Environmental Fate, Transformation, and Interactions of Agrochemicals and Micro-Nano Plastics in Agricultural Ecosystem

Abstract

The extensive use of agrochemicals and plastic materials has led to the accumulation of persistent pollutants in agricultural soils, raising concerns about agroecosystems through posing potential risks to soil and environmental health. This review synthesizes recent knowledge on these pollutant sources, including their distribution, fate, transformation pathways, and detection methods, as well as their impacts on soil physicochemical properties, microbial populations, plants, and ecosystems. Existing findings indicate that agrochemicals and micro-nano plastics (MPs-NPs) can significantly impede the stability of soil aggregation, increase soil water holding capacity (WHC) and porosity, reduce bulk density and infiltration, alter soil structure, and affect soil pH, cation exchange capacity (CEC), electrical conductivity (EC), and nutrient retention capacity. Moreover, exposure to these pollutants alters soil microbial communities, enzymatic activity, nitrification and denitrification processes, and arbuscular mycorrhizal fungi (AMF), thereby affecting carbon pools and fluxes as well as nutrient cycling. However, the magnitude and direction of these effects are strongly influenced by soil type, pollutant class, concentration, and physicochemical properties. Furthermore, terrestrial and aquatic ecosystems are negatively affected due to the presence of such persistent pollutants by impairing their physiological processes. Despite these findings, mechanistic understanding remains limited due to a lack of long-term field investigation and proper detection methods, particularly regarding NPs. A comprehensive understanding of agrochemical and MP-NP interactions is essential for developing sustainable soil management strategies and agroecosystems. Future studies should address the development of standardized NP detection methods and the conducting of long-term field studies to elucidate MP-NP and agrochemical interactions, soil impacts, and crop uptake mechanisms.

1. Introduction

Agriculture is at the forefront of addressing global food security needs, especially with a growing world population. However, the widespread application of agrochemicals, including fertilizers, herbicides, pesticides, nematicides and seed-treated chemicals, along with plastic materials such as greenhouse films, plastic mulching films, irrigation pipes and agrochemical packaging, has become integral to modern farming practices. The intensifications of these inputs involve a wide range of chemical compounds with varying degrees of stability, mobility and bioavailability, which are vital for enhancing agricultural productivity and nutrient use efficiency. However, plastic materials introduced into agroecosystems are susceptible to degradation through physical (e.g., UV radiation and mechanical fragmentation), chemical (e.g., oxidation), and biological processes, leading to the formation of micro- and nano-plastics (MPs-NPs) [1,2,3]. Agrochemicals can further modulate these degradation pathways by altering soil physicochemical conditions and microbial activity, thereby influencing the fragmentation, transport and fate of plastic particles within the soil matrix [4]. Consequently, the accumulation and persistence of MPs-NPs in agricultural soils have emerged as significant environmental concerns due to their potential impacts on soil health, biochemical cycles and ecosystem functioning [5,6].

In parallel with the increasing use of traditional agrochemicals alongside plastic materials, MPs-NPs have emerged as a pervasive environmental pollutant in agroecosystems. Weber and Opp [7] estimated that 14% of plastics moved to agricultural soils from different sources due to suboptimal management. Previous studies have shown that agricultural soils, notably those exposed to plastic materials, are heavily contaminated with MPs-NPs. For instance, a recent study in China revealed that agricultural soils are heavily contaminated with plastic mulch films, averaging 83.6 kg ha⁻¹ across 19 provinces [8]. Agrochemicals and plastic materials exposed to soil microorganisms and ultraviolet light undergo natural degradation and weathering, ultimately leading to the formation of MPs-NPs [9,10]. MPs have a diameter smaller than 5 mm, whereas NPs have diameter less than 100 nm [11,12]. Quantitative evidence shows that plastic contamination in agricultural soils in regions like the United Kingdom (UK) and Canada has reached alarming levels. In the UK, MP concentrations have been found to average 3680 ± 129.1 particles kg⁻¹ soil due to plastic mulch, and concentrations in Canada have soared to as high as 1.4 × 10⁴ items kg⁻¹, largely associated with the application of recycled biosolids [13,14]. Due to their extremely low biodegradability, MPs-NPs accumulate in soil and water bodies, exacerbated by human activities and the ubiquity of plastics (Figure 1) [15,16]. The presence of MPs-NPs in soil can alter soil physicochemical properties, including texture, structure, porosity, bulk density, water holding capacity (WHC), infiltration, evapotranspiration, aeration, soil pH, organic carbon and cation exchange capacity (CEC). These changes negatively impact soil health and crop production [17,18,19]. Additionally, MPs-NPs act as vectors for hazardous elements like heavy metals and pesticides through adsorption and transportation, thus intensifying an environmental persistence that creates eco-toxicity and disrupts soil-plant ecosystems [20].

Soil health, which plays a vital role in the sustainability of agroecosystems, is directly affected by MPs-NPs. The rhizosphere, which serves as a vital interface within agroecosystems, plays a central role in interactions among plants, microorganisms and pollutants. Due to the natural degradation and weathering of plastic materials, MPs-NPs can be sequestered in the rhizosphere [21,22]. Additionally, agrochemicals contribute indirectly to MP-NP pollution through polymer-coated fertilizers, carriers and plastic-based packaging materials [1,23]. Several recent studies have reported that plant uptake of MPs-NPs via endocytosis or any opening impairs photosynthesis by damaging stromata and chlorophyll content and alters the acquisition and uptake of essential nutrients [24,25,26,27,28]. These impacts are not only confined to plant systems but also extend to the soil microbial dynamics. MPs-NPs can affect key soil biogeochemical processes by changing microbial activity and function, creating localized microbial hotspots, and modifying soil properties dependent on soil type as well as plastic particle size and concentration [29,30]. Their effects on plant growth and performance may result from the direct toxicity of MPs-NPs or indirect changes in soil structure and microbial communities [31]. Lin et al. [32] highlighted that MPs-NPs exert pronounced cascading effects on soil agroecosystems, including microorganisms, especially nematodes, by altering microbial activity and thereby impacting soil chemical processes.

Despite the growing recognition of MPs-NPs as emerging pollutants in agricultural soils, their interactions with different agrochemicals remain poorly understood. With few review papers on MPs-NPs, research has largely focused on the synergistic interactions between agrochemicals and MPs-NPs in soil and plant ecosystems. This knowledge gap impedes a holistic understanding of their fate, transport, and cumulative toxicity effects on agroecosystems, specifically in terms of long-term plant and soil health and ecological sustainability. Moreover, the lack of standardized NP detection methods and the paucity of field-scale investigations hinder reliable risk assessments. This review synthesizes the current research on the interactions between agrochemicals and MPs-NPs in agricultural soils. We intended to comprehensively understand the effects of MPs-NPs on agricultural ecosystems. This review provides better insight into the interactions between agrochemicals and MPs-NPs in agricultural soil, thereby shaping their impact on soil health, crops, and agroecosystems.

2. Aim, Scope and Methodology of the Review

The pervasive presence of agrochemicals and MPs-NPs in the environment has become a global concern for sustainable agriculture. This review was motivated by the concern over the combined impact of these pollutants on terrestrial and aquatic ecosystems. It aims to explore the synergistic interactions between these pollutants and their impact on various components of ecosystems. Given the rising prevalence of these pollutants, understanding their combined effects is crucial for developing sustainable agricultural practices and safeguarding environmental health.

This review primarily focuses on the impact of agrochemicals and MPs-NPs on soil-plant systems, with particular emphasis on the analytical techniques for their detection and quantification and the associated challenges. Moreover, this review addresses the following areas: the source of these pollutants, their fate and behavior in soil, and their effects on soil health and microbial diversity. It also discusses the ecotoxicological effects of these pollutants and their movement through the food web, leading to bioaccumulation and biomagnification at higher trophic levels.

The narrative review involved a literature search using well-known scientific databases like Google Scholar, Scopus, Springer, Nature and ScienceDirect. The main keyword search terms included “microplastics”, “nanoplastics”, “agrochemicals”, “pesticides”, “herbicides”, “fungicides”, “ecotoxicology”, “terrestrial”, “aquatic”, and “higher trophic”. Boolean operators such as AND/OR were used to combine keywords and refine the search. Relevant studies were screened based on title, abstract, and full-text relevance to the objectives of the review. The inclusion criteria were as follows: (1) priority was given to articles published between 2015 and 2026; (2) the articles must be peer-reviewed and written in English; (3) the articles describe the source and detection methods of agrochemicals and MPs-NPs; (4) the articles indicate the impacts of MPs-NPs and agrochemicals on the bioavailability, persistence, or toxicity in agroecological ecosystems. The exclusion criteria were: (1) articles explaining only molecular effects, without focusing on chemical interactions and (2) studies dealing exclusively with human health effects, without direct linkage to agricultural ecosystems. As this was a narrative review, no formal bibliometric or systematic review framework was applied. The review synthesizes the findings to present a comprehensive understanding of how these pollutants impact soil ecosystems, plant health, and higher trophic levels.

3. Sources and Characteristics of Pollutants in Agricultural Soils

Agricultural soil systems are increasingly threatened by anthropogenic inputs, such as agrochemicals and MPs-NPs. These diverse classes of pollutants are pervasive in agricultural ecosystems and enter the soil matrix through distinct, possibly interactive pathways. A critical understanding of the source and characteristics of these pollutants is essential for predicting their environmental behavior and impact on agricultural soils.

3.1. Sources and Characteristics of Agrochemicals

Agrochemicals, including fertilizers, pesticides, and growth regulators, are intentionally added to the natural environment to improve productivity [33,34]. However, their long-term fate depends on their chemical nature and the associated soil environmental properties. These compounds enter the soil through direct and indirect pathways. Direct inputs include soil incorporation of agrochemical-formulations, whereas indirect inputs include incidental deposition, such as foliar wash-off and spray drift [35,36]. Prophylactic seed coatings have become the dominant route for systemic pesticides. For instance, studies indicate that only 2–20% of the active ingredient of seed-treated chemicals is typically assimilated by the crop, leaving the remaining 80–98% as a soil pollutant [37,38].

Pesticides constitute one of the largest classes of bioactive organic compounds applied to the biosphere [39]. Their behavior in soil is governed by their intrinsic physicochemical properties. The development of persistent and bioaccumulative organochlorines in the mid-20th century introduced a significant long-term risk to soil ecosystems [40]. For instance, organochlorines contaminate soil and alter its pH, which can drastically change the soil biota, resulting in the massive degradation of soil nutrients [41,42]. Pesticides such as chlorpyrifos are highly hydrophobic (Log K_ow ~ 4.7) and strongly sorb to organic carbon, limiting leaching but encouraging topsoil accumulation [43]. Conversely, carbamates are more water-soluble and mobile, often displaying lag phases in persistence in sterile soils [44]. Neonicotinoids, such as imidacloprid (IMD), are the most frequently used pesticides on a global scale [45,46]. Unlike their forerunners, they are extremely water-soluble and exceptionally persistent, with half-lives exceeding 1000 days, creating toxic reservoirs that pose risks to non-target invertebrates, including earthworms and ground-nesting bees [47]. In contrast, synthetic pyrethroids (e.g., bifenthrin) are characterized by extreme hydrophobicity (K_oc > 100,000) and have a moderate amount of binding and stability in soil, which increases their acute toxicity in the top layer of soil [48,49].

Herbicides form the highest proportion of pesticide mass, a trend fueled by conservation tillage and herbicide-tolerant crops [50]. Glyphosate (GLY), being the most used herbicide, is a zwitterionic compound, which, despite its high-water solubility, competitively binds with iron and aluminum oxides, similar to phosphate, thereby immobilizing it [51,52,53]. Moreover, a major pesticide, atrazine, used in maize and sorghum, has moderate water solubility (33 mg L⁻¹ at 20 °C) and low soil organic carbon partitioning (K_oc ≈ 100–200), resulting in high mobility and leaching potential, notably in soils with low organic matter content [54]. The field dissipation half-lives of atrazine, i.e., DT₅₀, range from 13 to 100 days, depending upon environmental factors [55,56]. The vertical migration of atrazine has also been confirmed in different agroecosystems, especially in wastewater-irrigated areas [57,58]. A lysimeter experiment conducted in Rhodic Ferralsols under simulated heavy rainfall conditions revealed atrazine leaching after 48 h of application [59].

Although fungicides are indispensable for the management of crop pathogens, they may exert significant non-target effects on beneficial soil microbial communities, particularly on arbuscular mycorrhizal fungi (AMF) [60]. Therefore, distinct chemical classes present unique environmental challenges. Triazoles (e.g., tebuconazole) disrupt fungal ergosterol biosynthesis and exhibit extreme persistence (with field half-lives exceeding 600 days), frequently surpassing environmental safety thresholds in agricultural top soils across Europe and Asia [61,62]. Strobilurins (azoxystrobin) inhibit mitochondrial respiration via the cytochrome bc₁ complex and strongly sorb to soil organic matter [63]. Organic amendments such as compost may unintentionally prolong their retention and elevate runoff risks [64]. Furthermore, the historical reliance on copper-based fungicides in perennial systems has driven severe, non-degradable elemental copper accumulation [65]. Soil concentrations often reach phytotoxic levels (200–500 mg kg⁻¹), suppressing earthworm populations, diminishing microbial biomass, and impairing essential nutrient-cycling enzymes [66,67]. The physicochemical properties and environmental persistence of representative agrochemicals in soils are presented in Table 1.

The intensive application of inorganic fertilizers has shifted from securing global food production to driving severe soil chemical degradation [68]. Chronic overuse of synthetic nitrogen (N), phosphorus (P), and potassium (K) fundamentally alters ecological stoichiometry by narrowing the soil C:N and N:P ratios. This nutrient imbalance favors fast-growing copiotrophic bacteria over stress-tolerant oligotrophic taxa [69,70], thereby reducing microbial diversity and disrupting critical nutrient cycling pathways [71]. Excessive N fertilization compromises chemical integrity by inducing soil acidification through nitrification-mediated H⁺ release [72]. This pH decline solubilizes phytotoxic Al³⁺ and accelerates the leaching of essential base cations such as Ca²⁺ and Mg²⁺ [73]. Similarly, sustained P over-application generates immobilized legacy phosphorus, precipitated as Fe/Al-phosphates in acidic soils or Ca-phosphates in alkaline soils [74]. Although largely unavailable to plants, this legacy pool poses a substantial eutrophication risk via surface runoff and erosion [75]. These macronutrient disruptions are further complicated by micronutrient toxicity and ionic antagonism. The historical application of Zn- and Cu-enriched livestock manure has driven the progressive accumulation of heavy metals in croplands [76]. Furthermore, elevated soil PO₄³⁻ levels competitively inhibit the root uptake of Zn and Fe [77], while excess Cu suppresses essential extracellular enzymes, such as urease, bottlenecking nitrogen mineralization [78].

