1 Introduction

The terrestrial environment serves as a significant reservoir for microplastics, receiving 4–23 times more plastic waste annually compared to the marine environment (de Souza Machado et al. 2018a; Nizzetto et al. 2016). However, due to the challenges associated with separating microplastics from soil organic matter (SOM) and minerals, it was not until Rillig’s call for research in 2012 that scientific attention began to focus on surveying microplastics in soil (Rillig 2012). Subsequently, the scientific community discovered the presence of microplastics in various soils worldwide, including home gardens, greenhouses, agricultural lands, coastal areas, industrial sites, and floodplain soils (Table S1). Based on our global inventory of microplastics in soil environments, recovered microplastic concentrations range widely from baseline levels up to 20 mg kg–1 in inhabited areas and reach as high as 67,500 mg kg–1 in industrial soil (Figs. 1 and S1), while some studies have suggested even more severe levels of microplastic contamination that may occur in specific soils (Huerta Lwanga et al. 2017). The sources of soil microplastics primarily encompass sewage sludge amendment, irrigation, composting, plastic mulching, dry and wet deposition (i.e., rain and snow) from the atmosphere, fragmentation and degradation processes, the mismanaged runoff by sewer systems from roads or littering activities, landfills, as well as plastics-processing plants (Table S1) (Chen et al. 2022c; Chen et al. 2020; Chen et al. 2023a; Feng et al. 2020; van den Berg et al. 2020). Microplastics exhibit remarkable resistance to environmental degradation (de Souza Machado et al. 2018a; He et al. 2005; Tisdall and Oades 1982). In addition, soil aggregates also significantly affect soil porosity, which in turn influences the movement of gases and water and the activities of associated microbial communities (Rillig and Lehmann 2020; Rillig et al. 2017). The size and stability of soil aggregates regulate soil erodibility (López et al. 2000; Planchon et al. 2000; Somaratne and Smettem 1993). A decrease in soil aggregates might decrease the diversity of soil microenvironments, thereby impoverishing the soil structure (Six et al. 2006; Zheng et al. 2016). PA beads were more prone to be incorporated into the soil matrix (de Souza Machado et al. 2018b) and soil minerals (Chen et al. 2009) and has a central role in numerous physical, chemical, and biological processes in soil (Kalbitz et al. 2000). Although only accounting for < 0.25% of the total SOM, DOM is crucial in regulating the turnover of SOM, the transport of nutrients, the solubility and mobility of heavy metals and organic pollutants, and the activity of microbes (Kalbitz et al. 1997, 2003; Temminghoff et al. 1997). Microplastic input could alter the quantity of DOM in soil, which depends on the imbalance between the production and mineralization of DOM. For example, a majority of studies have observed increased DOM content after the addition of different microplastics (Liu et al. 2017; Meng et al. 2022; Shi et al. 2022a; Zhou et al. 2020a). This is understandable, as the increased activity of enzymes involved in the degradation of recalcitrant (phenolic) compounds may lead to the decomposition of the poorly dissolved large compounds in SOM into easily dissolved small compounds (Keuskamp et al. 2015). However, some studies have also suggested a decrease in DOM content in microplastic-introduced soils (Liu et al. 2019), which was potentially due to the sorption of DOM by microplastics and the degradation of DOM (Chen et al. 2018). The divergence may also be associated with polymer type, microplastic concentration, and incubation time (Ren et al. 2020). Moreover, microplastics may also influence the chemodiversity of soil DOM, such as increasing DOM aromaticity and molecular weight and deepening the DOM humification degree. (Chen et al. 2022a, 2023b; Feng et al. 2022; Li et al. 2022b; Liu et al. 2017). For instance, the introduction of biodegradable microplastics could enhance the relative abundance of labile compounds in soil DOM, such as lipid-like, protein/amino sugar-like, and carbohydrate-like compounds, while conventional polystyrene (PS) microplastics may decrease the relative abundance of stable compounds like lignin-like and more condensed aromatic-like compounds (Sun et al. 2022). As suggested by Qiu et al. (2024), labile components underwent degradation and transformation after microplastic addition, leading to increased aromaticity and oxidation degree, reduced molecular diversity, as well as altered nitrogen and sulfur contents within soil DOM. The changes in soil DOM might be the combined results of microplastic-derived DOM input, organo-organic and organo-mineral interactions, and microbial degradation. These factors warrant further investigation. In fact, chemodiversity is an important indicator to evaluate the environmental reactivity and destination of soil DOM. In general, DOM of lower molecular weight and aromaticity has higher bioavailability (Fouché et al. 2020; Ye et al. 2020). Additionally, DOM contains polar groups and phenolic structures, which have a stronger binding force with heavy metals (Dong et al. 2020; Wang et al. 2018). Therefore, the shifts in DOM compositions after microplastic input are an important topic.

