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Article

The Effects of Microplastics and Heavy Metals Individually and in Combination on the Growth of Water Spinach (Ipomoea aquatic) and Rhizosphere Microorganisms

by
Jing-Yi Wang
1,
Meng Wang
1,
Jian-Wei Shi
2,
B. Larry Li
3,
Ling Liu
4,
Peng-Fei Duan
1,* and
Zhao-Jin Chen
3,*
1
Collaborative Innovation Center of Water Security for Water Source Region of Mid-Route Project of South-North Water Diversion of Henan Province, College of Life Science, Nanyang Normal University, Nanyang 473061, China
2
Nanyang Ecological Environment Monitoring Center in Henan Province, Nanyang 473000, China
3
Overseas Expertise Introduction Center for Discipline Innovation of Watershed Ecological Security in the Water Source Area of the Mid-Line Project of South-to-North Water Diversion, Nanyang 473061, China
4
College of Agriculture, Henan University of Science and Technology, Luoyang 471023, China
*
Authors to whom correspondence should be addressed.
Agronomy 2025, 15(6), 1319; https://doi.org/10.3390/agronomy15061319
Submission received: 15 April 2025 / Revised: 23 May 2025 / Accepted: 26 May 2025 / Published: 28 May 2025
(This article belongs to the Special Issue Microplastics and Nanoplastics in Agroecosystems: Strategies to Mitigate Plastic Pollution for Sustainable Agriculture)
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Abstract

:
Microplastics (MPs) and heavy metals are commonly present in soil at significant concentrations and can interact in complex ways that pose serious threats to environmental and ecological systems. The effects of combined contamination by different types of heavy metals and microplastics on plants, as well as on soil microbial communities and their functions, remain largely unexplored. In this study, a series of pot experiments was conducted to investigate the effects of composite contamination involving two heavy metals (Cd and Pb) and two types of microplastics polylactic acid (PLA) and polybutylene succinate (PBS) at varying concentrations (0.1% and 0.5%, w/w). The impacts on water spinach (Ipomoea aquatica) growth and heavy metal accumulation were evaluated, and the rhizosphere bacterial and fungal community structure and diversity were analyzed using high-throughput sequencing. The presence of Cd, Pb, and microplastics significantly inhibited the growth of water spinach, reducing both its length and biomass. Under combined microplastic–heavy metal contamination, phytotoxicity increased with rising concentrations of PLA and PBS. Microplastics were found to alter the mobility and availability of heavy metals, thereby reducing their accumulation in plant tissues and decreasing the levels of available potassium and phosphorus in the soil. Furthermore, microplastic–heavy metal interactions significantly influenced the composition and diversity of soil microbial communities, leading to an increased abundance of heavy-metal-tolerant and potential plastic-degrading microorganisms. A strong correlation was observed between microbial community structure (both bacterial and fungal), soil physicochemical properties, and plant growth. Functional predictions using PICRUSt2 suggested that the type and concentration of microplastics significantly affected rhizosphere microorganisms’ metabolic functions. In conclusion, the present study demonstrates that combined microplastic and heavy metal contamination exerts a detrimental effect on soil nutrient availability, resulting in alterations to soil microbial community composition and function. Furthermore, this study shows that these contaminants can inhibit plant growth and heavy metal uptake. The findings provide a valuable contribution to the existing body of knowledge on the ecotoxicological impacts of microplastic–heavy metal composite pollution in terrestrial ecosystems.

1. Introduction

Soil is a fundamental component for sustaining terrestrial life. In recent years, large-scale industrialization and urbanization have led to increased concentrations of heavy metals in soil, resulting in growing environmental concern regarding soil heavy metal pollution [1]. Once released into the soil, heavy metals tend to accumulate in the plough layer and persist in the soil matrix. Due to their non-degradable nature and resistance to microbial decomposition, these pollutants are difficult to remove and pose stress on plant growth. Furthermore, heavy metals can enter the human body through the food chain, posing significant risks to human health. According to survey reports, the concentration of heavy metals in Chinese soils exceeded standard limits in 19.4% of samples, with exceedance rates of 7% and 1.4% for cadmium (Cd) and lead (Pb), respectively [2]. Globally, surveys conducted in countries such as the United States and across Europe have shown that Cd and Pb are also major components of heavy metal pollution [3]. Cd2+ in particular, exhibits high water solubility and mobility and can be readily absorbed by plants via non-selective ion channels and subsequently accumulate through translocation pathways, presenting potential health risks [4]. Excessive Pb levels in soil have been shown to inhibit plant growth and development and interfere with the uptake of essential mineral nutrients, ultimately reducing crop yield and quality [5]. Studies have demonstrated that exposure to Cd and Pb can have harmful effects on various organs, including the bones, kidneys, and immune system. For example, Cd tends to accumulate in the kidneys, leading to damage of epithelial cells and, ultimately, kidney failure [6]. Even more concerning is that heavy metals in natural environments often coexist with a complex mixture of other pollutants [7].
Microplastics (MPs), as a new type of environmental pollutant, refer to plastic fragments or particles with sizes ranging from 0.1 µm to 5 mm [8]. Due to the use of agricultural plastic films, wastewater irrigation, biosolids, and organic fertilizers, MPs have become a common emerging contaminant in soils, posing serious global concern and attracting widespread attention from researchers. In China, the annual production of microplastics has exceeded 80 million tons, with approximately 20% directly released into the environment. In Europe, the concentration of MPs in the Danube River has been reported to reach 500 particles per cubic meter [9]. The accumulation of MPs in soil can disrupt soil structure, reduce fertility, and negatively impact ecosystem functions [10]. Studies have shown that high concentrations of plastic film residues in soil reduce saturated hydraulic conductivity, alter microbial activity and abundance, and further affect soil fertility [11]. Moreover, due to their large specific surface area and diverse surface functional groups, MPs are capable of adsorbing metal pollutants, thereby serving as effective carriers for heavy metals [12]. Research has demonstrated that MPs can adsorb metals such as Cu, Cd, and Pb, intensifying their ecological impacts [13,14,15]. The accumulation of MPs in plant tissues can block cell junctions or pore walls, thereby hindering the absorption and transport of nutrients [16]. These findings suggest that the combined pollution of MPs and heavy metals can severely disrupt soil microbial communities, negatively affect the development of plants and animals in terrestrial ecosystems, and ultimately impair soil multifunctionality [17,18]. Although research on the interactions between MPs and organic pollutants is relatively abundant, studies focusing on the co-contamination of MPs and heavy metals remain limited.
The rhizosphere serves as a critical interface for plant–microbe interactions. Microbial colonization of the rhizosphere can establish symbiotic relationships with host plants, promoting nutrient acquisition, enhancing growth, and increasing tolerance to abiotic stress and pathogens [14,19]. High-throughput sequencing is a powerful tool for investigating the composition and structure of soil microbial communities [20]. By sequencing the 16S rRNA and ITS gene regions, researchers can analyze the species composition, abundance, and functional potential of bacterial and fungal communities in detail [21]. Studying the responses of soil microbiota to co-contamination by MPs and heavy metals is crucial for assessing ecological risks, promoting green remediation technologies, and supporting sustainable soil management. Although numerous studies have examined the effects of either heavy metals or MPs individually, fewer have addressed the more realistic scenario of multiple pollutants coexisting in the environment. While the responses of bacterial communities to combined microplastic and heavy metal stress have received increasing attention, the effects on fungal communities remain largely understudied.
This study aims to address the following scientific questions: (1) How do different types of MPs and heavy metals interact to affect plant growth and soil microbial communities? (2) Do fungal communities respond differently to composite pollution compared to bacterial communities? (3) What are the primary factors driving these microbial responses? To answer these questions, this study investigates the effects of compound contamination by two heavy metals (Cd and Pb) and two types of MPs—polylactic acid (PLA) and polybutylene succinate (PBS)—on the growth and metal accumulation in water spinach (Ipomoea aquatica). Additionally, it evaluates the impact of combined pollution on soil microbial community structure and function using high-throughput sequencing. The findings provide a scientific basis for understanding and managing the ecological effects of MP–heavy metal composite pollution in soil.

