Earthworm gut microbiome promotes biodegradation of albendazole in soil

TL;DR

Earthworms' gut microbes help break down the drug albendazole in soil, with specific bacteria and genes playing key roles in the process.

Contribution

Identified specific earthworm gut microbes and genes involved in albendazole biodegradation, and the role of zinc oxide nanoparticles in reducing drug accumulation.

Findings

Albendazole enrichment increased specific degradation genes (hmr, ami) and activated sulfur reduction and hydroxylation pathways.

Bacteria like Sphaerobacter and Nocardia were identified as potential hosts for biodegradation genes in earthworm guts.

Zinc oxide nanoparticles reduced albendazole bioaccumulation in earthworms and accelerated its dissipation in soil.

Abstract

The excretion of the anthelmintic drug albendazole (ALB) from treated animals into the soil, as well as its widespread application as a fungicide, poses a serious ecological risk to the soil environment. In this study, we investigated the degradation of ALB in soil and its bioaccumulation in earthworms, changes in the microbiome and degradation genes, and the effect of zinc oxide nanoparticles on the degradation and enrichment behaviors of ALB and microbial community structure and function. Our findings showed that ALB selectively enriched specific albendazole degradation genes (i.e., hmr and ami) in the earthworm, preferentially activating the pathways associated with sulfur reduction, amination of ALB sulfone, and hydroxylation of ALB. Metagenomic analysis revealed that the relative abundances of ppo, xylA, cutC, and nfsl in the earthworm gut were 0.19–52.64-fold higher in the ALB treatment than in the control, indicating their potential dominance in ALB biodegradation. Network analysis further identified potential bacterial hosts carrying biodegradation genes and albendazole degradation genes. Notably, Sphaerobacter, Saccharothrix, Actinomadura, and Nocardia were recognized as potential dual hosts of biodegradation genes and albendazole degradation genes, displaying a 0.05–1.32-fold elevation in relative abundance in ALB-treated earthworm guts compared to the control. Additionally, ZnO nanoparticles were found to reduce ALB bioaccumulation in earthworms and accelerate its dissipation in soil. These findings provide novel insights into the bioremediation mechanisms of pesticides in soil-earthworm ecosystems.

The online version contains supplementary material available at 10.1007/s44297-026-00068-5.

Figures

Bioaccumulation concentration a and bioaccumulation factor b of albendazole in the earthworm gut

Biodegradation gene profiles in the earthworm gut. a Relative abundance of BDGs and ADGs in different treatments. b Heatmap of the BDGs and ADGs based on common logarithmic transformed abundance. Abbreviations: BDGs, biodegradation genes; ADGs, albendazole degradation genes

Changes in microbial metabolic functions in the earthworm gut. a Relative abundance of microbial metabolic pathways at level 1 among different treatments. b Heatmap of the microbial metabolic pathways at level 2 and the corresponding relative abundance based on common logarithmic transformed abundance among different treatments

Links between BDGs and MGEs in the earthworm gut. a Total relative abundance of MGEs among different treatments. b Linear regression analysis between the relative abundance of BDGs and MGEs. c Proportions of BCCs with MGEs among different treatments. d The mapping relationship between BDGs and MGEs in the E-ALB samples. The bars on the inner ring represents the co-occurrence of BDGs and MGEs on BCCs. Abbreviations: BDGs, biodegradation genes. MGEs, mediated by mobile genetic elements. BDGs, biodegradation genes. BCCs, BDG-carrying contigs. E-ALB, earthworm samples treated with albendazole. E-Z

Shifts in the microbiome, BDG and ADG hosts in the earthworm gut. a The structural composition of the soil microbial community among the different treatments. The co-occurrence network of BDGs b and ADGs c with genera (top 100) in the earthworm gut. Data in all samples with significant correlation (|ρ|> 0.7) and significance (*p* < 0.05) were screened based on Spearman analysis. The red line represents negative correlation, and the green line represents positive correlation. d Visualization based on the relative abundance and taxonomic assignment of the recovered MAGs carrying BDGs. e Relative

