Organic Residues and Soils: History
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Subjects: Agriculture, Dairy & Animal Science
Contributor:
María Sonia Rodríguez-Cruz

The management of large volumes of organic residues generated in different livestock, urban, agricultural and industrial activities is a topic of environmental and social interest. The high organic matter content of these residues means that their application as soil organic amendments in agriculture is considered one of the more sustainable options, as it could solve the problem of the accumulation of uncontrolled wastes while improving soil quality and avoiding its irreversible degradation. However, the behavior of pesticides applied to increase crop yields could be modified in the presence of these amendments in the soil.

  • soil amendment
  • organic matter

1. Introduction

The use of large quantities of pesticides in today’s intensive agricultural systems is a widespread practice for controlling pests, diseases and weeds. This increases the yield per hectare, ensuring the food supply for the world’s ever-growing population [1][2], which currently stands at over 7.7 billion people, and is estimated to rise above 9.6 billion by 2050, and reach nearly 11 billion around 2100 [3]. The application of a wide range of pesticides is considered a regular and required practice in agriculture, as almost 45% of annual food production is lost due to pest infestation or the competition between crops and weeds for soil nutrients [4]. In fact, 3.5 million tons of pesticides are being used, of which 47.5% are herbicides, 29.5% are insecticides, 17.5% are fungicides, and 5.5% are other pesticides [5]. The global pesticide market recorded a value of nearly USD 84.5 billion in 2019, increasing at an annual growth rate of 4.2% since 2015, and it is likely to reach 11.5% with a value of nearly USD 130.7 billion by 2023 [6]. The ten countries consuming the most pesticide in the world are China, USA, Argentina, Thailand, Brazil, Italy, France, Canada, Japan, and India [7].

However, this extensive use of pesticides over recent decades is now of considerable environmental concern because of the release of mobile and/or persistent pollutants into the environment, and the potential accumulation of these toxic substances in soils and/or waters [8][9][10]. The fate of pesticides and their degradation products determines the contamination of the soil, water and air ecosystems over time. Moreover, if agrochemicals remain in the crops, they could finally enter the food chain, posing a threat to human, animal, and plant welfare [11][12][13][14].

The contamination of agricultural soils with pesticides could lead to changes in their chemical and biological properties, affecting their quality and causing a negative impact on crop yields [15]. They may impair soil microbial biodiversity and enzymatic activity (a vital indicator of soil tolerance to pollutants), and the associated degradation of soil organic matter (OM) [16][17]. Many reports are available on these negative effects on soil microbial communities [17][18], and on the processes associated with microbial activities [19].

A recent study involving 317 agricultural topsoil samples from the European Union and 76 pesticide residues as target compounds has revealed that 83% of the soils have been contaminated by one or more residues [9]. The contamination of surface and ground waters by pesticides has also been detected in recent years, probably due to deficient pesticide management, and increased by precipitation and/or irrigation that give rise to the runoff or leaching process of these compounds through the soil [20][21][22][23]. In fact, the contamination of water by pesticides is increasing in agricultural areas across different countries, and a broad range of pesticide concentrations has been found, in some cases exceeding the limit established for drinking water by European Union (EU) legislation (0.1 µg·L−1) [24][25][26].

These environmental contamination data highlight the need to roll out strategies to optimize agricultural sustainability by maximizing crop productivity and reducing or preventing soil and water contamination by pesticides. This has been widely addressed in recent years due to the requirement to meet European Community regulations [27]. One of these strategies is based on the in-situ application of organic residues as organic amendments [28]. This method is a common agricultural practice which allows increasing soil OM content, and it can be used to control soil and water contamination by pesticides: (i) promoting the immobilization of pesticides in soil OM, enhancing their subsequent biodegradation, and preventing or reducing their potential mobility into water resources [28][29][30], and (ii) delivering nutrients to the soil by increasing OM content to promote soil fertility and plant growth and stimulate ecological restoration with concomitant benefits for the health of the soil ecosystem [11]. In addition, organic materials require minimal pre-treatment before their application to the soil because of their biological origin [31].

Large amounts of organic residues are generated from livestock, urban, agricultural and industrial activities, and their management is a topic of environmental and social interest in many countries today due to the problems surrounding their disposal [32][33]. In general, these wastes have a high OM content, and they could be used as organic amendments in agriculture, with this being one of the most sustainable options and with greater environmental advantages. Moreover, numerous organic residues could perform as possible sorbents for pesticides [34][35][36]. These studies have assessed the effects that organic carbon (OC) from exogenous sources have on the behavior and environmental fate of pesticides in soils due to the affinity of pesticides, which are generally hydrophobic substances, by these organic materials. The OC of the amendments, depending on their nature, composition and content, can modify the main physicochemical processes of pesticides (adsorption–desorption, dissipation and leaching) in soils. These processes determine their efficiency as well as the dissipation or persistence of these compounds in the soil and their effects as potential environmental contaminants of the soil and surface or ground waters [37].

2. Organic Residues as Soil Amendments

2.1. Origin, Characteristics and Impact on Soil Properties

The use of organic residues as soil amendments to improve soil quality and fertility dates back thousands of years. Animal manure and human sewage was already being applied to the soil by Greeks and Romans [38]. Waste products such as crop residues, seashells, farmyard manure and others were previously used to enhance crop growth [39]. The use of organic residues as soil organic amendments is still a widespread and common practice in modern agriculture. In fact, this agronomic practice is on the rise in Europe, the US, and many other countries. Its application improves soil properties, maintaining soil health and productivity, while reducing the disposal of organic wastes into landfills, with the consequent environmental benefits [40]. The restoration and maintenance of soil OM content is one of the main benefits of the application of organic residues to agricultural soil because they add exogenous OM, which contributes greatly to soil fertility and long-term use [41]. OM is the main soil property, as it supports and interrelates the biological, chemical and physical dimensions of soil fertility and health [42]. Furthermore, because OM also contributes to the process of sequestering carbon dioxide from the atmosphere (soil C uptake, and thus climate change mitigation), its improvement has been championed at international forums on food security and climate change [43].
Regarding the benefits that organic amendments have for soil biological properties, they directly stimulate microbial growth and biomass by providing energy and essential nutrients (especially N, P, K, Ca and Mg), or indirectly by promoting plant growth, and consequently the amount of root exudates in the rhizosphere [44]. Moreover, the presence of diverse substrates susceptible to enzymatic hydrolysis within the amendments stimulates soil microbial activities [45]. Soil microbial diversity and composition could also be affected due to a higher availability of nutrients and growth substrates present in the amendments, which increase the number of ecological niches and promote a variety of ecological interactions, such as competition and/or antagonism between organisms [46]. The beneficial effects of organic amendments on the biomass, activity, and diversity of soil organisms have a long-term beneficial impact on soil health [47][48] and also contribute to different ecosystem services (C and nutrient cycling, disease suppression, etc.). However, it is important to stress that microbial responses to the application of organic amendments vary greatly depending on the nature and lability of the amendments’ OM [49].
Organic amendments also influence soil chemical properties in a positive way. Indeed, their favorable effects on soil microbial communities are often linked to changes in soil chemical characteristics [50][51]. Several organic amendments may have a direct effect on soil fertility by providing a wide variety of macro- and micro-nutrients, which support plant and microbial growth [52]. In addition, they may affect soil pH and enhance the cation exchange capacity, thus indirectly influencing nutrient availability, microbial activity and, therefore, soil fertility [53]. Nutrient availability may be influenced by the amendment’s biochemical composition, and in particular by its carbon–nitrogen (C/N) ratio, which may limit soil microbial growth and activity, thereby influencing the rate of OM decomposition and the patterns of nutrient release [49][54].
Soil physical characteristics can also be positively influenced by the application of organic amendments. Accordingly, the addition of exogenous OM directly improves soil structure (better porosity, aggregation and structural stability) [55] and water retention capacity [56], with the associated positive effects previously noted for soil performance and crop productivity. In turn, the stimulation of soil microbial communities through the use of organic amendments may also indirectly improve soil structure, as microbial activity (e.g., through the secretion of exopolysaccharides) and particularly hyphal growth significantly influence soil aggregation and aggregate stability [57]. The increase in soil porosity often reduces soil crusting and bulk density, thus favoring the movement of air and water through the soil matrix, the exploratory capacity of plant root systems, and the development of a suitable environment for soil biological communities [58]. Moreover, organic amendments influence particle size distribution, connectivity and the total surface area within the soil, increasing the number and types of available niches for biological colonization [49].
The potential positive effects of organic amendments on the soil ecosystem depend on many factors, such as their origin (forestry, farming, urban or industrial wastes, etc.), material stage (solid, semisolid, liquid), subjected or not to treatments (composting, anaerobic digestion, etc.), composition, stability, maturity, and application frequency, method and rate. Moreover, soil type, cropping system, and weather conditions are also important factors. In order to identify the different properties, agronomic potential, and limitations of any given organic amendment for soil and crop health, a thorough characterization of both the organic amendment and the agroecosystem itself needs to be performed before its application [49]. In this sense, however, it is also important to stress that the use of organic residues as a soil organic amendment is not devoid of risk. This agricultural practice may sometimes have unwanted effects on the environment depending on the factors previously cited (nature, origin, dose of application, etc.). Some of the potential negative effects analyzed and determined in the literature are the increase in soil electrical conductivity and its salinization, negative impacts on the sensitivity and resilience of soil bacteria communities, and the release of pollutants into the soil and/or waters (nitrate, heavy metals, antibiotics, polycyclic aromatic hydrocarbons, dioxins, PCBs, etc.) [51][59][60][61][62][63].