Plant growth regulators (PGRs), encompassing both synthetic and naturally occurring compounds, are widely applied to modulate plant development, especially for optimizing crop architecture and reproductive timing [79,80]. Among them, synthetic auxins such as 2,4-dichlorophenoxyacetic acid (2,4-D) are commonly employed as selective herbicides, disrupting growth regulation in broadleaf weeds [81]. Gibberellin biosynthesis inhibitors, including chlormequat chloride, suppress stem elongation by inhibiting ent-kaurene oxidase, thereby improving lodging resistance in cereals and other crops [82]. Among the PGRs, the environmental persistence of ethephon and paclobutrazol differs in degree. Ethephon, used as a ripening agent, hydrolyzes rapidly in the environment, yielding ethylene gas, with low environmental persistence, characterized by DT₅₀ < 10 days and low residue levels [82,83]. On the other hand, paclobutrazol, used as a triazole-based PGR in orchards to inhibit vegetative growth and induce flowering, shows strong environmental persistence, characterized by DT₅₀ > 365 days, mainly because of its strong sorption by organic matter and clay [84]. The environmental persistence of paclobutrazol shows considerable carryover effects, as its presence in the soil from previous applications can significantly inhibit the growth of rotational crops like potato and taro, characterized by thickened stem development [85]. In addition, paclobutrazol disrupts the health of the soil microbiome and reduces dehydrogenase activity and biomass C, important measures of the functional microbiome, thus affecting the fertility of the soil [86].

Emerging agrochemicals such as nano-fertilizers and controlled-release fertilizers (CRFs) enhance nutrient use efficiency but introduce novel ecotoxicological risks [87]. Nano-fertilizers (1–100 nm), such as ZnO-nano particles or hydroxyapatite, leverage high surface-area-to-volume ratios to optimize targeted root delivery [88]. Engineered nanoparticles can penetrate microbial membranes and generate reactive oxygen species (ROS), disrupting symbiotic functions such as nitrogen fixation and phosphorus solubilization in Rhizobium, Azospirillum, and AMF, ultimately impairing soil biological health [89,90]. Concurrently, CRFs also modulate nutrient release kinetics via polymeric encapsulation, reducing nitrate leaching by up to 40% [91]. However, these coatings frequently utilize non-biodegradable, petroleum-derived polymers like polyethylene (PE) and polysulfone (PSU) [92,93]. Post-release, these inert shells persist, accumulating over 10⁴ MP particles Kg⁻¹ of soil after repeated cycles, a burden exacerbated by mechanical tillage fragmenting shells into secondary MPs < 1 mm [94,95]. Furthermore, contaminated organic amendments like compost and manure may act as environmental vectors for MPs and antibiotic resistance genes, thereby contributing to their introduction into agroecosystems [94]. Balancing the agronomic benefits of these advanced systems with their ecological consequences remains a critical challenge. While ZnO nanoparticles can alleviate abiotic stress and enhance yields [96], their long-term soil bioactivity remains unclear. Similarly, while biodegradable starch/poly (butylene adipate-co-terephthalate) (PBAT) composite coatings are emerging, their field-scale efficacy and complete degradation remain unvalidated. The sustainable integration of next-generation agrochemicals demands rigorous life-cycle assessments, standardized ecotoxicology protocols, and the advancement of genuinely biodegradable carriers to prevent agricultural efficiency from inadvertently accelerating soil degradation [97,98].

The physicochemical properties of agrochemicals control their fate in the soil system, affecting adsorption, mobility, and degradation [99]. The molecular structure of a compound influences its functional groups, which in turn affects the ability of compounds to interact. Polar functional groups, such as the carboxyl group in glyphosate, increase the solubility of the compound but also affect adsorption. Polarity and aqueous solubility affect the potential leaching of compounds [100]. Highly soluble compounds, such as paraquat, are likely transported via runoff. The ionization potential is affected by the ability of a compound to bind at different pH values in the soil system [101]. Acidic pesticides tend to ionize in alkaline soils, thereby increasing their mobility. The ability of a compound to bind to soil organic matter (SOM) and clay minerals is affected by the octanol-water partition coefficient (K_ow) [102]. Higher K_ow values imply a higher hydrophobic potential, such as in chlorpyrifos. The degradation half-life (DT₅₀) is a measure of the compound’s persistence in the system. DT₅₀ is affected by microbial, photolytic, and hydrolytic degradation pathways, which tend to produce metabolites that are more persistent and toxic than the parent compound [103].

Table 1. Physicochemical properties and environmental persistence of representative agrochemicals in soils.

Agrochemical Category	Representative Compound	Molecular Weight (g mol−1)	Water Solubility (mg L−1)	Log Kow	Soil Half-Life (DT50, Days)	Primary Environmental Concern
Herbicide	Glyphosate	169.07	12,000	−3.2 to −4.3	2–197 (typically 30–60)	Strong soil binding but potential leaching in certain conditions [104]
Herbicide	Atrazine	215.68	33	2.6	60–100	Groundwater contamination due to moderate persistence and mobility [104]
Insecticide (organophosphate)	Chlorpyrifos	350.59	1.4	4.7–5.0	30–120	Bioaccumulation in soil organisms and high toxicity to non-target species [105]
Insecticide (neonicotinoid)	Imidacloprid	255.66	600	0.57	30–120	Leaching to groundwater and high toxicity to aquatic invertebrates [105,106]
Fungicide (triazole)	Tebuconazole	307.82	36	3.7	60–150	Persistence affecting soil microbiota and enzyme activities [107]
Fungicide (strobilurin)	Azoxystrobin	403.39	6	2.5	10–30	Runoff to aquatic systems, with moderate persistence [105]
Insecticide (carbamate)	Carbaryl	201.22	110	2.4	7–14	Rapid degradation but high acute toxicity to beneficial insects [108,109]
Insecticide (pyrethroid)	Cypermethrin	416.3	0.004	5.5–6.6	15–60	Strong soil adsorption limiting mobility but persistent in sediments [110]
Herbicide	Paraquat	257.16	620,000	−5.7 to −6.0	Highly variable (30–2000+)	Extreme persistence when bound to clay minerals, limited leaching [111]

Table 1. Physicochemical properties and environmental persistence of representative agrochemicals in soils.

Agrochemical Category	Representative Compound	Molecular Weight (g mol−1)	Water Solubility (mg L−1)	Log Kow	Soil Half-Life (DT50, Days)	Primary Environmental Concern
Herbicide	Glyphosate	169.07	12,000	−3.2 to −4.3	2–197 (typically 30–60)	Strong soil binding but potential leaching in certain conditions [104]
Herbicide	Atrazine	215.68	33	2.6	60–100	Groundwater contamination due to moderate persistence and mobility [104]
Insecticide (organophosphate)	Chlorpyrifos	350.59	1.4	4.7–5.0	30–120	Bioaccumulation in soil organisms and high toxicity to non-target species [105]
Insecticide (neonicotinoid)	Imidacloprid	255.66	600	0.57	30–120	Leaching to groundwater and high toxicity to aquatic invertebrates [105,106]
Fungicide (triazole)	Tebuconazole	307.82	36	3.7	60–150	Persistence affecting soil microbiota and enzyme activities [107]
Fungicide (strobilurin)	Azoxystrobin	403.39	6	2.5	10–30	Runoff to aquatic systems, with moderate persistence [105]
Insecticide (carbamate)	Carbaryl	201.22	110	2.4	7–14	Rapid degradation but high acute toxicity to beneficial insects [108,109]
Insecticide (pyrethroid)	Cypermethrin	416.3	0.004	5.5–6.6	15–60	Strong soil adsorption limiting mobility but persistent in sediments [110]
Herbicide	Paraquat	257.16	620,000	−5.7 to −6.0	Highly variable (30–2000+)	Extreme persistence when bound to clay minerals, limited leaching [111]

3.2. Sources and Characteristics of MPs-NPs in Agricultural Soils

In parallel to the chemical burden of agrochemicals, agricultural soils are increasingly characterized by the presence of synthetic polymer particles, known as MPs-NPs. MPs are generally defined as plastic particles smaller than 5 mm, whereas NPs are less than 100 nm in diameter. Both types of plastic particles are introduced into agricultural soils through various direct and indirect sources. The different sources, polymer types, and characteristics of MPs detected in agricultural soils are presented in

Table 2. Sources, polymer types, and characteristics of MPs detected in agricultural soils.

Source Category	Polymer Type	Shape	Size Range	Typical Agricultural Origin	References
Mulch films	Polyethylene (LDPE), Polypropylene (PP)	Films, Fragments	0.03–5 mm	Fragmentation of weed control films; photodegradation residues	[112,113]
Sewage sludge/biosolids	Polyester (PES), Polyamide (PA), Polypropylene (PP)	Fibers (90–97%), Fragments	10 µm–5 mm	Washing machine effluent (textiles); industrial wastewater	[114]
Controlled-release fertilizers	Polyethylene (PE), Polyacrylate, Polyacrylamide	Fragments, Shells	<5 mm (fragmented)	Non-biodegradable encapsulating shells remaining after nutrient release	[115]
Irrigation infrastructure	Polyethylene (PE), Polyvinyl chloride (PVC)	Fragments, Particles	0.1–5 mm	Wear and degradation of drip tapes, pipes, and emitters	[116]
Seed coatings	Polyvinyl acetate, Polypropylene (PP), polyethylene terephthalate (PET)	Fragments, Dust	<1 mm	Abrasion of polymer binders during sowing and handling	[115]
Atmospheric deposition	Polyester, Polystyrene (PS)	Fibers, Dust	<0.5 mm	Wind-blown transport from urban/industrial centers	[117]

.

3.2.1. Primary Sources: Intentional and Unintentional Inputs

The intentional introduction of MPs-NPs into agricultural systems through the application of agronomic technologies has been observed to create a paradox in which the efficiency of these technologies is associated with the creation of permanent pathways for pollution. A good example of the trade-off in the application of agronomic technologies is the use of polymer-coated controlled-release fertilizers (PC-CRFs). Although these fertilizers are efficient in reducing nutrient leaching in the soil, they leave behind non-biodegradable “ghost capsules” made of polymers such as polyethylene (PE) and polyurethane (PU). Results from longitudinal studies in the field suggest that the residue from these fertilizers has been observed to build up considerably, affecting the organo-mineral composition of the soil [118]. The application of these fertilizers has also been observed to increase the fragmentation of these capsules into MPs-NPs through mechanical tillage and osmotic stress [119]. These plastic particles are difficult to biodegrade, even in the presence of biofilms on polymer surfaces [120]. In recent years, fertilizers, particularly polymer-coated slow-release formulations and organic fertilizers containing residual plastic contaminants, have been identified as contributors to MPs-NPs in soil environments, adding to their accumulation alongside more prominent sources such as plastic mulches and sewage sludge [121]. Polymer-coated seeds, made from polymers such as polyvinyl acetate (PVAc), also create MP-NP dust through abrasion during pneumatic sowing. Although the contribution of MPs-NPs from these sources is negligible, their cumulative effects have been observed to target the rhizosphere, which is the most biologically active part of the soil. The rhizosphere is the part of the soil where the roots of plants are located, making it the most important part of the soil [122]. Furthermore, drip irrigation infrastructure composed of high-density polyethylene (HDPE) and polyvinyl chloride (PVC) degrades under UV radiation, thermal cycling, and mechanical wear, shedding micro-fragments that migrate into deeper soil layers [123].

3.2.2. Secondary Sources: Fragmentation and Waste Stream Transfer

The sources of MPs-NPs in the agricultural domain are mainly influenced by secondary sources, with the fragmentation of low-density polyethylene (LDPE) mulch films being the most significant. Comprehensive field surveys indicate an overall prevalence of MPs in the domain, with 96% of the samples containing residual plastic film fragments and mass concentrations reaching up to 78.5 kg ha⁻¹ in managed agroecosystems [124]. The fragmentation process stops once the plastic films are buried, and the process of mechanical and biological degradation begins, breaking down the macro-films into MP-NP forms that affect nutrient and water retention in agroecosystems [125,126]. The use of alternatives, including biodegradable plastics, also results in large quantities of persistent fragments, even after the first growing season [127]. The application of sewage sludge, also known as biosolids, acts as an important transfer mechanism for MPs-NPs from the waste domain to the agroecosystem, with 57–99% of the influent MPs, mainly synthetic textile fibers, being concentrated in the wastewater treatment process and ending up in the sludge [128]. Long-term field studies have shown a strong correlation between application and MP-NP content in soils, with significant quantities of plastics recorded in biosolid-amended soils compared to controls [129,130], including 1 µm MPs [131]. Simultaneously, the application of contaminated organic amendments, including compost and manure, also results in the transfer of MPs-NPs in agroecosystem [132]. The transfer mechanism via the atmosphere also ensures that no agroecosystem remains untouched by the problem, with wind currents carrying micro-fibers and films from the industrial domain and disseminating them in the agroecosystem [133], with virtually no soil being free from the influence of MPs-NPs.

3.2.3. Physicochemical Characteristics and Ecological Behavior

The ecological kinetics of MPs-NPs in soil ecosystems are driven by their physicochemical properties. The size of MPs controls their vertical distribution, with smaller MPs (<0.5 mm) penetrating deeper (up to 20–30 cm) due to water percolation and bioturbation (e.g., earthworm casting), while MPs > 2 mm remain in the tilled surface layer (0–10 cm) [134,135,136]. Significantly, as MPs-NPs fragment, their surface area/volume ratio increases exponentially, increasing their sorption of co-occurring pollutants (e.g., heavy metals, pesticides) and ingestion risks by soil mesofauna [137,138]. MP-NP morphology also controls physical effects in soil ecosystems. The input of MPs-NPs as agricultural mulch can impede water infiltration and enhance evaporation, altering soil aggregate stability [139,140]. The presence of rigid MPs-NPs in soil may indicate the weathering of thicker infrastructure, irrigation tapes, or fertilizer coatings [141]. Chemically, LDPE and polypropylene (PP) are dominant MPs in agricultural matrices, with their low density facilitating their transportation by water during wetting events [142,143]. Significantly, while biodegradable plastic products offer a promising solution for their degradation, they are highly dependent on environmental factors (e.g., temperature, moisture, biomass); therefore, they may also behave as transient micro-fragments, creating physical hazards and resulting in reduced porosity and reduced hydraulic conductivity [144,145].

