUSCRIPT In the search for new or alternative uses for spent mushroom substrates, García-Delgado et al. (2013) studied the metal adsorption capacity of two of them (P. ostreatus and A.
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bisporus), which were originally formed of wheat straw and wheat straw in combination
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with poultry litter and grape marc, respectively. According to adsorption isotherms,
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both materials showed high Cd and Pb adsorption potential, which was related to the presence of carboxyl acidic groups that increase the CEC of these materials and induced the formation of heavy metal complexes. Spent substrate used as an A. bisporus
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inoculum carrier was described as being useful for soil bioremediation, producing the
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immobilization of heavy metals and the enhancement of PAH biodegradation (GarcíaDelgado et al., 2013). In agreement with this, Shuman (1998) reported that spent mushroom compost was able to lower Cd and Pb availability in an artificially
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contaminated (spiked) soil, by redistribution of the metals from exchangeable and/or OM related fractions to less available fractions in the soil. In the same experiment,
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cotton gin litter (short fibers and plant trash combed from the cotton before processing for spinning) caused much lower immobilization of Cd and Pb.
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Another example of a byproduct with potential use as an amendment in trace elements contaminated soils is rice husk, whose addition to a Cd contaminated soil provoked a significant reduction in the concentrations of Cd accumulated in the leaves of Amaranthus caudatus and improved plant yield to levels comparable with those obtained in non-contaminated sites (Singh & Prasad, 2014). Finally, rapeseed residue, in addition to being a source of biofuel, can be used as an amendment to improve soil fertility (Tejada, Hernández, & García, 2009). Lee et al. (2013) treated soils from a rice paddy area adjacent to a mine site in Korea, contaminated by Cd and Pb (11.3 and 1233 mg kg−1, respectively), with rapeseed residue and eggshells. This increased soil pH values and decreased significantly the
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ACCEPTED MANUSCRIPT extractable (acetic acid-TCLP and diluted HCl) Cd and Pb concentrations, that declined with increasing rates of eggshell addition (Lee et al., 2013). Eggshell has been used in
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soil remediation experiments as a lime-based waste material, as it is very readily
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obtained and contains >95% CaCO3 (Ahmad et al., 2012). It can be a good alternative to
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commercial limes, with many advantages for immobilizing heavy metals in soils (Ok et al., 2010).Application of rapeseed residue and eggshell waste was thus shown to be a cost-effective way to remediate the soil contaminated with these metals, mitigating soil
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acidity and immobilizing the metals in the soil (Lee et al., 2013).
5. Final considerations and aims for the future.
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Agricultural land is being significantly reduced (through degradation, erosion, salinization and contamination processes) or diverted to non-agricultural uses, while
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total food production is notably increasing, this generating an environmental footprint which represents a major issue that must be urgently addressed (Lal, 2013). Agri-food
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industry is one of the largest production sectors worldwide, and is thus of primary importance for the global economy. One of the main issues challenging this sector is the huge generation of byproducts and wastes of many different kinds. Only in the EU the total amount of food byproducts produced has been estimated at roughly 222 million tons per year (AWARENET, 2004) and approximately 30% of global agricultural products end up as residues and refuses (Ajila et al., 2012). In the last few years, new methods and policies for waste treatment have been introduced, and the global perception of agricultural wastes is changing rapidly in response to the need for environmental preservation and sustainability, and for global food security (Ajila et al., 2012). In addition, the efficiency of food production can be
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ACCEPTED MANUSCRIPT enhanced by reusing and recycling the byproducts generated, turning them from wastes into resources (Lal, 2013).