2.2. European Legislation on the Use of Organic Residues as Soil Amendments

Economic growth in the EU continues to increase the volume of residues generated. As a result, waste disposal has become a social and environmental concern because it causes the unnecessary loss of materials and energy, environmental damage, and negative effects on health and quality of life. Landfills, for example, occupy space and can pollute air, water and soil, while incineration leads to emissions of air pollutants. The long-term goal of EU waste management policies is to decrease the amount of residues generated, and when their generation is unavoidable, promote them as a resource and achieve higher levels of recycling and safe waste disposal, reducing the negative impacts on the environment and health and ushering in an efficient “Recycling Society” [64].
Two common EU targets for 2030 are recycling 65% of municipal solid waste (MSW) and reducing the corresponding landfill to a maximum of 10% of this figure [65]. In 2018, 5.2 tons of residues were generated per EU inhabitant, and 38.7% of waste in the EU was landfilled and 38.1% was recycled [64]. According to the European Compost Network [66], between around 118 and 138 million tons of bio-residues are generated annually across the EU, of which only about 40% (equivalent to 47.5 million tons/year) is efficiently recycled into high quality compost and digestate. Based on the Status 2019 report [66], a total of 47.5 million tons of bio-residues are treated in 4274 plants, and the predominant treatment process is composting.
Considering that up to 50% of MSW is organic, the bio-residue fraction has a significant role in recycling and developing the circular economy. Most of the MSW generated in the EU is still landfilled (24%) or incinerated (27%), and less than half is recycled (31%) or composted (17%) [64]. According to the data available, there has not been any increase in bio-residue recycling in recent years [67]. Moreover, waste management practices vary widely across EU member countries, with many continuing to send large amounts of MSWs to landfills.
Soil health and quality have been seriously compromised in recent years by constant changes in land use and the depletion of soil OM [68]. Indeed, approximately 45% of Europe’s topsoil (0–30 cm) has a low OM (<3.5%) content [69][70], and the soils in Mediterranean regions are highly susceptible to its loss, with almost 75% recording a low (≤2%) or very low (≤1%) OM content [71], whereby most of them are considered degraded. The reuse of organic residues as soil amendments in agriculture is an ancient but increasingly popular practice that not only helps to reduce the dependency on agrochemicals, but also constitutes an ecologically, economically, and socially acceptable alternative to landfill disposal and incineration, contributing at the same time to the objectives of the EU policies of “Zero Waste”, “End-Of-Waste”, and the “Circular Economy Strategy” [46][72].
It should nonetheless be stressed that an organic residue must meet a series of requirements for its potential use as a soil amendment. According to Commission Decision (EU) 2015/2099 of 18 November 2015 establishing the ecological criteria for the award of the EU Ecolabel for growing media, soil improvers and mulch [65], the following definitions apply: “1. Soil amendment means a fertilizer product incorporated into the soil in situ whose function is to maintain, improve or protect the physical or chemical properties, structure or biological activity of the soil, with the exception of limestone amendments, and 2. Organic soil amendment means a soil amendment that contains carbonaceous materials whose main function is to increase the OM content of the soil”. Among the organic residues potentially applicable to soil are those from the following activities: urban (sewage sludge (SS) or MSW), agricultural (crop residues), livestock (manure and slurry), and agro-industrial (wine, beer, sugar and olive production, and mushroom cultivation) [38].
The addition of OM through organic amendments plays a major role in the fate of xenobiotic compounds including pesticides [73][74]. The solid organic matter (SOM) and dissolved organic matter (DOM) of these amendments applied to the soil may modify the physicochemical behavior of pesticides (e.g., adsorption–desorption, persistence, bioavailability, degradation, and mobility), affecting soil quality and surface and ground waters [34][75][76][77][78][79] (Figure 1).
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Figure 1. Summary of the organic amendments’ (solid organic matter (SOM) and dissolved organic matter (DOM)) effects on the processes controlling the fate of pesticides in soil.

