While soil management is not a soil property in itself, management modifies many soil properties and soil profile morphology, and hence affects the fate and mobility of contaminants in the soil matrix (Yaron, Dror and Berkowitz, 2012); thus its inclusion in this section is particularly relevant.
The main sources of contaminants to agricultural soils are excessive use and/or misuse of agrochemicals (synthetic fertilizers, pesticides and other chemicals), amendment with sewage sludge, manure and other organic amendments, irrigation with contaminated wastewater, and misuse of agroplastics. Soils under other land uses are also exposed to contaminants, and changes in land use can greatly contribute to contaminant mobilization. For example, it is widely accepted that land use change from forest or grassland to agriculture leads to a decrease in SOM and a loss of biodiversity, contributing to the mobilization of contaminants (Ondrasek et al., 2019). In addition, other management practices contribute to the availability and mobilization of contaminants in soil. Tillage, liming, drainage, irrigation, and changes in crop species all influence physical, chemical and biological processes in the soil that affect the speciation of contaminants and the redistribution of certain soil fractions and compounds (Yaron, Dror and Berkowitz, 2012).
Despite the positive effects of tillage on crop yield, it is widely accepted that tillage (i) disrupts soil structure and breaks down soil aggregates, mainly macro-aggregates (Janušauskaite, Kadžienė and Auškalnienė, 2013; Munkholm, Hansen and Olesen, 2008; Zheng et al., 2018), (ii) alters the habitat for soil organisms (Frouz et al., 2014; Janušauskaite, Kadžienė and Auškalnienė, 2013), (iii) increases soil erosion risk, and (iv) accelerates the mineralization of labile SOC and reduces recalcitrant forms in the long term (Jemai et al., 2012; Zhao et al., 2012). The conjunction of these processes increases soil erodibility (Withers et al., 2007) and mobilizes soil contaminants (Figure 19) (Rampazzo Todorovic et al., 2014). The breakage of the macro-aggregates exposes the contaminants that had been physically occluded to microbial degradation and mobilization, increasing the bioavailability (Withers et al., 2007). Soil porosity shifts to an increase in macropores after tillage, increasing aeration that favours mineralization of SOC and reduced soil moisture (Li et al., 2015; Wang et al., 2017). Despite the increasing popularity of no-till or minimum tillage practices in the last decades, these remain only the third most common practice: conventional tillage10 (41 percent of total cropland) and traditional tillage11 (35 percent of total cropland) are still the most applied practices globally (Porwollik et al., 2019). Although there are some contradictory studies on the effects on soil management on contaminant transport (Flury, 1996), Gjettermann et al. (2009) pointed out that soil tillage before pesticide application facilitates contaminant mobility, due to the alteration of soil structure and the mobilization of colloids along with the strongly adsorbed contaminants.
Rotary tillage and mouldboard ploughing have a contaminant dilution effect in the soil profile due to vertical contaminant migration. Contaminants adsorbed in the upper profile layers are buried (He et al., 2015), while deeper soil layers are exposed to wet-dry and freeze-thaw cycles on the surface, disrupting the soil structural organization (Urbanek, Horn and Smucker, 2014) and exposing contaminants previously migrated down the soil profile to microbial action and plant uptake. Li, Gong and Komatsuzaki (2019) reported mobilization of 137Cs in a soil profile under rotary tillage and homogenization with depth under mouldboard plough. However, in the particular case of 137Cs, tillage is considered a beneficial and effective agricultural practice to reduce the transfer of this contaminant to crops (Li, Gong and Komatsuzaki, 2019; Rosén, 1996; Whicker et al., 2007).
On the other hand, in the last decades, conservation tillage practices have been promoted on the basis of environmental benefits (Karayel and Šarauskis, 2019). Conservation tillage practices reduce the physical disruption of soil structure, but are frequently associated with surface compaction and emergence of weeds that could therefore result in an increased use of herbicides (Alletto et al., 2011). Tillage is also used to reduce the pest pressure from previous crops, so no-tillage may require increased use of fungicides, nematicides and other pesticides (Alletto et al., 2011). The absence or limitation of tillage also favours pore connectivity, which has been associated with higher leaching losses of veterinary drugs and other contaminants compared with tilled soils (Aust et al., 2010; Burkhardt and Stamm, 2007). Moreover, since conservation tillage enhances the content of SOM and is often defined by the use of a green mulch over at least 30 percent of the surface, increased adsorption and persistence of hydrophobic contaminants such as pesticides, antimicrobial agents or trace elements may occur (Alletto et al., 2011; Yang et al., 2017a). This reduces the contaminant bioavailability and mobility, but increases the persistence in the soil matrix, with the potential of further release if soil conditions change. Soil biodiversity indices (microbial biomass, diversity and activity, and the abundance and diversity of macro-organisms), however, are also improved with conservation practices, and contribute to the degradation of organic contaminants but immobilization of inorganic contaminants (Alletto et al., 2011; Derpsch et al., 2010; Yang et al., 2017a). Thus, the selection of appropriate practices is essential to ensure conservation of healthy soils in the long term, rather than increased pollution load (Yang et al., 2017a).
Certain irrigation practices increase drainage, alter redox conditions, and thereby influence soil contaminant fate and transport. Water quality influences the ionic strength of the soil solution and can change the hydraulic properties of the soil (Pontoni et al., 2016). For example, the leaching of veterinary drugs from contaminated manure has been observed to increase with increased irrigation rate (Aust et al., 2010). In addition, irrigation with sodic water produces degradation of the soil structure due to clay dispersion caused by the replacement of divalent cations (Ca2+ and Mg2+) by Na+. Contaminants bound to or occluded inside aggregates are therefore released. Rodríguez-Liébana, Mingorance and Peña (2014) found that the role of wastewater quality on desorption of pre-adsorbed pesticides in soils was significant, especially for hydrophobic pesticides.