4. Interactions Between Agrochemicals, MPs and NPs

4.1. Mechanisms Governing Agrochemical Interaction with MPs and NPs

4.1.1. Adsorption and Desorption Mechanisms (MPs vs. NPs)

The adsorption of agrochemicals on MPs and NPs is a complicated interplay of physicochemical forces, and the basic differences depend on particles size and the properties of the polymer [146]. In the case of MPs, pore-filling is prevalent, and agrochemicals can access the interior porous structure of larger particles (Figure 2A). Conversely, NPs are characterized by surface-dominated adsorption because of their outstanding specific surface area, with the effects of surface charges and colloidal stability being of central importance (Figure 2B) [147].

The major driving force among non-polar agrochemicals is hydrophobic interactions governed by the octanol-water partition coefficient (log K_ow). This aligns with the polymer–pesticide partitioning theory, which predicts that hydrophobic compounds preferentially partition into plastic matrices. For instance, polyethylene (PE) and polypropylene (PP), owing to their non-polar nature, show high affinities for hydrophobic pesticides such as chlorpyrifos (log K_ow = 4.7) and cypermethrin (log K_ow = 6.6) [148,149]. For more polar compounds such as glyphosate (log K_ow = −3.2) and imidacloprid (log K_ow = 0.57), secondary interactions including π–π stacking (significantly for aromatic pesticides on polystyrene), hydrogen bonding with oxygenated functional groups, and van der Waals forces become increasingly important [150].

Plastic aging, which is caused by UV weathering, oxidation and biofilm formation, significantly changes the adsorption properties. The radical chain reaction in the formation of photoinitiation and hydroperoxide in polyethylene is as follows:

RH + hν → R^• + H^•
R^• + O₂ → ROO•
ROO• + RH → ROOH + R^•

This process introduces oxygen-containing functional groups (e.g., carbonyl and hydroxyl), which increase the surface polarity and roughness [151,152]. The subsequent carbonyl (ketone) formation via hydroperoxide decomposition proceeds as follows:

ROOH → RO• + •OH
RO• + RH → Ketone + Chain Scission Products

Structurally, this transforms the polymer backbone [153]:

…–CH₂–CH(OOH)–CH₂–… → …–CH₂–C(=O)–CH₃ + H₂O

In the case of atrazine, the adsorption capacity of UV-aged polyethylene MPs reached 0.646–0.742 mg g⁻¹ compared to 0.405–0.504 mg g⁻¹ for pristine MPs; the adsorption kinetics were best characterized by pseudo-second-order models, which implies chemisorption [154]. Plastic aging (weathering/UV-induced photo-oxidative degradation) also results in an increase of the negativity of NP surfaces (e.g., zeta potential of −32 mV for aged vs. −26 mV for pristine PE), which increases electrostatic interactions with polar or ionizable agrochemicals [154].

The hysteresis of sorption and irreversible binding is stronger for NPs since they have a large surface area and can have stronger interactions with the chemicals. The polymer-specific effects were significant. With S-metolachlor, PVC had the highest adsorption capacity (17.67 µg g⁻¹), followed by PP (14.71 µg g⁻¹) and PE (14.41 µg g⁻¹) [155]. Polymers such as polystyrene (PS) and PVC have a higher diffusion rate and adsorption capacity for high-molecular-weight contaminants than PE and PP because they are not rubbery but glassy [156]. The adsorption processes of biodegradable plastics differ owing to their initial hydrophilicity and rapid surface modification during decomposition.

4.1.2. MPs and NPs as Agrochemical Vectors

MPs and NPs are widespread pollutants in arable communities, mostly as a result of plastic mulch breakdown, pesticide packaging, biosolid use, and polymer-coated fertilizers [12,13]. These particles accumulate in soil layers, particularly in areas with high agrochemical use, and act as long-term stores of organic contaminants [17].

The tendency of MPs and NPs to act as vectors can be explained by their high surface area and heterogeneous surface chemistries. The adsorption of pesticides, herbicides, and polycyclic aromatic hydrocarbons (PAHs) occurs via sorption processes, such as hydrophobic interactions, van der Waals forces, and electrostatic bonding [157]. The stability of the resulting complexes of the vectors is determined by the specific interaction mechanisms. In the case of hydrophobic substances, non-polar interactions dictate the partitioning effect, as in the case of chlorpyrifos on PE:

Chlorpyrifos (aq) + PE (surface) ⇌ [Chlorpyrifos·PE] complex

Polar or charged agrochemicals, on the other hand, interact through hydrogen bonding and electrostatic force, such as in glyphosate. As an example, the process of glyphosate adsorption to cationic polystyrene NPs is as follows [158]:

Glyphosate⁻ + NPs–NH₃⁺ ⇌ [Glyphosate·NPs–NH₃] complex

Although the sorption has been well reported, the ability of such particles to release agrochemicals into deeper layers of the soil is controversial. An important ratio to assess this potential as a vector is the Damköhler number (Da), which compares the time period of contaminant desorption to the time period of particle transportation [156]:

$$\mathrm{D}\mathrm{a}={\displaystyle \frac{{\tau }_{\mathrm{transport}}}{{\tau }_{\mathrm{desorption}}}}$$

In this framework, Da < 0.01 indicates decoupled transport, where MPs-NPs facilitate relocation, whereas Da > 100 indicates full equilibrium, where desorption occurs too rapidly for significant vector effects. Modeling studies suggest that for most organic contaminants, desorption is too rapid to facilitate significant transport under typical, slow soil water flow (e.g., 1 m year⁻¹). However, under preferential flow conditions (e.g., 1 m h⁻¹ in macropores), MPs larger than 10 µm can contribute to the relocation of highly hydrophobic contaminants with log K_ow > 5 [156].

The Trojan horse effect has become a relevant issue in bioavailability despite the restriction of the long-distance transport of goods using physical means. Specifically, NPs reach equilibrium quickly and become active as reactive colloids rather than carriers of transport [156]. For example, cationic polystyrene NPs (NPs-NH₂) were shown to adsorb glyphosate, which causes an antagonistic effect on the growth of algae by decreasing the immediate bioavailability of the herbicide and at the same time improving NP stability [151].

The major threat highlighted by recent studies is the biochemical consequence of this comorbidity. According to Lagos et al. [159], the combination of MPs and fungicides has a more significant adverse effect on the soil nitrogen cycle and the organization of the microbial community than the effects of individual contaminants. Similarly, MPs-NPs can act as localized reactive phases that regulate the bioavailability of agrochemicals in the rhizosphere [160]. Even though MPs-NPs typically occur in lower amounts than natural vectors, such as soil organic matter [150], their surface chemistries and persistence present a distinct and potentially disruptive constituent of the soil chemical matrix [147,161].

4.1.3. Influence of Soil Physicochemical Factors

Soil physicochemical properties exert a profound and differential influence on MP and NP interactions with agrochemicals. NPs demonstrate substantially greater sensitivity to soil chemistry than MPs due to their higher surface energy and dominant colloidal behavior [162].

Soil pH has a decisive influence on electrostatic interactions, especially with regard to the ionizable agrochemicals and charged NPs. The augmented MP-NP transport is reported as increasing pH and a concomitant decrease in ionic strength, where increasing pH decreases cation bridging and increases electrostatic repulsion between negatively charged particles and soil surfaces [163]. Divalent cations (Ca²⁺, Mg²⁺) suppress MP-NP mobility through compression of the electrical double layer and the formation of aggregation or adhesion through bridging with cations [164].

The SOM controls the movement of MPs-NPs through competitive adsorption and hetero-aggregation processes. It may also assist in transport by allowing adsorption onto mobile organic colloids, or alternatively slow down mobility by increasing retention within the soil matrix [165]. MPs-NPs are renowned retention sites under electrostatic attraction and surface complexation using clay minerals and iron/aluminum oxides. The lower the Fe/Al oxide content, the higher the mobility of MPs-NPs [164].

Sandy soils have larger and more interconnected pore spaces, which allow MPs-NPs to move vertically and laterally in greater amounts as compared to clay-rich soils that have smaller pore sizes, and have greater specific surface area, which intensifies physical straining and adsorption [166]. Preferential flow paths (macropores) formed due to bioturbation (earthworms and root growth) can pass through the filtering capacity of the soil matrix, thus providing faster transport of MPs-NPs to depths of 40–50 cm [167].

4.2. Fate and Transport of Agrochemical and MP-NP Complexes in the Soil Profile

4.2.1. Vertical Migration and Transport (MPs vs. NPs)

The vertical migration of MP–NP–agrochemical complexes is governed by a combination of particle properties, soil hydrology, and biological activity, with MPs and NPs following distinct pathways (Figure 3). MPs (1 µm–5 mm) are primarily transported via physical processes like infiltration through macropores and preferential flow paths, where their movement is often size-limited by pore throat constrictions [167]. In contrast, NPs (<100 nm) undergo colloidal transport, governed by DLVO theory (Derjaguin–Landau–Verwey–Overbeek), where their mobility is controlled by electrostatic repulsion, van der Waals attraction, and diffusion [147,156].

In such pathways, larger MPs (>10–100 µm) can be carried away under the influence of high-velocity flow regimes, which may carry along with them hydrophobic contaminants [156]. Field experiments showed that there was little vertical transport of polyethylene terephthalate (PET) fragments and fibers (65–125 µm) within two years, with only approximately 1% of the particles going to depths greater than 4 cm, suggesting that the advective transport of MPs was slow without substantial bioturbation [167].

Earthworms are a significant force in vertical movement. They move MPs-NPs to deeper layers (up to 40–50 cm) through ingestion and egestion of soil and release particle-rich casts into their burrows, known as drilospheres [168]. This is a size and shape dependent transport in which smaller particles and microfibers are mobilized effectively [169]. Distribution is also regulated by plant roots; it has been found that maize roots transport MPs to the 6–12 cm layer, which is likely due to the influence of pores and buoyancy, and the roots of the ryegrass retain them [170].

Due to their small size, NPs can move through the soil matrix; however, they are vulnerable to aggregation and retention by soil solution chemistry, including soil pH and ionic strength. They are believed to move optimally at a diameter of around 1 µm [156]. In turn, macro-particles are more likely to be physically filtered and strained, leading to their concentration in the layer at or close to the surface, unless bioturbation or facilitation occurs at large, connected macropores [147].

Vertical and lateral transport processes are determined by the intensity of irrigation and rainfall. Wet–dry cycles may cause cracking and stimulate vertical movement upwards, but heavy rains can trigger favorable circulation and surface run-off and can redistribute MPs-NPs [37]. The fate, transport, and biological effects of agrochemical–MP–NP complexes are presented in

Table 3. Fate, Transport, and Biological Effects of MP–NP–Agrochemical Complexes.

Agrochemical Complex	Dominant Transport Mechanism	Effect on Degradation/Persistence	Bioaccumulation and Uptake Pathway	References
MPs-Atrazine	Facilitated adsorption	Aged MPs have higher adsorption (0.742 mg g−1), increasing herbicide persistence	Root growth inhibition due to direct contact; dry or wet mass reductions	[154]
MPs-Chlorpyrifos	Biogenic transport	MPs reduce chlorpyrifos (CPF) degradation, resulting in lower 3,5,6-Trichloro-2-pyridinol (3,5,6-TCP) formation, increasing soil persistence	Earthworm-mediated transport; 62.5% mortality and 17.6% weight loss in earthworms	[171]
MPs-S-Metolachlor	Carrier effect	Adsorption varies by polymer: PVC > PP > PE; interaction alters environmental transfer	Vector-mediated distribution changes exposure for aquatic or soil organisms	[155]
MP-Simazine	Leaching potential	Mineralization decreases (lower 14CO2), enzyme activities decrease 20–46%; half-life increases	Residual risk to surface or groundwater due to slower degradation	[172]
MP-Imidacloprid	Polymer-dependent	Aged MPs adsorb more, slowing vertical migration; polymer type affects transport	Oxidative stress in soil invertebrates (GSH/GSSG ratio drops 20–40%)	[173]
NP-Imidacloprid	Charge-selective transport	Co-exposure shifts metabolic pathways (IMI-NTG and 5-OH-IMI)	PS-COOH accumulates in roots/shoots; PS-NH2 inhibits growth	[174]
NPs-Atrazine	Eco-corona modulation	Humic acid alters NP surface charge, reducing hetero-aggregation and bioavailability	Synergistic toxicity mitigated by HA; oxidative stress genes regulated (SOD, catalase, GR)	[175]
NPs-Glyphosate	Colloidal transport	Strong adsorption reduces migration 24–44%; half-life extended	Reduced immediate toxicity (~44% lower algal growth inhibition)	[151]

.

4.2.2. Degradation Kinetics and Persistence

MPs and NPs can significantly alter the degradation kinetics and environmental persistence of co-occurring agrochemicals through several mechanisms, often leading to extended residence times (Figure 3 and Figure 4).

Figure 3. Schematic representation of hydrolysis [176,177,178,179], photolysis [180], and oxidative degradation [181] of some common agrochemicals.

Figure 3. Schematic representation of hydrolysis [176,177,178,179], photolysis [180], and oxidative degradation [181] of some common agrochemicals.

By adsorbing agrochemicals, plastics shield them from key degradation processes. This includes protection from microbial enzymatic attacks, UV photolysis on the soil surface, and hydrolysis in the aqueous phase [182]. The hydrolysis of atrazine proceeds as follows [183]:

C₈H₁₄ClN₅ + H₂O → C₈H₁₅N₅O + HCl

Polyethylene MPs have been shown to enhance the persistence of atrazine, whereas polystyrene MPs have been shown to delay the degradation of 2,4-D [182].

MPs can alter the structure and function of soil microbial communities, potentially inhibiting the microbial consortia responsible for pesticide degradation. Microbial carbon-phosphorus (C–P) cleavage of glyphosate is represented by [184]:

C₃H₈NO₅P → CH₆NO₃P (AMPA) + C₂H₂O₃ (Glyoxylate)

PVC-MPs, in particular, have been noted to exert strong effects on carbon-cycling enzymes, which can indirectly slow herbicide breakdown [31,150].