3. Effect of Organic Residues on the Fate of Pesticides in Soil

3.1. Effect of Organic Residues on the Adsorption-Desorption of Pesticides

The combined application of pesticides and organic residues in soils modifies the former’s physicochemical behavior, mainly through their adsorption-desorption by the amended soils [78][79]. Organic amendments increase soil OC content, and this parameter is the most relevant factor influencing the adsorption process and the affinity of hydrophobic pesticides by soils [80][81]. Adsorption-desorption determines the environmental fate of any organic pollutants in the soil-water environment [82]; it directly or indirectly controls the availability of pesticides to be transported to surface waters by runoff or to groundwaters by leaching, to the air by volatilization, to be degraded/transformed by microbial attack, or be taken up by plants [83]. Thus, the weak adsorption and/or strong desorption of pesticides promotes leaching, run-off, volatilization, biodegradation and even ecotoxicological impacts on non-target organisms, including human beings, while strong adsorption prevents losses of pesticides by such processes [84].
Accordingly, the addition of organic amendments to soil could lead to a greater or lesser degree of pesticide immobilization in the amended soil. This effect has consequences for pesticide degradation, persistence or mobility, enhancing a pesticide’s subsequent chemical, physical, and biological transformation or degradation, decreasing its transport through the soil profile, and consequently reducing groundwater pollution in some cases [28][78]. However, it could also affect the final concentration bioavailable for absorption by the targeted weeds [37]. Therefore, adsorption and desorption processes help to understand how to predict the mobility and availability of pesticides in unamended and amended soils. Numerous references report the ability that organic amendments have to adsorb pesticides [29][85][86][87].
Adsorption is a physicochemical process in which pesticide molecules are retained on a solid surface (especially by the soil colloidal fraction) within a solution through hydrophobic interactions, van der Waals forces, π–π interaction, and covalent, ionic or hydrogen bonds [88][89]. Soil OM and its more active components, humic acids (HAs) and fulvic acid (FAs), are the principal adsorbents for pesticides, followed by clay colloids and oxyhydroxides of iron and manganese, which interact with pollutants when they reach the soil [90][91]. HA is the OM fraction with the highest reactivity (determined by the number and type of functional groups) and largest surface area. OM’s highly variable composition means it can interact with neutral or ionizable molecules [75]. The pesticide adsorption capacity varies, in general, according to the physicochemical characteristics of the adsorbent and pesticide properties, mainly its water solubility and its hydrophobic, polar, or ionic character [37][92].
The nature and composition of the amendment’s OM vary, with the consequent difficulty in predicting its efficiency for adsorbing pesticides [34][76]. The addition of organic amendments to soil introduces not just SOM but also DOM (Figure 1). The influence of the SOM and DOM content of organic residues on the adsorption of pesticides with different characteristics by amended soils has been frequently studied [34][76][77]. The DOM content in unamended soil is usually very low, but it could become relevant if the organic amendment has a high content in this fraction [93]. DOM is a diverse mixture of complex compounds with different chemical structures and molecular weights that might enhance the formation of multiple interactions with organic pesticides, controlling their distribution in the soil [94]. This is why DOM may modify the movement of pesticides, generally decreasing their adsorption by SOM and increasing their leaching, leading to groundwater contamination [78][95][96], although other authors have indicated that DOM could also be adsorbed by the soil, increasing the adsorption of pesticides and decreasing their leaching [97].
Different processes have been proposed to explain the decreased adsorption of pesticides in the presence of DOM [94][98][99][100] (Figure 1). These include the competition between DOM and pesticide molecules for the adsorption sites in soil, the saturation of soil adsorption sites by DOM, masking these sites for the adsorption of pesticides, the co-sorption of pesticides by DOM, and the formation of mobile DOM-pesticides complexes. Some authors have also indicated that DOM has characteristics similar to surfactants with the capacity to decrease surface tension and increase the solubility of pesticides, reducing their adsorption [101]. The extent and nature of DOM-pesticide interactions depend on factors such as pesticide molecular weight and polarity [102]. These relative effects of DOM will be greater for more hydrophobic chemicals, and will be influenced by the concentration, source, size, polarity, and molecular configuration of the organic colloids [95].
Table 1 includes a summary of the main results obtained from the recent literature on the adsorption and/or desorption of pesticides by some organic residues and by soils amended with different organic residues used as amendments. An organic material widely studied for these purposes is biochar (BC). It is an efficient adsorbent and a potential material for soil amendment [30]. It is a carbonaceous and porous product generated from the partial combustion of biomass, and its effects as soil amendment in the adsorption–desorption of pesticides has been assessed from different points of view. The effects of different types of BC, treatments or aging periods, and their different doses or application forms as organic amendments in soils have been reported for pesticides with different characteristics. Parlavecchiaetal. [103] have investigated the effect of two types of BC from grape vine pruning residues (BC-G) and spruce wood (BC-S) and two vermicomposts (VC) involving digestates from a mixture of manure and olive mill wastewater (VC-M) and buffalo manure (VC-B) in the sorption-desorption capacity of the fungicide metalaxyl-M. Both types of amendments (BC and VC) have a significant capacity to adsorb the high-water-soluble fungicide. However, BC has recorded a much higher sorption efficiency than VC and lower desorption, which is explained by the composition and structural differences in OM between the two (VC has less aromatic carbon and a higher content of hydrophilic functional groups interacting with polar compounds and solvents than BC). Metalaxyl-M is adsorbed to a similar extent on the two VCs, while a different sorption behavior is observed in BC-G and BC-S due to their different porous structures. Likewise, Wu et al. [89] have assessed the effects of different types of BC from peanuts (BCP), chestnuts (BCC), bamboo (BCB), maize straw (BCM), and rice husk (BCR), and the effects of BCR aging on the sorption, degradation and bioavailability of the herbicide oxyfluorfen in various amended soils. The sorption capacity of the five BC differs significantly due to their physicochemical properties. The sorption capacity of BC for oxyfluorfen is significantly correlated with the specific surface area and elemental composition, but it decreases with longer aging time. BC reduces the bioavailability of oxyfluorfen in amended soils, but a higher bioavailability is recorded with an increase in the aging period of BC. Nevertheless, the sorption capacity of amended soil for oxyfluorfen after six months is still better than the unamended soil, highlighting that BCR is an effective way of reducing the risk of contaminating soil with oxyfluorfen, although it could also diminish the herbicide’s bioavailability and efficacy. Deng et al. [104] have studies the effect that BC obtained from cassava residues at 750 °C (MS750) applied at different rates between 0% and 5% has on the sorption-desorption and mobility of atrazine. The MS750 application significantly enhances the sorption capacity and decreases the sorption reversibility of atrazine in the amended soil compared to the unamended soil, due to the larger surface area and greater aromaticity of MS750 (with favorable sorption domains for organic compounds). Moreover, sorption affinity increases with higher BC application rates, although it is also influenced by solution pH, ambient temperature, and contact time between soil and BC (equilibrium time). The entrapment of atrazine in micropore or pore deformation could lead to desorption hysteresis in BC-amended soils.