2. Materials and Methods

2.1. Experimental Materials

The seeds of water spinach (Ipomoea aquatica) used in this study were obtained from Zhuzhou Nongzhizi Seed Industry Co., Ltd. (Zhuzhou, China). Uniform, undamaged seeds with full grain development were selected for the experiment. The MPs used were polylactic acid (PLA) and polybutylene succinate (PBS), each with a particle size of 100 μm, sourced from the Dongguan Plastic Business Department (Dongguan, China). Soil samples were collected near Moon Season Garden at Nanyang Normal University, located in Nanyang City, Henan Province, China. The topsoil (0–20 cm) from the cultivated layer was collected and subsequently air-dried. The soil was manually cleaned to remove visible debris and then ground and sieved through a 20-mesh sieve to ensure uniform particle size.

2.2. Pot Experiment

Following sieving, the soil was spiked with CdSO4·8H2O and Pb(NO3)2 to achieve final concentrations of 10 mg·kg−1 for Cd and 200 mg·kg−1 for Pb. The soil was thoroughly mixed and subsequently incubated beginning in January 2024. The experimental design included two types of MPs (PLA and PBS) and two concentrations (0.1% and 0.5% w/w). A control group (CK), without the addition of heavy metals or MPs, was established alongside treatment groups exposed to either individual contaminants (heavy metals or MPs alone) or their combinations. The detailed experimental setup is summarized in Table 1. Each treatment was conducted in triplicate. Germinated water spinach seeds were sown into pre-treated soil, and the soil was pre-weighed to ensure that each pot contained 0.75 kg of soil. The pots were irrigated regularly to maintain adequate soil moisture, and humidity and temperature were controlled throughout the plant’s growth period. When the plants reached a defined growth stage, three uniformly developed individuals were retained in each pot. All pots were arranged randomly and periodically rotated to ensure uniform light exposure. After a 60-day growth period, plant and soil samples were collected for subsequent analysis.

2.3. Sample Processing

Plant samples were harvested, and rhizosphere soil was collected using the shake-off method. The plants were rinsed with deionized water and then immersed in a 0.01 mol·L−1 EDTA-2Na solution for 20 min. Subsequently, they were placed in paper envelopes and dried in an oven at 80 °C for 24 h. The aboveground and belowground parts were then separated and measured for length and dry weight. Both portions were ground into powder and subjected to acid digestion. The concentrations of Cd and Pb in the plant tissues were quantified using inductively coupled plasma–optical emission spectrometry (ICP-OES; Optima 2100 DV, PerkinElmer, Springfield, IL, USA). The accumulation of Cd or Pb in aboveground and belowground plant tissues was calculated using the following formula: Cd or Pb accumulation = Cd or Pb concentration × plant dry weight.
Soil samples were air-dried and sieved through a 1 mm mesh. A 5.00 g aliquot of soil (±0.01 g) was placed into a 25 mL centrifuge tube, to which 20 mL of extraction solution was added. The extraction solution was a mixture of DTPA-CaCl2-TEA, c(DTPA) = 0.005 mol·L−1, c(CaCl2) = 0.01 mol·L−1, and c(TEA) = 0.1 mol·L−1. The tubes were sealed and shaken at 160–200 rpm for 2 h at 25 °C. Afterward, the samples were centrifuged for 10 min, and the supernatant was transferred to a clean tube and centrifuged again for another 10 min. The final extract was analyzed for DTPA-extractable Cd and Pb using ICP-OES. Soil physicochemical properties were determined as follows: soil pH was measured using a pH meter; available phosphorus was determined by the molybdenum–antimony colorimetric method after leaching; available potassium was determined by flame photometry; total phosphorus was determined by the NaOH fusion–molybdenum–antimony colorimetric method; and total potassium was determined by the NaOH fusion–flame photometry method.

