Uncontrolled Disposal of Plastic Agrochemical Packaging: Double-Trouble Ecotoxicological Effects in Soil and Water

Abstract

The widespread use of agrochemicals, including inorganic and organic pesticides and fungicides, has contributed to the persistence of hazardous residues in agricultural environments, particularly through their accumulation in plastic packaging and containers. High-density polyethylene (HDPE), polypropylene (PP), and other polymer types commonly employed for agrochemical storage and transport retain significant quantities of active substances even after standard rinsing procedures. This phenomenon raises concerns over improper disposal practices, environmental contamination, and potential ecotoxicological impacts. Recent studies demonstrate that both inorganic and organic pesticide residues exhibit strong interactions with plastic polymers, influenced by factors such as polymer chemistry, surface aging, pH, ionic strength, and dissolved organic matter. These interactions not only delay degradation but also facilitate secondary release into soils and aquatic systems, where they may impair soil microorganisms, alter plant physiology, and disrupt aquatic food webs, including phytoplankton, fish, and microbial assemblages. Despite regulatory frameworks and container management schemes in some regions, major knowledge gaps remain regarding the long-term fate of pesticide residues on plastics, their transfer to ecosystems, and cumulative effects on agroecosystem sustainability. This review synthesizes current evidence on the chemical characteristics of pesticide residues in plastic packaging, their environmental mobility, and ecotoxicological effects. It further identifies urgent research needs, including long-term field assessments of polymer–pesticide interactions, improved recycling technologies, and the development of safer container designs. Effective management strategies, coupled with strengthened international stewardship programs, are essential to reduce risks to environmental health, agricultural productivity, and human safety.

1. Introduction

The continuous rise in global plastic production over recent decades has emerged as one of the most pressing environmental challenges of our time. Current estimates indicate that global plastic production exceeds 400 million tons annually, and projections by the OECD suggest that plastic use could nearly triple by 2060 if no additional policies are implemented [1].

In agriculture, plastics are extensively utilized in various applications, including pesticide containers, mulching films, greenhouse covers, and irrigation systems [2,3]. Although these materials offer significant advantages such as flexibility, durability, and ease of use, the majority are not properly managed at the end of their life cycle. Inadequate disposal practices and accidents contribute to the persistent accumulation of plastics in terrestrial and aquatic ecosystems, causing environmental adverse effects over time.

Microplastics (MPs), defined as plastic particles smaller than 5 mm, are increasingly recognized as pollutants of emerging concern. They can be intentionally produced at microscopic scales (primary MPs) for industrial and commercial uses or generated from the degradation and fragmentation of larger plastics through ultraviolet radiation, mechanical abrasion, and weathering processes (secondary MPs) [4]. Due to their extremely low biodegradability, MPs accumulate in soil and water bodies [5]. In addition to their physical presence, MPs are capable of adsorbing and transporting hazardous substances, including pesticides and heavy metals, thereby intensifying their environmental persistence and potential eco-toxicity [6]. Although microplastic research initially focused on marine ecosystems, recent studies have highlighted the widespread occurrence and impact of MPs in terrestrial environments, particularly agricultural soils [7]. The breakdown of agricultural plastic films and agrochemical packaging has been identified as a significant source of MPs in these systems. Once released, MPs can migrate vertically through the soil profile, with the potential to reach groundwater resources [8]. This movement is influenced by multiple factors, including particle size and shape, polymer type, and biological processes such as root growth and soil organism activity (Figure 1) [9]. However, the behavior of MPs as vectors for chemical pollutants remains a subject of debate. While some studies report a strong affinity between MPs and pesticide molecules, others suggest that contaminants may desorb quickly upon entering the soil environment [10]. Another means of transport could be through waterways like rivers, streams and pipelines in that case MPs enter various marine environments [11].

Figure 1. Main sources and transport pathways of microplastics (colorful beads) in soil, posing potential threats to groundwater.

Global assessments confirm that plastics are embedded across agricultural value chains, with approximately 12.5 Mt of plastic products used annually in plant and animal production and 37.3 Mt in food packaging [12]. Within this total, agrochemical packaging constitutes a significant stream that grows in parallel with crop protection intensity and input use. National and regional stewardship reports demonstrate a consistent long-term increase in container recovery, suggesting that packaging flows entering collection systems are steadily rising [12].

At the regional scale, Latin America provides some of the most robust data. In Brazil, the Campo Limpo program coordinated by inpEV has processed more than 700,000 t of agrochemical packaging since 2002, with ~52,500 t collected in 2022 alone, and reports coverage exceeding 90% of primary containers placed on the market [13]. These volumes make Brazil the global leader in pesticide container recovery. In North America, the Ag Container Recycling Council (ACRC) reports cumulative recycling of >257 million lb of HDPE since inception, with ~11 million lb processed annually [14]. Similarly, Cleanfarms Canada [15] documents cumulative returns of 58.5 million small containers and ~578,000 large drums/totes by the end of 2024.

In Europe and Africa, coordinated programs show uneven but increasing success. CropLife International reports that South Africa achieved ~80% collection coverage of agrochemical containers in 2021, while France (Adivalor), Spain (Sigfito), and Germany (Pamira) operate mature extended producer responsibility (EPR) schemes [16,17]. Across these regions, coverage remains variable, but the trend is clearly toward harmonized stewardship frameworks.

Global trends in agrochemical packaging can therefore be summarized as follows: first, container flows increase proportionally with agrochemical sales; second, recovery schemes are expanding through EPR-driven mandates and industry-led take-back networks; third, recycling markets for non-food-grade HDPE have stabilized, with outputs going into products such as drainage pipe or fence posts [2] and fourth, while reporting transparency has improved, methodological inconsistencies (e.g., unit counts vs. mass-based estimates) complicate direct international comparisons [12,16].

With respect to materials, rigid containers for agrochemicals are overwhelmingly composed of high-density polyethylene (HDPE), chosen for its chemical resistance and cost-effectiveness. Polypropylene (PP) is widely used for closures and fitments, while fluorinated HDPE provides added resistance against aggressive formulations in some markets [14]. Flexible agrochemical packaging typically involves polyamide (PA) sachets or multilayer films, which offer enhanced stability but pose greater challenges for post-consumer recycling [2]. These material choices directly affect rinsability, recyclability, and the potential for residual agrochemical retention, making them central to both environmental risk and management discussions.