The high surface area and strong adsorption affinity of aged NPs can lead to encapsulation, making the agrochemical less accessible for degradation. Such encapsulation can alter degradation pathways, potentially leading to the accumulation of different and sometimes more toxic metabolites [150]. The net effect is an increase in the observed half-life (DT₅₀) of agrochemicals. Comparative studies indicate that agrochemical half-lives can be significantly longer in MP-associated systems than in plastic-free soil. The persistence of plastics is also notable; estimates suggest MP half-lives of 2.5–4 years in agricultural soils [185].

4.2.3. Bioaccumulation and Trophic Transfer

The incorporation of MP–NP–agrochemical complexes into biological systems represents a critical exposure pathway, with implications for ecosystem health and food safety.

Earthworms and other soil invertebrates readily ingest MPs-NPs. This ingestion is dose-dependent and can lead to reduced growth and survival [150,171]. The associated agrochemicals can be absorbed by the organism, with bioavailability influenced by the desorption rate from the plastic carrier. The dioxygenation of phenanthrene [a polycyclic aromatic hydrocarbon (PAH) model] typically proceeds as follows [186]:

C₁₄H₁₀ + O₂ → C₁₄H₁₂O₂ (diol) → C₁₁H₈O₃

Studies on phenanthrene have suggested that while MPs-NPs reduced overall bioaccumulation by sequestering the contaminant, they can also alter its organ-specific distribution within the organism [187].

NPs have an increased probability of entering cells through endocytosis or other mechanisms owing to their particle size; hence, they may provide a direct route for agrochemicals to biological tissues. This enhanced bioavailability is a major problem compared to that of larger MPs [150].

Plants can absorb NPs and potentially small MPs. The primary route is often the “crack-entry” pathway at the root tips, with particles then translocating via the apoplastic or symplastic transport systems of the xylem to the aerial tissues [188]. Confocal microscopy has confirmed the presence of PS-NPs in the root epidermis, cortex, and xylem of crops like rice [150]. This uptake provides a direct route for MPs or NP-bound agrochemicals to enter the food chain.

The transfer of these complexes from soil to plants, herbivores, and potentially predators create a pathway for human exposure. This risk is indicated by the presence of plastic additives, including phthalates, in human urine and MPs in edible food products [150]. Agrochemicals in combination with MPs-NPs have been linked to either synergistic or antagonistic toxicological effects and are related to an increased expression of antibiotic resistance genes in soil fauna, thus posing a broader ecological threat [171].

5. Impact of Agrochemicals and MPs-NPs on Soil Health

5.1. Impacts of Agrochemicals on Soil Properties

5.1.1. Effect of Agrochemicals on Soil Physical Properties

Different agrochemical applications on soil can influence soil physical properties by altering soil texture, structure, porosity, bulk density, water retention, and infiltration rate, which are essential to determine soil health and crop production (Figure 5). For example, the non-judicious use of chemical fertilizers may affect soil aggregate formation and reduce soil organic matter, leading to weakened soil structure and changes in porosity that limit aeration and water movement [189]. Massah et al. [190] stated that excessive fertilizer use caused soil compaction, increasing bulk density (1.34–1.80 Mg m⁻³) and root penetration resistance (0.89–3.54 MPa) while drastically limiting permeability, available water, and crop yield. However, the combined application of organic and inorganic fertilizers enhanced the aggregate stability and porosity of soil [191]. Pesticide degradation can modify soil physical properties, especially infiltration rates and water retention [42]. Furthermore, residual by-products formed during pesticide degradation can alter soil texture and structural integrity by interfering with the aggregation of soil particles [192]. Such interactions may lower soil porosity and obstruct pore spaces, ultimately restricting water movement and reducing infiltration rates [42,193]. In addition, Yasir et al. [42] reported that pesticide breakdown can alter microbial activity and deplete organic matter (OM) pools, which are essential for maintaining soil structure and WHC.

5.1.2. Effect of Agrochemicals on Soil Chemical Properties

Agrochemicals, including fertilizers and pesticides, strongly influence soil chemical properties depending on their type, application rate, and duration of application (Figure 5). Long-term use of ammonium-based nitrogen fertilizers accelerates soil acidification through nitrification, leading to base cation leaching and reduced nutrient availability [194], whereas continuous phosphorus inputs can cause P accumulation and fixation, increase eutrophication risk, and induce micronutrient deficiencies such as Zn and Fe [17]. Excessive fertilizer application also increases soil EC and salinity, markedly in irrigated systems, negatively affecting nutrient uptake and crop performance [195,196]. Pesticides further modify soil chemistry by altering nutrient transformations, enhancing organic matter mineralization, and contributing to soil acidification and the accumulation of toxic residues and heavy metals [197]. These processes can disrupt the soil carbon cycle and impair soil organic matter formation and its structural stability [198]. Merouani et al. [199] revealed that pesticide application reduces soil organic carbon, alters P dynamics through mineralization and plant uptake, and affects soil pH, with varietal differences in tolerance. Moreover, agrochemical use, especially macronutrient fertilizers combined with glyphosate, reduces soil microbial populations, micronutrient availability including Fe and Zn, and the nutritional quality of crops [200]. Additionally, pesticide degradation products and residue-induced microbial shifts may reduce soil CEC by limiting nutrient exchange sites, thereby constraining nutrient retention and reducing long-term soil fertility [201].

5.1.3. Effect of Agrochemicals on Soil Biological Properties

Agrochemicals directly or indirectly affect microbial diversity, richness, and evenness, thereby limiting nutrient availability (Figure 5). Pesticides exert direct effects on soil microbes by altering enzyme activities and metabolic functions [170,198,202,203]. For instance, Filimon et al. [204] reported that thiamethoxam decreased soil phosphatase activity by 6.5% and nitrifying bacteria by 58.1%, whereas Cypermethrin reduced dehydrogenase activity by 32.8% and nitrifying bacteria by 74%. Similarly, organophosphorus insecticides such as malathion, diazinon, and dimethoate significantly reduced microbial population growth depending on their concentration and exposure time [205]. Repeated applications of an insecticide such as malathion, endosulfan, chlorpyrifos, and lindane markedly suppressed soil nitrification and denitrification processes, even at field-recommended application rates [42,206]. Likewise, Schäffer et al. [207] stated that cyfluthrin and imidacloprid reduced nitrification and denitrification by enhancing sulfur oxidation. Pignatello et al. [208] reported that captan, anilazine, chlorothalonil, and benomyl increased nitrogen mineralization, likely by modifying the interaction between fungi and bacteria. However, mancozeb, prosulfuron, quinalphos, and chlorothalonil have been reported to inhibit nitrification [42,209], whereas nitrification was enhanced by metalaxyl and mefenoxam [210]. Herbicides decrease the functions of soil microbes when applied along with fertilizer. The application of phosphorus fertilizer at higher doses suppresses the diversity of AMF and their colonization [211,212]. Furthermore, herbicides also affect nitrogen fixing and nitrogen transforming by free-living bacteria in soil. Terbutryn, simazine, premteryn, and benzoate herbicide inhibit nodulation and decrease nitrogen fixation in legumes, whereas butachlor primarily enhances nitrogen fixation but later causes a decline owing to toxic buildup [42,213]. Organochlorines such as 2,4-D suppress Rhizobium activity and reduce nitrogenase and phosphatase enzyme activities [214,215]. Moreover, glyphosate also decreases azotobacter and phosphatase enzyme activity [216,217].

5.2. Impacts of MPs-NPs on Soil Properties

5.2.1. Effect of MPs-NPs on Soil Physical Properties

As previously indicated, soil is recognized as the ultimate environmental sink for MPs-NPs [218]. Due to their limited infiltration capacity, MPs-NPs are predominantly retained in the surface soil layer [219,220]. The existence of MPs-NPs in soil reduces the soil bulk density (BD) owing to their lower density compared to soil particles, thus altering soil structure, pore space, infiltration rate, and nutrient and water retention capacities [4,18,19,221,222]. These structural changes may further facilitate the downward transport of MP-NP-associated pollutants into deeper soil horizons, with the magnitude of these effects being strongly governed by soil texture [157,160]. The impact of MPs-NPs on soil bulk density varies with soil type. For example, bulk density decreased by 9% in Vertisol following the addition of polyester (PES) fibers, whereas no significant changes were observed in Entisol and Alfisol soils [223]. Beyond bulk density, MPs-NPs significantly influence soil WHC, aeration, and root penetration, depending on soil texture, particle size, and MP-NP concentration [224]. At higher concentrations (2%), PE particles markedly altered WHC: small-sized fibers (150 µm) enhanced WHC in loamy soils, while larger particles (950 µm) reduced WHC in sandy soils. In contrast, no significant effects were observed at lower concentrations (0.5%) [225]. In addition, they significantly affect water-stable aggregates, with outcomes depending on polymer type and particle shape [226]. Several studies have reported that PES fibers increase water-stable macroaggregates [227,228], whereas polyethylene terephthalate microfragments and polystyrene microspheres reduce macroaggregate formation and promote microaggregate development [229]. Overall, the incorporation of MPs-NPs tends to increase soil porosity and aeration and alter water retention, evapotranspiration, and aggregate stability.

5.2.2. Effect of MPs-NPs on Soil Chemical Properties

Soil chemical properties can be negatively influenced when MPs-NPs alter soil pH, electrical conductivity (EC), and nutrient cycling (Figure 5). The effect of MPs-NPs on the various chemical properties of soil is presented in

Table 4. Effect of MPs-NPs on the chemical properties of soil.

Type of Soil	MP-NP Type and Size/Concentration	Effect on Soil Chemical Properties	References
Sandy loam	Types: PE, PLA Size: 100–154 µm	Soil pH: decreased by PE; increased by PLA	[247]
Sandy	Types: LDPE, Bioplastic mulch films Size: 50 µm–1 mm, 4–10 mm	Soil pH: enhanced	[248]
Sandy loam	Types: PE, PA, PS, PHB, PLA, PBS Size: 39–80 µm	Soil pH: remains unchanged at 0.2% concentrations	[236]
Laterite	Types: Plastic mulch film Size: 0.2%, 0.6% Films; 0.5 cm × 0.5 cm, thickness: 0.008 mm	SOM: reduced SOC and SOM contents	[249]
Paddy field (Abandoned)	Types: PLA Size: 20–50 µm	SOM: total DOC content decreased Soil nutrients: soil NH4+-N was negatively affected, while NO2− and NO3 levels increased, with no significant change in inorganic phosphorus	[29]
Farmland	Types: Plastic mulch Shape: Films	Soil nutrient: N content decreased	[250]
Fluvo-aquic soil	Types: PE, PLA Doses: 0.5%, 1.0%, 1.5%, 2.0%, 2.5% w/w	SOM: increased SOC and SOM content Soil Nutrient: PE at higher concentrations increased NH4+-N content	[251]
-	Type: PE Dose: 10%	CEC: decreased	[252]
-	Type: LDPE Dose: 1.0%	EC: decreased	[232]
-	Type: LDPE Dose: 2.0%	EC: unchanged
-	Type: PE Dose: 5.0%	Soil nitrogen: increased	[253]
-	Type: PVC Dose: 0.5%	Soil nitrogen: increased	[254]
-	Types: PE and PS Doses: 0.2% and 2.0%	Soil nitrogen: decreased	[236]
-	Types: PS, PE, PVC Dose: 1.0%	Potassium: decreased	[255]
-	Types: LDPE and PS	Potassium: decreased	[256]
-	Type: PVC Doses: 0.1%, 1.0%	Phosphorus: decreased	[257]
-	Type: PE Dose: 0.8%	Phosphorus: decreased	[258]

. MPs-NPs’ influence on soil pH varies depending on the type, shape, dose, and exposure time. Several studies have reported that the presence of MPs in soil increased, decreased, or did not vary soil pH [1,230,231,232]. For example, HDPE and tire debris MPs decreased the soil pH value significantly [233,234]. MPs-NPs also affect soil organic carbon (SOC) dynamics and organic matter (OM) content due to a slow degradation rate in the environment. MPs-NPs were found to increase dissolved organic carbon (DOC) and total organic carbon (TOC) levels and to elevate emissions of CH₄, N₂O, and CO₂ [235,236,237,238]. These effects may arise because MPs-NPs act as a source of SOC and interact with the plant-soil ecosystem, thereby altering global carbon cycling dynamics [235,239,240]. For instance, Gao et al. [241] reported that soil CO₂ emission was increased by 28.7% due to MPs-NPs. Moreover, MPs-NPs may disrupt nutrient availability in soil and various biochemical processes. MPs-NPs at higher concentrations markedly enhanced humus, fulvic acid, dissolved organic matter (DOM), dissolved nitrogen content (DON), NO³⁻, and PO₄³⁻ [18]. Chen et al. [29] and Iqbal et al. [242] reported that biodegradable plastics, polylactic acid (PLA)-MPs, enhanced NO³⁻ and NO²⁻ concentration while reducing NH⁴⁺ contents. In addition, MPs-NPs accelerate fluorescein diacetate hydrolase (FDAase) activity, increase nutrient content, and enhance the deposition of humic acid and the mineralization of nutrients, thus influencing the activity of soil enzymes and soluble nutrient storage [243]. Jiang et al. [244] and De Souza Machado et al. [228] found that the presence of MPs-NPs significantly influences P and N dynamics in rice and onion fields, respectively. MPs-NPs increase the C:N ratio and enzymatic activity, thus affecting carbon turnover [1,232,245,246].