2.4. High-Throughput Sequencing

Rhizosphere soil samples were weighed and sent to Shanghai Major Biomedical Technology Co., Ltd. (Shanghai, China) for DNA extraction and sequencing. After genomic DNA was extracted, the DNA quality was assessed using 1% agarose gel electrophoresis. PCR amplification was performed using the TransGen Aap-221-02 system, which includes TransStart FastPfu DNA polymerase (Beijing, China). A 20 μL reaction volume was prepared with primers 341F (5′-CCTACGGGNGGCWGCAG-3′) and 805R (5′-GACTACHVGGGGTATCTAATCC-3′) targeting the V3–V4 hypervariable regions of the bacterial 16S rRNA gene. For fungal community analysis, primers ITS1F (5′-CTTGGTCATTTAGAGGAAGTAA-3′) and ITS2 (5′-GCTGCGTTCTTCATCGATGC-3′) were used to amplify the ITS1 region. PCR was conducted using an ABI GeneAmp 9700 thermal cycler (Waltham, MA, USA) under the following conditions: initial denaturation at 95 °C for 3 min; 30 cycles of 95 °C for 30 s, annealing at 55 °C for 30 s, and extension at 72 °C for 45 s; followed by a final extension at 72 °C for 10 min. All samples were processed under standardized experimental conditions, with triplicate reactions per sample. PCR products were verified by 2% agarose gel electrophoresis, and products of the same sample were pooled. Gel purification was performed using the AxyPrep DNA Gel Recovery Kit (AXYGEN, Union City, CA, USA), and DNA was eluted in Tris-HCl buffer. DNA quantification was conducted using the QuantiFluor™-ST Blue Fluorescence Quantification System (Promega, Madison, WI, USA). The purified amplicons were pooled in equimolar concentrations based on sequencing requirements and subjected to paired-end sequencing using the Illumina platform (San Diego, CA, USA). After sequencing, raw reads were merged using paired-end joining strategies, followed by quality control and filtering. Amplicon Sequence Variant (ASV) clustering and taxonomic classification were performed. Alpha and beta diversity indices were calculated, and taxonomic compositions at various levels were statistically analyzed. Multivariate statistical analyses and significance testing were applied to explore the community structure and phylogenetic patterns. Redundancy analysis (RDA) and Mantel tests were conducted and visualized using R software (version 4.2.0). The functional profiles of the bacterial community were predicted using PICRUSt2 (version 2.2.0; https://github.com/picrust/picrust2, (accessed on 25 May 2025)) based on the Kyoto Encyclopedia of Genes and Genomes (KEGG; http://www.genome.jp/kegg/, (accessed on 25 May 2025)) database [22].

2.5. Data Analysis

All data were analyzed based on three biological replicates. Statistical analysis was performed using SPSS 23.0 software. Differences among groups were tested using one-way ANOVA and independent-sample t-tests, with significance set at p < 0.05.

3. Results

3.1. Effects of Different Treatments on Soil Physicochemical Properties and Bioavailable Metal Content

Compared with the CK group, the Cd treatment group showed a significant reduction of 20% in available potassium (AK), while the Pb treatment group showed a 19% decrease (Table 2). The combined treatments of MPs and heavy metals also exhibited an overall decreasing trend in soil AK, with reductions ranging from 6% to 18%. The Cd treatment led to a 28% decrease in available phosphorus (AP) compared to the CK group. Similarly, the Pb treatment group exhibited a 36% reduction in AP. The addition of PLA and PBS at varying concentrations resulted in a decline in AP content ranging from 10% to 28%. In addition, the combined treatments of Cd and Pb with PLA or PBS led to reductions in total phosphorus (TP) and potassium (TK) contents, with the most pronounced declines observed in the co-contaminated groups (Table 2).
In terms of bioavailable Cd (DTPA-Cd), the addition of 0.1% PLA and 0.1% PBS reduced DTPA-Cd content by 9% and 13%, respectively, compared to the Cd-only group. In contrast, the 0.5% PLA treatment resulted in an 8% increase in DTPA-Cd content. For bioavailable Pb (DTPA-Pb), the 0.1% PLA + Pb and 0.5% PLA + Pb treatments increased DTPA-Pb content by 44% and 47%, respectively, compared to the Pb-only group. Similarly, the 0.1% PBS + Pb and 0.5% PBS + Pb treatments increased DTPA-Pb content by 41% and 6%, respectively. Overall, DTPA-Pb levels were higher in the PLA-supplemented groups than in the PBS-supplemented groups, with increases of 2% and 28% at concentrations of 0.1% and 0.5%, respectively.

3.2. Effect of Different Treatments on the Biomass of Water Spinach

Compared with the CK group, the lengths of the aboveground and belowground parts of water spinach were reduced by 35.50% and 32.27%, respectively, under Cd stress, and by 16.63% and 18.24%, respectively, under Pb stress (Figure 1a). In the MPs-treated groups, both aboveground and belowground lengths also decreased relative to the CK group, with the reductions becoming more pronounced at higher MPs concentrations. Under combined MPs and Cd contamination, the lengths of the aboveground and belowground parts decreased by 17.75% and 33.34%, respectively, in the 0.1% PLA + Cd and 0.5% PLA + Cd treatments. In the 0.1% PBS + Cd and 0.5% PBS + Cd treatments, the reductions were 24.74% and 42.15%, respectively.
As shown in Figure 1b, changes in dry weight followed a pattern similar to that of plant length. Compared to the CK group, the dry weights of the aboveground and belowground parts decreased by 40.95% and 35.96%, respectively, under Cd stress, and by 19.05% and 34.21% under Pb stress. Treatments with MPs alone also reduced the dry weight of water spinach, and this inhibitory effect was more severe with PLA and PBS at higher concentrations. Specifically, the dry weight of the aboveground parts decreased by 6.67% and 29.52% in the 0.1% and 0.5% PLA treatments, while the belowground parts decreased by 6.14% and 50.88%, respectively. For the PBS treatments, reductions in aboveground dry weight were 38.10% and 58.10%, and in belowground dry weight, they were 40.35% and 64.51%, respectively. The effect of PLA + Cd combined treatment on dry weight was greater than Cd alone, with the 0.5% PLA + Cd and 0.5% PBS + Cd treatments reducing aboveground dry weight by 32.26% and 46.77%, respectively. A similar trend was observed under Pb stress, where the combination of MPs and Pb resulted in substantial reductions in aboveground biomass.
Overall, the reductions in length and dry weight indicate that both heavy metals (Cd and Pb) and MPs individually induce phytotoxicity in water spinach. However, the toxicity was markedly amplified under combined contamination, particularly at higher concentrations of PLA and PBS.