There are several hypothesized pathways by which microplastics affect the turnover of native SOM (Fig. 2). (1) Microplastics could change soil physicochemical properties such as soil aggregates (habitats for SOM stabilization) and therefore affect microbe growth and activity, thus changing the turnover of native SOM. (2) Microplastics could serve as an inert carbon source and establish a unique ecological niche for microbial settlement and growth, thus affecting SOM turnover by directly altering the microbial community. (3) Microplastics can adsorb SOM, interact with soil minerals, and affect the interaction between minerals and SOM, thus affecting SOM accessibility to microbes. (4) Microplastics may impact SOM mineralization through negative priming effects stemming from dilution and substrate switching or positive priming effects arising from co-metabolism, etc. The latter is likely to have marginal importance given the inert nature of microplastic carbon. Alternatively, these effects might be initiated by the more easily metabolizable bioplastics or microplastic leachates (Zhang et al. 2023a). (5) Conventional PE microplastics can influence CO2 emissions solely by altering DOM electron transfer capabilities. In contrast, the application of biodegradable microplastics (e.g., PLA) can impact gas emissions by increasing both the quantities and transfer capabilities of soil DOM (Shi et al. 2023).

3.4 Microbial response and contribution

Microorganisms have two critical, contrasting roles in controlling terrestrial canbon fluxes: promoting release of canbon to the atmosphere through their catabolic activities, but also preventing release by stabilizing canbon into a form that is not easily decomposed (e.g., the accumulation and stabilization of microbial necromass) (Liang et al. 2017). It is of significant importance to understand the adaptive and evolutionary responses of soil microorganisms to microplastics.

3.4.1 Microbial community structure

The soil microbial community is an important player in regulating nutrient cycling, maintaining soil structure, and detoxifying noxious chemicals. In addition, microbial diversity serves as a sensitive indicator of soil quality, which can reflect subtle changes and soil function (He et al. 2015; Sebiomo et al. 2011). Microplastics exhibit certain filtration effects on soil biota (Guo et al. 2020), and recent studies have reported shifts in community structure, diversity, and evolutionary consequences of microbes in soil in the presence of microplastics (Han et al. 2024; Lu et al. 2023; Rillig et al. 2019a). Microplastics have been proven to serve as substrates for microbial colonization and assemblage in soil environments, leading to the formation of a unique environment termed the plastisphere (Rillig et al. 2023). Distinct differences were noted for the microbial communities between microplastic surface and ambient soils. For example, the PE microplastic surface was found to carry more plastic-degrading bacteria and pathogens than the surrounding soils (Huang et al. 2019). A similar colonization of plastic-degrading bacteria on microplastic surfaces was also observed in an e-waste dismantling field (Chai et al. 2020). The selection phenomena of microplastics on soil biota shed light on screening microbes for microplastic biodegradation.

Many studies have also reported alterations in bulk microbial diversity and composition in soils after microplastic addition. For example, the bacterial composition in microplastic-amended soils showed significant variations from the control after 90 days of incubation (Huang et al. 2019). Membranous PE and fibrous PP could raise the alpha diversities of the soil microbial community (Yi et al. 2021). In contrast, plastic film residues sharply decreased the soil microbial community and its diversity (Wang et al. 2016), which was attributed to the negative effects of lipophilic phthalate ester additives on the soil biota by destroying cell membrane fluidity (**e et al. 2010; Zhou et al. 2005). Additionally, the soil microbial diversity was lowered in the presence of PE and PVC microplastics (Fei et al. 2020). Until now, the effects of microplastics on microbial community structures have remained unclear. The contradictory results may be attributed to different microplastic types, concentrations, and additive contents in soil. In general, high microplastic concentrations may induce a quick response in the soil microbiota, whereas lower microplastic concentrations may have insignificant effects on the microbial communities. For example, the addition of PE, polyethylene terephthalate (PET), and PVC microplastics at rather low concentrations (< 1%, w/w) after 9 months barely altered the microbial community structure in soil (Judy et al. 2019).

To ease traditional microplastic contamination, biodegradable microplastics have received wide attention and application. As biodegradable microplastics have short lifespans, they are supposed to have more significant effects on soil ecosystems regarding changes in soil biota, greenhouse gas emissions, plant performance, etc. For example, more significant effects on wheat growth were observed for biodegradable microplastics compared to traditional PE microplastics, potentially attributed to microbial immobilization (Qi et al. 2018). More significant variations in bacterial communities, such as the increase in the relative abundance of the genera Bacillus and Variovorax, were noted for biodegradable microplastics than for traditional PE microplastics (Qi et al. 2020b). In addition, PES and PP were observed to increase root symbiosis by arbuscular mycorrhizal fungi, whereas PET had the opposite effect (de Souza Machado et al. 2019). Notably, microplastics are N-limited carbon materials. Thereby, Zhang et al. (2023a) proposed that bioplastics may promote the proliferation of fast-growing r-strategists through co-metabolism, thereby fostering a positive priming effect in the short term. Conversely, under conditions of severe nitrogen deficiency and labile carbon exhaustion, this may lead to the active growth of K-strategists alongside microbial necromass from r-strategists.