Table 1. Adsorption–Desorption of pesticides by organic residues and amended soils.
Pesticide Soil Characteristics Organic Amendment/Dose Experimental Design Results Reference
Metalaxyl-M Silt loam soil
(pH 6.70, OC 2.90%)		 Biochar from grape vine pruning residues (BC–G) (pH 9.9, OC 75.1%) and spruce wood (BC–S) (pH 9.1, OC 83.8%). Vermicomposts (VC) from manure and olive mill wastewater (VC–M) (pH 7.9, OC 31.6%) and buffalo manure (VC–B) (pH 7.8, OC 36.6%)
Biochar/soil: 2% (w w–1)		 Sorbent/Solution: 25 mg biochar/5 mL
or 3 g soil/8 mL water solution
Herbicide concentration:
1–20 mg L−1
Shaken: 24 h, T: 20 °C
Analytical determination: HPLC		 Metalaxyl sorption order: non–amended soil < soil–VC–M ≤ soil–VC–B < soil–BC–S < soil–BC–G
Much higher sorption efficiency by BC than by VC and a lower extent of metalaxyl desorption due to composition and structural differences of the organic matter of BC.		 Parlavecchia et al. [103]
Oxyfluorfen Loamy clay soil
(pH 4.85, OC 0.84%)
Sandy loam soil
(pH 7.55, OC 0.98%)
Clay loam soil
(pH 6.59, OC 2.23%)		 Biochar from peanut (BCP) (pH, 7.05, C 49.17%), chestnut (BCC) (pH 6.08, C 58.07%), bamboo (BCB) (pH 7.45, C 63.25%), maize straw (BCM) (pH 6.83, C 43.36%), rice hull (BCR) (pH 6.96, C 33.60%)
BCR/soil: 0.5%, 1%, or 2% (w w–1)		 Sorbent/Solution: 0.1 g biochar/40 mL
or 2 g soil/200 mL 0.01 M CaCl2
Herbicide concentration: 0.05–10 mg L−1
Shaken: 6 days, T: 25 °C
Aging time of BCR-soil: 1, 3, 6 months
Analytical determination: GC/MS		 BC sorption capacities followed the order: BCR > BCB > BCM > BCC > BCP owing to differences in physicochemical properties.
BCR sorption capacity decreased with aging time.		 Wu et al. [89]
Atrazine Krasnozem soil
(pH 7.05, OC 0.89%, clay 28.2%, silt 37.8%)		 Biochar from cassava wastes (pH 9.55, C 62.38%) obtained at 750 °C (MS750). SSA: 430.4 m2/g, MP: 0.144 m3/g
Biochar/soil:
0%, 0.1%, 0.5%, 1%, 3% and 5% (w w–1)		 Sorbent/Solution: 0.2–2 g/10 mL 0.01 M CaCl2
Herbicide concentration:
0.5–20 mg L−1
Shaken: 24 h
T: 15, 25, 35 °C, pH: 3,5, 7, 9
Analytical determination: HPLC		 Great sorption capacity for atrazine of MS750 in soil due to high surface area and micropore volume. High degrees of aromaticity and hydrophobicity (H/C: 0.02, N + O/C: 0.09) of MS750 supplied numerous sorption sites. Deng et al. [104]
Hexazinone
Metribuzin
Quinclorac		 Sandy loam soil
(pH 6.9, OC 0.52%, clay 15.1%, silt 3.3%)		 Bone char (BC) (pH 9.72, C 11%)
BC/soil: 5% (w w–1) or 60 t ha−1		 Sorbent/Solution: 10 g/10 mL 0.01 M CaCl2
Herbicide concentration:
0.63–3.13 mgL−1, 1.60–8 mgL−1, 0.31–1.56 mgL−1
Shaken: 24 h, T: 20 °C
Analytical determination: Liquid scintillation		 High sorption of herbicides by BC, regardless of the application form of the material (topsoil or incorporated in the surface layer in leaching columns). Mendes et al. [105]
Aminocyclopyrachlor
Mesotrione		 Clay soil
(pH 6.44, OC 2.73%), clay 50.9%, silt 19.6%)		 Bone Char (BC) (pH 9.72, C 11%)
BC/soil: 0%, 1%, 5%, 10%, and 100% (w w−1) or 0, 12, 60, 120, and 1200 t ha−1
BC particle size groups: 0.3–0.6 and 0.15–0.3 mm		 Sorbent/Solution: 10 g/10 mL 0.01 M CaCl2
Herbicide concentration:
0.051 mg L−1 (0.32 Bq L−1) aminocyclopyrachlor
5.0 mg L−1 (1.13 Bq L−1) mesotrione
Shaken: 24 h, T: 20 °C
Analytical determination: Liquid scintillation		 Higher BC rates (regardless of the particle size) increased both herbicides adsorption and decreased their desorption. Mendes et al. [106]
Linuron
Alachlor
Metalaxyl		 Sandy loam soil
(pH 6.3, OC 0.51%, clay 11.8, silt 13.6%),
Sandy clay soil (pH 6.9, OC 1.04%, clay 38.1%, silt 5.8%)		 Pine Wood (OC 41.6%, DOM 1.62%, lignin 24.4%), oak wood (OC 38.5%, DOM 6.86%, lignin 18.2%)
Wood/soil: 5% and 50% (w w–1) (40 and 400 t C ha–1)		 Sorbent/Solution: 5 g/10 mL water solution
Herbicide concentration:
1–25 mg L−1 (100 kBq L−1)
Shaken: 24 h, T: 20 °C
Incubation times: 0, 5 and 12 months
Analytical determination: Liquid scintillation		 Pesticide adsorption increased with high wood dose but OC nature was not relevant. Adsorption did not change after incubation times. The adsorption irreversibility decreased in presence of wood for alachlor and increased that of linuron and metalaxyl. Marín–Benito et al. [107]
Aminocyclopyrachlor
Mesotrione		 Clay soil
(pH 6.0, OC 2.21%, clay 60.5%, silt 11.3%)		 Sewage sludge (SS) (pH 6.8, OC 16.64%)
SS/soil: 0.1%, 1%, and 10% (w∙w–1) or 1.2, 12, and 120 t∙ha–1		 Sorbent/Solution: 10 g/10 mL 0.01 M CaCl2
Herbicide concentration:
0.08–0.64 Bq·L−1 (aminocyclopyrachlor)
0.28–2.27 Bq·L−1 (mesotrione)
Shaken: 24 h, T: 20 °C
Analytical determination: Liquid scintillation		 SS slightly affected sorption–desorption of both herbicides (lowest Kd at soil-SS1%). Kd for mesotrione was ~3.5–fold higher than for aminocyclopyrachlor (higher water solubility). Mendes et al. [108]
Imazapic
Atrazine
Hexazinone
Diuron
Metribuzin		 Red Ferrusol
(pH 7.1, OC 2.1%, clay 41%,
silt 23%),
Grey Dermosol (pH 5.7, OC 0.9%, clay 30%, silt 22%),
Red Kandosol (pH 6.5, OC 3.5%, clay 22%, silt 8%)		 Eleven mill muds/ash from different sugar mills (pH 6.04–7.26, OC 27.7–37.8%)
Mill muds/soil: 5–25% (w w–1)		 Sorbent/Solution: 1 g/5 mL 0.01 M CaCl2
Herbicide concentration:
0.5 mg L−1
Shaken: 24 h, T: 25 °C
Analytical determination: Q-TOF		 Sorption order: diuron > atrazine = metribuzin > hexazinone = imazapic (consistent with herbicide properties). Mill muds at 5% dose increased herbicide retention up to tenfold. Amendments reduced desorption of mobile herbicides in low OC soils. Duhan et al. [109]
MCPA
Diuron
Clomazone
Terbuthylazine		 Sandy loam soil
(pH 7.93, OC 0. 54%, clay 6.7%, silt 16.8%)
Loam soil (pH 6.77, OC 1.77%, clay 22.1%, silt 34.2%) Clay loam soil (pH 8.14, OC 1.38%, clay 31.1%, silt 26.8%)		 Mucilage extracted from chia seeds (Salvia hispanica L.)
Organic residue/soil: 10% (w w–1)		 Sorbent/Solution: 0.5 g unamended
or amended soil/8 mL water solution
Herbicide concentration: 1 mg L−1
Shaken: 24 h, T: 20 °C
Analytical determination: HPLC		 Soil porosity decreased by mucilage amendment. Sorption of herbicides increased in amended soils (sandy–loam < loam < clay–loam). Diuron recorded the highest Kd value and desorption was observed only for terbuthylazine. Marsico et al. [110]
Dichlorvos
Chlorpyrifos		 Sandy soil (pH 8.52, OC 0.7%, clay + silt 9.3%) Compost (C) from mixed wastes (pH 6.61, OC 29.5%, DOM 354 mg L−1), and dried goat organic manure (OM) (pH 8.67, OC 14.4%, DOM 620 mg L−1)
Organic residues/soil: 2.5 and 5% (w w–1)		 Sorbent/Solution: 5 g soil/100 mL in
C-DOM or 0.01 M CaCl2
Herbicide concentration:
0.1–10 mg L−1 (chlorpyrifos)
0.25–100 mg L−1 (dichlorvos)
Shaken: 24 h, T: 25 °C
Analytical determination: GC		 C–and OM–DOM increased dichlorvos sorption (S < S–OM–DOM< S–C–DOM) and decreased chlorpyrifos sorption (S > S–C–DOM> S–OM–DOM). Humified and aromatic nature of DOM determines the interactions with pesticides with different hydrophobic character. Gaonkar et al. [111]
Triasulfuron
Prosulfocarb
Chlorotoluron
Flufenacet		 Sandy loam soil (pH 7.36, OC 1.20%, clay 17%, silt 25%)