5.2.3. Effect of MPs-NPs on Soil Biological Properties

The most significant impacts of MPs-NPs on microbial communities include OM decomposition, nutrient cycling, and overall soil health (Figure 5). MP-NP exposure affects microbial communities and their diversity in the soil rhizosphere [254,259,260]. For instance, microbial communities were reduced by approximately 268 to 60% in Mollisols and Inceptisols soil over a year of MP-NP exposure, reducing the decomposition of OM and nutrient cycling [261]. Similarly, MPs-NPs significantly altered soil bacterial diversity when soils were applied with 0.5 or 5 μm MPs-NPs in 100 and 1000 μg PS MPs-NPs g⁻¹ soil [262]. In contrast, polyhydroxyalkanoates (PHAs) and PS further increase microbial community growth and density due to their composition and biodegradability in the environment [262,263]. Different nitrogen-fixing bacteria, nematodes, and earthworms showed decreased growth due to MP-NP exposure [32]. Fei et al. [253] reported that MPs-NPs also reduced nitrogen-fixing bacteria at a 0.5% concentration and disturbed nutrient cycling by 15–25%. In addition, MPs-NPs alter enzyme activity, including alkaline phosphatase, N-acetyl-β-glucosaminidase, urease, dehydrogenase, β-D-glucosidase, and cellobiosidase in soil [264,265]. Plastic particles reduce microbial activity as they compete with microbes for physicochemical habitats, and as a result, extracellular enzyme activity declines [266]. The different forms and types of MPs-NPs may also have a significant effect on enzyme activities in soil. For instance, PP, PS, and PVC may either increase or decrease fluorescein diacetate hydrolase activity in soil, based on their polymer type, whereas PE and PVC enhance urease and acid phosphatase enzyme activity [4,253,267]. Furthermore, MPs-NPs have a notable impact on the AMF for root colonization. DeSouza Machado et al. [228] revealed that PS and PP fragments increased root colonization by 8- and 1.4-fold, respectively, whereas PET fragments decreased by approximately 50%. However, PA beads, PEHD fragments, and PS fragments did not show any effect on root colonization. Additionally, a recent study indicates that MPs-NPs help to disseminate antibiotic resistance genes (ARGs) by altering microbial diversity and structure [268,269]. These alterations would notably affect the carbon pools, flux, and nutrient cycling, with responses varying owing to the MP-NP type and rate of application. The effect of MPs-NPs on the soil microbial community is presented in

Table 5. Effect of MPs-NPs on soil microbial community.

Types of Soil	MP-NP Type and Size/Concentration	Effect on Soil Biological Properties	References
Dry Soil	Type: PE Size: 2 mm × 2 mm × 2 mm	Increased soil bacterial turnover	[6]
Field Soil	Type: LDPE Size: <400 µm	Caused earthworm mortality	[270]
Pot Soil	Types: PVC, PE, PS Size: 200 µm	PE reduced rhizosphere bacterial community more than PS and PVC	[254]
Paddy field (Abandoned)	Type: PLA Size: 20–50 µm	Microbial community remained unchanged	[29]
Field Soil	Type: PHAs Size: N/A	Enhanced microbial turnover and nutrient-use efficiency	[263]
Farmland Soil	Type: Plastic mulch Shape: Films	Actinomycetes decreased, while Proteobacteria increased	[250]
-	Type: PS Size: 0.082, 42, 182 μm	Damaged intestinal tissues and cells and disrupted gut microbial communities	[271]
Loamy Sand Soil	Type: HDPE Size: 15.94 µm	Springtails’ growth, reproduction, and survival rate were reduced	[272]
Fluvo-aquic Soil	Types: PE, PLA Doses: 0.5%, 1.0%, 1.5%, 2.0%, 2.5% w/w	Lower concentrations increased the microbial community, while it was reduced at higher concentrations	[251]
Field Soil	Type: PS Size: 0.05, 0.5, and 5 μm	Reduced Rhizomicrobium bacteria	[262]
Mangrove ecosystem Soil	Types: PLA and PE	Decreased Proteobacteria abundance in the subsoil, with increased Sulfurovum and Desulfatiglans abundance in the subsoil and changed interaction of soil microbial communities	[273]
Farmland	Types: PE and PP Doses: 1% and 5% w/w	Enhanced Bacteroidetes and Acidobacteria abundance, while reducing Chloroflexi and Deinococcus thermus abundance	[274]
Loamy Sand	Types: PA, PE, PS, and PMMA Sizes: <5 mm, <0.1 μm	PA and PE enhanced microbial activity, while PS and PMMA reduced microbial activity	[4]
Various soils	Types: PE and PET	Altered microbial diversity	[275]
Field soil	Types: PP and PE Size: 150 μm	Change in microbial population	[276]
Field soil	Types: PBS and PLA Size: 150–180 μm	Proteobacteria diversity increased, and Actinobacteria diversity decreased	[276]
Arable Agricultural Topsoil	Types: PP and LDPE Size: 200–630 μm	Decreased microbial biomass	[277]
Cropland Soils	Type: PE Doses: 0.1, 0.5, 1% w/w	Stimulated or suppressed microbial activity depending on soil amendment	[278]
-	Types: PS, PE, and PA Size: 5 μm, 20 μm, 50 nm	Low concentrations of NPs/MPs stimulated bacterial and viral growth	[278]
-	Type: PE Size: 50–200 nm	Suppressed microbial activity and disrupted phosphorus-removing microbial communities	[279]
Constructed wetlands	Type: PS Size: 70–110 nm	Reduced microbial activity, impairing nutrient removal and increasing greenhouse gas emissions, while AMF inoculation mitigated these effects	[280]
Tourist areas at Alqueva’s reservoir edges	Type: PHAs	Soil microbes, including plastic-degrading bacteria, remained active and potentially mitigated POP	[281]
-	Type: PE (white microbeads, W and fluorescent blue microbeads, FB)	C. elegans showed minimal sensitivity, and V. fischeri showed approximately 27% and 6% inhibition with W-MPs and FB-MPs, respectively	[282]

.

Although several laboratory studies employ high concentrations of MPs-NPs, these exposure levels are often much higher than those observed in real-world environmental conditions. Therefore, for a more accurate understanding of the risks posed by MPs-NPs in agroecosystems, it is essential to distinguish between high-dose laboratory experiments and the more realistic, environmentally relevant concentrations typically encountered in field conditions. High-dose studies, while valuable for understanding mechanisms, may overestimate the potential impacts and risks to ecosystems, particularly in terms of long-term exposure and cumulative effects.

6. Ecotoxicological Effects of Agrochemicals and MPs-NPs on Ecosystems

Agrochemicals and MPs-NPs are increasingly recognized in terrestrial and connected aquatic environments. Agrochemicals and MPs-NPs are largely used in agricultural systems and are directly dumped into ecosystems. Previous studies have focused on aquatic environments, and the present research highlights that agrochemical and MP-NP pollutants in agricultural fields cause severe effects on terrestrial ecosystems (Figure 6).

6.1. Impact on Terrestrial Ecosystems

Due to improper management, agrochemicals and MPs-NPs can be potentially accumulated in soil-plant systems, ultimately enter into the food chain, and cause human and animal health concerns through food ingestion [283]. Soil microorganisms represent the first line of biological response to co-contamination and can significantly influence the redistribution and fate of agrochemicals and MPs-NPs in soils. Different invertebrates present in soil, especially earthworms, facilitate MP-NP movement by internalizing particles and excreting them into diverse substrate horizons. However, MP-NP accumulation in biological systems mainly depends on the size, shape and concentration [270]. However, NPs are considered smaller particles and more easily transmitted than MPs. Recent studies suggest that with the MP-NP ingestion, accumulated agrochemicals could also be introduced into soil living organisms, with the resultant toxicity. For instance, the singular or combined effects of MPs-NPs (PVC, PP, PE) and agrochemicals such as dufulin, 2,4-D, glyphosate, atrazine, etc., significantly alter the metabolic pathways and metabolites of different species of earthworms [284,285,286], which increases oxidative stress and leads to abnormal gene expression. Some other researchers have stated that higher concentrations of glyphosate and PE negatively influence earthworm activity while reducing their body weight and burrow volume [287]. Moreover, interactions between agrochemicals and MPs-NPs in soil can alter the soil environment by reducing oxygen availability and changing the physicochemical properties, which may lead to an imbalance in microbial populations. As a result, some microbes thrive, while others, notably those involved in nutrient cycling, such as nitrogen-fixing bacteria, may decline [32,253]. The effects of MPs-NPs on different earthworm species are presented in

Table 6. Effects of MPs-NPs on different earthworm species.

Earthworm Species	MP-NP Type and Size	Effects	References
Eisenia fetida	Type: PS Size: 58 µm	Burrowing activity increased	[288]
	Types: PE and PS Size: 50–150 µm	Acts as a vector for hydrophobic organic pollutants	[289]
	Type: LDPE Size: <400 µm	Caused skin damage	[290]
	Type: PS Size:100–1300 nm	Increased GSH content and decreased SOD content	[291]
	Type: LDPE Size: 550–1000 µm	CAT, SOD, and GST activities decreased, and MDA content increased	[284]
	Type: PS Size: 100–1300 nm	Caused DNA damage	[292]
	Type: PES Size: 0.1, 1, 10, and 100 µm	CAT and GST genes decreased, and TCTP, SOD, and MT genes remained unchanged by large MPs	[292]
	Type: LDPE Size: 550–1000 µm	TCTP, HSP70 genes increased	[284]
	Type: LDPE Size: <400 µm	Induced neurotoxicity and stimulated acetylcholine esterase activity	[290]
	Type: PS Size: 100 nm	Affected gut microbiome	[292]
	Types: PS + Phenanthrene Size: 100 nm	Decreased phenanthrene-degrading bacteria in the gut	[292]
	Types: LDPE + atrazine Size: 550–1000 µm	Oxidative stress	[284]
	Types: MPs + dufulin Size: 40–50 µm	Increased oxidative damage	[285]
Lumbricus terrestris	Type: PES	Metallothionein gene expression changed	[293]
	Type: PE Size: <150 µm	Decreased growth and increased mortality rate	[270]
	Type: PE Size: <150 µm	Burrowing activity increased	[294]
	Types: PS, PP, PET, LDPE Size: 250 µm	Caused physical damage of mucus membranes	[295]
	Types: LDPE + glyphosate Size: <150 µm	The volume of galleries and the weight of earthworms decreased	[296]
Eisenia andrei	Types: PE and UV-aged PE	PE enhanced reproduction, whereas UV-aged PE reduced reproduction	[297]
	Type: PE Size: 250 and 1000 µm	Molecular changes and damage to the gut	[298]
	Type: PE Size: 180–212 µm and 250–300 µm	Serious damage of intestinal tissue and sperm plasma membrane as well as reduced sperm density	[299]

.

The impact of MPs-NPs on plants is closely tied to the way they interact with the agrochemicals present in the soil; the effect of MPs-NPs has been described in different ecotoxicological studies. Previous studies have shown variable effects ranging from inhibition to negligible effects or no effect of MPs-NPs on seed germination [300]. MPs-NPs may inhibit seed germination by blocking voids and seed pores, ultimately inhibiting water imbibition, or enhancing germination via the nano-priming effect of MPs-NPs, which facilitates seed coat penetration and upregulates water transport genes [268,301,302]. Lian et al. [303] reported that PS-NPs have no effect on wheat seed germination. However, Wang et al. [304] investigated the effects of 6.5 and 13 μm PE-MPs at different concentrations on Glycine max and Vigna radiata and reported that medium concentrations (50 or 100 mg L⁻¹) of PE-MPs significantly reduced Glycine max germination, with a decrease of approximately 20% compared to the control when exposed for 7 days. In contrast, 6.5 μm PE-MPs showed no consistent inhibitory effect at different concentrations; at low concentrations (10 mg L⁻¹), the germination rate decreased by approximately 75% and then increased by 83% at a high concentration (500 mg L⁻¹) compared to the control. In contrast, the germination rate of Vigna radiata was not significantly affected by different sizes and concentrations of μm PE-MPs, while an 88–100% increase in germination speed compared to the control was observed. Additionally, Pignattelli et al. [305] conducted a study and demonstrated that exposure to PP, PE, PVC and PE + PVC mixtures inhibited the germination of Lepidium sativum, and reported 55% inhibition of germination relative to the control. Hydroponic experiments were conducted by Li et al. [300] to evaluate the germination and growth of Pisum sativum under exposure to different particle sizes and concentrations of PS-MPs. They demonstrated that small particle sizes (0.08 μm) combined with high concentrations (2000 mg L⁻¹) exerted the strongest suppression of germination and growth of Pisum sativum. In contrast, low concentrations or large particle sizes produced weaker or occasionally stimulatory effects.

Persistent agrochemicals and MPs-NPs accumulate through root tissues and subsequently partition within upper ground plant tissues, consequently disrupting plant growth, physiological mechanisms and nutrient uptake [306]. The ability of MPs-NPs to accumulate in plant tissues depends largely on particle size, transport mechanism, exposure time and crop type [307,308]. The presence of MPs-NPs can increase the persistence and bioavailability of agrochemical (OC and OP) residues in the rhizosphere by the absorption of hydrophobic agrochemicals. These processes induce oxidative stress and toxicity in plants and limit food safety [11,301,309,310]. Previous studies have indicated that at soil-relevant concentrations, PE and PE particles reduce plant biomass and root length, particularly when coupled with persistent herbicides like glyphosate and atrazine or insecticides like imidacloprid [20,160]. An integrated molecular bioassay was used by Fajardo et al. [311] to assess the effects of PE and the interactions with OCs. They reported that Zea mays exhibited inhibited leaf growth and progressively enhanced root growth inhibition, especially under contaminants and mixed treatments. These effects were accompanied by the altered expression of antioxidant-responsive genes, indicating oxidative stress-mediated responses [301]. In contrast, another research group aimed to assess the single and dual impacts of PE-MPs and contaminants on different organisms and reported that PE-MPs have the capability to adsorb contaminants, with dual impacts ultimately resulting in reduced leaf development in Lactuca sp. plants [282].

6.2. Impact of Aquatic and Soil-Water Interface Ecosystems

Agrochemicals and MPs-NPs create contamination pathways, as they are frequently transported from soils to freshwater and coastal environments through runoff and leaching [312]. Similar to terrestrial systems, MPs-NPs can act as vectors for persistent agrochemicals in aquatic systems, potentially increasing their bioavailability and toxicity. MPs and NPs alone often exhibit low or negligible toxicity toward certain aquatic organisms. Dual exposure to MPs-NPs and agrochemicals has been shown to reduce algal growth and photosynthetic efficiency and modify microbial diversity through metabolic alteration, neurotoxicity and genotoxicity [313,314]. Tang [315] further suggested that interactions between agrochemicals and MPs-NPs may modify the toxicity of agrochemicals toward target pests and crops, indicating that the particles of MPs and NPs can influence agrochemicals’ behavior beyond simple co-occurrence.