Some agricultural soils, such as heavy clay soils or poorly draining soils found in areas that were formerly wetlands, require artificial drainage to increase crop productivity. These soils can retain high concentrations of contaminants due to the high content of 2:1 clays and soil organic carbon, and redox conditions, respectively (Rosolen et al., 2015). The drainage water can contain high concentrations of agrochemicals and other contaminants and becomes a source of diffuse pollution if not properly treated before release into the environment (Rieke et al., 2018; Tournebize, Chaumont and Mander, 2017). In addition, the drainage of agricultural soils leads to a more oxidizing environment (Chen et al., 2016), in which organic matter is mineralized faster (Ma et al., 2016). Electron acceptors such as sulphur or iron are also oxidized, changing potential sorption of contaminants (Streeter and Schilling, 2015). In general, soil pH decreases after drainage (Chen et al., 2016; Streeter and Schilling, 2015). The decrease in SOM and soil biodiversity and changes to the pH of drained agricultural soils can lead to mobilization of sorbed soil contaminants within the soil profile, and increased bioavailability for plant uptake.
Several organic amendments are used to stabilize trace elements and other contaminants. While biochar has been observed to efficiently sorb inorganic and organic contaminants, compost and peat can increase the concentration of dissolved organic carbon and organo-metallic colloids, increasing the bioavailability of trace elements (Egene et al., 2018; Park et al., 2016). However, well matured compost can reduce trace metal solubility due to chelation with humic substances (Clemente et al., 2015; Clemente and Bernal, 2006). Overall, mobilization and bioavailability of trace elements decreases in amended soils due to the positive influence of the newly added organic matter (Abad-Valle, Iglesias-Jiménez and Álvarez-Ayuso, 2017; Chen et al., 2016).
Lime is an alkaline material that effectively contributes to increasing soil pH. It is widely applied to acidic agricultural soils to improve soil fertility and crop productivity and as a remediation treatment for trace element polluted soils, contributing to the immobilization of many trace elements by precipitation (Chen et al., 2016; Zhu et al., 2016). However, lime amendment has a limited efficiency over time (Ruttens et al., 2010) and efficacy is highly site-specific (Yang et al., 2018).
Almost all sources of pollution are associated with several groups of contaminants so that few soils are polluted with a single contaminant. The interactions among contaminants also influences contaminant fate and transport in the soil environment.
Ionic contaminants can increase either the acidity or alkalinity of the soil solution. Soil organisms cannot metabolise organic contaminants if toxicity is sufficiently high and these contaminants become part of the organic carbon pool, and hence can retain other organic and inorganic contaminants, such as trace elements. This occurs by increasing the CEC, through surface complexation mechanisms, or by forming colloids or organometallic complexes that will behave differently to the original species (Zhu et al., 2018). Organic contaminants can also modify the repulsion of soil water (Yaron, Dror and Berkowitz, 2012). Similarly, the presence of trace elements may lead to the predominant formation of organo-metallic complexes instead of organic contaminant adsorption to the soil minerals. For example, the presence of copper in agricultural soils due to the spray of copper-based fungicides increases the half-life of some pesticides, such as pyrethroid insecticides and DDT (Gaw et al., 2003; Liu et al., 2007). This increase in pesticide half-life in soil is related to the inhibitory effect of copper on soil microbial activity, reducing the ability of soil organisms to degrade the organic contaminants (Gaw et al., 2003).
Studies on the impact of irrigating soils with wastewater, which contains a wide variety of contaminants, can provide good examples of the interactions among contaminants and influences on the fate of other contaminants. Wastewater is usually rich in dissolved organic matter and nutrients, but also in salts, trace elements and organic contaminants such as pharmaceuticals or microplastics. The increased salt content produces an increase in the competition for sorption sites, so trace elements are displaced to the soil solution, enhancing leaching potential to groundwater (Mapanda et al., 2005; Tarchouna et al., 2010). Microplastics can be associated with pathogens and other organisms or enriched in pesticides and trace elements, in addition to all the chemicals that form the plastic polymers and additives, for an increase in ecotoxicity (de Souza Machado et al., 2018).
Climate change aggravates the impacts of soil pollution (Figure 20). Climate change exacerbates the frequency and severity of extreme climatic events such as droughts, floods, and dust storms, and also increases the incidence of wildfires (IPCC, 2019). These extreme events contribute to changes in the moisture and temperature regimes of soils and groundwater (Feng, 2016). In addition, depending on the nature of the event, extreme weather can increase rates of movement of soil contaminants via soil erosion (wind or water), soil runoff, leaching and volatilization (Boxall et al., 2010; IPCC, 2019; Jarsjö et al., 2020).
Climate change also affects the distribution of living organisms and their biomass, induces community changes and alters crop and pest cycles (Classen et al., 2015; IPCC, 2019). These changes can alter contaminant biotransformation processes to produce more active metabolites or reduce the biomass of soil-dwelling organisms that degrade contaminants (Noyes et al., 2009).
It is widely accepted by the scientific community that the effects of climate change on soils under current use and management will lead to a decrease in SOM, especially in temperate and cold regions but also already depleted agricultural soils in warmer areas (Wiesmeier et al., 2016). Higher inputs of agrochemicals, both pesticides and fertilizers, may be required with the increasing effects of climate change, due to the loss of soil fertility and increased incidence of pests and pathogens in regions previously unaffected by certain pests from warmer environments (Boxall et al., 2010). Depletion of SOM, decrease in plant cover and loss of soil biodiversity can greatly contribute to soil contaminant mobilization (Ondrasek et al., 2019).