Chemical studies often investigate the sorption and desorption behavior of pollutants on MP surfaces [18], while soil-focused research addresses changes in soil structure, porosity, and microbial dynamics induced by the presence of MPs [19]. Concurrently, eco-toxicological studies examine the synergistic and cumulative impacts of MPs and associated pollutants on both terrestrial and aquatic organisms [18]. Despite the increasing volume of research, findings are often fragmented and discipline-specific, complicating efforts to develop an integrated understanding of MP dynamics and risks. The management of agricultural plastics remains insufficient to address the escalating environmental implications. Many of these materials have short functional lifespans, and current collection and management systems are often ineffective or poorly implemented, leading to the continued accumulation of plastic waste in the environment [20]. Addressing this challenge requires the development and adoption of improved frameworks, the promotion of biodegradable alternatives, and enhanced awareness and stewardship among agricultural stakeholders [21]. Research on MPs is highly interdisciplinary, bridging fields such as environmental engineering, soil science, analytical chemistry, and ecotoxicology [22].

The aim of this review is to consolidate and critically assess the current knowledge related to microplastics derived from agrochemical packaging, with a specific focus on pesticide adsorption and desorption mechanisms and their ecotoxicological implications for soil and water systems. Additionally, this work seeks to bridge existing disciplinary gaps by offering a more integrated and multidisciplinary perspective and by identifying key areas for future research that are essential for advancing sustainable agricultural practices. An artificial-intelligence tool, ChatGPT (OpenAI), was employed to generate preliminary graphic illustrations for Figures 1 and 3–6. All outputs were manually checked, refined, and validated by the authors. ChatGPT did not contribute to data interpretation, experimental design, or writing of the scientific conclusions. AI-assisted figure generation was used to clearly depict the complex processes discussed in this review, including microplastic formation from agrochemical packaging and their roles in pesticide adsorption–desorption and ecotoxicological impacts on soil and water systems. These visualizations were necessary to synthesize multidisciplinary information and support the integrated perspective offered in this work.

2. Sources of MPs in Agricultural Ecosystems

The increasing use of plastics in agriculture, estimated by the FAO at 12.5 million tons annually globally, is crucial for enhancing crop productivity and resource conservation, with nearly 60% attributed to film-based products (e.g., greenhouse covers, mulching) that create controlled microclimates and reduce water loss [2,12]. Agricultural plastics are classified into five primary groups (protective films, nets, packaging, irrigation components, and other products) and typically comprise polymers such as polyethylene variants (LDPE, LLDPE, HDPE), polypropylene (PP), and PVC [23]. While offering agronomic benefits like yield enhancement and protection from extreme weather [12,23], these materials, particularly polyethylenes which exhibit higher resistance to abiotic degradation than PET or PS [24], pose significant environmental risks, including difficult residue removal, slow degradation, leaching of hazardous additives (e.g., phthalates), and the accelerated formation of microplastics from thinner films [23,25,26]. To address these challenges, the “6R” framework (Refuse, Redesign, Reduce, Reuse, Recycle, and Recover) is advocated for strategic environmental management [12]. In the EU-28+2, the agricultural sector used approximately 1.74 million tonnes in 2018 (3–4% of total European plastic converter demand), with PP and PE being predominant, and 0.71 million tonnes specifically dedicated to non-packaging agricultural purposes [27,28] (Figure 2).

Figure 2. Annual plastics use in Europe in (a) livestock production and (b) crop production, readapted from APE Europe.

2.1. Plastics Used in Agriculture

Agricultural plastic films (mulching sheets, greenhouse films, and low tunnel covers) represent the dominant category, with mulching alone accounting for approximately 50% of the total mass [25]. Polyethylene (LDPE and LLDPE) is the primary material, valued for its flexibility and low cost [2,12], and these films provide significant agronomic benefits, including accelerating crop production and controlling weeds [23]. Critically, the durability of these films influences environmental contamination; Steinmetz et al. [26] showed that thinner perforated foils (40 μm) release substantial microplastics (PE up to 35 mg/kg of soil), while thicker films (50 μm) show significantly lower emissions, underscoring the need for more robust materials [26]. In contrast, nets (e.g., anti-hail, shading) form a distinct category, predominantly manufactured from HDPE with UV stabilizers (or PP for nonwoven covers), which are essential for protecting crops from environmental stressors like hail and intense solar radiation, thereby helping to reduce pesticide use [29]. Irrigation systems (pipes and drip tapes) are typically made from LDPE or HDPE and help reduce water losses through controlled delivery [2]. In livestock farming, silage films are mainly produced from LDPE/LLDPE with multilayer structures and additives to enhance durability [12]. Bale wraps and silage nets are essential for maintaining the quality of animal feed.

This category includes fertilizer sacks made of PP, pesticide containers made of HDPE or PET, as well as plastic crates and nursery trays made of PP or PS [12]. According to Tziourrou & Golia [25], packaging plastics and geotextiles contribute to soil pollution if they are not properly removed or recycled. Briassoulis [2] emphasizes the need for collection and recycling systems specifically designed for these materials.