However, toxicity responses are not uniform across aquatic organisms. Liu et al. [316] reported that bioluminescent gram-negative bacteria such as Aliivibrio fischeri demonstrated no significant reduction in bioluminescence or acute toxicity upon exposure to pesticides (diethofencarb and pyrimethanil) and MPs (PVC, PS and PET). In contrast, another biotoxicity study revealed that MPs inhibited Escherichia coli growth, particularly after agrochemical adsorption [317,318]. Moreover, in aquatic macrophytes such as Salvinia cucullata, glyphosate and PS-MPs exhibited synergistic effects at high concentrations, manifested as increased leaf yellowing [255]. Previous studies have indicated that combined exposure to PE-MPs and agrochemicals (GYL) can reduce photosynthetic pigments, disrupt nutrient uptake and alter antioxidant enzyme (SOD, POD, CAT) activity in aquatic plants [319].

Beyond microbial and plant responses, numerous studies have examined the effects of agrochemicals and MPs-NPs on planktonic organisms and aquatic fauna (crustaceans, bivalves and fish) to evaluate the broader ecological consequences in aquatic ecosystems. Chronic exposure of Daphnia magna to PVC-MPs, PET-MPs and PP-MPs has been shown to disrupt the first brood, reducing offspring numbers and altering antioxidant defense mechanisms [320]. Additionally, metabolomic analysis revealed that MP ingestion alters energy and lipid metabolism in Daphnia magna, contributing to decreased survival rate and physiological performance [260]. Some other researchers also reported that the activity of deltamethrin and PE-MPs significantly reduced the number of surviving females and the brood number of Daphnia magna by 51% and 46%, respectively [321].

Higher trophic levels of aquatic ecosystems also display sensitivity to agrochemical and MPs-NP exposure. For instance, PE-MPs have been associated with a negative influence in Danio rerio and Carassius auratus, leading to decreased hatching rates and larval survival rates in Danio rerio and physiological impairments in Carassius auratus along with body weight reduction and gastrointestinal tissue damage [322]. Additional studies concluded that exposure to MPs-NPs can induce oxidative stress in fish gill tissues and disrupt essential physiological processes [323]. Under laboratory conditions, a group of researchers examined the effects of pristine PS-MPs pellets and Chlorpyrifos (CPF) PS-MPs pellets on Onchorhynchus mykiss and highlighted notable alterations in histomorphometric markers of fish skin, gills and guts under separate and co-administered concentrations of CPF PS-MP pellets, while pristine PS-MP pellets induced toxicity and enhanced the adverse effects of CPF significantly in the gill tissue [324]. Moreover, MPs-NPs have the putative ability to alter the toxicity of agrochemicals. Hanachi et al. [314] demonstrated that PE-MPs enhance the bioavailability of CPF to fish and toxicity in amino acid profiles, lipid composition and protein levels in Onchorhynchus mykiss. Moreover, the small terrestrial crabs Leptuca festae and Minuca ecuadioriensis also demonstrated higher mortality rates in the presence of OPs and MPs [325]. A study conducted by Bringer et al. [326] evaluating the effect of chlortoluron and HDPE (separately or in combination) on Crassostrea gigas reported that Crassostrea gigas shell growth was reduced when exposed for 24 days to cocktail conditions (chlortoluron, 97 µg L⁻¹ + HDPE, 108 MPs mL⁻¹). A distinct cohort of researchers reported similar findings on the enzymatic activities of Scinax squalirostris regarding MP–herbicide interactions and stated that PE-MPs increase glutathione-S-transferase (GST) activities but decrease carboxylesterase (CbE) activities, whereas GLY-PE-MPs increase GST activities and GA-PE-MPs reduce acetylcholinesterase and CbE activities [327].

However, minor or antagonistic interactions between MPs-NPs and agrochemicals have been observed. For instance, in Porcellio scaber, MP-NP exposure reduced the bioaccumulation of chlorpyrifos and enhanced immune parameters [328]. Similarly, other researchers reported that MP exposure reduces the bioavailability of chlorpyrifos and difenoconazole while ameliorating the biotoxicity of difenoconazole in Danio rerio [320,329]. Moreover, the combined effect of PS-MPs and IMD did not modify the DNA integrity and reproduction rate of another crustacean, Ceriodaphnia dubia [330]

These findings collectively highlight that agrochemicals and MPs-NPs impact key aquatic organisms at multiple trophic levels, from zooplankton to fish, through mechanisms involving oxidative stress, metabolic disruption, reproductive impairment and altered biochemical responses. Thus, these impacts are determined by the type of plastic particles and agrochemicals, exposure duration, and aquatic species.

6.3. Trophic Transfer and Food Web Implications

The mobility of agrochemicals and MPs-NPs through terrestrial and aquatic food webs ultimately leads to bioaccumulation and biomagnification of these pollutants, impacting organisms at higher trophic levels, including humans and animals [331,332]. Plants serve as the primary entry point for these pollutants, acting as a key link in the food web and connecting soil, water, animals and humans [275]. Previously, it was believed that MPs could not enter plants due to their larger size, but in recent years researchers have reported that MPs-NPs can accumulate through root tissues, with subsequent partitioning within upper ground plant tissues [11,306]. The bioavailability and toxicity of agrochemicals like glyphosate is further amplified when MPs-NPs are adsorbed by terrestrial organisms and plants [284,320]. Once MPs-NPs are incorporated into plant tissue, they subsequently can be consumed by herbivores and thereby introduce contaminants into higher trophic levels. Moreover, recent studies have investigated fruits and vegetables from markets, and it is evident that different sizes of MPs are present in all fruit and vegetable samples. Researchers also reported MP-NP particles in Malus domestica (195,500 particles for gram or p/g), Pyrus communis (189,550 p/g), Brassica oleracea var. italica (126,150 p/g), Daucus carota (101,950 p/g) and Lactuca sativa (50,550 p/g) [333]. In addition, Cheng et al. [284], Sun et al. [285], and Boughattas et al. [286] reported accumulation of MPs-NPs in soil invertebrates, which serve as crucial prey for higher organisms like birds and small mammals, leading to bioaccumulation of these pollutants in predators [334].

In aquatic systems, the process of trophic transfer is more complex. MPs-NPs act as carriers of agrochemicals and enhance their bioavailability in aquatic food webs. Aquatic microorganisms, plants and phytoplankton adsorb these pollutants, which are then consumed by predators, and the pollutants continue to move up the food chain. The bioaccumulation of these pollutants in terrestrial and aquatic organisms and their subsequent biomagnification in top predators pose significant health risks. Studies have shown that exposure to these pollutants can lead to altered metabolic pathways, liver toxicity, kidney damage, genotoxicity, cytotoxicity and neurotoxicity in humans and animals [301,323,324,335,336]. The effect of MPs-NPs on animal/human health is presented in

Table 7. Effect of MPs-NPs on Animal/Human health.

MP-NP Type and Size	Effect on Animal/Human Health	References
Type: Polyolefin Size: <50 μm	Induced genotoxicity and inflammation	[337]
Type: PS Size: 2 μm–80 nm	Cytotoxicity and genotoxicity	[338]
Type: PS Size: 70, 200, and 500 nm	Clathrin- and caveolae-induced endocytosis mechanisms and bioaccumulation of MPs-NPs	[301]
Types: PE, PP, PS and PVC Size: 6.5–100 μm	Enhanced systemic inflammation and increased the deposition of senescent cells	[339]
Types: PS, PP, and PE Size: 5 mm	Caused pulmonary inflammation, thereby increasing inflammatory cytokines and chemokines in mice	[340]
Type: PS Size: 20 nm	Increased MUC2 expression and induced intestinal toxicity	[341]
Types: Terephthalante, PP, and PE Size: 100 and 600 nm	Induced genotoxicity, production of ROS, and change in the cell cycle	[342]
Type: PS Size: 0.1–5 μm	Membrane damage and disrupted cell membrane functions	[301]
Type: PS Size: 5 μm	DSS mediated colitis in mice and caused intestinal inflammation, increasing the risk of liver injury	[335]
Type: PS Size: 5 μm–50 nm	Increased IL-8 levels, with smaller aminated particles being most harmful to hepatocytes, while larger particles mainly induced apoptosis or inflammation	[343]
Type: PS Size: 1 μm	Changed glucose tolerance, increased accumulation of lipids in the liver, and impaired the composition of hepatic lipid	[260]
Type: PS Size: <15 μm	Liver injury	[344]
Type: PS Size: 3 μm	Acute kidney injury	[259]
Type: PS Size: 1 μm	Altered cell metabolism	[345]
Type: PS Size: 100 nm	MNPLs intensified lipopolysaccharide-induced apoptosis by activating the oxidative stress-mediated IRE1/XBP1 endoplasmic reticulum stress pathway	[346]
Types: PS-MPs with a high-fat diet (HFD) Size: 1 μm	Kidney damage in mice	[347]
Types: MPs and Bisphenol A (BPA) Size: 1–4 μm	Triggered overexpression of oxidative stress and inflammatory mediators, potentially resulting in kidney damage	[348]
Types: CPE, PVC, PMMA, PTFE, SBS, PP, PS, PE Size: 20–100 μm	Effect on oocyte maturation	[349]
Type: MPs-NPs	Disruptede endocrine chemicals in both humans and animals	[350]
Types: MPs and POPs	Caused cytotoxicity, cell damage, endocrine disruption, neurotoxicity, and bioaccumulation of persistent organic pollutants (POPs)	[336]
Type: MPs	Smaller MPs threatened lungs, while larger MPs affected the gastrointestinal tract	[351]

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7. Analytical Challenges and Future Perspectives

The most widely used agrochemicals in agriculture include pesticides (insecticides, herbicides, and fungicides), inorganic fertilizers (N, P, K, and essential micronutrients) and plant growth regulators. Additionally, emerging agrochemical inputs, such as nano-fertilizers and slow- or controlled-release formulations designed to improve nutrient use efficiency, are also frequently applied in farming systems. Among these, pesticides are widely regarded as environmental pollutants due to their persistence, potential for bioaccumulation, and associated ecological and human health risks. In contrast, fertilizers and growth regulators are not inherently pollutants; however, excessive or improper application can cause soil, water, and ecosystem contamination.

In parallel, the excessive use of plastic-based materials in agriculture has led to the accumulation of plastic residues in agroecosystems, resulting in MPs-NPs as a considerable environmental risk to sustainable agriculture. These pollutants have been widely reported across terrestrial and aquatic ecosystems, causing several health concerns for humans, animals, plants, and the environment [19,352]. Therefore, accurate analytical techniques for the detection and quantification of agrochemicals and plastic particles are crucial for defining dose–response relationships, understanding ecological impacts, and developing effective mitigation strategies.

Detection and quantification techniques for pesticides have been well developed and are widely established worldwide. In contrast, analytical methods for the detection and quantification of MPs-NPs remain underdeveloped and lack methodological standardization. The detection and quantification of these particles is often challenging due to their small size and the complexity of the environmental matrices in which they occur [353].

The analysis of MPs-NPs in soil generally requires sequential steps of physical separation, the digestion of organic matter, and chemical identification. As pre-treatment, extraction, digestion of organic matter, and identification protocols differ widely owing to differences in soil composition; the limitations are outlined in

Table 8. Limitations of physical separation and organic matter digestion in MP-NP analysis.

Process	Technique Used	Limitations	Reference
Density separation	Soil is mixed with high-density salt solutions such NaCl, ZnCl2, and CaCl2 to isolate plastic particles from soil particles based on density difference.	High-density polymers PET and PVC may not float in salt due to their sinking behavior, whereas low-density polymers float easily.	[354,355,356,357]
Organic matter digestion	Soil treated with strong acids, alkalis, and hydrogen peroxide (H2O2), to break down organic matter of soil.	These aggressive chemicals can be destructive to the plastic particles in a sample.	[358]

.

Although extensive studies on MPs in aquatic ecosystems have driven the development of several analytical techniques, these methods are mainly applicable to aquatic systems [359]. However, studies on terrestrial ecosystems, particularly agricultural soil-plant systems, remain inadequate. For NPs, the analytical challenge is even greater due to their nanoscale size, strong interaction with soil colloids, and lack of standardized detection and quantification in soils. Appropriate and efficient methods for MP analysis in soil have not yet been established owing to the complex nature of soil matrices, including heterogeneity, high organic matter content, diverse particle sizes, digestion-related issues, and interference from non-plastic materials [360,361]. Furthermore, the complexity of identifying solid analytes in a complex solid matrix substantially complicates MP analysis in soils [169,362]. A comparison of the early methods for MP detection is presented in

Table 9. Comparison of different early identification methods of MPs.

Detection Methods	Technique Used	Polymer Type	Limitations	References
Light microscopy	Optical microscope	Indeterminable	Error ranging from 20% to 70%	[363,364,365]
Fluorescence microscopy	Samples are stained with fluorescent dye (Nile Red is mostly used)	Indeterminable	Soils with high organic content require appropriate pretreatments to prevent false positives from organic matter in soil	[366,367]
Electron microscopy (able to characterize NPs)	Electron microscope (SEM, TEM, REM)	No chemical identification	Requires additional sample preparation steps such as drying, coating, or vacuum conditions	[359,368]

TEM: Transmission Electron Microscopy; SEM: Scanning Electron Microscopy; REM: Reflection Electron Microscopy.

.

7.1. Advancing Laboratory Research and Detection of Pollutants in Plant-Soil Systems

7.1.1. Identification and Quantification Methods of MPs

Early MP analysis methods (Table 9) were essential for particle detection and visualization. However, microscopic observation alone is considered inadequate for reliable MP identification and often results in false positives, where organic fibers are commonly mistaken as MPs. Polymer validation is crucial for overcoming the issue of false positives [369]. Fourier-transform infrared spectroscopy (FTIR) and Raman spectroscopy are commonly used methods for the identification of MP polymers [370,371]. These methods can successfully identify the chemical composition of MPs, making it easier to differentiate polymers from other chemicals in the sample. A comparison of modern methods for MP detection is presented in

Table 10. Comparison of different modern identification methods of MPs in soil.