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In this sense, the use of food byproducts, fresh or transformed, alone or in combination
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with other (residual) materials, in the remediation of trace elements contaminated soils
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(summarized in Table 2) has been clearly shown in this review to represent an attractive alternative to conventional landfill or incineration treatments, especially when further acquisition of valuable products is not feasible (e.g., after biogas production or
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extraction of valuable chemical compounds from waste materials). Land disposal of
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organic and/or inorganic food byproducts may alter trace elements solubility and bioavailability in the soil, likely provoking their immobilization in the soil and improving plant establishment and growth by improving the nutrient status and the
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physico-chemical characteristics of the soil. This may offer a low-cost, sustainable solution for the remediation of trace elements contaminated soils and the establishment
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of a healthy and self-sustaining vegetation cover, which would protect the soil against further degradation and prevent the dispersion of the contaminants in the ecosystem
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(Clemente et al., 2012; van Herwijnen et al., 2007). Nevertheless, some aspects may have to be carefully considered in order to elude any possible negative effects in the environment; namely, the type (stability) of organic matter and the amount (dose) of byproduct added may have to be adjusted to control the effects on soil pH and redox conditions that could cause mobilization of certain elements like As or Mn. In any case, it has also been evidenced here that byproducts recycling in trace elements contaminated soils would restore or improve soil health and quality, contributing significantly to the stimulation of the growth and activity of microorganisms, as well as to the completion of the cycles of the principal nutrients, which are usually inhibited in these soils. This would also positively influence the overall ecosystem functioning and
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the potential environmental impact they may pose.
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Acknowledgements
The authors thank Dr. David J. Walker for the English corrections. Part of R.
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Clemente‘s salary is paid by the European Social Fund of the EU.
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ACCEPTED MANUSCRIPT Madejón, P., Burgos, P., Cabrera, F., & Madejón, E. (2009b). Phytostabilization of amended soils polluted with trace elements using the Mediterranean Shrub:
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Rosmarinus officinalis. International Journal of Phytoremediation, 11, 542–557.
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(2010). Do amended, polluted soils require re-treatment for sustainable risk reduction? — Evidence from field experiments. Geoderma, 159, 174–181. Mahmoud, M., Janssen, M., Haboub, N., Nassour, A., & Lennartz, B. (2010). The
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impact of olive mill wastewater application on flow and transport properties in soils.
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soil. Soil & Tillage Research, 124, 24–31. Martínez-Fernández, D., & Walker, D. J. (2012). The effects of soil amendments on the
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growth of Atriplex halimus and Bituminaria bituminosa in heavy metal-contaminated soils. Water, Air, & Soil Pollution, 223, 63–72.
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McGrath, S. P., Chaudri, A. M., & Giller, K. E. (1995). Long-term effects of metals in sewage sludge on soils, microorganisms and plants. Journal of Industrial Microbiology, 14, 94-104. Medina, A., Vassilev, N., Barea, J. M., & Azcón, R. (2005). Application of Aspergillus niger-treated agrowaste residue and Glomus mosseae for improving growth and nutrition of Trifolium repens in a Cd-contaminated soil. Journal of Biotechnology, 116, 369–378. Medina, A., Vassilev, N., Barea, J. M., & Azcón, R. (2006). The growth-enhancement of clover by Aspergillus-treated sugar beet waste and Glomus mosseae inoculation in Zn contaminated soil. Applied Soil Ecology, 33, 87–98.
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ACCEPTED MANUSCRIPT Ministerio de la Presidencia. (2013). Real Decreto 506/2013, de 28 de junio, sobre productos fertilizantes. BOE 164, 51119-51207.
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Mirabella, N., Castellani, V., & Sala, S. (2014). Current options for the valorization of
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food manufacturing waste: a review. Journal of Cleaner Production, 65, 28-41.