3.3. Effects of Different Treatments on the Content and Accumulation of Cd and Pb in Water Spinach

As shown in Figure 2a, compared with the single Cd-contaminated treatment group, the Cd content in the aboveground and belowground parts of water spinach was reduced in the 0.1% PBS + Cd and 0.5% PBS + Cd combination groups, with reductions of 24–27% and 12–18%, respectively. In contrast, the combined PLA + Cd contamination treatment showed an increasing trend in Cd content in both plant parts, with a more pronounced rise observed at the 0.5% PLA level compared to 0.1%. As illustrated in Figure 2b, Cd accumulation in the aboveground part was reduced by 29.99%, 41.84%, and 46.62% in the 0.5% PLA + Cd, 0.1% PBS + Cd, and 0.5% PBS + Cd treatment groups, respectively, compared to the Cd-only group. However, Cd accumulation in the belowground part increased by 5.45% and 45.45% in the 0.1% and 0.5% PLA + Cd groups, respectively. In contrast, the combined PBS + Cd contamination reduced belowground Cd accumulation by 13.52% to 41.59%.
Compared with the single Pb-contaminated group, both PLA + Pb and PBS + Pb treatments led to a decrease in Pb content in the aboveground parts of water spinach, with the reduction becoming more pronounced at higher MP concentrations. Specifically, the combination of 0.5% PLA or 0.5% PBS with Pb resulted in a 33% and 35% decrease in aboveground Pb content, respectively. Similarly, the content of Pb in the belowground part was reduced by 16% and 20% in the 0.5% PLA + Pb and 0.5% PBS + Pb treatments, respectively. As shown in Figure 2d, co-contamination with MPs and heavy metals led to a further reduction in Pb accumulation compared with Pb treatment alone. The 0.1% and 0.5% PLA + Pb treatments reduced Pb accumulation in the aboveground part by 37% and 51%, respectively and in the belowground part by 28% and 39%, respectively. A similar trend was observed in the PBS + Pb treatments, with both concentrations (0.1% and 0.5%) reducing Pb accumulation by approximately 25% in both plant parts.

3.4. Effect of Different Treatments on the Rhizosphere Soil Bacterial and Fungal Communities

High-throughput sequencing of soil microbial communities in the rhizosphere of water spinach revealed a total of 42 bacterial phyla and 1090 genera. As shown in Figure 3a, the dominant bacterial phylum was Actinobacteriota (23.06–30.87%), followed by Proteobacteria (17.68–30.82%), Acidobacteria (7.73–20.35%), Chloroflexi (10.41–19.59%), and Firmicutes (2.90–5.89%). The relative abundance of Actinobacteriota increased in the Cd-treated group, Pb-treated group, 0.1% PLA group, 0.1% PLA + Pb group, 0.1% PBS + Cd group, 0.1% PBS + Pb group, and 0.5% PBS + Pb group compared to the CK group. At the genus level, dominant taxa included RB41 (1.68–5.31%), Arthrobacter (1.09–3.30%), Ramlibacter (0.8–5.36%), and Bacillus (1.67–4.45%) in the Cd-treated group. Compared to the CK group, the relative abundance of RB41, Arthrobacter, and Ramlibacter decreased by 22%, 13%, and 16%, respectively, while Bacillus increased by 22% (Figure 3b). In the Pb treatment group, RB41, Ramlibacter, and Bacillus increased by 6%, 29%, and 36%, respectively, whereas Arthrobacter decreased by 7%. The addition of PLA and PBS at varying concentrations enhanced the abundance of both Arthrobacter and Ramlibacter relative to the CK group.
Rhizosphere fungal communities were classified into 16 phyla and 676 genera (Figure 3c). The dominant fungal phyla were Ascomycota (36.6–70.50%), Basidiomycota (8.11–49.30%), Mortierellomycota (4.44–11.58%), and Glomeromycota (0.35–4.57%). Compared with the CK group, the abundance of Ascomycota, Mortierellomycota, and Glomeromycota decreased under Cd stress, while Basidiomycota increased. Under Pb stress, Ascomycota and Glomeromycota decreased, whereas Basidiomycota and Mortierellomycota increased. MPs exposure resulted in an 11–48% reduction in Ascomycota and a corresponding increase in Basidiomycota. Combined MPs and heavy metal contamination led to further decreases in Ascomycota and Glomeromycota abundance. As illustrated in Figure 3d, predominant fungal genera included Condenascus (3.57–29.95%), Mortierella (4.14–11.42%), and Peziza (0–21.31%). Heavy metals, MPs, and combined treatments reduced Condenascus abundance. PLA and PBS treatments caused a 21–67% decrease in Condenascus and an increase in Mortierella compared to the single Cd group. Combined MPs and heavy metal contamination reduced Condenascus by 27–63% and Mortierella by 19–35% compared to Pb treatment alone.
Compared with the CK group, the bacterial ACE, Chao, and Shannon indices increased in the Cd-contaminated group (Figure 4a). Similarly, ACE and Chao indices increased in the Pb group. In the 0.5% PLA group, all three indices (ACE, Chao, Shannon) declined, while the Simpson index increased. In the composite contamination groups, the ACE, Chao, and Shannon indices of the 0.1% PBS + Cd, 0.1% PBS + Pb, 0.5% PBS + Cd, and 0.5% PBS + Pb groups all exhibited increasing trends.
For fungi, Cd and Pb stress caused declines in the ACE, Chao, and Sobs indices, while the coverage index increased relative to CK. As shown in Figure 4b, the 0.5% PLA and 0.5% PBS groups exhibited decreased ACE and Chao indices, with the 0.5% PLA group showing a significant increase in the coverage index and the 0.5% PBS group showing a decrease in the Shannon index. In the composite contamination groups (e.g., 0.1% PLA + Cd, 0.1% PLA + Pb, 0.5% PLA + Cd, 0.5% PLA + Pb), ACE and Chao indices were lower than CK, while the coverage index increased. Similar patterns were observed in the 1% and 0.5% PBS + Cd groups.

3.5. RDA

RDA demonstrated that the first two axes contributed 18.51% of the community variation, with RDA1 explaining 14.75% and RDA2 3.76% (Figure 5a). The results of the knot analysis demonstrated a statistically significant correlation between the bacterial community and the following parameters: DTPA-Cd (R2 = 0.3136, p = 0.001), pH (R2 = 0.3493, p = 0.001), AK (R2 = 0.6202, p = 0.001), TP (R2 = 0.2516, p = 0.001), and TK (R2 = 0.1836, p = 0.001). Additionally, a significant negative correlation was observed between pH and AK, TP, and TK. AK, TP, and TK exhibited a similar directional relationship with the arrows of both axes, suggesting a potential synergistic effect among these elements. The Actinomycetes were found to be positively correlated with both pH and DTPA-Cd. Proteobacteria were demonstrated to be positively correlated with DTPA-Cd. Chloroflexi was negatively correlated with TK and TP.
RDA of the five most abundant fungal phyla indicated that the first two axes explained 25.19% of the variation in community composition (Figure 5b). The fungal community was significantly correlated with DTPA-Pb (R2 = 0.2401, p = 0.003), TK (R2 = 0.1620, p = 0.015), and pH (R2 = 0.1879, p = 0.010). There was a positive correlation between soil total potassium and total phosphorus content and a negative correlation between pH and DTPA-Pb and TK.