Loamy sand soil (pH 7.61, OC 0.9%, clay 13%, silt 6%)		 Spent mushroom substrate (pH 7.9, C 26.4%, DOM 1.29%), green compost (pH 7.2, C 23.6%, DOM 0.69%), manure (C 18.5%, DOM 1.32%), sewage sludge (pH 7.6, C 28.9%, DOM 1.18%)
Organic residues/soils: 10% (w w–1)		 Sorbent/Solution: 5 g soil or 0.1 g
organic residues/10 mL 0.01 M CaCl2
Herbicide concentration:
1–25 mg L−1 (TSF, CTL, FNC)
0.25–10 mg L−1 (100 Bq mL−1) (PSC)
Shaken: 24 h, T: 20 °C
Analytical determination: HPLC/MS
and Liquid scintillation		 Highest adsorption for prosulfocarb (lowest water solubility and highest Kow) in all materials. Aliphatic and aromatic structures optimize adsorption and O-alkyl and N-alkyl groups enhance desorption hysteresis.
Mendes et al. [105] have recently studied the effect that BC from cow bone applied to the topsoil or incorporated into the surface layer has on the sorption-desorption of the herbicides hexazinone, metribuzin, and quinclorac in an unamended soil, pure BC, and BC-amended soil under laboratory conditions. The results indicate low values of Kf adsorption and desorption constants in the unamended soil. BC increases these Kf values, stimulating the retention of all the herbicides in the surface soil. The low C content of BC has a minimal impact on the total OC of amended soils, with the main changes occurring in pore size (up to 60,000 nm), volume (0.225 cm3 g−1), and area (133 m2 g−1) to increase herbicide adsorption by the soil. The desorption data for all herbicides were consistent with the values found for adsorption in the unamended soil, where quinclorac was the herbicide with the highest Kfoc (adsorption) and lowest Kfoc (desorption). The desorbed amount of herbicides was close to zero after the addition of BC in the soil or in pure BC, confirming the high adsorption potential of BC regardless of the material’s method of application. Similar high sorption has been reported for two weak acid herbicides, aminocyclopyrachlor and mesotrione, in amended soils with the same cow bone char [106]. The comparison of organic residue application at varying rates and with two different particle size groups (0.3–0.6 and 0.15–0.3 mm) indicates that higher BC application rates increases the adsorption and decreases the desorption of both herbicides, regardless of particle size.
The effectivity and rate of application of other organic residues as adsorbents of pesticides has also been reported in different studies. Marín-Benito et al. [107] have studied the effect of large amounts of lignocellulosic residues from forestry and industrial activities on the adsorption–desorption of certain pesticides by soils. The study involves two wood wastes (pine and oak wood) at two different doses (5% and 50%) and various incubation times (0, 5 and 12 months) in two soils with different textures (sandy loam and sandy clay). The effect on the adsorption–desorption of two herbicides and one fungicide (linuron, alachlor, and metalaxyl) has revealed that the application of oak or pine wood to soils increases the adsorption of linuron and metalaxyl by both soils, and of alachlor by the sandy loam soil at a lower dose (5%), while the adsorption of the three pesticides increases under all conditions at the highest dose (50%). The results also indicate the influence of soil type on alachlor desorption and/or its possible bioavailability from wood-soils, but not for linuron and metalaxyl, although this behavior changes with incubation time. The role of the nature of the OC (Koc values) for sorption has been evidenced for alachlor and metalaxyl, but not for linuron. Other residues, such as the SS applied to the soil at various rates (0.1%, 1%, and 10% w∙w–1), have a non-significant effect on the sorption–desorption of aminocyclopyrachlor and mesotrione [108]. Both herbicides follow a similar adsorption behavior in all treatments, although Kd for mesotrione is ≈3.5-fold higher than for aminocyclopyrachlor due to the latter’s higher water solubility. This leads to a higher bioavailability of aminocyclopyrachlor in soil solution for its absorption by weeds and crops.
Fewer studies have addressed the adsorption–desorption of several pesticide-organic residue combinations or the amendment effect on soil physical properties. Duhan et al. [109] have studied the behavior of five herbicides commonly used in sugarcane production (imazapic, atrazine, hexazinone, diuron, and metribuzin) by eleven waste materials (mill muds) and by three soils amended with them at different rates (5–25%, dry weight basis). The authors have observed that all the amendments enhance the adsorption efficiency for four of the five herbicides, depending on the rate of application, especially in the soil with low OC. Even at the lowest application rate, the adsorption of the herbicides increases from two to ten times. Mill muds in soil also reduce the rate and extent of herbicide desorption, especially at a 5% application rate and for mobile herbicides such as metribuzin and atrazine. Marsico et al. [110] have studied the effect that the mucilage extracted from Chia seeds (Salvia hispanica L.) has as a soil amendment on soil physical properties and on the sorption-desorption behavior of four herbicides (MCPA, diuron, clomazone, and terbuthylazine) used in cereal crops. The assessment of the changes in the microstructural characteristics caused by the reactions between the mucilage and soil particles in three soils indicates that mucilage amendment reduces soil porosity due to a decrease in larger pores (radius > 10 μm) and a significant increase in finer pores (radius < 10 μm), as well as in particle surface. Higher herbicide adsorption has been observed in the amended soils than in the unamended ones. Moreover, herbicide desorption is severely inhibited in the amended soils.
Although many organic amendments have proven to be effective adsorbents of pesticides, only a few studies have evaluated the functional groups involved in the adsorption process [112][111][113][114]. Accordingly, Gaonkar et al. [111] have used spectroscopy to characterize the DOM from two organic amendments (mixed waste compost and dried goat manure) and the amended soils, and assess their influence on the sorption of the insecticides dichlorvos and chlorpyrifos. The DOM contained large amounts of highly humified and aromatic molecules. DOM led to a non-significant increase in dichlorvos adsorption (hydrophilic pesticide), due mainly to the additional sites provided by the adsorbing DOM and no interactions between DOM and the insecticide in solution. However, a significant reduction in chlorpyrifos adsorption (hydrophobic pesticide) was observed, probably due to interactions between DOM and the insecticide mostly in solution, and to some extent at the soil/solution interface, increasing the solubilization of chlorpyrifos. This reduction in adsorption depended on the nature and concentration of the DOM, as well as on insecticide properties. In agreement with the adsorption results, chlorpyrifos desorption was significantly increased by the DOM residue. In another recent study, García-Delgado et al. [114] have determined the OC functional groups from four organic amendments (spent mushroom substrate (SMS), GC, manure, and SS) by elemental analysis and 13C-NMR, and their effects on the adsorption of four herbicides with different structures (triasulfuron, chlorotoluron, flufenacet, and prosulfocarb) by two unamended and amended soils with different textures. The chemical composition and structure of the organic amendments (especially OC content and structural C type), and external factors such as herbicide polarity (hydrophobicity) and soil properties controlled the adsorption process. The adsorption of herbicides was promoted by carbon-rich organic amendments with aliphatic and aromatic structures, while the irreversible adsorption (hysteresis) of herbicides in the amended soils was enhanced by the abundance of O-alkyl and N-alkyl groups of organic amendments.