In general, microplastics possess a range of effects on soil microbial community, which are still largely unknown. To better underline the ecosystem effects of microplastics, the adaptive and evolutionary responses of soil biota to microplastic stresses should be further addressed in future studies (Rillig 2018; Rillig et al. 2019a).

3.4.2 Enzyme activity

Soil enzymes are useful for monitoring soil health because of their sensitivity to soil stress, energy flow, and nutrient availability (Wang et al. 2015b). Microplastics could affect the excretion of various soil enzymes by microbes, and extracellular enzymes could attach to microplastic surfaces or affect other soil substrates, thus regulating microplastic degradation and turnover of SOM. de Souza Machado et al. (2018b) first evaluated the effects of microplastics on the hydrolysis of fluorescein diacetate enzyme activity and observed a significant correlation between microplastic concentration and microbial activity irrespective of polymer type. This is consistent with the increased hydrolysis of fluorescein diacetate enzyme activity after PE addition (Liu et al. 2017). Similarly, the addition of membranous PE, fibrous PP, and microsphere PP all increased the urease, dehydrogenase, and alkaline phosphatase enzyme activities in soil (Yi et al. 2021). Additionally, the addition of PP microplastics at a rather high application rate (28%, w/w) could significantly increase the activities of fluorescein diacetate hydrolase and phenol oxidase in sandy loam soils (Yang et al. 2018), thus affecting soil C, N, and P cycling and increasing nutrient availability to plants by enhancing microbial hydrolytic activity on SOM (Liu et al. 2017; Yang et al. 2018). However, a lower application rate (7%, w/w) of PP only had marginal effects on the enzyme activities of fluorescein diacetate hydrolase, urease, and phenol oxidase (Yang et al. 2018). In addition, PE and PVC microplastics could increase the activity of urease and acid phosphatase, whereas the activity of fluorescein diacetate hydrolase activity was inhibited (Fei et al. 2020). Decreases in dehydrogenase activity and enzyme activities involved in the C-(β-glucosidase and cellobiohydrolase), N-(leucine-aminopeptidase), and P-(alkaline-phosphatase) cycles were also observed in soils after 28 days of incubation under 100 and 1000 ng g–1 PS nanoplastic treatments, indicating a broad and detrimental impact of PS nanoplastics on soil microbial activity (Awet et al. 2018). Moreover, the residues of plastic film (67.5 kg ha–1) significantly lowered the activity of fluorescein diacetate hydrolase and dehydrogenase by 10% and 20%, respectively (Wang et al. 2016), which may be attributed to the negative effects of concomitant plastic additives. The contradictory results of enzyme activities after microplastic application are potentially attributed to microplastic concentrations and constitutions, plastic additives, and soil properties, the effects of which should be further addressed to better understand the mechanisms. Song et al. (2023) elucidated that PVC elevated β-glucosidase, leucine aminopeptidase, and acid phosphatase activities in both hot- and coldspots within the rice rhizosphere. In contrast, PLA influenced enzyme activities exclusively in the hotspot soil, showing no impact in the coldspot soil. These variations arose from changes in microbial enzyme systems favoring nutrient mining, potentially mitigating some of the adverse effects of microplastics on soil nutrient processes. The close association between soil enzyme activities and soil carbon dynamics necessitates greater emphasis on investigating the response of enzyme activity to microplastic addition, as well as its correlation with other biotic and abiotic processes in soil.

3.4.3 Microbes and enzymes for microplastic degradation

Microplastics can establish a unique ecological niche for certain microbes by providing habitat for microbial settlement and growth. In return, these microbes may contribute to the degradation of microplastics by utilizing polymer carbon as an inert carbon source, consequently impacting the mineralization process of native SOC. As summarized in Table 1, both bacteria and fungi have been proven to have the potential to promote the degradation of microplastics (Russell Jonathan et al. 2011; Shah et al. 2008; Zafar et al. 2013). Some studies have also shown that several bacteria and fungi can use plastic as the sole carbon source, including either in solid or liquid matrices, such as soil (Mohan et al. 2016), compost (Jeon and Kim 2013), and sea water (Harshvardhan and Jha 2013), thus highlighting the potential of such microbes for plastic remediation. Other microbes involved in plastic degradation include Alcaligenes faecalis, Comamonas acidovorans TB-35, Pseudomonas putida, Pseudomonas stutzeri, Saccharomycopsis, Streptomyces sp., and Staphylococcus sp. (Akutsu et al. 1998; Benedict et al. 1983; Caruso 2015; Ghosh et al. 2013).