3.6. Effect of Combined MPs and Heavy Metal Contamination Treatment on the Prediction of Bacterial Function

To further investigate bacterial functional potential, PICRUSt2 was employed to predict functional differences based on COG categories among treatment groups subjected to different types of combined MPs and heavy metal contamination. The analysis identified 23 functional groupings, categorized into five major classes: organismal systems, metabolism, cellular processes, genetic information processing, environmental information processing, and human diseases. Notable functional differences were mainly observed in the domains of energy production and conversion, amino acid transport and metabolism, carbohydrate transport and metabolism, transcription, signal transduction mechanisms, cell wall/membrane/envelope biogenesis, and replication, recombination, and repair. Among these, amino acid transport and metabolism, carbohydrate transport and metabolism, cell wall/membrane/envelope biogenesis, and replication, recombination, and repair were significantly enriched in the groups exposed to compound contamination.
As illustrated in Figure 6a, the 15 treatment groups were clustered into two primary branches: one comprising the 0.5% PLA and 0.5% PBS groups, and the other comprising all remaining treatments. Compared with the CK group, gene expression was altered in the Cd- and Pb-treated groups. High concentrations of MPs (0.5% MPs) also elicited active expression across various functional pathways, whereas lower concentrations (0.1% MPs) induced comparatively weaker gene expression. Notably, the 0.1% PLA group displayed elevated cellular activity and ecological risk, while the 0.5% PLA + Cd group showed enhanced expression in cellular translation, sugar biosynthesis, and metabolism. In contrast, most other treatment groups demonstrated relatively low levels of functional gene expression. Both the 0.5% PLA and 0.5% PBS groups exhibited markedly higher functional activity compared to the other groups.
As shown in Figure 6b, the expression levels of genes in the 0.5% PBS, 0.1% PBS + Cd, 0.5% PBS + Cd, and 0.5% PBS + Pb groups were higher than those in the PLA-treated groups. Comparative analysis of pathway enrichment revealed significant upregulation of genes associated with quorum sensing, ABC transporters, two-component systems, and secondary metabolite biosynthesis in the 0.5% PLA and 0.5% PBS groups. Subsequent KEGG analysis of gene function differences among treatment groups further highlighted the 0.5% PLA and 0.5% PBS treatments as having the most pronounced functional shifts. These groups showed increases of 15% and 19% in genes associated with cellular processes and 17% and 19% in genes related to environmental information processing, respectively, compared with the CK group. Of particular concern were the elevated expression levels of genes linked to human diseases: 19% and 25% increases in disease-related genes, 34% and 46% increases in oncogenic pathways, 42% and 48% increases in cardiovascular-related functions, and 42% and 53% increases in genes associated with viral infections, respectively.

3.7. Mantel Test Analysis

To examine the relationships between key bacterial and fungal taxa, plant growth parameters, Cd content and accumulation, and soil physicochemical properties, a Mantel test was conducted (Figure 7). The results showed that bacterial communities were significantly negatively correlated with the dry weight of belowground plant parts and pH (p < 0.05), and positively correlated with AK (p < 0.01). In contrast, fungal communities exhibited a significant positive correlation (p < 0.01) with the length of belowground parts, as well as with AK, TP, and TK.

4. Discussion

4.1. Combined Microplastic and Heavy Metal Contamination Can Exert Greater Phytotoxic Effects on Water Spinach Growth Compared to Single Pollution

Plant biomass, which reflects both plant growth and heavy metal accumulation, is a key indicator in phytoremediation research [23]. The presence of Cd and Pb in the growth medium resulted in significant reductions in the length and biomass of water spinach. Specifically, Cd exposure led to marked decreases in plant growth. Additionally, the dry weight declined by 40.95% and 35.96%, which is consistent with previous findings [24]. Under Pb stress, the lengths and dry weights of the aboveground and belowground parts decreased by 16.63–34.21%. A comparison between the two single-metal treatments revealed that Pb exerted a less detrimental effect on water spinach biomass than Cd. Heavy metals can be absorbed by plant roots, leading to growth inhibition. Previous studies have shown that both Pb and Cd can suppress root elongation and interfere with the uptake of water and nutrients [25]. Moreover, the accumulation of heavy metals can disrupt essential physiological processes, including photosynthesis, respiration, and nutrient transport [26]. These disruptions often manifest as physiological dysfunctions, such as stunted growth, chlorosis, and reduced flowering and fruiting rates [27,28].
Furthermore, MPs were also found to inhibit water spinach growth. Both PLA and PBS MPs showed increasing toxicity to the biomass of aboveground and belowground parts with increasing concentrations, which is consistent with the findings of Zhao et al. [29]. It has been demonstrated that MPs exhibit enhanced toxicity when combined with heavy metals, compared to either pollutant alone. For instance, treatments with 0.5% PLA + Cd and 0.5% PBS + Cd reduced the dry weight of the aboveground portion of water spinach by 32% and 47%, respectively. Similarly, 0.5% PLA + Pb and 0.5% PBS + Pb reduced dry weight by 27% and 31%, respectively, confirming the synergistic toxicity of combined contamination, as supported by other studies [10,18,25].
It is evident that MPs can act as carriers of specific elements, thereby influencing the transport and transformation of heavy metals in contaminated soils. This influence is mediated through a variety of physical, chemical, and biological mechanisms, which in turn can affect the uptake of heavy metals by plants [30]. Research has demonstrated that the coexistence of MPs and heavy metals can alter the accumulation of metals in plant tissues. Moreover, the effect of MPs on metal uptake varies depending on both the type and concentration of the MPs involved [31]. MPs have been shown to significantly affect the bioavailability and accumulation of heavy metals in plants via several mechanisms. The surfaces of MPs contain functional groups such as hydroxyl and carboxyl groups, which can adsorb heavy metals through electrostatic interactions, complexation, and ion exchange. These processes reduce the free ionic forms of metals or lead to co-precipitation, thus promoting the immobilization of heavy metals in soils [13]. MPs can also indirectly influence metal uptake by altering the structure of the microbial community. For example, MPs may become colonized by heavy-metal-resistant or phosphate-solubilizing bacteria, which contribute to metal immobilization through the secretion of chelating agents or the formation of biofilms [32,33]. In the present study, the addition of PBS significantly reduced the content and accumulation of Cd in both the aboveground and belowground parts of water spinach compared to the Cd-only treatment. Similarly, the combination of PBS and Pb yielded comparable effects. These findings suggest that the impact of MPs on heavy-metal-contaminated soil–plant systems depends on the characteristics of the MPs, the specific heavy metals involved, and the prevailing environmental conditions [34,35]. Conversely, PLA addition led to increased Cd and Pb concentrations in water spinach. PLA was also observed to exhibit adsorption properties for both organic matter and heavy metals. Due to its large specific surface area and hydrophobicity, PLA may facilitate the entry of Cd and Pb into plant tissues by damaging the root epidermis. Additionally, MPs–heavy metal complexes may enhance metal accumulation in plants in a less predictable manner [36]. Furthermore, increasing MPs concentrations were associated with elevated levels of heavy metals in water spinach samples. The coexistence of MPs and heavy metals has been increasingly recognized as a significant factor affecting plant health. The physicochemical properties of MPs play a crucial role in determining the extent of their impact on plant growth and metal uptake [37]. MPs can also alter soil porosity and permeability, while their adsorption capacity facilitates the movement of heavy metals through the soil, ultimately imposing stress on plant development and quality.
Upon entering the soil, MPs interact with surrounding materials and become incorporated into soil aggregates. This process alters the physical properties of the soil, including changes in bulk density and water-holding capacity, which in turn affect water and nutrient uptake by plant roots, ultimately inhibiting plant growth [38]. In the present study, both PLA and PBS significantly reduced the contents of AK and AP in the soil, with more pronounced effects observed under combined MPs–heavy metal treatments. This reduction in nutrient availability was associated with decreased biomass of water spinach.