2.2. Indirect Sources

The pervasive issue of microplastic (MP) contamination extends significantly into agricultural soils, with a substantial portion of this pollution introduced through indirect sources rather than just the fragmentation of plastic mulch. These pathways are primarily linked to the widespread use of organic soil amendments and external transport mechanisms. The application of sewage sludge (biosolids) is one of the most critical indirect routes. Wastewater treatment plants, while removing the bulk of plastics from water, concentrate microplastics—predominantly microfibers from synthetic clothing and fragments from personal care products—into the resulting solid sludge [30]. When this sludge is applied to farmland as a fertilizer, it can introduce vast quantities of MPs, with estimates suggesting annual inputs of hundreds of thousands of tons across continents [22,30]. Studies in sludge-amended soils have reported highly variable, yet significant, concentrations, often dominated by Polyethylene (PE) and Polypropylene (PP) polymers [22]. However, the complexity of contamination is highlighted by findings that even untreated soils can contain high MP levels, suggesting that biosolids are a major, but not the sole, contributor [31]. Another key source is the use of compost, particularly that derived from municipal solid waste (MSW) or biowaste. Inefficient sorting allows plastic packaging, bags, and fragments to be incorporated and fragmented during the composting process, leading to MP-contaminated compost that is then spread on fields [32]. Long-term compost application has been shown to increase soil MP concentrations over time [22]. Animal wastes and manure constitute a further pathway, as livestock often ingest plastic litter, feed packaging, and synthetic materials like silage nets, leading to the excretion of MPs onto agricultural land [33,34]. Beyond amendments, atmospheric deposition contributes a ubiquitous baseline of contamination, with microplastics from urban dust, tire wear, and textile fibers transported by wind over long distances before settling on farmland [35]. Finally, wastewater irrigation and surface runoff from adjacent contaminated areas or roads act as transport vectors, further introducing microplastics into the soil matrix [9]. Collectively, these indirect sources ensure that agricultural soils act as a massive, long-term sink for microplastics, raising concerns about potential impacts on soil health, crop quality, and the transfer of contaminants into the food chain [36,37].

2.3. Agricultural Plastics to Microplastics

Plastics commonly employed in agriculture—such as mulching films, greenhouse covers, and irrigation system components—undergo various degradation processes once introduced into the soil environment. These processes are multifactorial, encompassing physical, chemical, and biological stressors, and ultimately lead to the formation of secondary microplastics and nanoplastics. Such particles exhibit high persistence in the environment and have the potential to interact with soil physicochemical properties and co-occurring contaminants [38,39]. The rate and extent of degradation are strongly influenced by polymer type, environmental conditions, usage patterns, and the presence of additives, while complete mineralization is rarely achieved under typical agricultural conditions [40,41].

Mechanical stress is primarily induced by agricultural practices, including plowing and the use of machinery. These activities cause abrasion, cracking, and fragmentation of plastic materials into progressively smaller particles. Field surveys have documented polyethylene fragments up to 28 cm² in agricultural soils, indicating the onset of degradation via physical breakage [38,42]. Prolonged and repeated mechanical stress facilitates further size reduction, ultimately yielding microplastic particles.

Photodegradation, driven by exposure to sunlight—particularly ultraviolet (UV) radiation—initiates photo-oxidative reactions that weaken polymer chains. Sa’adu and Farsang [43] report that UV exposure results in discoloration, increased brittleness, and surface cracking. Lan et al. [44] further demonstrated that aged plastics exhibit increased surface roughness and the formation of polar functional groups which enhance their susceptibility to further degradation. This process predominantly occurs at the soil surface where sunlight penetration is greatest.

Chemical degradation arises from oxidative and hydrolytic processes, as well as from interactions with agrochemicals containing sulfur, chlorine, or iron. These compounds can accelerate plastic aging, promoting surface erosion and cracking, especially when acting synergistically with UV radiation or microbial activity [43]. The physicochemical alterations induced by such reactions—such as the introduction of polar functional groups—further facilitate polymer breakdown [44].

Biodegradation involves the activity of microorganisms, including bacteria and fungi, capable of modifying or partially degrading plastics. Masciarelli et al. [38] confirmed the role of microbial communities in the partial degradation of biodegradable polymers, while Wanner [40] noted that materials such as polylactic acid (PLA) and polyhydroxybutyrate (PHB) can decompose within months. In contrast, conventional polymers such as high-density polyethylene (HDPE) may persist in soils for centuries to millennia. Despite these limitations, biodegradation remains a crucial pathway for mitigating the accumulation of plastic residues in agricultural soils.

2.4. Vertical Transport of Microplastics in Agricultural Soils and Changes in Physicochemical Soil Properties

Vertical migration of microplastics (MPs) in soil is a complex, multifactorial process governed by the interplay of physical, chemical, and biological factors [45]. While earlier assumptions suggested that MPs predominantly accumulate in surface soil layers, recent experimental and field studies in agricultural contexts have demonstrated otherwise. MPs have been detected at depths exceeding 50 cm—and in some cases over 100 cm—raising concerns about their potential to reach groundwater reservoirs [46,47,48]. Long-term observations by Kato et al. [49], Heinze et al. [46], and Li et al. [47] under realistic irrigation and rainfall regimes confirm that MPs can migrate to substantial depths after years of agricultural application, indicating that vertical transport is not only possible but likely more common than previously assumed.

Intrinsic physical characteristics of MPs—such as particle size, shape, density, and sur-face roughness—play a pivotal role in determining their vertical mobility. Smaller particles tend to penetrate more deeply, moving more readily through soil pores [50,51]. However, migration potential does not always increase linearly with decreasing size; certain studies have identified particles in the 25–147 μm range as exhibiting the highest down-ward mobility [52]. Particle shape also exerts a significant influence. Spherical MPs, owing to their geometry and hydrophobicity, typically move downward more readily, whereas fibrous MPs are more prone to entrapment within soil pore networks [52,53]. Nonetheless, fibers can also migrate downward under specific conditions, particularly when they attach to plant roots or follow preferential flow pathways [46,54]. Density affects settling velocity but is not always predictive of final penetration depth, as hydrophobicity and other surface properties can exert greater control [51]. Higher surface roughness generally enhances adhesion to soil particles, thereby reducing mobility [55], while bio-film formation on MP surfaces can further modify transport dynamics [56].