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Morillo, J. A., Antizar-Ladislao, B., Monteoliva-Sánchez, M., Ramos-Cormenzana, A.,
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& Russell, N. J. (2009). Bioremediation and biovalorisation of olive-mill wastes. Applied Microbiology and Biotechnology, 82, 25–39. Mosen, A. (2007). Sugar-beet Handbook. Hoboken: John Wiley & Sons, Inc., (Chapter
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Ok, Y. S., Oh, S. E., Ahmad, A., Hyun, S., Kim, K. R., Moon, D. H., Lee, S. S., Lim, K. J., Jeon, W. T., & Yang, J. E. (2010). Effects of natural and calcined oyster shells on Cd and Pb immobilization in contaminated soils. Environmental Earth Sciences, 61, 1302-1308. Oreopoulou, V., & Russ, W. (2007). Utilization of by-products and treatment of waste in the food industry. New York: Springer, 316 pp. Pardo, T., Bernal, M. P., & Clemente, R. (2014a). Efficiency of soil organic and inorganic amendments on the remediation of a contaminated mine soil: I. Effects on trace elements and nutrients solubility and leaching risk. Chemosphere, 107, 121128.
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ACCEPTED MANUSCRIPT Pardo, T., Clemente, R., & Bernal, M. P. (2011). Effects of compost, pig slurry and lime on trace element solubility and toxicity in two soils differently affected by mining
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activities. Chemosphere, 84, 642-650.
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Pardo, T., Clemente, R., Alvarenga, P., & Bernal, M. P. (2014b). Efficiency of soil
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the phytostabilisation efficiency in a trace elements contaminated soil using soil
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health indicators. Journal of Hazardous Materials, 268, 68-76. Pardo, T., Martínez-Fernández, D., Clemente, R., Bernal, M. P., & Walker, D. J. (2014d). The use of olive-mill waste compost to promote the plant vegetation cover
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in a trace element-contaminated soil. Environmental Science and Pollution Research,
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Paredes, C., Cegarra, J., Roig, A., Sánchez-Monedero, M. A., & Bernal, M. P. (1999). Characterization of olive mill wastewater (alpechín) and its sludge for agricultural
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purpose. Bioresource Technology, 67, 111-116. Paredes, C., Roig, A., Bernal, M.P., Sánchez-Monedero, M.A., Cegarra, J. (2000). Evolution of organic matter and nitrogen during co-composting of olive mill wastewater with solid organic wastes. Biology and Fertility of Soils, 32, 222-227. Pérez de Mora, A., Ortega-Calvo, J. J., Cabrera, F., & Madejón, E. (2005). Changes in enzyme activities and microbial biomass after ―in situ‖ remediation of a heavy metalcontaminated soil. Applied Soil Ecology, 28, 125-137. Pérez-de-Mora, A., Burgos, P., Cabrera, F., & Madejón, E. (2007). ―In situ‖ amendments and revegetation reduce trace element leaching in a contaminated soil. Water, Air and Soil Pollution, 185, 209–222.
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ACCEPTED MANUSCRIPT Pérez-de-Mora, A., Burgos, P., Madejón, E, Cabrera, F., Jaeckel, P., & Schloter, M. (2006a). Microbial community structure and function in a soil contaminated by
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heavy metals: effects of plant growth and different amendments. Soil Biology &
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Biochemistry, 38, 327–341.
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Pérez-de-Mora, A., Madejón, E, Burgos, P., & Cabrera, F. (2006b). Trace element availability and plant growth in a mine-spill-contaminated soil under assisted natural remediation: II. Plants. Science of the Total Environment, 363, 38-45.
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Pérez-de-Mora, A., Madejón, E., Burgos, P., & Cabrera, F. (2006c). Trace element
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availability and plant growth in a mine-spill contaminated soil under assisted natural remediation I. Soils. Science of the Total Environment, 363, 28–37. Pérez-de-Mora, A., Madejón, P., Burgos, P., Cabrera, F., Lepp, N. W., & Madejón, E.
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(2011). Phytostabilization of semiarid soils residually contaminated with trace elements using by-products: Sustainability and risks. Environmental Pollution, 159,
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Maestri, D. (2012). Physico-chemical and toxicological assessment of liquid wastes from olive processing-related industries. Journal of the Science of Food and Agriculture, 92, 216–223. Piotrowska, A., Iamarino, G., Antonietta, M., & Gianfreda, L. (2006). Short-term effects of olive mill wastewater (OMW) on chemical and biochemical properties of a semiarid Mediterranean soil. Soil Biology & Biochemistry, 38, 600–610. Quilty, J. R., & Cattle, S.R. (2011). Use and understanding of organic amendments in Australian agriculture: a review. Soil Research, 49, 1–26.