Identification Method	Polymer Type	Sample Size	Advantages	Limitations	References
SEM-energy-dispersive X-ray	Polymer identification is very limited but provides elemental composition of MPs	Not determined	Simultaneous surface morphology analysis and elemental composition of MPs	Complex pre- treatment procedures required, costly, time-consuming and inefficient	[364,372]
Fourier-Transform Infrared Spectroscopy (FTIR)	PP, PE, PS, PVC, PVA, PU, PTFE, PET, others	MPs larger than 500 µm	Non-destructive, surface-based method that identifies polymer accurately with minimal sample preparation	Requires complete drying of samples and difficult to analyze, opaque or black MPs	[373,374,375]
µ-Fourier-Transform Infrared Spectroscopy (µ-FTIR)	PE, PET, PAN	MPs larger than 10 µm	Quick bulk analysis, automated analysis of particle sizes and polymer types	Full-filter scans frequently require more than one hour per sample	[376,377]
Attenuated total reflectance FTIR (ATR-FTIR)	PE, PP, PS, PET, PVC, PA, PMMA, PU, PTFE	Particles larger than 500 µm	High-quality imaging spectra with high accuracy, less interference from impurities	Highly expensive, not suitable for large-scale MP analysis	[376,378]
Focal plane array FTIR (FPA-FTIR)	PE, PP, PS, PET, PVC, PA, PMMA, PU, PTFE,	MPs larger than 20 µm	Large-area and high-efficiency detection of MPs	Multi-step chemical digestion and long imaging times (up to ~9 h per 47 mm filter)	[379]
Raman Spectrometry	PS, PE, PMMA, PA, PP	MPs smaller than 20 µm	Analysis of non-transparent and dark-colored particles, low sensitivity to water, and reduced dependency on particle thickness and shape	Interference of organic/inorganic and biological contaminants with fluorescence, affecting spectra and identification, pre-purification of samples, time-consuming	[334,368,380]
Hyperspectral imaging technology (HIS)	PE, PP, PVC	300–5000 µm	Simple sample preparation, effective automated and rapid identification	Sensitive to contaminations	[381]
Micro-Raman Spectrometry (µ-RS)	PE, PP, PS, PVC, PET, PA, PC, PU, PMMA, ABS	Up to >1 μm	Spatial and chemical characterization of particles (≥10 µm) at 1 µm resolution	Instrumentation is difficult	[353]
Photoluminescence spectroscopy	PS, PE, PP, PET, PMMA, PC, and PVDF	≥200 µm	Fast optical measurements through simple laser diodes provide characteristic emission spectra	Emission spectra vary with excitation wavelength, requiring determination of optimal wavelength for each polymer	[382,383]

PE: polyethylene; PP: polypropylene; PS: polystyrene; PVC: polyvinyl chloride; PET: polyethylene terephthalate; PVA: polyvinyl alcohol; PU: polyurethane; PTFE: polytetrafluoroethylene; PA: polyamide (e.g., nylon); PMMA: polymethyl methacrylate; PC: polycarbonate; PAN: polyacrylonitrile; PVDF: polyvinylidene fluoride; ABS: acrylonitrile butadiene styrene.

.

7.1.2. Quantification Methods of MPs in Soil

Accurate quantification of MPs is vital for understanding MP pollution in the environment. Mass and quantitative concentrations are the two indicators frequently used to assess the amount of MPs in the environment [374]. The mass of MPs is not easily affected by environmental factors; therefore, mass concentration is more reliable than quantitative concentration for MP quantification [384]. Spectroscopy is a method used for the quantitative analysis and qualitative classification of MPs without damaging the samples. Currently, spectroscopy is often combined with other techniques to count MPs. According to Chen et al. [385] and Prata et al. [378], μ-Fourier-transform infrared spectroscopy (FTIR spectroscopy with microscope) and μ-Raman spectroscopy (Raman spectroscopy with microscope) are used to identify and quantify MPs > 10 μm and >1 μm particle size in environmental samples, respectively, but do not directly provide mass concentration. Several other quantification methods are used to measure practical mass-based MP concentrations, which are listed in

Table 11. Comparison of various quantification methods of MPs in soil.

Quantification Method	Polymer Type	Units	Advantages	Limitations	References
Pyrolysis gas chromatography mass spectrometry (Pyr GC-MS)	PE, PP, PS, PET, PVC	mg g−1	Rapid qualitative and quantitative analysis, simultaneous analysis of whole sample	Extensive sample cleanup for matrix-rich samples, interference from contaminants	[386,387]
Thermal extraction desorption-gas chromatography- mass spectrometry (TED GC-MS)	PE, PP, PS, PET, PVC	µg kg−1	No pretreatment other than grinding and mixing, processing time requires 2–3 h	Destructive; number, size, and morphology of the plastic particles cannot be obtained	[388]
TGA with mass spectrometry (MS)	PET, PVC, PE, PS	mg kg−1	Simple, low cost and no pretreatment	Information on shapes, sizes and morphologies was missing	[389]
Short-wave infrared (SWIR) imaging spectroscopy	PE, PP, PS, PET	Particles kg−1	Minimal analysis time and scanning of large area	large lower size limit (only large particles)	[390]
Proton Nuclear Magnetic Resonance Spectroscopy (1H NMR)	PE, PS, PET	mg kg−1	cost-efficient and fast (approximately 1 min per measurement)	100% removal of organic matter from soil required	[391]
Near Infra-red spectroscopy (NIR)	LDPE, PET, PVC	Particles kg−1	Time and labor saving, as no extractions are required	low accuracy and high detection limit	[355]
Differential Scanning Calorimetry (DSC)	PE, PP, PS	mg kg−1	rapid, simple, and low-cost technology	Information on particle size, shape, and number is missing	[392,393]
Gel Permeation Chromatography (GPC)	PE, PP, PS	mg kg−1	Provides information on molecular weight, distribution, and environmental aging of polymer	destructive, provides no particle-level information (count, shape, size)	[300]
Ultraviolet–Visible spectrometry (UV–Vis)	PS	%	Direct measurement without filtration	Cannot measure floated particles	[394]
Accelerated Solvent Extraction (ASE)	PA, PS, PE, PET, PVC	µg g−1	Analysis of total mass content	no particle-level information	[395]
Pressurized Liquid Extraction (PLE)	HDPE, PS, PVC, PET, PP	µg g−1	Simple, cost-effective, and rapid analysis, with uniform concentration reporting	unable to measure total mass (only pure mass of polymers) and size fractions of MPs	[396]

PE: Polyethylene; PP: Polypropylene; PS: Polystyrene; PA: Polyamide; PET: Polyethylene terephthalate; PVC: Polyvinyl chloride; HDPE: High-density polyethylene; LDPE: Low-density polyethylene.

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7.1.3. Identification and Quantification Methods of NPs in Soil

Contamination of NPs in ecosystems has raised global concerns recently [397]. However, the actual concentrations of these particles are unknown due to the limitations of actual detection methods. Much of the earlier research investigated the environmental effects of NPs using laboratory-scale experiments; however, there is still a lack of standard methodological detection methods for NPs in environmental samples [398]. Detection technologies for NPs in environmental samples require further advancement to enable robust assessment of their ecological impacts. A comparison of different NP detection methods is presented in

Table 12. Comparison of various detection methods of NPs.

Detection Method	Size Range	Advantages	Limitations	References
Dynamic light scattering (DLS)	1 nm–3 mm	Fast, cheap, non-invasive	Large particles, polydispersity affects results	[399]
Optical-photothermal infrared micro spectroscopy (OPTIR)	0.6–332 µm	Provides sub-micron chemical imaging	NPs <100 nm may still be below the resolution limit	[400]
Multiangle laser scattering (MALS)	10–1000 nm	Online coupling, large size ranges, easy, fast	Requires clean sample and only spherical models	[401]
Laser diffraction (LD)	10 nm–10 mm	Easy, fast, and automated	Only spherical model	[402,403]
Estrogen receptor-based nano-plasmonic sensor	100 nm	Ultra-sensitive detection	Require specific NP surfaces, costly	[404]
Nanoparticle tracking analysis (NTA)	10 nm–2 μm	Size and number concentration	Complex in operation	[405]
Transmission Electron Microscopy (TEM)	0.1 nm–100 μm	Chemical information and images	Not effective for identifying MPs	[406]
Atomic force microscopy (AFM)	Up to 0.1 nm	Surface morphology, mechanics	Expensive, limited scan area (100 μm)	[407]
CPE-ICP/MS	269 nm	High sensitivity	Metal-dependent, extraction needed	[408]
Protein Corona-Mediated Extraction–Py-GC/MS	50 nm–100 nm	Reduces matrix interference	Complex preparation, costly	[263]
AFM-IR	Smaller than 20 nm	Expensive	Expensive particle morphology, chemical identification	[409]
Thermo desorption–pyrolysis GC/MS (TD-PTR-MS)	Smaller than 20 μm	High sensitivity and rapid analysis due to proton-transfer ionization	Matrix interferences from organic matter, high instrumental cost	[410]
AFM-Raman	Single molecule detection	Particle morphology and chemical fingerprint	Expensive, under development for rapid nanoscale chemical analysis	[411]
Confocal Raman Microscopy (CRM)	Smaller than 100 nm	Point-by-point and line-by-line scanning, analysis of chemical components in NP	Presence of co-contaminants, diffraction spot size limitation	[412]

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7.1.4. Quantification Methods of NPs

The main bottleneck for NP quantification is the lack of validated analytical methods and standards. Currently, research focuses on the characterization of NPs, while information on quantification is limited. Although the integration of optical techniques, such as microscopy, and spectroscopic methods, like Fourier-transform infrared spectroscopy (FTIR), has significantly advanced MP analysis by providing resolutions down to 20 μm [413], these techniques remain inadequate for detecting NPs. Researchers have turned to more advanced methodologies for the analysis of NPs. A comparison of the different NP quantification methods is presented in

Table 13. Comparison of different quantification methods of NPs.

Quantification Method	Sample Size/Unit	Advantage	Limitations	References
LIBS	ng dm−1	Multi-element microanalysis; little or no sample preparation	Unable to discriminate between different sizes of NPs	[414]
GF-AAS	ng L−1	Direct NP injection; little or no sample preparation	Only applicable for NPs containing detectable metals	[415,416]
TRPS	Particles mL−1	Very high resolution, low cost per sample	Aggregation of particles with high-salinity buffer	[417]
CF3-MALS	Particles mL−1	Indirect information on morphology, moderate cost per sample	Possible low particle recovery during fractionation
Pyrolysis-Gas-Chromatography-Mass-Spectrometry (Py- GC–MS)	µg L−1	No sample pre-treatment	Destructive technique, sample cannot be reused	[418]
Thermal Desorption-Gas-Chromatography-Mass-Spectro Metry (TDS-GC–MS)	up to 100 mg	Minimal contamination from solvent impurities	Destructive techniques, qualitative analysis is very challenging	[406]
Single particle inductively coupled with plasma mass spectroscopy (Sp-ICP-MS)	800 nm–5 μm particles mL−1 concentrations (mg L−1)	Provides data on size, elemental composition, density, concentration	Multiple steps of sample pretreatment	[419,420]
Inductively coupled plasma optical emission spectroscopy (ICP-OES)	µg L−1	Simultaneous NP analysis, relatively low detection limits	No information on morphology	[421]
Centrifugal Liquid Sedimentation (CLS)	µg mL−1	Very high resolution, indirect information on density	Calibration is needed before each measurement, calibrant size and density affect results	[417]

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7.1.5. Identification and Quantification Methods of Pesticides

The intensive application of pesticides in modern agriculture has enhanced crop yield and food security but has introduced serious health concerns due to the accumulation of pesticide residues in fruits, vegetables and the environment [422]. Pesticides can be classified as insecticides, herbicides, and fungicides on a broader level based on the target pest. These are categorized by their chemical structure, mode of action, risk and use [423,424]. Most pesticide residue analyses rely on common sample preparation techniques, such as extraction, cleanup, and instrumental measurements, regardless of the pesticide type. Therefore, analytical methods are generally categorized based on the technique rather than the pesticide class. The analytical methods for analyzing pesticide residues in different fruit samples are listed in

Table 14. Analytical methods for analyzing pesticide residues in various food samples.