4.2. Combined Microplastic and Heavy Metal Contamination Has a More Pronounced Effect on the Composition and Structure of the Water Spinach Rhizosphere Microbial Community Compared to Single Pollution

Heavy metal contamination also impacts the activity, biomass, community structure, and ecological function of soil microorganisms. Due to their sensitivity to environmental changes, soil microbes are widely used as indicators of soil health and quality [39]. Our results showed that different contamination treatments led to significant changes in bacterial community composition. Consistent with previous studies, Actinobacteriota and Acidobacteria, were the predominant phyla in soils contaminated with heavy metals [40]. These groups exhibited strong adaptability and resistance to stressors such as Cd and Pb. Actinobacteriota exhibited the highest relative abundance, underscoring its key role in mediating the effects of MPs–heavy metal co-contamination [41]. This phylum is known for its tolerance to heavy metals and its functional contribution to bioremediation [42]. The PLA + Cd treatment increased the relative abundance of Actinobacteriota, a dominant taxon in polluted acidic soils. This shift may be attributed to the acidic microenvironment generated by PLA degradation and Cd presence, which supports the proliferation of acidophilic microbes and enhances the decomposition of complex organic matter, such as humus and cellulose [43]. At the genus level, Arthrobacter, Bacillus, Sphingomonas, and RB41 were dominant. Bacillus, in particular, is known for its heavy metal resistance and its ability to form biofilms on MPs, thereby playing a role in MPs degradation [44]. Notably, elevated MPs combined with heavy metals significantly increased the abundance of Ramlibacter spp., a metabolically diverse taxon capable of utilizing a wide range of organic and inorganic compounds. Ramlibacter contributes to ecological remediation through heavy metal adsorption, transformation, and immobilization, and its high reproductive capacity supports rapid colonization [45].
The combined contamination of MPs and heavy metals exerted a more significant effect on the dominant fungal phyla and genera than on bacteria, as evidenced by substantial alterations in their relative abundance before and after treatment. The relative abundance of the Ascomycota and Coccidioides significantly decreased in response to combined MPs and heavy metal contamination, which is consistent with previous findings [46]. These fungi are known for their tolerance to elevated heavy metal concentrations; however, fluctuations in soil quality can negatively impact their abundance.
The impact of PLA and PBS, as biodegradable plastics, on rhizosphere microbial communities is multifaceted. Due to their biodegradability, both plastics decompose in soil, releasing carbon sources that promote microbial proliferation [47]. However, differences in degradation products, rates, and carbon source availability between PLA and PBS may affect the structure, function, and diversity of rhizosphere microbial communities differently [48]. PLA is composed of lactic acid monomers, which polymerize and release lactic acid during degradation. This process has been observed to promote the growth of lactic acid-utilizing organisms, thereby altering the microbial community structure in the rhizosphere [48,49]. The present study compares the effects of various PLA and PBS doses on soil bacteria involved in the degradation of these plastics. The results show that in soil samples contaminated with both materials, the abundance of Actinobacteriota and Bacillus was higher in PLA-contaminated soils than in PBS-contaminated soils. These findings are consistent with those of Zhang et al. [50], who reported higher levels of Actinobacteriota and Bacillus in PLA-contaminated soils compared to those with PBS. PBS, synthesized from butanedioic acid and butanediol, degrades into products that may also foster the proliferation of specific microorganisms. As biodegradable MPs, PLA and PBS share similar effects on microbial community structure during degradation. Degradation of PLA results in the production of lactic acid, which lowers soil pH. Lactic acid has been shown to inhibit acid-sensitive microorganisms while promoting the growth of lactic acid-utilizing organisms. Concurrently, it induces competitive inhibition of other microbial species [51,52,53]. Consequently, PLA degradation leads to a reduction in soil microbial diversity. In this study, the effect of PLA on soil microbial diversity was found to be more pronounced than that of PBS. This increase in microbial diversity was primarily due to PLA’s rapid release of carbon sources, creating ecological niches that support a broader range of functional microorganisms [54].
The Mantel test analysis revealed a significant correlation between the physicochemical properties of soil and the dominant bacterial and fungal populations. Dominant groups, including Actinomycetes, Ascomycetes, and Bacilli, were observed to be recruited in response to the stress caused by combined MPs and heavy metal contamination. These tolerant microorganisms secreted metabolites that altered the soil environment [55]. It has been established that the release of organic acids by Actinomycetes and other organisms reduces the mobility of heavy metals in soil. This process has been shown to decrease soil pH by 0.3–0.8 units, facilitating the formation of complexes between heavy metals and organic acids [56]. Furthermore, other organisms have been shown to secrete extracellular polymers that enhance soil aggregate stability. Changes in enzyme activities have been observed to accelerate the decomposition of organic matter, which directly influences soil chemical indicators [57]. The combined stress of MPs and heavy metal pollution has been shown to promote the recruitment of beneficial microorganisms, with metal-tolerant and MPs -degrading bacteria and fungi playing roles in the breakdown and transformation of these pollutants, ultimately influencing the soil microbial structure and ecological stability.