Table 1 Microbes involved in microplastics biodegradation
Full size table

During the degradation process, it is also very important to identify the enzymes involved (Auta et al. 2018; Jaiswal et al. 2020). Extracellular enzymes secreted by microbes are prone to depolymerize microplastics through hydrolysis reactions (Shah et al. 2008). Lipases, cutinases, carboxylesterases, and laccases have been proven to efficiently degrade microplastics (Lucas et al. 2008). The polymer chain can be cleaved into micromolecular water-soluble intermediates, which may be absorbed by the cells and undergo a special metabolism (Gewert et al. 2015). The final degradation products may end up as CO2, H2O and CH4 and are released into the ambient environment (Tokiwa et al. 2009).

3.4.4 Contribution of microbial biomass and necromass to soil carbon pool

Microbial biomass represents the living or actively growing microorganisms, which is associated with the decomposition efficiency of SOM. To date, data on the effects of microplastics on the accumulation of soil microbial biomass are far from sufficient. A 28-day laboratory incubation showed that the addition of 100 and 1000 ng g–1 PS nanoplastics both significantly decreased soil microbial biomass (Awet et al. 2018), suggesting potentially broad antimicrobial activity of PS nanoplastics on soil microbiota. Interestingly, a gradual increase in soil microbial biomass with time was observed for 10 ng g–1 PS nanoplastic-amended soil throughout 28 days of incubation (Awet et al. 2018). At day 1, the soil microbial biomass remained almost unchanged, whereas the enzyme activities, basal respiration rate, and metabolic quotient decreased, suggesting a sublethal effect with 10 ng g–1 PS nanoplastic amendment. At day 28, the soil microbial biomass was significantly higher, potentially due to the increased antimicrobial activity of PS nanoplastics against some microbial genera over time. Thus, dead cells might provide substrates for resistant microorganisms and thus result in cryptic growth (PostGate 1967). Similarly, several studies also observed a decreased microbial biomass in 1% PP-, PE-, or PLA-amended soils (Blöcker et al. 2020; Shi et al. 2022a). In this case, the decrease in microbial biomass was unlikely caused by toxicity effects, as the used plastics were free of antimicrobial additives. However, Zhang et al. (2023a) noted that bioplastics, such as PHA, PBS, and PLA, elevated microbial biomass carbon and dissolved organic carbon levels. This implies that biodegradable microplastics can concurrently expedite microbial assimilation and the transformation of SOM into dissolved organic substrates.

Microbial necromass refers to the non-living remnants of microorganisms, constituting over 50% of SOC pools and approximately 40 times the amount of live microbial biomass carbon (Liang et al. 2017). Chen et al. (2019). As mentioned above, microplastics can affect the water and nutrient supply to plant roots. Correspondingly, several detoxification mechanisms are involved in alleviating such adverse effects. Long-term stressed plants are prone to secrete extracellular enzymes to degrade microplastics and coexisting pollutants and additives. For example, exposure to PS microplastics leads to shifts in the antioxidant defenses of Vicia faba by secreting antioxidant enzymes such as superoxide dismutase and peroxidase enzymes (Jiang et al. 2019). The increased production of H2O2 in plants, as well as the triggered production of low molecular weight compounds with antioxidant action, also supported the antioxidant defenses of plants to microplastic stressors (Pignattelli et al. 2020). Moreover, the microorganisms and microanimals in plant roots could alleviate the toxic effects by degrading and utilizing microplastics and coexisting intermediates. Unfortunately, the influence of microplastics on plant health and microplastic biodegradation in the rhizosphere of plants is still largely unknown and merits further study.

5 Future research directions

Due to the irresistible increase in plastic production and the lack of effective waste disposal countermeasures, the current pressure of microplastic contamination in soil is expected to continue for many years to come. To better understand the ecosystem effects and risks of microplastics in the soil environment, there are still some critical unknowns that need to be addressed in future work: (1) elucidating the correlation between alterations induced by microplastics in physicochemical properties and soil carbon cycling; (2) clarifying the influence of microplastics on microbial stabilization and mineral protection of soil carbon, with particular attention given to microbial necromass accumulation and carbon storage within mineral-associated fractions; (3) determining the source of carbon emissions from native SOC and microplastics through utilization of 13C isotope technology; (4) investigating the effects of microplastics on rhizosphere dynamics, particularly microbial activity and function, as well as microbial stabilization and mineral protection mechanisms for rhizodeposits within soils; (5) exploring ecosystem-level consequences associated with so-called "eco-friendly" bioplastics, since microbioplastics may exert more pronounced effects on soil biophysical properties, which should be considered during their safe management within agricultural contexts.