Instruments	Sample	Analytes	LOD	LOQ	References
GC–MS	Orange, lemon (juices)	8 multiclass pesticide residues	0.000018–0.000096 mg mL−1	0.00006–0.00032 mg mL−1	[425]
	Orange (juice)	19 multiclass pesticide residues	5.0 × 10−7–1.0 × 10−6 μg g−1	–	[426]
	Fruits + vegetables	439 multiclass pesticide residues	–	0.001–0.0015 mg g−1	[427]
LC–MS/MS	Tea	Triazoles	4 × 10−6–3.16 × 10−5 μg g−1	–	[428]
GC–MS/MS	Orange (pulp)	360 multiclass pesticide residues	–	0.001–0.05 mg g−1	[429]
GC–ECD	Citrus	Fluazinam	0.003 mg mL−1	0.01 mg g−1	[430]
GC–ECD	Orange (juice)	Chlorpyrifos, hexaconazole	0.00067–0.00089 mg mL−1	0.00222–0.00294 mg mL−1	[431]
GC–MS, FPD	Pomegranate, orange (juice)	7 multiclass pesticide residues	0.0008–0.00116 mg mL−1	0.0000028–0.000004 mg mL−1	[432]
GC–FPD	Pomegranate, orange (juice)	9 multiclass pesticide residues	0.00032–0.00076 mg g−1	0.0011–0.0026 mg g−1	[422]
	Orange (juice)	7 multiclass pesticide residues	0.00030–0.00061 mg g−1	0.0010–0.0020 mg g−1	[432]
GC–FID	Pomegranate, orange (juice)	6 multiclass pesticide residues	0.00030–0.00061 mg g−1	0.0010–0.0020 mg g−1	[432]
	Orange (juice)	Fenitrothion, malathion, ethion, chlorpyrifos, diazinon	0.00003–0.00011 mg mL−1	0.00011–0.00038 mg mL−1	[433]
HPLC–UV	Orange (whole fruit) Orange (whole fruit)	Multiclass pesticide	0.0097–0.010 mg g−1	0.039–0.32 mg g−1	[434]
		115 multiclass pesticide residues	0.000001–0.000007 mg g−1	0.000003–0.000019 mg g−1	[435]
	Citrus, lemon (whole fruits)	Thiabendazole	0.004–0.009 mg g−1	-	[436]
	Kumquat (whole fruit)	Bifenthrin	0.003 mg g−1	0.01 mg g−1	[437]
HPLC–MS/MS	Orange (whole fruit)	115 multiclass pesticide residues	0.000001–0.000007 mg g−1	0.000003–0.000019 mg g−1	[435]
	Orange (whole fruit)	200 multiclass pesticide residues	–	0.000010–0.000100 mg g−1	[438]
	Orange (whole fruit)	Etoxazole	–	0.005 mg g−1	[439]
	Citrus (whole fruits)	Spirodiclofen–pyridaben	0.001 mg g−1	0.005 mg g−1	[440]
	Lemon (essential oil)	256 multiclass pesticide residues	–	≤0.010 mg mL−1	[441]
	Orange (whole fruit)	8 multiclass pesticide residues	0.00002–0.00032 mg g−1	0.00007–0.00106 mg g−1	[442]
	Citrus (peel, albedo, pulp)	165 multiclass pesticide residues	-	0.71–5.97 mg g−1	[443]
	Orange (peeled fruit)	Carbendazim	0.03 mg mL−1	0.10 mg mL−1	[444]
	Orange (whole fruit)	20 organophosphorus pesticides (OPPs)	0.000002–0.000063 mg g−1	0.001 mg g−1	[303]
	Orange (juice)	10 multiclass pesticides	0.000001–0.0001 mg g−1	0.000003–0.0003 mg g−1	[445]
	Orange (pulp)	74 multiclass pesticide residues	0.0003–0.067 mg g−1	0.001–0.222 mg g−1	[446]
	Apple, grapes, peach, kiwi, orange (whole fruit)	5 multiclass pesticide residues	0.000003–0.00018 mg g−1	0.00001–0.00059 mg g−1	[447]
	Orange (whole fruit)	5 fungicides	0.00067–0.00125 mg g−1	0.00224–0.00415 mg g−1	[448]
	Orange (whole fruit)	Chlorpyrifos, triazophos	0.0001–0.0003 mg g−1	0.20–0.51 mg g−1	[449]
	Orange (juice)	7 fungicides	0.00052–0.00183 mg mL−1	–	[419]
	Citrus (whole fruits)	287 multiclass pesticide residues	–	0.001–0.01 mg g−1	[450]
	Citrus (whole fruits)	Forchlorfenuron, paclobutrazol, uniconazole	0.00009–0.00017 mg g−1	0.00029–0.00056 mg g−1	[451]
	Citrus (whole fruits)	Albendazole	0.000001–0.01354 mg mL−1	0.000003–0.04513 mg mL−1	[283]
	Citrus (whole fruits)	Abamectin, spinosad, imidacloprid, difenoconazole	–	0.01–0.05 mg g−1	[452]
HPLC–DAD	Orange (juice)	Carbendazim	8.7–15 mg mL−1	102–110 mg mL−1	[453]
	Citrus (whole fruits)	Thiacloprid	0.03 mg mL−1	0.05 mg mL−1	[454]
	Orange (juice)	Methamidophos, parathion, phoxim	0.00006–0.00013 mg mL−1	0.00021–0.00044 mg mL−1	[455]
	Lemon (juice)	5 benzimidazoles	0.0025–0.0029 mg mL−1	0.0088–0.0097 mg mL−1	[456]
	Food samples (solid/liquid)	Thiabendazole	0.009–0.017 mg mL−1	0.028–0.052 mg mL−1	[457]
HPLC–FD	Tangerine, grapefruit (whole fruits)	Chlorpyrifos, carbendazim	0.6–0.7 mg g−1	0.19–0.22 mg g−1	[458]
	Orange (pulp)	Thiabendazole, carbendazim, fuberidazole	0.00003–0.00968 mg mL−1	0.00012–0.03236 mg mL−1	[459]
HPLC–UV	Lemon (whole fruit)	Carbendazim, thiabendazole	0.00045–0.00054 mg mL−1	0.00150–0.00180 mg mL−1	[460]
UHPLC–MS/MS	Orange (whole fruit)	5 multiclass pesticide residues	-	0.0001–0.0015 mg g−1	[461]
GC, HPLC–MS	Orange (whole fruit)	93 multiclass pesticide residues	-	<0.005 mg g−1	[462]
GC, HPLC–MS/MS	Lemon (whole fruit, juice, essential oil)	16 multiclass pesticide residues	–	0.01–0.10 mg mL−1	[452]
	Lemon (fruit and juice)	355 multiclass pesticide residues	–	0.01 mg g−1	[463]
HPLC-Chiral Detection	Orange (juice)	Imazalil	0.54–0.94 mg mL−1	1.80–3.18 mg mL−1	[464]

LOD: limit of detection; LOQ: limit of quantification; GC–MS: Gas Chromatography–Mass Spectrometry; LC–MS/MS: Liquid Chromatography–Tandem Mass Spectrometry; GC–MS/MS: Gas Chromatography–Tandem Mass Spectrometry; GC–ECD: Gas Chromatography–Electron Capture Detection; GC–MS, FPD: Gas Chromatography–Mass Spectrometry with Flame Photometric Detection; GC–FPD: Gas Chromatography–Flame Photometric Detection; GC–FID: Gas Chromatography–Flame Ionization Detection; HPLC–UV: High-Performance Liquid Chromatography with Ultraviolet Detection; HPLC–MS/MS: High-Performance Liquid Chromatography–Tandem Mass Spectrometry; HPLC–DAD: High-Performance Liquid Chromatography with Diode Array Detection; HPLC–FD: High-Performance Liquid Chromatography with Fluorescence Detection; UHPLC–MS/MS: Ultra-High-Performance Liquid Chromatography–Tandem Mass Spectrometry; GC, HPLC–MS: Gas Chromatography and High-Performance Liquid Chromatography–Mass Spectrometry; GC, HPLC–MS/MS: Gas Chromatography and High-Performance Liquid Chromatography–Tandem Mass Spectrometry; HPLC–Chiral detection: High-Performance Liquid Chromatography with Chiral Detection.

.

In addition to the chromatographic and mass spectrometric techniques summarized in Table 14, different non-separative detection techniques have been developed for pesticide residue analysis. These methods do not require initial chromatographic separation and are mainly designed for rapid, non-destructive, and on-site screening. These include spectroscopy-based techniques, electroanalytical sensors, fluorescence-based systems, aptasensors, biosensors, lateral flow assays, and emerging AI-assisted detection platforms. These methods offer shorter sample preparation times, lower operational costs, portability, and minimal solvent consumption. However, they may face challenges related to mixed pesticide discrimination, matrix interference and quantitative accuracy [422].

Recent Advances in Rapid Detection Techniques for Pesticide Residue

In recent years, significant progress has been made in the development of rapid pesticide detection technologies, and various new methods have emerged. Rapid detection technologies offer significant advantages over conventional methods. Most importantly, they are less time-consuming and less costly. Some innovative and novel rapid detection technologies for pesticide residues in different matrices are discussed below in

Table 15. Novel rapid detection methods of pesticide residues in different matrices.

Techniques	Samples	Mechanisms	Type of Pesticides	References
Pesticide enzymatic and immunoassay test kits	Water, vegetables, fruits	Qualitative colorimetric acetylcholinesterase (AChE) inhibition-based tests	Organophosphate and carbamate	[465]
ELISA kits (immunoassays)	Water, soil, vegetables	Commercial immunoassay kits introduced in 1988 for pesticide analysis. Suitable for detection of metals, pesticides, organic chemicals	Atrazine; polychlorinated biphenyls; polyaromatic hydrocarbons	[466,467]
Nanoparticle-based electrochemical, optical, and magnetic environmental sensors	Apple fruit samples, water, air, dirt	Antigen-coated filter paper with single-walled carbon nanotubes (SWNTs). Portable, sensitive, inexpensive, and ~28× faster than ELISA	Phenoxy organophosphates, carbamates, pyrethroids, atrazine, neonicotinoids, organochlorines	[468,469,470]
Mass spectrometry (MS)	Fruits, vegetables	Ion trap MS using PDMS membranes for analyte diffusion. Wastewater sampling every 12 min	Carbaryl, imidacloprid, deltamethrin, cypermethrin, malathion, acetamiprid, monocrotophos, chlorpyrifos-methyl, diazinon	[471,472]
Azo-coupling reaction-based method	Fruits, vegetables	Color change from yellow to orange	Carbaryl	[429,473]
Microfluidic arrays sensor	Fruits, vegetables	Paraoxon pre-inhibits AChE followed by incubation; color change from red to blue indicates detection	Organophosphate	[474,475]
Gold nanoparticle-based colorimetric aptasensor	Food, water	AuNP colorimetric assay for rapid detection; solution cooler changes to blue	Organophosphorus, carbamate	[476,477]
AChE inhibitor detection using AuNPs with dynamic light scattering	Food, water	Color changes from red to colorless or bright pink	Organophosphorus	[478,479]
Citrate-capped AuNP method	Food, water samples	Citrate-capped AuNPs produce color change from wine-red to purple-blue	Dithiocarbamate	[478,480]

ELISA: Enzyme-Linked Immunosorbent Assay; AuNPs: Gold Nanoparticles; AChE: Acetylcholinesterase; PDMS: Polydimethylsiloxane; SWNTs: Single-Walled Carbon Nanotubes.

.

7.2. New Emerging Technologies

Recent progress in machine learning (ML) and automated high-throughput systems provides promising solutions to overcome persistent analytical challenges in MP-NP detection. The integration of machine learning with FTIR and Raman spectroscopy enables accurate spectral pattern recognition.

Recently, artificial intelligence has emerged as a powerful tool for the automated analysis of MP-NP detection. This approach can enhance various analytical methods, including spectroscopic methods (Raman, FTIR) or image-based methods (microscopy, fluorescence, SEM). AI has already been employed to support the detection of MPs via FTIR [481], Raman [451], SEM [259], and fluorescence microscopy [225]. However, the application of AI to NP detection remains challenging. AI-enhanced chemical imaging workflows have emerged as powerful tools for reliable NP mapping and quantification through algorithmic image processing and Gaussian surface fitting [482].

Although these methods remain relatively unexplored in soil matrices compared to aquatic systems, their continued advancement is expected to play a vital role in future environmental monitoring.

7.3. Methodological Gaps/Limitations in Detection and Quantification

The widespread application of agrochemicals and plastic-based agricultural materials has enhanced the accumulation of pesticide residues and MPs-NPs in agricultural soils. While pesticide detection and quantification methods are well established and highly sensitive, analytical approaches for MPs-NPs in soil remain underdeveloped and non-standardized. Despite significant progress in environmental monitoring, major methodological limitations remain in the analytical methods for MPs, NPs, and pesticides, particularly in complex agricultural plant-soil systems.

The analytical methods used for MP-NP detection are largely similar in both soil and plant studies; however, plant samples require customized extraction protocols due to their complex biological matrix [21].

Several sequential steps, including sampling, pretreatment, and extraction, are required to ensure the accurate detection and quantification of MPs or NPs, and the reliability of analytical methods strongly depends on these preparatory stages. Although pesticide residues can be analyzed using well-established chromatographic and mass spectrometric methods, there is currently no universally accepted protocol for the sampling, extraction, identification, and quantification of MPs, and especially NPs, in soils [363,483,484,485,486]. The heterogeneous nature of agricultural soils (minerals, organic matter, and colloids) further complicates plastic isolation and identification [360,361]. Therefore, future efforts should prioritize the development of matrix-specific standardized protocols, along with certified reference materials and calibration standards, specifically for nanoscale plastics. In addition, harmonized reporting metrics (e.g., particle number, size distribution, and mass concentration) are essential to improve comparability across studies. Moreover, current methodologies predominantly investigate soil and plant systems separately. However, in real agricultural settings, soil and plants function as coupled systems. These limit the understanding of combined transport, bioavailability and transformation in soil-plant systems. Future studies should therefore incorporate system-level approaches to better understand interaction mechanisms, including sorption–desorption behavior, transport processes, and the influence of MPs/NPs on pesticide persistence and bioavailability in soil-plant systems.

Present analytical methods typically assess MPs, NPs and pesticides separately, despite their known interactions (e.g., plastics act as sorbents of agrochemicals) [292]. MPs often coexist with fertilizer, pesticides, heavy metals and other pollutants in agricultural soil [160]. Further research is required to understand the impact of MPs on pollutant persistence, bioavailability, and the associated ecological risks. A key challenge lies in the co-extraction and simultaneous analysis of plastics and pesticides, as current protocols are typically designed for single contaminant groups. To address this limitation, future research should focus on developing harmonized co-extraction strategies, including careful selection of solvent systems and density separation techniques that are compatible with both MPs/NPs and pesticide residues while preserving analytical recovery.

Traditional visualization tools (optical, fluorescence and electron microscopy) are beneficial for morphological and structural observations but lack chemical characterization. Scanning and transmission electron microscopy (TEM) enables high-resolution imaging of MP location and morphology at both the tissue and cellular levels [285]. However, these methods require cross-sectional sample preparation, which increases the analysis time.

Thermal techniques such as Py-GC–MS, TED-GC–MS, and TGA–MS provide total mass quantification but are destructive in nature and cannot provide morphological information on particle size, shape or abundance [487]. Conversely, imaging and spectroscopic methods provide particle-level data but do not directly yield mass concentrations. Nuclear magnetic resonance (NMR) spectroscopy provides additional insights into polymer structure, but its application is limited due to its high technical demands [391]. Other spectroscopic methods such as µ-FTIR and µ-Raman spectroscopy are widely applied for polymer identification and semi-quantification; however, these methods are limited by their respective size detection thresholds (~10 μm and ~1 μm) [488].

Rapid field-deployable sensors are required for the quantitative analysis of MP or NP materials. MPs-NPs are mainly detected using laboratory-based analytical instruments. However, these systems are neither portable nor suitable for on-site applications. Many portable devices have been developed for water samples; however, they also rely on principles similar to those of laboratory instruments and therefore suffer from several drawbacks [489].

To address these methodological gaps, future research should emphasize integrating multi-technique workflows that combine imaging, thermal analysis, and spectroscopy for simultaneous particle and mass-based quantification. Efforts should also aim to establish analytical frameworks capable of assessing the interactions and co-occurrence of plastics and pesticides in agroecosystems.

8. Conclusions

The existence of MPs-NPs and agrochemicals in soil heightens the concern that persistent pollutants in agriculture threaten plant, soil, animal, and aquatic health and significantly reshape agroecosystems. The accumulation of MPs-NPs along with persistent pollutants in soil is due to their slow biodegradability. The presence of MPs-NPs and agrochemicals in soil has been shown to impede soil structure, reduce bulk density, alter microbial communities, and influence nutrient cycling in soil. Moreover, their ability to adsorb and vector contaminants introduces additional concerns regarding food safety and human health. However, mechanistic understanding remains constrained by limited long-term field data and the absence of standardized, high-resolution detection methods for NPs in complex soil matrices. Addressing these challenges will be critical for developing sustainable agricultural management strategies and informing regulatory policies aimed at mitigating the environmental and economic risks associated with agrochemicals and MPs-NPs coupling in intensively managed agroecosystems.