4.3. Combined Microplastic and Heavy Metal Contamination Has a More Pronounced Effect on the Functional Composition of the Water Spinach Rhizosphere Microbial Community Compared to Single Pollution

MPs–heavy metal composite contamination has been demonstrated to drive microbial population reconstruction and loss of diversity through multiple pathways, including physical adsorption, chemical complexation, and biological metabolism. This process forms a nonlinear feedback loop with certain soil chemical indicators, representing ecological risk and offering a target for microbial regulation-based soil remediation.
To investigate the effects on functional genes within the microbial community, community prediction analyses were conducted using PICRUSt2 software, which accurately predicts the function and abundance of functional genes. Linking microbial variability to biological function through comparison with the KEGG database supports the study of bacterial ecological functions and their responses to environmental change [58]. Studies have shown that Cd inhibits methylase activity, causing DNA damage and increasing the expression of DNA repair proteins and protein synthesis. To remove metal ions from cells and reduce toxicity, the expression of transport proteins is upregulated [59]. Studies found that MPs can affect bacterial transport, transfer antibiotic resistance genes, and influence microbial metabolism [60,61]. Functional prediction of different types of composite pollution treatment groups in this study revealed that the main functional genes involved included energy production and conversion, amino acid transport and metabolism, carbohydrate transport and metabolism, transcription, signal transduction mechanisms, and cell wall/membrane/envelope biogenesis. These findings indicate functional richness. Metabolism, particularly through carbohydrates, energy, and amino acids, was the most dominant bacterial function, consistent with previous studies [62]. The present study showed that soil metabolic functions were enhanced by high concentrations of MPs pollution, particularly in terms of ABC transporters, two-component systems, and metabolite production, which were significantly increased. This may be related to the recruitment of beneficial soil microorganisms. The presence of MPs alters the structure of surrounding soil aggregates and has a negative effect. MPs -degrading microorganisms and other microorganisms receive signals that enhance their metabolic activity, facilitating MPs degradation in response to stress [63]. At the same time, studies have shown that MPs pose potential risks to human health, damaging the cardiovascular system and intestinal barrier and upregulating functional genes associated with human diseases due to high levels of MPs contamination [64].

5. Conclusions

Previous studies have extensively examined the individual effects of MPs and heavy metals on soil–plant ecosystems; however, few have explored the combined impacts of degradable MPs (PLA and PBS) with Cd and Pb on plant growth and soil microbial communities. The results of this study demonstrate that both heavy metals and MPs alone reduce the dry weight of water spinach, with their combined contamination exhibiting greater phytotoxicity. The coexistence of MPs and heavy metals also reduced the availability of soil potassium and phosphorus, further hindering plant growth. Unlike previous research, this study conducted a comprehensive comparison of how soil bacterial and fungal community composition and functions respond to both single and combined pollution. It was found that MPs–heavy metal interactions significantly altered microbial community structure, diversity, and function, promoting the enrichment of heavy-metal-tolerant and potential plastic-degrading microorganisms. PLA and PBS had differential impacts on microbial community composition and function, likely due to their distinct physicochemical properties. These findings provide theoretical insights and experimental references for understanding the ecotoxicological risks of MPs–heavy metal co-contamination and for developing future soil remediation strategies.

Author Contributions

Z.-J.C. and P.-F.D. designed the experiments; J.-Y.W., Z.-J.C., P.-F.D. and B.L.L. participated in writing the paper; J.-Y.W., M.W., J.-W.S. and L.L. performed the experiments and analyzed the data; J.-Y.W., Z.-J.C., P.-F.D., L.L. and B.L.L. reviewed the manuscript. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the National Natural Science Foundation of China (Grant No. U2004145), the Program for Science & Technology Innovation Talents in Universities of Henan Province (No. 23HASTIT018), the Key Research and Development Projects of Henan Province (Grant Nos. 221111520600, and 231111113000), the Key Scientific and Technological Project of Henan Province (Grant Nos. 242102521067, and 232102320252), the Research Project of Colleges and Universities of Henan Province Education Department (Grant No. 24B210008), the National Natural Science Foundation of China Cultivation Project of Nanyang Normal University (Grant No. 2025PY038) and the Youth Program of Nanyang Normal University (Grant No. 2024QN029).

Data Availability Statement

The data that support the findings of this study are available on request from the corresponding authors.

Conflicts of Interest

The authors declare no conflicts of interest.