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ACCEPTED MANUSCRIPT Romero, E., Benítez, E., & Nogales, R. (2005). Suitability of wastes from olive-oil industry for initial reclamation of a Pb/Zn mine tailing. Water, Air, & Soil Pollution,
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Rubio, R., Perez-Murcia, M. D., Agullo, E., Bustamante, M. A., Sanchez, C., Paredes,
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C., & Moral, R. 2013. Recycling of agro-food wastes into vineyards by composting: agronomic validation in field conditions. Communications in Soil Science and Plant Analysis, 44, 502-516.
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Shaheen, S. M, Shams, M. S., Ibrahim, S. M., Elbehiry, F. A., Antoniadis, V., & Hooda,
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P. S. (2014). Stabilization of sewage sludge by using various by-products: effects on soil properties, biomass production, and bioavailability of copper and zinc. Water, Air and Soil Pollution, 225:2014.
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Shaheen, S. M., & Tsadilas C. (2013). Phosphorus sorption and availability to canola grown in an alfisol amended with various soil amendments. Communications in Soil
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Science and Plant Analysis, 44, 89–103. Shuman, L. M. (1998). Effect of organic waste amendments on cadmium and lead in
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soil fractions of two soils. Communications in Soil Science and Plant Analysis, 29, 2939-2952.
Singh, A., & Prasad, S. M. 2014. Effect of agro-industrial waste amendment on Cd uptake in Amaranthus caudatus grown under contaminated soil: An oxidative biomarker response. Ecotoxicology and Environmental Safety, 100, 105-113. Svendson, A., Henry, C., & Brown, S. (2007). Revegetation of high zinc and lead tailings with municipal biosolids and lime: greenhouse study. Journal of Environmental Quality, 36, 1609–1617. Tejada, M., Hernández, M. T., & García, C. (2009). Soil restoration using composted plant residues: effect on soil properties. Soil & Tillage Research, 102, 109–117.
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ACCEPTED MANUSCRIPT Torres, J. L., Varela, B., García, M. T., Carilla, J., Matito, C., Centelles, J.J., Cascante, M., Sort, X., & Bobet, R. (2002). Valorization of grape (Vitis vinifera) byproducts.
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Antioxidant and biological properties of polyphenolic fractions differing in
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procyanidin composition and flavonol content. Journal of Agriculture and Food
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Chemistry, 50, 7548-7555.
Van Dyk, J. S., Gama, R., Morrison, D., Swart, S., & Pletschke, B. I. (2013). Food processing waste: Problems, current management and prospects for utilization of the
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lignocellulose component through enzyme synergistic degradation. Renewable and
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Sustainable Energy Reviews, 26, 521–531.
van Herwijnen, R., Hutchings, T. R., Al-Tabbaa, A., Moffat, A. J., Johns, M. L., & Ouki, S. K. (2007). Remediation of metal contaminated soil with mineral-amended
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composts. Environmental Pollution, 150, 347-354. Villaescusa, I., Fiol, N., Martínez, M., Mirrales, N., Poch, J., & Seralocs, J. (2004).
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Removal of copper and nickel ions from aqueous solutions by grape stalks wastes. Water Research, 38, 992–1002.
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Walker, D. J., Clemente, R., Roig, A., & Bernal, M. P. (2003). The effects of soil amendments on heavy metal bioavailability in two contaminated Mediterranean soils. Environmental Pollution, 122, 303–312.
Figure legends
Figure 1. Effects of the addition of AL, AL-compost (CM), and their corresponding water extracts (WE-AL and WE-COM) on Mn and Zn availability (sequential
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treatment.
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Figure 2. Principal component analysis (PCA) of the physico-chemical (nutrients and
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trace elements availability) and biochemical (enzymatic activities) parameters of a contaminated soil treated with AL compost (CM) and other amendments (CT: control;
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HL: hydrated lime; PS: pig slurry) in a glasshouse column experiment.