This entry is adapted from the peer-reviewed paper 10.3390/environments8040032

References

  1. Imfeld, G.; Vuilleumier, S. Measuring the effects of pesticides on bacterial communities in soil: A critical review. Eur. J. Soil Biol. 2012, 49, 22–30.
  2. Nyamwasa, I.; Li, K.; Rutikanga, A.; Rukazambuga, D.N.T.; Zhang, S.; Yin, J.; Ya-Zhong, C.; Zhang, X.X.; Sun, X. Soil insect crop pests and their integrated management in East Africa: A review. Crop Prot. 2018, 106, 163–176.
  3. United Nations. Global Issues. Our Global Population. 2020. Available online: (accessed on 10 September 2020).
  4. Ahmad, A.; Shahid, M.; Khalid, S.; Zaffar, H.; Naqvi, T.; Pervez, A.; Bilal, M.; Ali, M.A.; Abbas, G.; Nasim, W. Residues of endosulfan in cotton growing area of Vehari, Pakistan: An assessment of knowledge and awareness of pesticide use and health risks. Environ. Sci. Pollut. 2018, 26, 20079–20091.
  5. Sharma, A.; Kumar, V.; Shahzad, B.; Tanveer, M.; Preet, G.; Sidhu, S.; Handa, N. Worldwide pesticide usage and its impacts on ecosystem. SN Appl. Sci. 2019, 1, 1–16.
  6. Pesticide Reports. Pesticides Global Market Opportunities and Strategies to 2023. 2020. Available online: (accessed on 10 September 2020).
  7. Worldatlas. 2018. Available online: (accessed on 10 September 2020).
  8. Hvězdová, M.; Kosubová, P.; Košíková, M.; Scherr, K.E.; Šimek, Z.; Brodský, L.; Šudoma, M.; Škulcová, L.; Sáňka, M.; Svobodová, M.; et al. Currently and recently used pesticides in Central European arable soils. Sci. Total Environ. 2018, 613–614, 361–370.
  9. Silva, V.; Mol, H.G.J.; Zomer, P.; Tienstra, M.; Ritsema, C.J.; Geissen, V. Pesticide residues in European agricultural soils—A hidden reality. Sci. Total Environ. 2019, 653, 1532–1545.
  10. Picó, Y.; Alvarez-Ruiz, R.; Alfarhan, A.H.; El-Sheikh, M.A.; Alshahrani, H.O.; Barceló, D. Pharmaceuticals, pesticides, personal care products and microplastics contamination assessment of Al-Hassa irrigation network (Saudi Arabia) and its shallow lakes. Sci. Total Environ. 2020, 701, 135021.
  11. Vangronsveld, J.; Herzig, R.; Weyens, N.; Boulet, J.; Adriaensen, K.; Ruttens, A.; Thewys, T.; Vassilev, A.; Meers, E.; Nehnevajova, E.; et al. Phytoremediation of contaminated soils and groundwater: Lessons from the field. Environ. Sci. Pollut. 2009, 16, 765–794.
  12. Sánchez-González, S.; Pose-Juan, E.; Herrero-Hernández, E.; Álvarez-Martín, A.; Sánchez-Martín, M.J.; Rodríguez-Cruz, S. Pesticide residues in groundwaters and soils of agricultural areas in the Águeda River Basin from Spain and Portugal. Int. J. Environ. Anal. Chem. 2013, 93, 1585–1601.
  13. Pose-Juan, E.; Sánchez-Martín, M.J.; Andrades, M.S.; Rodríguez-Cruz, M.S.; Herrero-Hernández, E. Pesticide residues in vineyard soils from Spain: Spatial and temporal distributions. Sci. Total Environ. 2015, 514, 351–358.
  14. Herrero-Hernández, E.; Pose-Juan, E.; Sánchez-Martín, M.J.; Andrades, M.S.; Rodríguez-Cruz, M.S. Intra–annual trends of fungicide residues in waters from vineyard areas in La Rioja region of northern Spain. Environ. Sci. Pollut. Res. 2016, 23, 22924–22936.
  15. AL-Ahmadi, M.S. Pesticides, anthropogenic activities, and the health of our environment safety. IntechOpen 2019.
  16. Baxter, J.; Cummings, S.P. The degradation of the herbicide bromoxynil and its impact on bacterial diversity in a top soil. J. Appl. Microbiol. 2008, 104, 1605–1616.
  17. Arora, S.; Sahni, D.; Sehgal, M.; Srivastava, D.; Singh, A. Pesticides use and its effect on soil bacteria and fungal populations; microbial biomass carbon and enzymatic activity. Curr. Sci. 2019, 116, 643–649.
  18. Satapute, P.; Kamble, M.V.; Adhikari, S.S.; Jogaiah, S. Influence of triazole pesticides on tillage soil microbial populations and metabolic changes. Sci. Total Environ. 2019, 651, 2334–2344.
  19. Zhang, Q.; Saleem, M.; Wang, C. Effects of biochar on the earthworm (Eisenia foetida) in soil contaminated with and/or without pesticide mesotrione. Sci. Total Environ. 2019, 671, 52–58.
  20. Li, Y.; Niu, J.; Shen, Z.; Zhang, C.; Wang, Z.; He, T. Spatial and seasonal distribution of organochlorine pesticides in the sediments of the Yangtze Estuary. Chemosphere 2014, 114, 233–240.
  21. Herrero-Hernández, E.; Rodríguez-Cruz, M.S.; Pose-Juan, E.; Sánchez-González, S.; Andrades, M.S.; Sánchez-Martín, M.J. Seasonal distribution of herbicide and insecticide residues in the water resources of the vineyard region of La Rioja (Spain). Sci. Total Environ. 2017, 609, 161–171.
  22. Sousa, J.C.G.; Ribeiro, A.R.; Barbosa, M.O.; Pereira, M.F.R.; Silva, A.M.T. A review on environmental monitoring of water organic pollutants identified by EU guidelines. J. Hazard. Mater. 2018, 344, 146–162.
  23. Barizon, R.R.M.; Figueiredo, R.O.; de Souza Dutra, D.R.C.; Regitano, J.B.; Ferracini, V.L. Pesticides in the surface waters of the Camanducaia River watershed, Brazil. J. Environ. Sci. Health B 2020, 55, 283–292.
  24. Kapsi, M.; Tsoutsi, C.; Paschalidou, A.; Albanis, T. Environmental monitoring and risk assessment of pesticide residues in surface waters of the Louros River (N.W. Greece). Sci. Total Environ. 2019, 650, 2188–2198.
  25. Herrero-Hernández, E.; Simón-Egea, A.B.; Sánchez-Martín, M.J.; Rodríguez-Cruz, M.S.; Andrades, M.S. Monitoring and environmental risk assessment of pesticide residues and some of their degradation products in natural waters of the Spanish vineyard region included in the Denomination of Origin Jumilla. Environ. Pollut. 2020, 264, 114666.
  26. Marsala, R.Z.; Capri, E.; Russo, E.; Bisagni, M.; Colla, R.; Lucini, L.; Gallo, A.; Suciu, N.A. First evaluation of pesticides occurrence in groundwater of Tidone Valley, an area with intensive viticulture. Sci. Total Environ. 2020, 736, 139730.
  27. Directive 2008/105/EC of the European Parliament and of the Council of 16 December 2008 on environmental quality standards in the field of water policy, amending and subsequently repealing Council Directives 82/176/EEC, 83/513/EEC, 84/156/EEC, 84/491/EEC, 86/280/EEC and amending Directive 2000/60/EC of the European Parliament and of the Council. Off. J. Eur. Union L 2008, 348, 84–97.
  28. Marín-Benito, J.M.; Sánchez-Martín, M.J.; Rodríguez-Cruz, M.S. Impact of spent mushroom substrates on the fate of pesticides in soil, and their use for preventing and/or controlling soil and water contamination: A review. Toxics 2016, 4, 17.
  29. Álvarez-Martín, A.; Rodríguez-Cruz, M.S.; Andrades, M.S.; Sánchez-Martín, M.J. Application of a biosorbent to soil: A potential method for controlling water pollution by pesticides. Environ. Sci. Pollut. Res. 2016, 23, 9192–9203.
  30. Khorram, M.S.; Zhang, Q.; Lin, D.; Zheng, Y.; Fang, H.; Yu, Y. Biochar: A review of its impact on pesticide behavior in soil environments and its potential applications. J. Environ. Sci. 2016, 44, 269–279.
  31. Barrow, C.J. Biochar: Potential for countering land degradation and for improving agriculture. Appl. Geogr. 2012, 34, 21–28.
  32. Dickerson, G. A Sustainable Approach to Recycling Urban and Agricultural Organic Wastes; Guide H–159, Cooperative Extension Service; College of Agriculture and Home Economics: Belen, NM, USA, 2000.
  33. Moral, R.; Moreno-Caselles, J.; Perez-Murcia, M.; Perez-Espinosa, A.; Rufete, B.; Paredes, C. Characterization of the organic matter pool in manures. Bioresour. Technol. 2005, 96, 153–158.
  34. Marín-Benito, J.M.; Andrades, M.S.; Rodríguez-Cruz, M.S.; Sánchez-Martín, M.J. Changes in the sorption-desorption of fungicides over time in an amended sandy clay loam soil under laboratory conditions. J. Soils Sediments 2012, 12, 1111–1123.
  35. Rodríguez-Cruz, M.S.; Herrero-Hernández, E.; Ordax, J.M.; Marín-Benito, J.M.; Draoui, K.; Sánchez-Martín, M.J. Adsorption of pesticides by sewage sludge, grape marc, spent mushroom substrate and by amended soils. Int. J. Environ. Anal. Chem. 2012, 92, 933–948.
  36. Castillo, J.M.; Beguet, J.; Martin-Laurent, F.; Romero, E. Multidisciplinary assessment of pesticide mitigation in soil amended with vermicomposted agroindustrial wastes. J. Hazard. Mater. 2016, 304, 379–387.