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Figure 1. Length (a) and dry weight (b) of water spinach under different treatment conditions. Different lowercase letters indicate significant differences between treatments within the same row (p < 0.05).
Figure 1. Length (a) and dry weight (b) of water spinach under different treatment conditions. Different lowercase letters indicate significant differences between treatments within the same row (p < 0.05).
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Figure 2. Cd content (a), Cd accumulation (b), Pb content (c), and Pb accumulation (d) in water spinach under different treatment groups. Different lowercase letters indicate significant differences between treatments within the same row (p < 0.05).
Figure 2. Cd content (a), Cd accumulation (b), Pb content (c), and Pb accumulation (d) in water spinach under different treatment groups. Different lowercase letters indicate significant differences between treatments within the same row (p < 0.05).
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Figure 3. Relative abundance of microbial communities in different treatment groups at the phylum and genus levels. (a) Bacterial phylum level; (b) bacterial genus level; (c) fungal phylum level; (d) fungal genus level.
Figure 3. Relative abundance of microbial communities in different treatment groups at the phylum and genus levels. (a) Bacterial phylum level; (b) bacterial genus level; (c) fungal phylum level; (d) fungal genus level.
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Figure 4. Diversity indices of microbial communities. (a) Comparison of ACE, Chao, and coverage indices of bacterial community; (b) comparison of ACE, Shannon, and Simpson indices of fungal community. * (p < 0.05), ** (p < 0.01) indicate significant differences between groups.
Figure 4. Diversity indices of microbial communities. (a) Comparison of ACE, Chao, and coverage indices of bacterial community; (b) comparison of ACE, Shannon, and Simpson indices of fungal community. * (p < 0.05), ** (p < 0.01) indicate significant differences between groups.
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Figure 5. Redundancy analysis (RDA) of bacterial (a) and fungal (b) communities and soil chemical characteristics. Blue arrows represent dominant phyla; red arrows represent environmental variables.
Figure 5. Redundancy analysis (RDA) of bacterial (a) and fungal (b) communities and soil chemical characteristics. Blue arrows represent dominant phyla; red arrows represent environmental variables.
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Figure 6. Predicted bacterial functions. (a) Heatmap of bacterial functions in different treatment groups. The x-axis represents treatment groups, and the y-axis represents functional genes. The color blocks on the left indicate the associated functions. (b) Bar chart showing the relative abundance of predicted bacterial functions.
Figure 6. Predicted bacterial functions. (a) Heatmap of bacterial functions in different treatment groups. The x-axis represents treatment groups, and the y-axis represents functional genes. The color blocks on the left indicate the associated functions. (b) Bar chart showing the relative abundance of predicted bacterial functions.
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Figure 7. Relationships among water spinach growth indices, soil chemical properties, bacterial communities, and fungal communities. Growth indices of water spinach are displayed in a butterfly plot on the left, while soil physicochemical properties and microbial community characteristics are shown on the right.
Figure 7. Relationships among water spinach growth indices, soil chemical properties, bacterial communities, and fungal communities. Growth indices of water spinach are displayed in a butterfly plot on the left, while soil physicochemical properties and microbial community characteristics are shown on the right.
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Table 1. Design of experiments.
Table 1. Design of experiments.
Experimental TreatmentMPS MPs Concentrations/% (w/w)Cd Concentrations/mg·kg−1Pb Concentrations/mg·kg−1
CK 0000
Cd00100
Pb000200
0.1%PLAPLA0.1%00
0.5%PLAPLA0.5%00
0.1%PBSPBS0.1%00
0.5%PBSPBS0.5%00
0.1%PLA + CdPLA0.1%100
0.1%PLA + PbPLA0.1%0200
0.5%PLA + CdPLA0.5%100
0.5%PLA + PbPLA0.5%0200
0.1%PBS + CdPBS0.1%100
0.1%PBS + PbPBS0.1%0200
0.5%PBS + CdPBS0.5%100
0.5%PBS + PbPBS0.5%0200
Table 2. Physicochemical properties of the soil under different treatment conditions.
Table 2. Physicochemical properties of the soil under different treatment conditions.
Experimental TreatmentpHAK/(mg⋅kg−1)AP/(mg⋅kg−1)TK/(mg⋅kg−1)TP/(mg⋅kg−1)DTPA-Cd (mg⋅kg−1)DTPA-Pb (mg⋅kg−1)
CK7.13 ± 0.00 cdef78.23 ± 0.28 ab17.82 ± 0.15 a7.99 ± 0.22 b140.22 ± 12.30 a00
Cd7.39 ± 0.0 1 ab62.27 ± 0.76 g12.83 ± 0.72 cdef6.73 ± 0.12 cd97.13 ± 8.69 e2.26 + 0 ab0
Pb7.09 ± 0.00 defg63.33 ± 0.24 fg11.33 ± 0.25 fg6.35 ± 0.24 d99.61 ± 4.82 e012.33 + 0.20 c
0.1%PLA6.84 ± 0.01 b76.87 ± 0.15 b14.26 ± 0.59 bcd7.57 ± 0.16 b136.33 ± 1.91 ab00
0.5%PLA6.85 ± 0.01 b77.6 ± 0.69 ab15.36 ± 0.65 bc7.23 ± 0.06 bc133.58 ± 5.73 abc00
0.1%PBS7.14 ± 0.02 cde76.53 ± 0.35 b16.00 ± 0.24 b7.65 ± 0.01 b131.38 ± 6.26 abcd00
0.5%PBS6.87 ± 0.01 gh79.80 ± 1.79 a12.75 ± 0.06 cdef7.58 ± 0.07 b133.64 ± 8.36 abc00
0.1%PLA + Cd7.50 ± 0.00 a66.23 ± 0.82 e10.53 ± 0.21 fg6.71 ± 0.08 cd114.88 ± 10.98 cde2.06 + 0.001 ab0
0.1%PLA + Pb7.23 ± 0.02 bcd69.10 ± 0.98 d12.2 ± 0.25 defg5.09 ± 0.02 f98.88 ± 2.90 e017.75 + 0.010 a
0.5%PLA + Cd7.09 ± 0.01 defg64.73 ± 0.58 ef12.67 ± 0.04 cdef6.31 ± 0.05 d107.42 ± 11.71 e2.43 + 0.003 a0
0.5%PLA + Pb6.90 ± 0.01 fgh68.97 ± 1.61 d11.80 ± 0.02 efg5.51 ± 0.05 ef116.69 ± 11.93 bcde018.18 + 0.32 a
0.1%PBS + Cd7.0 ± 0.01 defgh64.33 ± 0.98 efg14.1 ± 0.35 bcde6.07 ± 0.11 de111.43 ± 9.39 de1.97 + 0.012 b0
0.1%PBS + Pb7.0 ± 0.00 defgh65.70 ± 0.49 e13.7 ± 0.35 bcde5.09 ± 0.08 f110.10 ± 13.80 e017.34 + 0.01 a
0.5%PBS + Cd6.94 ± 0.00 efgh65.23 ± 1.95 ef12.83 ± 0.18 cdef6.25 ± 0.08 de115.12 ± 7.46 bcde2.27 + 0.032 ab0
0.5%PBS + Pb7.16 ± 0.00 bcde73.73 ± 0.63 c12.51 ± 0.00 def6.31 ± 0 d103.44 ± 14.36 e013.04 + 1.27 b
Different lowercase letters indicate significant differences between treatments within the same column (p < 0.05).
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Wang, J.-Y.; Wang, M.; Shi, J.-W.; Li, B.L.; Liu, L.; Duan, P.-F.; Chen, Z.-J. The Effects of Microplastics and Heavy Metals Individually and in Combination on the Growth of Water Spinach (Ipomoea aquatic) and Rhizosphere Microorganisms. Agronomy 2025, 15, 1319. https://doi.org/10.3390/agronomy15061319

AMA Style

Wang J-Y, Wang M, Shi J-W, Li BL, Liu L, Duan P-F, Chen Z-J. The Effects of Microplastics and Heavy Metals Individually and in Combination on the Growth of Water Spinach (Ipomoea aquatic) and Rhizosphere Microorganisms. Agronomy. 2025; 15(6):1319. https://doi.org/10.3390/agronomy15061319

Chicago/Turabian Style

Wang, Jing-Yi, Meng Wang, Jian-Wei Shi, B. Larry Li, Ling Liu, Peng-Fei Duan, and Zhao-Jin Chen. 2025. "The Effects of Microplastics and Heavy Metals Individually and in Combination on the Growth of Water Spinach (Ipomoea aquatic) and Rhizosphere Microorganisms" Agronomy 15, no. 6: 1319. https://doi.org/10.3390/agronomy15061319

APA Style

Wang, J.-Y., Wang, M., Shi, J.-W., Li, B. L., Liu, L., Duan, P.-F., & Chen, Z.-J. (2025). The Effects of Microplastics and Heavy Metals Individually and in Combination on the Growth of Water Spinach (Ipomoea aquatic) and Rhizosphere Microorganisms. Agronomy, 15(6), 1319. https://doi.org/10.3390/agronomy15061319

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