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ACCEPTED MANUSCRIPT Table 1. Chemical composition of olive mill wastewater (OMWW, fresh weight), and alperujo (AL) and experimental and industrial solid-OMWs composts (dry weight).
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Adapted from Alburquerque et al. (2004), Morillo et al. (2009), Chowdhury, Akratos,
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Vayenas, & Pavlou (2013) and Álvarez de la Puente (2006).
Dry matter (%) pH -1
EC (dS m )
Industrial
Composts
Composts
AL
6.33-7.19
49.6-71.4
n.d.
24.8-45.3
4.2-5.2
4.9-6.8
5.4-9.5
7.8-9.1
0.98-5.2
1.2-7.3
1.7-5.3
848-976
260-900
n.d.
5.5-12.0 -1
Experimental
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OMWW
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Parameters
46.5-62.1
-1
34.2-39.8
495-539
110-580
n.d.
TN (g kg )
0.61-2.10
7.0-18.5
11-54
16-29
C/N
52.3-54.3
28.2-72.9
9-36
14-20
0.16-0.31
0.7-2.2
1-30
0.8-2.2
1.97-8.97
7.7-29.7
6-44
6.5-24
0.11-0.42
0.5-1.6
2-41
6-49
0.20-0.64
1.7-9.2
7-72
19-72
0.04-0.22
0.7-3.8
0.9-57
1.6-12.7
18.3-120
78-1462
100-4100
991-1188
1.5-6
12-29
1.4-79
14-126
1.1-12
5-39
13-131
56-199
a
39-321
b
n.d. n.d.
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Organic Matter (g kg ) TOC (g kg )
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-1
P (g kg ) -1
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K (g kg ) -1
-1
Ca (g kg ) -1
Mg (g kg ) -1
Fe (mg kg ) -1
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Cu (mg kg )
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Na (g kg )
-1
Mn (mg kg ) -1
Zn (mg kg )
2.4-12
10-37
38-138
Phenols (%)
0.98-10.7
0.5-2.4
0.1-3.8
Lipids (%)
1.64-12.2
3.7-18
0.2-8.1 b
Carbohydrates (%) 1.4-16.1 9.6-19.3 0.1-1.2 n.d. n.d.: not determined; a Alburquerque, personal communication; b water soluble fractions.
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Table 2. The main agri-food byproducts used in the remediation of trace elements contaminated soils (OMWW: olive mill wastewater; AL: solid
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olive mill waste (alperujo); SL: sugar beet lime). Byproduct
Soils
Effects
Olive oil
OMWW
Artificially contaminated soils,
Acidification of soil. Increased trace elements solubility (especially Fe and Mn) and leaching.
agricultural soils
Inhibition of plant growth.
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Source
References Madrid & Díaz-Barrientos, 1998; Mahmoud et al., 2010, 2012; Nogales et al., 1997; Piotrowska et al., 2006
Increased soil aggregates stability and soil hydrophobicity. Acid and calcareous contaminated soils
In acid soil, Pb and Zn immobilization and slight Cu and Fe mobilization. In calcareous soil, slight increase in Cu, Pb and Zn availability.
AL
Mine soils, contaminated calcareous soils
Acidification of soil. Increased Cu and Mn solubility. Immobilization of some metals (like Pb and Zn) after alteration of carbonate/bicarbonate equilibrium.
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Humic acids from OMWWcompost
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Inhibition of plant growth and phytotoxicity.