  37. Worrall, F.; Fernandez-Perez, M.; Johnson, A.C.; Flores-Cesperedes, F.; Gonzalez-Pradas, E. Limitations on the role of incorporated organic matter in reducing pesticide leaching. J. Contam. Hydrol. 2001, 49, 241–262.
  38. Goss, M.J.; Tubeileh, A.; Goorahoo, D. A review of the use of organic amendments and the risk to human health. Adv. Agron. 2013, 120, 275–379.
  39. Scotti, R.; Bonanomi, G.; Scelza, R.; Zoina, A.; Rao, M.A. Organic amendments as sustainable tool to recovery fertility in intensive agricultural systems. J. Plant. Nutr. Soil Sci. 2015, 15, 333–352.
  40. Medina, E.; Paredes, C.; Bustamante, M.A.; Moral, R.; Moreno-Caselles, J. Relationships between soil physico-chemical, chemical and biological properties in a soil amended with spent mushroom substrate. Geoderma 2012, 173–174, 152–161.
  41. Lugato, E.; Bampa, F.; Panagos, P.; Montanarella, L.; Jones, A. Potential carbon sequestration of European arable soils estimated by modelling a comprehensive set of management practices. Glob. Change Biol. 2014, 20, 3557–3567.
  42. Hijbeek, R.; Van Ittersum, M.K.; ten Berge, H.F.; Gort, G.; Spiegel, H.; Whitmore, A.P. Do organic inputs matter—A meta–analysis of additional yield effects for arable crops in Europe. Plant Soil 2017, 411, 293–303.
  43. UNFCCC, United Nations Framework Convention on Climate Change. Join the 4/1000 Initiative. In Soils for Food Security and Climate; UNFCCC: Rio de Janeiro, Brazil; New York, NY, USA, 2015; Available online: (accessed on 10 September 2020).
  44. Donn, S.; Wheatley, R.E.; McKenzie, B.M.; Loades, K.W.; Hallett, P.D. Improved soil fertility from compost amendment increases root growth and reinforcement of surface soil on slope. Ecol. Eng. 2014, 71, 458–465.
  45. Singh, R.P.; Singh, P.; Ibrahim, M.H.; Hashim, R. Land Application of sewage sludge: Physicochemical and microbial response. Rev. Environ. Contam. Toxicol. 2011, 214, 41–61.
  46. Gómez-Sagasti, M.T.; Hernández, A.; Artetxe, U.; Garbisu, C.; Becerril, J.M. How Valuable Are Organic Amendments as Tools for the Phytomanagement of Degraded Soils? The Knowns, Known Unknowns, and Unknowns. Front. Sustain. Food Syst. 2018, 2, 68.
  47. Fließbach, A.; Oberholzer, H.R.; Gunst, L.; Mäder, P. Soil organic matter and biological soil quality indicators after 21 years of organic and conventional farming. Agric. Ecosyst. Environ. 2007, 118, 273–284.
  48. Montiel-Rozas, M.M.; Domínguez, M.T.; Madejón, E.; Madejón, P.; Pastorelli, R.; Renella, G. Long-term effects of organic amendments on bacterial and fungal communities in a degraded Mediterranean soil. Geoderma 2018, 332, 20–28.
  49. Urra, J.; Alkorta, I.; Garbisu, C. Potential benefits and risks for soil health derived from the use of organic amendments in agriculture. Agronomy 2019, 9, 542.
  50. Li, R.; Khafipour, E.; Krause, D.O.; Entz, M.H.; de Kievit, T.R.; Fernando, W.D. Pyrosequencing reveals the influence of organic and conventional farming systems on bacterial communities. PLoS ONE 2012, 7, 51897.
  51. Memoli, V.; De Marco, A.; Baldantoni, D.; De Nicola, F.; Maisto, G. Short- and long-term effects of a single application of two organic amendments. Ecosphere 2017, 8, e02009.
  52. Ventorino, V.; de Marco, A.; Pepe, O.; Virzo de Santo, A.; Moschetti, G. Impact of Innovative Agricultural Practices of Carbon Sequestration on Soil Microbial Community. Carbon Sequestration in Agricultural Soils. A Multidisciplinary Approach to Innovative Methods; Piccolo, A., Ed.; Springer: Berlin/Heidelberg, Germany, 2012; pp. 145–177.
  53. Abbott, L.K.; Macdonald, L.M.; Wong, M.T.F.; Webb, M.J.; Jenkins, S.N.; Farrell, M. Potential roles of biological amendments for profitable grain production—A review. Agric. Ecosyst. Environ. 2018, 256, 34–50.
  54. Edmeades, D.C. The long-term effects of manures and fertilizers on soil productivity and quality: A review. Nut. Cycl. Agroecosyst. 2003, 66, 165–180.
  55. Leroy, B.L.M.; Herath, H.M.S.K.; Sleutel, S.; De Neve, S.; Gabriels, D.; Reheul, D.; Moens, M. The quality of exogenous organic matter: Short-term effects on soil physical properties and soil organic matter fractions. Soil Use Manag. 2008, 24, 139–147.
  56. Scotti, R.; Conte, P.; Berns, A.E.; Alonzo, G.; Rao, M.A. Effect of organic amendments on the evolution of soil organic matter in soils stressed by intensive agricultural practices. Curr. Org. Chem. 2013, 17, 2998–3005.
  57. Six, J.; Paustian, K. Aggregate–associated soil organic matter as an ecosystem property and a measurement tool. Soil Biol. Biochem. 2014, 68, A4–A9.
  58. Ingelmo-Sánchez, F.; Rubio-Delgado, J. Efecto de la aplicación del compost sobre las propiedades físicas y químicas del suelo. In Compostaje; Moreno-Casco, J., Moral-Herrero, R., Eds.; Mundi-Prensa: Madrid, Spain, 2008.
  59. Hernández, T.; García, E.; García, C. A strategy for marginal semiarid degraded soil restoration: A sole addition of compost at a high rate. A five-year field experiment. Soil Biol. Biochem. 2015, 89, 61–71.
  60. Zhang, H.; Luo, Y.; Wu, L.; Huang, Y.; Christie, P. Residues and potential ecological risks of veterinary antibiotics in manures and composts associated with protected vegetable farming. Environ. Sci. Pollut. Res. Int. 2015, 22, 5908–5918.
  61. Pan, M.; Chu, L.M. Leaching behavior of veterinary antibiotics in animal manure-applied soils. Sci. Total Environ. 2017, 579, 466–473.
  62. Godlewska, P.; Ok, Y.S.; Oleszczuk, P. The dark side of black gold: Ecotoxicological aspects of biochar and biochar-amended soils. J. Hazard. Mater. 2021, 403, 123833.
  63. Tao, R.; Li, J.; Hu, B.; Chu, G. Ammonia-oxidizing bacteria are sensitive and not resilient to organic amendment and nitrapyrin disturbances, but ammonia-oxidizing archaea are resistant. Geoderma 2021, 384, 114814.
  64. EUROSTAT, Statistical Office of the European Communities. Waste Statistics. 2020. Available online: (accessed on 10 September 2020).
  65. EC, European Commission Decision 2015/2099. Review of Waste Policy and Legislation. 2020. Available online: (accessed on 10 September 2020).
  66. ECN, European Compost Network. Bio-Waste in Europe. 2020. Available online: (accessed on 10 September 2020).
  67. EEA, European Environment Agency. Waste Generation in Europe. 2020. Available online: (accessed on 10 September 2020).
  68. Vimal, S.R.; Singh, J.S.; Arora, N.K.; Singh, S. Soil-plant-microbe interactions in stressed agriculture management: A review. Pedosphere 2017, 27, 177–192.
  69. EC (European Commission). Impact Assessment of the Thematic Strategy on Soil SEC 620. 2006. Available online: (accessed on 10 September 2020).
  70. Martínez-Blanco, J.; Lazcano, C.; Christensen, T.H.; Muñoz, P.; Rieradevall, J.; Møller, J.; Antón, A.; Boldrin, A. Compost benefits for agriculture evaluated by life cycle assessment. A review. Agron. Sustain. Dev. 2013, 33, 721–732.
  71. Schreuder, R.; De Visser, C. EIP AGRI Focus Group Protein Crops: Final Report. European Innovation Partnership for Agricultural Productivity and Sustainability (EIP AGRI). Brussels. 2014. Available online: (accessed on 10 September 2020).
  72. Saveyn, H.; Eder, P. End of Waste Criteria for Biodegradable Waste Subjected to Biological Treatment (Compost and Digestate): Technical Proposal; © European Commission in 2014, EUR 26425; Joint Research Centre—Institute for Prospective Technological Studies: Luxembourg, 2014; p. 310.
  73. Büyüksönmez, F.; Rink, R.; Hess, T.; Bechinski, E. Occurrence, degradation and fate of pesticides during composting. Part I. Composting, pesticides, and pesticides degradation. Compost Sci. Util. 1999, 7, 66–82.
  74. Büyüksönmez, F.; Rynk, R.; Hess, T.; Bechinski, E. Occurrence, degradation and fate of pesticides during composting. Part II. Occurrence and fate de pesticides in compost and composting systems. Compost Sci. Util. 2000, 8, 61–81.
  75. Briceño, G.; Palma, G.; Durán, N. Influence of organic amendment on the biodegradation and movement of pesticides. Crit. Rev. Environ. Sci. Technol. 2007, 37, 233–271.
  76. Marín-Benito, J.M.; Rodríguez-Cruz, M.S.; Andrades, M.S.; Sánchez-Martín, M.J. Assessment of spent mushroom substrate as sorbent of fungicides: Influence of sorbent and sorbate properties. J. Environ. Qual. 2012, 41, 814–822.