Abiotic soil parameters—including pH, ionic strength, dissolved organic matter (DOM), and the presence of metal oxides—also exert strong control over MP transport. Soil pH in-fluences the surface charge of both MPs and soil particles, affecting aggregation behavior. At elevated pH, increased electrostatic repulsion reduces aggregation and favors the formation of smaller, more mobile particles [45,52]. Conversely, ionic strength modulates electrostatic interactions: high ionic strength compresses the electric double layer, enhancing attraction and sedimentation, thus reducing mobility, whereas low ionic strength promotes repulsion and facilitates vertical migration [51,57]. DOM, particularly humic and fulvic acids, can alter MP hydrophobicity and surface charge, enhancing hydrophilicity and dispersion. This can improve suspension stability and promote downward movement [58,59], although the magnitude of this effect varies with DOM composition and origin [43,47]. Metal oxides (e.g., of iron, manganese, aluminum) can increase soil surface roughness and promote MP adsorption via electrostatic and physicochemical interactions, restricting mobility [57,60]. In some agricultural soils, iron oxides in the rhizo-sphere can further immobilize MPs [46,61]. These abiotic factors rarely act in isolation; their combined influence ultimately determines whether MPs remain in surface layers or migrate to greater depths [49,53,62]. Hydrological dynamics—such as rainfall intensity, water flow, and moisture fluctuations—are critical drivers of vertical MP transport. Soil macropores, including worm burrows and shrinkage cracks, serve as conduits for rapid downward movement [46]. Heavy rainfall under neutral pH conditions has been shown to enhance vertical migration, particularly in organic matter–rich soils [57]. In sloped or erosion-prone landscapes, lateral displacement and redeposition may occur, leading to localized MP accumulation [53]. Moisture fluctuations, especially repeated wetting–drying cycles, have been identified as particularly influential. O’Connor et al. [50] reported that such cycles increased MP mobility in sandy soils, with ~21 μm polyethylene particles reaching 7.5 cm depth after multiple cycles. Similarly, Gao et al. [51] observed penetration depths up to 10.5 cm after 16 wet–dry cycles, with humic acid reducing particle hydrophobicity and enhancing transport. Freeze–thaw cycles, relevant in colder climates, can also facilitate MP migration. Koutnik et al. [48] demonstrated that consecutive winter freeze–thaw events in stormwater biofilter systems generated microcracks and released colloids, promoting deeper MP penetration compared to warmer wet–dry conditions. While most particles remained within the top 3–5 cm, the progressive disruption of soil structure over time may eventually allow MPs to reach deeper horizons and groundwater.

Overall, vertical migration of MPs in soils emerges from the combined effects of morpho-logical particle traits, abiotic soil parameters, and hydrological processes (Figure 3). Climate-driven fluctuations in soil moisture and temperature are likely to intensify these processes, especially in permeable soils and for smaller MPs. These findings underscore the need for long-term field studies to better understand MP distribution patterns under changing climatic conditions and to assess associated risks to soil and water resources.

Figure 3. Influencing factors on the vertical migration of microplastics in soil.

2.5. Agricultural Practices

Agricultural practices exert a substantial influence on the spatial distribution of microplastics (MPs) within soil profiles. Activities such as plowing, irrigation, and the application of organic amendments not only enhance MP accumulation in surface horizons but also facilitate their vertical redistribution (Figure 4). For example, a study conducted in the Shouguang region of China reported nearly equivalent MP concentrations in the 0–20 cm and 20–40 cm soil layers, a pattern largely attributed to tillage-induced mixing [63].

Figure 4. Agricultural practices and microplastics.

Irrigation—particularly in intensively managed greenhouse and polytunnel systems—further promotes vertical transport. In some cases, up to 19% of the total detected MPs occurred at depths of 40–60 cm, with the highest concentrations observed in greenhouse soils subjected to high-frequency irrigation regimes. Liu et al. [58] similarly reported that irrigation with river water not only enhanced downward migration but also altered the compositional profile of MPs, with fibrous particles exhibiting notably high mobility.

Agricultural management practices also affect the overall MP load in soils. In a long-term field experiment, the application of sewage sludge as an organic amendment resulted in an eightfold increase in MP mass and a twofold increase in particle number relative to plots treated solely with inorganic fertilizers [46]. Likewise, in agricultural fields subjected to 32 years of continuous plastic mulch film use, MPs were detected at depths reaching 100 cm—providing compelling evidence of long-distance infiltration under conditions of sustained plastic input [47].

Overall, agricultural activities influence not only the depth distribution of MPs but also their morphological characteristics and transport behavior. These practices frequently interact with environmental drivers, thereby amplifying the potential for microplastic migration into subsoil layers and, ultimately, deeper environmental compartments.

2.6. Biological Drivers of Microplastic Transport in Soil

Biological systems and functions—including plant roots, soil fauna, and microorganisms—can significantly influence the redistribution and fate of microplastics (MPs) in soils (Figure 5). Plant roots physically disturb the soil matrix and can either transport MPs deeper into the profile or immobilize them within the rhizosphere [64]. Evidence from several studies has revealed the presence of small MPs within plant tissues, suggesting possible uptake and translocation to aboveground organs [65]. Arbuscular mycorrhizal fungi (AMF) may further mediate MP movement—either facilitating the transport of smaller particles across soil pores or, conversely, acting as a barrier to larger MPs [61].

Figure 5. Biological contribution to microplastic transport in soils.

Soil invertebrates, particularly earthworms, contribute to MP redistribution by ingesting particles and excreting them at different soil depths (Figure 5). Their activity can also fragment certain biodegradable plastics, such as polylactic acid (PLA) and polybutylene adipate terephthalate (PBAT), thereby increasing particle dispersion [66]. Smaller invertebrates, including springtails and mites, can transport MPs through ingestion or by passive adherence to their exoskeletons [67,68]. Microorganisms can also influence MP mobility via adhesion, biofilm formation, and surface chemical modification. AMF hyphae, for instance, are capable of transporting small MPs through soil pore networks [61], while microbial secretions may alter MP surface chemistry, affecting aggregation and hydrophobicity [69]. However, the spatially localized nature of microbial activity means that their overall contribution to MP movement is generally more limited compared to larger soil fauna [69]. Collectively, biological processes can enhance MP redistribution in soils, often acting synergistically with agricultural practices and natural soil dynamics. Although their influence may be less pronounced than that of physical forces such as hydrological flow or tillage, these biotic interactions are a critical component of the overall transport and transformation pathways of MPs in agroecosystems. Vertical migration of MPs in soils is inherently multifactorial, influenced by particle morphology and chemistry, soil properties, climatic conditions, agricultural management, and biological activity. As highlighted by Heinze et al. [46], underestimating the potential for deep migration can lead to incomplete assessments of environmental exposure in both terrestrial and connected aquatic systems. Similarly, Kato et al. [49] emphasize that characterizing MP distribution across soil profiles is essential for accurately evaluating ecological risks and safeguarding crop productivity and human health.