Clemente & Bernal, 2006
Burgos et al., 2010; Clemente et al., 2007a, 2007b; de la Fuente et al., 2008, 2011; Pierantozzi et al., 2012; Romero et al., 2005
AL-compost
Mine soils, soils indirectly contaminated by mine activities, and soils affected by pyritic sludge spill
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Short-term negative effects on soil microbial communities. Longterm stimulation of soil microbial biomass and activity. Slow mineralisation of OM. Increase in soil pH and content of OM and essential nutrients (N, K, P). Risk of nitrate and K leaching. Decreased trace elements solubility and leaching. Slight solubilization of As and Cu. Facilitation of plant establishment and growth, and reduction of trace elements uptake by plants. Stimulation of the development, activity and diversity of soil microbial communities. Reduction of soil direct and indirect toxicity to microorganisms, aquatic and terrestrial invertebrates, and plants.
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Alburquerque et al., 2011; Beesley et al., 2014; Clemente et al., 2006b, 2012; Curaqueo et al., 2014; de la Fuente et al., 2011; Fornes et al., 2009; MartínezFernández & Walker, 2012; MorenoJiménez et al., 2013; Pardo et al., 2011, 2014a, 2014b, 2014c, 2014d
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Mine soils, soils affected by pyritic sludge spill, agricultural acidic soils
Increase in soil pH and content of OM and essential nutrients (mainly K, P).
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SL
Decrease in trace elements availability and leaching. Solubilization of trace metals (Cd, Mn and Zn).
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Sugar beet
Increased plan establishment and reduced trace elements translocation in plants.
Mine soils, artificially contaminated soils
Decreased trace elements uptake and accumulation in plants. Increased plant growth.
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Fermented sugar beet pulp
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Improved growth and activity of the soil microbial communities.
Aguilar-Carrillo et al., 2009; Alvarenga et al., 2008; Brown et al., 2009; Burgos et al., 2010; Clemente et al., 2005, 2006a; de Sutter & Godsey, 2010; Garrido et al., 2006; Gómez-Paccard et al., 2013; González-Nuñez et al., 2012b; Hinojosa et al., 2008; Jiménez-Moraza et al., 2006; Madejón et al., 2006, 2009a; Pérez de Mora et al., 2005, 2007, 2011; Svendson et al., 2007; Shaheen et al, 2014 Alguacil et al., 2011; Medina et al., 2005, 2006
Stimulation of arbuscular mycorrhizal fungi populations. Aqueous contaminated solutions, agricultural soils
Decreased trace elements solubility (high adsorption capacity).
Winery wastescompost
Agricultural soils
Increase in soil nutrients content. Immobilization of trace elements. Increased crop production biomass and quality
Arvanitoyannis et al., 2006; Bustamante et al., 2010; García-Martínez et al., 2009; Karaka, 2004; Rubio et al., 2013
Mushroom cultivation
Spent substrate /compost
Alkaline soils, artificially contaminated soils
Slight decrease in soil pH. Increase in OM content.
García-Delgado et al., 2013; Karaka, 2004; Shuman ,1998
Eggs
Eggshell
Increase in soil pH. Immobilization of trace elements.
Rice
Rice husk
Contaminated soils, paddy soils Contaminated soils
Rapeseed
Rapeseed waste
Contaminated soils, paddy soils
Increase in soil pH. Decreased trace elements solubility. Enhancement of soil fertility
Lee et al., 2013; Tejada et al., 2009
Brewing industry
Spent brewing grain
Contaminated soils
Sulfate reduction. Removal of TEs.
Lindsay et al., 2011
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Winery waste sludge
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Enhancement of soil fertility vs. increase in soil electrical conductivity.
Reduction in trace elements solubility (high adsorption capacity). Degradation of PAH.
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Wine industry
Reduction of trace elements uptake by plants, increased plant growth
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Díaz et al., 2002; Ingelmo et al., 1998; Villaescusa et al., 2004
Ahmad et al., 2012; Lee et al., 2013; Ok et al., 2010 Singh & Prasad, 2014
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Huge amounts of byproducts are generated every year and cause environmental
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Highlights
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concern
Agro-wastes can be used as amendments in trace elements contaminated soils
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Food byproducts offer a low-cost and sustainable solution for soils remediation
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Type of byproduct and dose have to be tested to avoid negative effects
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This would help protecting the ecosystem and minimizing environmental risks
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