  77. Marín-Benito, J.M.; Andrades, M.S.; Sánchez-Martín, M.J.; Rodríguez-Cruz, M.S. Dissipation of fungicides in a vineyard soil amended with different spent mushroom substrates. J. Agric. Food Chem. 2012, 60, 6936–6945.
  78. Marín-Benito, J.M.; Brown, C.D.; Herrero-Hernández, E.; Arienzo, M.; Sánchez-Martín, M.J.; Rodríguez-Cruz, M.S. Use of raw or incubated organic wastes as amendments in reducing pesticide leaching through soil columns. Sci. Total Environ. 2013, 463–464, 589–599.
  79. Marín-Benito, J.M.; Herrero-Hernández, E.; Andrades, M.S.; Sánchez-Martín, M.J.; Rodríguez-Cruz, M. Effect of different organic amendments on the dissipation of linuron, diazinon and myclobutanil in an agricultural soil incubated for different time periods. Sci. Total Environ. 2014, 476–477, 611–621.
  80. Zolgharnein, J.; Shahmoradi, A.; Ghasemi, J. Pesticides removal using conventional and low–cost adsorbents: A review. Clean-Soil Air Water 2011, 39, 1105–1119.
  81. Garrido, I.; Vela, N.; Fenoll, J.; Navarro, G.; Pérez–Lucas, G.; Navarro, S. Testing of leachability and persistence of sixteen pesticides in three agricultural soils of a semiarid Mediterranean region. Span. J. Agric. Res. 2015, 13, 1104.
  82. Zhang, G.; Liu, X.; Sun, K.; Zhao, Y.; Lin, C. Sorption of Tetracycline to Sediments and Soils: Assessing the roles of pH, the presence of cadmium and properties of sediments and soils. FESE 2010, 4, 421–429.
  83. Gavrilescu, M. Fate of pesticides in the environment and its bioremediation. Eng. Life Sci. 2005, 5, 497–525.
  84. Jiang, L.; Lin, J.L.; Jia, L.X.; Liu, Y.; Pan, B.; Yang, Y.; Lin, Y. Effects of two different organic amendments addition to soil on sorption-desorption, leaching, bioavailability of penconazole and the growth of wheat (Triticum. aestivum L.). J. Environ. Manag. 2016, 167, 130–138.
  85. Loffredo, E.; Parlavecchia, M.; Perri, G.; Gattullo, R. Comparative assessment of metribuzin sorption efficiency of biochar, hydrochar and vermicompost. J. Environ. Sci. Health B 2019, 54, 728–735.
  86. Petter, F.A.; Ferreira, T.S.; Sinhorin, A.P.; Lima, L.B.; Almeida, F.A.; Pacheco, L.P.; Silva, A.F. Biochar increases diuron sorption and reduces the potential contamination of subsurface water with diuron in a sandy soil. Pedosphere 2019, 29, 801–809.
  87. García-Jaramillo, M.; Trippe, K.M.; Helmus, R.; Knicker, H.E.; Cox, L.; Hermosín, M.C.; Kalbitz, K. An examination of the role of biochar and biochar water-extractable substances on the sorption of ionizable herbicides in rice paddy soils. Sci. Total Environ. 2020, 706, 135682.
  88. Khan, S.U. Pesticides in the Soil Environment; Elsevier: New, York, NY, USA, 2016.
  89. Wu, C.; Liu, X.; Wu, X.; Dong, F.; Xu, J.; Zheng, Y. Sorption, degradation and bioavailability of oxyfluorfen in biochar-amended soils. Sci. Total Environ. 2019, 658, 87–94.
  90. Cox, L.; Cecchi, A.; Celis, R.; Hermosín, M.; Koskinen, W.; Cornejo, J. Effect of exogenous carbon on movement of simazine and 2,4–D in soils. Soil Sci. Soc. Am. J. 2001, 65, 1688–1695.
  91. Plaza, C.; Polo, A.; Brunetti, G.; Garcia-Gil, J.; D’Orazio, V. Soil fulvic acid properties as a means to assess the use of pig amendment. Soil Till. Res. 2003, 74, 179–190.
  92. Thorstensen, C.; Lode, O.; Eklo, O.; Christianse, A. Sorption of bentazone, dichlorprop, MCPA, and propiconazole in references soils from Norway. J. Environ. Qual. 2001, 30, 2046–2052.
  93. Cambier, P.; Pot, V.; Mercier, V.; Michaud, A.; Benoit, P.; Revallier, A.; Houot, S. Impact of long–term organic residue recycling in agriculture on soil solution composition and trace metal leaching in soils. Sci. Total Environ. 2014, 499, 560–573.
  94. Barriuso, E.; Andrades, M.S.; Benoit, P.; Houot, S. Pesticide desorption from soils facilitated by dissolved organic matter coming from composts: Experimental data and modelling approach. Biogeochemistry 2011, 106, 117–133.
  95. Huang, X.; Lee, S. Effects of dissolved organic matter from animal waste effluent on chlorpyrifos sorption by soils. J. Environ. Qual. 2001, 30, 1258–1265.
  96. Álvarez-Martín, A.; Sánchez-Martín, M.J.; Ordax, J.M.; Marín-Benito, J.M.; Rodríguez-Cruz, M.S. Leaching of two fungicides in spent mushroom substrate amended soil: Influence of amendment rate; fungicide ageing and flow condition. Sci. Total Environ. 2017, 584–585, 828–837.
  97. Rodríguez-Liébana, J.A.; Peña, A. Adsorption-desorption of dimethenamid and fenarimol onto three agricultural soils as affected by treated wastewater and fresh sewage sludge-derived dissolved organic carbon. J. Environ. Manag. 2018, 217, 592–599.
  98. Wanner, U.; Führ, F.; Burauel, P. Influence of the amendment of corn straw on the degradation behaviour of the fungicide dithianon in soil. Environ. Pollut. 2005, 133, 63–70.
  99. Wang, H.; Lin, K.; Hou, Z.; Richardson, B.; Gan, J. Sorption of the herbicide terbuthylazine in two New Zealand forest soils amended with biosolids and biochars. J. Soils Sediments 2010, 10, 283–289.
  100. Thevenot, M.; Dousset, S. Compost effect on diuron retention and transport in structured vineyard soils. Pedosphere 2015, 25, 25–36.
  101. Li, K.; Xing, B.S.; William, A.T. Effect of organic fertilizers derived dissolved organic matter on they sorption and leaching. Environ. Pollut. 2005, 134, 187–194.
  102. Spark, K.M.; Swift, R.S. Effect of soil composition and dissolved organic matter on pesticide sorption. Sci. Total Environ. 2002, 298, 147–161.
  103. Parlavecchia, M.; Orazio, V.D.; Loffredo, E. Wood biochars and vermicomposts from digestate modulate the extent of adsorption-desorption of the fungicide metalaxyl-m in a silty soil. Environ. Sci. Pollut. Res. Int. 2019, 26, 35924–35934.
  104. Deng, H.; Feng, D.; He, J.; Li, F.; Yu, H.; Ge, C. Influence of biochar amendments to soil on the mobility of atrazine using sorption-desorption and soil thin-layer chromatography. Ecol. Eng. 2017, 99, 381–390.
  105. Mendes, K.F.; de Sousa, R.N.; Takeshita, V.; Alonso, F.G.; Régo, A.P.J.; Tornisielo, V.L. Cow bone char as a sorbent to increase sorption and decrease mobility of hexazinone, metribuzin, and quinclorac in soil. Geoderma 2019, 343, 40–49.
  106. Mendes, K.F.; Hall, K.E.; Takeshita, V.; Rossi, M.L.; Tornisielo, V.L. Animal bonechar increases sorption and decreases leaching potential of aminocyclopyrachlor and mesotrione in a tropical soil. Geoderma 2018, 316, 11–18.
  107. Marín-Benito, J.M.; Herrero-Hernández, E.; Rodríguez-Cruz, M.S.; Arienzo, M.; Sánchez-Martín, M.J. Study of processes influencing bioavailability of pesticides in wood-soil systems: Effect of different factors. Ecotoxicol. Environ. Saf. 2017, 139, 454–462.
  108. Mendes, K.F.; Alonso, F.G.; Mertens, T.B.; Inoue, M.; Oliveira, M.G.D.; Tornisielo, V.L. Aminocyclopyrachlor and mesotrione sorption-desorption in municipal sewage sludge-amended soil. Soil Plant Nutr. 2019, 78, 131–140.
  109. Duhan, A.; Oliver, D.P.; Rashti, M.R.; Du, J.; Kookana, R.S. Organic waste from sugar mills as a potential soil ameliorant to minimize herbicide runoff to the Great Barrier Reef. Sci. Total Environ. 2020, 713, 136640.
  110. Di Marsico, A.; Scrano, L.; Amato, M.; Gàmiz, B.; Real, M.; Cox, L. Mucilage from seeds of chia (Salvia hispanica L.) used as soil conditioner, effects on the sorption-desorption of four herbicides in three different soils. Sci. Total Environ. 2018, 625, 531–538.
  111. Gaonkar, O.D.; Nambi, I.M.; Govindarajan, S.K. Soil organic amendments: Impacts on sorption of organophosphate pesticides on an alluvial soil. J. Soils Sediments 2019, 19, 566–578.
  112. Xing, B. Sorption of naphthalene and phenanthrene by soil humic acids. Environ. Pollut. 2001, 111, 303–309.
  113. Wang, T.; Zhang, Z.; Zhang, H.; Zhong, X.; Liu, Y.; Liao, S.; Yuea, X.; Zhouc, G. Sorption of carbendazim on activated carbons derived from rape straw and its mechanism. RSC Adv. 2019, 9, 41745–41754.
  114. García-Delgado, C.; Marín-Benito, J.M.; Sánchez-Martín, M.J.; Rodríguez-Cruz, M.S. Organic carbon nature determines the capacity of organic amendments to adsorb pesticides in soil. J. Hazard. Mater. 2020, 390, 122162.
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