Despite the increasing body of evidence demonstrating MP mobility under realistic agricultural conditions, there remains a pressing need for long-term field studies. Such investigations are vital for understanding the cumulative impacts of repeated plastic inputs, climate variability, and biological interactions, and for designing effective strategies to mitigate microplastic pollution in agricultural soils.

3. Agrochemical Plastic Packaging and Disposal

The term agrochemical generally refers to a broad range of pesticides, including insecticides, herbicides, and fungicides, but may also encompass synthetic fertilizers, hormones, other chemical growth agents, and concentrated stores of raw animal manure. Within this group, pesticides warrant particular attention due to their potential impacts on human health via contamination of groundwater, soils, food, and even air [70]. Pesticides not only target specific pest organisms but frequently persist as residues in soil, water, and food [71]. Although gaps in the available data hinder precise quantification of the scale and trends of pesticide use in Europe [70], evidence indicates that the issue remains both serious and growing [70]. Addressing the problem requires detailed data on pesticide usage, environmental conditions, and the toxicological, chemical, and physical properties of active ingredients, as well as effective management of the resulting empty containers [72].

According to Eurostat [73] pesticide sales in the European Union reached approximately 292,000 tonnes. Fungicides and bactericides were the largest category, representing 39% of total sales, followed by herbicides, haulm destructors, and moss killers (36%) and insecticides and acaricides (17%). The remaining groups—molluscicides, plant growth regulators, and other plant protection products—accounted for less than 10%. France, Spain, Germany, and Italy recorded the highest sales volumes across most major groups and together represented 52% of the EU’s utilized agricultural area and 49% of arable land. Inorganic fungicides—mainly copper compounds, inorganic sulfur, and other inorganic substances, several permitted in organic farming—comprised 62.9% of all fungicides and bactericides sold.

Agrochemical use across many European nations is subject to stringent regulation, often necessitating government-issued permits for product purchase and application. Several European Union (EU) Member States have implemented electronic monitoring systems to track the entire pesticide supply and use chain. On-farm practices are typically governed by strict rules concerning storage, labeling, emergency response, and the safe handling, application, and disposal of these chemicals.

Despite these regulatory efforts and the compilation of official pesticide use statistics, formal European data on agrochemical plastic packaging waste (APPW) remain absent. A previous estimation by Briassoulis et al. [2] approximated the generation of plastic APPW in seven European countries (Spain, Italy, France, Greece, Cyprus, Finland, and the United Kingdom) at approximately 18.5 Mt. Given that empty containers frequently retain pesticide residues, European legislation classifies APPW as hazardous waste. This classification complicates processes such as collection, transport, and disposal, simultaneously increasing management costs and raising significant environmental and public health concerns due to inherent risks of cross-contamination and accidental exposure.

Inadequate management of APPW is a pervasive global environmental concern, contributing to soil, water, and air pollution, and compromising agricultural product safety and human health. For example, poor APPW management in Thailand has led to high concentrations of organophosphate pesticides in sediments [74] and the detection of toxic residues in containers, soil, and surrounding water bodies [75]. To address this, Patarasiriwong et al. [76] proposed a comprehensive APPW management model for Thailand that incorporates multiple rinsing, community-based collection centers, producer-led recovery programs, and governmental oversight.

Within Europe, several national APPW management schemes have been established, including Pamira in Germany [77], Adivalor in France [78], and Sigfito in Spain [79]. Adivalor and Sigfito differ mainly in the legal classification of waste: Adivalor treats properly rinsed containers as non-hazardous, promoting recycling, while Sigfito has historically classified them as hazardous, leading to stricter disposal, often via incineration. Other initiatives, such as AIPROM in Romania, established in 2007, focus solely on the collection of plant protection product packaging. Conversely, countries like Greece, Italy, and Cyprus continue to operate without formal, centralized APPW management schemes, leading to uncontrolled disposal and heightened environmental risks.

To mitigate these challenges, the European AgroChePack project was initiated to develop an integrated, environmentally and economically sustainable APPW management system. This project aimed to transfer successful practices from established schemes and foster synergistic collaboration with the LabelAgriWaste program for general APW management [80]. The integrated approach was successfully piloted in five Member States: Greece, Italy, Cyprus, Spain, and France. Notably, Italy, Spain, and France represent major consumers of plant protection products, collectively accounting for nearly 60% of the 340,000t used across 20 EU countries surveyed [81]. The significant volume of homogeneous plastic waste generated in these high-use countries offers substantial potential for recycling or energy recovery, provided the materials are effectively decontaminated.

4. Residual Contaminants in Used Agrochemical Plastic Containers—Properties and Interactions

The use of plastic containers for agrochemicals introduces a critical technical failure at the waste management stage, creating a substantial ecotoxicological “dual problem”: the simultaneous disposal challenge of both the plastic residue and the highly toxic chemical residue it retains. This problem stems from the inherent difficulty in fully decontaminating the primary plastic packaging. The U.S. Environmental Protection Agency [82] estimates that a significant amount of pesticide formulation, ranging from 0.02% to 0.37%, persists inside containers after use, underscoring a fundamental failure in current emptying practices. These residual toxic substances, as emphasized by Eras et al. [83], present a substantial and uncontrolled risk for human health when containers are mismanaged. Monitoring studies confirm the prevalence and diversity of this contamination across various container types [84]. The prevalent use of plastics like high-density polyethylene (HDPE), polypropylene (PP), and fluorinated HDPE for packaging (chosen for their chemical resistance and durability [85]) exacerbates this failure. Even following procedures like triple rinsing, residual pesticides frequently linger on container walls. Laboratory studies show that the surfaces of materials like HDPE actively retain residues, demonstrating that standard rinsing protocols are often insufficient to remove all toxins. Consequently, these residual chemicals may be released into the environment during subsequent attempts at decontamination or when the container is improperly handled, directly contributing to the “dual problem” of plastic and chemical pollution.

4.1. Pesticides and Their Chemical Characteristics

Inorganic pesticides, including copper sulfate, basic copper oxychloride, elemental sulfur, lime–sulfur, and historically used arsenical and mercurial compounds, are mineral-based agents characterized by simple ionic or sparingly soluble salt and oxide forms [86]. They act primarily as contact protectants rather than systemic agents, meaning they must remain on the target surface to be effective [87]. Their ionic speciation and redox chemistry largely determine environmental mobility and persistence, with factors such as pH, dissolved organic matter (DOM), and competing ions influencing their fate. Aging processes, such as oxidation and photo-oxidation, can further modify their surface chemistry, potentially altering bioavailability and environmental reactivity. Research into the interaction between inorganic pesticide actives and plastics focuses primarily on microplastics representative of packaging residues—including PE, PP, PVC, PET, and PS. Sorption capacity depends on polymer polarity, surface chemistry, and roughness. In some studies, the sorption trend for divalent cations (Cu²⁺, Zn²⁺, Pb²⁺, Cr³⁺) appears to follow the order PVC ≥ PS ≥ PET/PE/PP. Copper(II) adsorption on PS and PET is typically spontaneous and endothermic, driven by electrostatic attraction, and enhanced by surface oxidation or DOM coatings [88]. Environmental conditions—including pH (near-neutral to alkaline favors adsorption), ionic strength (low values favor adsorption), and DOM content—strongly influence both adsorption and desorption. These findings suggest that residues from inorganic fungicides can persist on or within polymer packaging materials, creating potential for secondary contamination events during disposal, storage, or recycling.

4.2. Organic Pesticides and Their Chemical Characteristics

Organic pesticides encompass diverse chemical classes—organochlorines (OCs), organophosphates (OPs), carbamates, synthetic pyrethroids, neonicotinoids, azole fungicides (e.g., triazoles, imidazoles), strobilurins, phenylureas, and sulfonylureas—with physicochemical properties that govern environmental fate and interactions with plastics [89]. Hydrophobicity (log K_ow), ionization (pK_a), water solubility, and vapor pressure are primary determinants of sorption, transport, and persistence [90]. OCs (e.g., DDT and HCH legacy residues) show high log K_ow and persistence; OPs (e.g., chlorpyrifos, diazinon) are less persistent but partition strongly to organic phases; carbamates (e.g., carbaryl) hydrolyze more readily; pyrethroids (e.g., cypermethrin, permethrin) are highly hydrophobic with strong solid-phase affinity; and neonicotinoids (e.g., imidacloprid) are more polar and partially ionizable, giving distinct sorption behavior [89]. Formulations include solvents and surfactants (co-formulants) that can increase apparent solubility, modify interfacial tension, and thereby alter sorption kinetics and partitioning into polymers. Studies on polyethylene (PE; including HDPE/LDPE), polypropylene (PP), polystyrene (PS), polyvinyl chloride (PVC), and polyethylene terephthalate (PET) show that sorption of neutral organic pesticides is dominated by partitioning into the hydrophobic polymer phase and specific interactions at the surface [91]. Sorption generally increases with log K_ow and decreases with water solubility, with PS and PVC often exhibiting higher capacities than PE/PP because of π-π interactions (PS) and higher polarity (PVC) [92]. Aging (UV/thermal/oxidative) increases surface roughness and introduces O-containing groups, which can enhance sorption of many organics yet sometimes accelerate desorption of others via swelling and diffusion changes [36].

Reported parameters for representative pesticide–polymer pairs illustrate typical ranges: Pyrethroids on PE/PP/PS: pseudo-second-order k₂ ≈ 10⁻³−10⁻² g mg⁻¹ min⁻¹; Freundlich K_F ≈ 10⁻¹−10¹ (mg g⁻¹)(L mg⁻¹)^1/n; higher capacities on PS than PE/PP [92].

Triazole/strobilurin fungicides on PE/PET: Langmuir q_max ≈ 0.2–5 mg g⁻¹, K_L ≈ 0.05–1 L mg⁻¹ depending on temperature and aging; desorption 10–40% over 24–72 h with DOM present [36].

Neonicotinoids (ionizable) on PE/PVC: sorption increases near pH where neutral species dominates; q_max ≈ 0.05–0.5 mg g⁻¹, k₂ ≈ 10⁻³–10⁻² g mg⁻¹ min⁻¹; enhanced by ionic strength reduction and DOM coatings [36]. (Values are representative ranges drawn from the cited reviews/experimental papers; exact numbers depend on polymer grade, particle size, temperature, and formulation co-solvents.)

Surfactants and solvents in commercial formulations can increase apparent sorption by improving wetting and facilitating diffusion into amorphous regions and can also enhance desorption when rinsing water contains co-solvents [36,93]. pH and ionic strength modify ionization and electric double layers for ionizable pesticides, while DOM can either promote sorption (via co-sorption/bridging) or compete in solution, altering net fluxes [92]. Temperature generally increases diffusion coefficients and q_t rates [88].

5. Ecotoxicological Effects of Pesticide Residues and Plastic Interactions in Agroecosystems

The ecotoxicological consequences of pesticide residues, including those retained in plastic packaging and subsequently released into the environment, are increasingly recognized as complex and multifaceted. These effects extend beyond immediate target pests to plants, animals, soil properties, and soil microbial communities, with plastic–pesticide interactions acting as important mediators of fate and toxicity.

Persistent pesticides, including OCPs and OPs, threaten plant health by accumulating in root tissues and translocating to aerial parts, potentially disrupting vital physiological functions such as photosynthesis, transpiration, and nutrient uptake [94]. Additionally, fungicides like copper-based formulations can induce phytotoxicity at high concentrations, leading to leaf chlorosis, inhibited growth, and oxidative stress [94]. The presence of microplastics derived from degraded agrochemical packaging has also been demonstrated to influence plant uptake of pesticide residues. Evidence shows that microplastics can adsorb hydrophobic pesticides such as organochlorines and organophosphates, potentially inhibiting their degradation and increasing their persistence and bioavailability in the environment [95]. These processes increase the risk of sub-lethal toxicity, with implications for crop productivity and food safety.

Animals, particularly non-target terrestrial and aquatic species, are also vulnerable to pesticide–plastic interactions. Residues of hydrophobic pesticides such as pyrethroids and triazoles, once bound to plastic polymers, may be ingested by livestock or wildlife when containers are mismanaged or discarded in open environments [83]. Aquatic fauna are especially at risk, as microplastics act as vectors that enhance the persistence and transport of pesticides, leading to trophic transfer. Laboratory and field observations indicate that fish and invertebrates exposed to pesticide-laden microplastics exhibit endocrine disruption, oxidative stress, behavioral changes, and reduced reproductive output [92]. Birds and mammals may encounter pesticide-contaminated plastics indirectly through ingestion of contaminated prey or directly via misuse of discarded pesticide containers for water storage, a practice documented in low- and middle-income agricultural regions [84]. The chronic ingestion of pesticide residues has been linked to bioaccumulation, immunotoxicity, and neurological impairments in higher organisms, raising concerns about biodiversity loss and ecosystem stability.

Soil systems represent a major sink for pesticide residues and degraded plastic fragments, and their combined presence alters key physicochemical properties. Hydrophobic pesticides bound to microplastics can modify soil sorption dynamics by creating additional reservoirs for slow release, thereby prolonging pesticide half-lives in the soil environment [91,92]. Residual fungicides such as copper compounds increase soil metal burdens, which can negatively affect cation exchange capacity and alter pH buffering [88]. Persistent pesticide residues, in interaction with plastic fragments, have also been reported to change soil hydrophobicity, potentially reducing infiltration and enhancing runoff risks [36]. Moreover, aged polymers containing pesticide residues may release additives such as phthalates and bisphenols, further complicating soil chemistry by contributing endocrine-disrupting contaminants [85]. Collectively, these processes impair soil structure and fertility, jeopardizing long-term agricultural sustainability.

Soil microbial communities, which are essential for nutrient cycling, are highly sensitive to contamination [89]. Organophosphates and carbamates reduce microbial activity and biomass by inhibiting key enzymes [89]. Furthermore, fungicides disrupt symbiotic relationships with plants by reducing beneficial fungi, such as arbuscular mycorrhizal fungi [87]. Plastic–pesticide interactions exacerbate these effects: pesticide sorption to polymers can act as a reservoir for slow, chronic microbial exposure, while biofilms that form on microplastic surfaces serve as novel niches where pesticide residues accumulate at elevated concentrations [36]. The altered microhabitat conditions may select for resistant microbial strains, with potential implications for antimicrobial resistance development. Additionally, the combined stress of pesticide residues and plastic additives has been linked to reductions in soil respiration, nitrification, and decomposition processes, thereby undermining soil ecosystem services.

6. Mechanisms of Transport to Aquatic Ecosystems and Possible Ecotoxicological Effects

Plastic pesticide residues from agricultural use reach inland waters through a suite of hydrological and waste-management pathways that also transport plastic particles originating from films, irrigation components, and containers (Figure 6). Overland flow during storm events mobilizes both dissolved and particle-bound pesticides, while erosion and macropore flow entrain soil particles and microplastics (MPs) to ditches and streams; subsurface leaching contributes residues to tile drains and groundwater that later re-emerge as baseflow [30,96]. Additional inputs arise from rinsate produced during triple-rinsing or pressure-rinsing of containers, which, if applied to fields or poorly contained, can deliver concentrated mixtures of actives and co-formulants to receiving waters [12,82]. Where post-consumer containers are mismanaged, wind and surface runoff can disperse plastic fragments that sorb pesticides and then act as mobile vectors in rivers [83,85]. Atmospheric deposition provides a further route for both pesticides and microplastic fibers, which deposit on catchments and are rapidly flushed to waterways during rainfall [35,96]. Once in freshwaters, polymer–pesticide interactions evolve with aging (UV/oxidation), pH, ionic strength, and dissolved organic matter (DOM), which together control sorption–desorption equilibria, residence time on particles, and the timing of pulsed releases during changing flow and salinity conditions [36,91,92].

Figure 6. Transfer of (A) pesticide residues and plastics to Freshwaters and (B) ecotoxicological effects.

Aquatic macrophytes experience direct exposure to dissolved pesticides and to pesticide-laden MPs that settle in sediments or remain suspended in the photic zone. Contact fungicides (e.g., copper-based products) can provoke leaf chlorosis, inhibited growth, and oxidative stress, with severity increasing under low alkalinity and high light that favor redox cycling of copper [87]. Hydrophobic herbicides and insecticides partition into plant cuticles and biofilms, and their apparent bioavailability can increase when MPs deliver locally concentrated residues to plant surfaces; aging of polyethylene and polypropylene increases roughness and oxygenated functional groups, enhancing sorption and prolonging residence near roots and leaves [36]. Experimental exposures report reduced chlorophyll content, altered nutrient stoichiometry, and suppression of antioxidant enzymes in rooted and floating plants when co-exposed to MPs and typical organic actives (e.g., pyrethroids, triazoles), consistent with vector-mediated delivery and slow desorption that extend exposure duration [92]. Such sub-lethal impacts can attenuate primary production and macrophyte habitat quality, with knock-on consequences for oxygen dynamics and refugia for invertebrates and fish.

Phytoplankton are sensitive to biologically effective concentrations of both dissolved pesticides and colloid-bound residues. Laboratory and mesocosm studies demonstrate that micro- to nano-sized plastics alter algal exposure in two opposing ways: they can sorb and temporarily sequester hydrophobic pesticides, reducing freely dissolved fractions, yet they also deliver concentrated microenvironments where desorption near cell surfaces elevates local dose [91,92]. Interactions with DOM and photo-aging of plastics often tip this balance toward increased effective exposure by accelerating mass transfer to cells [36]. Consequences include impaired photosystem II efficiency, reduced growth rates, and elevated reactive oxygen species, reported for green algae and diatoms under co-exposure to MPs and common herbicides/insecticides; aggregation of algal cells with plastics additionally causes shading and sedimentation, restructuring species composition toward tolerant taxa [92]. Because phytoplankton regulates basal food-web productivity, these effects propagate to zooplankton and higher trophic levels.

Fish encounter pesticide residues through the water column, diet, and incidental ingestion of plastics, which can act as vectors for adsorbed chemicals (Figure 6). In a seminal field experiment, ingestion of degraded consumer plastics transferred sorbed priority pollutants to fish tissues and induced hepatic stress responses, establishing a causal link between plastic-borne chemicals and fish toxicity. Subsequent laboratory work shows that MPs combined with hydrophobic pesticides (e.g., pyrethroids, azoles) elevate internal concentrations and potentiate sub-lethal effects including oxidative stress, endocrine disruption, and altered swimming behavior relative to pesticide alone, particularly when particles are aged and coated with biofilms that promote desorption in the gut [92,97]. Early life stages are especially vulnerable: embryos and larvae exhibit impaired hatching, developmental malformations, and reduced growth under co-exposure scenarios consistent with time-dependent desorption from ingested particles and trophic transfer from contaminated prey [98].

Microbial communities govern nutrient cycling, pollutant transformation, and food-web base flows. Plastics rapidly develop biofilms (“plastisphere”) that differ from surrounding planktonic communities and often enrich for degraders and potentially pathogenic taxa [98]. Sorbed pesticides accumulate within these biofilms, where extracellular polymeric substances and steep micro-gradients of pH and redox modulate transformation pathways and stress responses [91,92]. Reported outcomes include suppressed respiration and nitrification, shifts in community composition and functional genes, and enhanced potential for horizontal gene transfer within particle-associated consortia [97]. Because biofilms colonize both pelagic particles and benthic substrates, pesticide–plastic complexes create persistent, spatially heterogeneous exposure niches that sustain chronic, low-level toxicity even after dissolved concentrations decline. Collectively, these processes alter microbial food-web efficiency and biogeochemical cycling, with ramifications for water quality and ecosystem resilience.

7. Conclusions

This review demonstrates that agrochemical plastics and their associated residues represent a critical but underexplored vector of environmental contamination. Inorganic pesticides, particularly copper- and sulfur-based fungicides, and organic pesticides such as carbamates, triazoles, and organophosphates, exhibit distinct chemical properties that govern their persistence and interactions with polymeric packaging materials. High-density polyethylene (HDPE), polypropylene (PP), and related polymers are widely used for agrochemical containers due to their resistance and stability, yet they retain measurable residues even after standard triple-rinsing procedures. These residues present risks of secondary release during recycling, disposal, or environmental exposure.

Adsorption and desorption processes strongly influence the environmental fate of both inorganic and organic pesticide residues associated with plastics. Metal-based species (Cu, Zn, As, Pb, Cr) interact with microplastics and packaging polymers through electrostatic attraction, complexation, and surface precipitation, with sorption capacity enhanced by polymer aging, roughness, and dissolved organic matter. Organic pesticides exhibit hydrophobic partitioning and surface sorption, with desorption driven by pH, ionic strength, and surfactant presence. These processes transform packaging plastics into long-term vectors for agrochemical transport in soils and aquatic systems.

In terrestrial environments, pesticide–plastic interactions exacerbate soil contamination, influencing physicochemical properties, microbial community composition, and plant uptake pathways. Transport mechanisms, including tillage, irrigation, biological activity, and preferential flow, promote vertical migration of microplastics with adsorbed residues into deeper soil horizons and potentially groundwater. Similarly, runoff, leaching, and colloid-mediated transfer extend contamination into aquatic ecosystems, where residues affect phytoplankton, aquatic plants, fish, and microbial assemblages. Documented effects include oxidative stress, impaired photosynthesis, endocrine disruption, growth inhibition, and altered community dynamics, collectively threatening ecosystem resilience and food security.

Agrochemical plastic packaging, primarily composed of rigid HDPE bottles, jerrycans, and bulk containers with polypropylene closures, plays a critical role in the safe storage and delivery of pesticides and fungicides. Despite standardized rinsing and disposal guidelines, studies show that residues frequently persist in container walls, especially for hydrophobic or aged products. These residues present risks during recycling, storage, or in cases where empty containers are misused for non-agricultural purposes. At the same time, recovery and recycling programs demonstrate that high-volume collection of cleaned containers is both feasible and economically viable, though uneven participation and inconsistent rinsing quality remain challenges.

Emerging concerns include PFAS leaching from fluorinated HDPE packaging, highlighting the need to evaluate barrier technologies not only for performance but also for long-term chemical safety. Additionally, plastics from packaging can interact with both organic and inorganic pesticide residues, sorbing and releasing them under varying environmental conditions, making packaging an active vector in agrochemical life cycles. However, research gaps remain on real-container kinetics, the sorption of inorganic actives, and harmonized methods for measuring plastics and sorbed residues under realistic agricultural conditions.

The path forward requires strengthening compliance with rinsing protocols, scaling up extended producer responsibility schemes, and creating transparent markets for recycled resins. Procurement choices should prioritize safer, non-fluorinated packaging options wherever possible, and stewardship programs should integrate standardized laboratory methods to monitor residues and recycled resin quality. By closing knowledge gaps and aligning farm practices, recycling logistics, and analytical methods, agrochemical packaging can transition toward safer, lower-risk, and circular solutions aligned with sustainable agricultural goals.

Taken together, the evidence highlights the urgent need for integrated management approaches addressing both agrochemical application and plastic packaging waste. Effective solutions require harmonized collection and recycling schemes, innovation in container materials and design, and comprehensive monitoring of pesticide–plastic interactions in real agricultural settings. Long-term, multidisciplinary field studies are essential to quantify cumulative impacts under realistic usage and climatic conditions. Without such coordinated efforts, the combined legacy of pesticides and plastic residues will continue to compromise soil fertility, water quality, biodiversity, and ultimately human health.