Full text

Introduction

Albendazole (ALB), a benzimidazole derivative and broad-spectrum anthelmintic, is widely used to treat parasitic infections in both humans and livestock [1, 2]. Recently, it has also been employed as a fungicide to prevent crop spoilage due to its ability to inhibit pathogen growth and spore germination [3]. However, ALB exhibits bio-toxicity, with studies reporting adverse effects on the nervous, hepatic, immune, reproductive, and hematopoietic systems following exposure [4–6]. ALB has been frequently detected in aquatic environments worldwide, including river water, seawater, domestic wastewater, and hospital wastewater treatment plants [7]. Although typically present at low concentrations (ng/L), residual ALB has even been found in residential drinking water, raising concerns about potential risks to non-target organisms [8]. Due to its high octanol–water partition coefficient (K_OW_), high organic carbon partition coefficient (K_OC_), and low water solubility, ALB tends to accumulate in soil and sediments [9, 10]. This persistence is particularly relevant in tropical and subtropical regions, where soil-borne helminthic infections are more prevalent and ALB is extensively used, leading to long-term environmental accumulation [11].

Microbial degradation is the primary mechanism driving the breakdown of persistent pollutants in soil, which is a process largely regulated by biodegradation genes (BDGs) [12]. Earthworms, a key component of soil ecosystems, are in direct contact with soil contaminants through their feeding activities, making their gut microbiomes particularly susceptible to pollutant exposure [13]. Furthermore, horizontal gene transfer (HGT), mediated by mobile genetic elements (MGEs), plays a vital role in spreading BDGs, enhancing microbial adaptation to soil contaminants [14]. Although previous studies have shown that the gut microbiota contributes to the efficacy of ALB and ivermectin against soil-transmitted helminthiases [15, 16], the response of earthworm gut microbiomes and their BDG profiles to ALB exposure remains poorly understood.

Zinc oxide nanoparticles (ZnO NPs), which typically range in size from 1 to 100 nm and exhibit diverse structural configurations, have shown great promise in nanomedicine for disease prevention, diagnosis, and treatment [15]. Due to their notable bioactivity and therapeutic potential, ZnO NPs have been widely explored in drug delivery systems and antimicrobial therapies. Previous studies have demonstrated that ZnO NPs combined with ALB exhibit strong therapeutic effects against cystic echinococcosis in both in vitro and in vivo experiments [16, 17]. Additionally, ZnO NPs are recognized as efficient and eco-friendly materials for pesticide degradation, suggesting their potential use in mitigating pesticide pollution [18]. Given these properties, we hypothesize that ZnO NPs can enhance the degradation of ALB in soil.

In this study, shotgun metagenomics was utilized to comprehensively investigate shifts in the BDG profiles and their potential mobility within the soil-earthworm ecosystem under ALB and ZnO NPs exposure. The objectives of this study were (i) to evaluate the dissipation and bioaccumulation patterns of ALB in soil and earthworms; (ii) to examine ALB-induced alterations in the gut microbiome, BDGs, and ADGs of earthworms; (iii) to elucidate the key role of HGT in the selective enrichment of BDGs; and (iv) to identify key indigenous bacterial taxa associated with ALB degradation. Collectively, these findings provide new insights into the bioremediation of ALB-contaminated soils.

Materials and Methods

Chemical, soil sampling, and earthworm

ALB (purity > 99%) was obtained from Guizhou Dao Yuan Biotechnology Co., Ltd. ZnO NPs (purity > 99%, particle size 50 nm) were purchased from Sangon Biotech (Shanghai) Co., Ltd. Soil samples (0–15 cm depth) were collected via a standardized five-point method from farmland in Huaian, Jiangsu Province, which had no prior history of ALB application. Key physicochemical properties are detailed in Table S1. The collected soil samples were air-dried, sieved through a 2 mm mesh, and then incubated at room temperature (75% relative humidity) under a 12:12-h dark/light photoperiod for 14 d. The earthworms (Eisenia fetida) used in the experiment were cultured under laboratory conditions for over one month. Before the pot experiment, sexually mature individuals were transferred to glass beakers lined with two layers of sterile filter paper saturated with deionized water. Earthworms were deprived of food for 24 h in darkness to evacuate gut contents.

Pot experiment and sample collection

ALB and ZnO NPs were spiked into soil at concentrations of 3 mg/kg and 100 mg/kg, respectively, based on the recommended dosage, environmental residue levels, and toxicity to E. fetida [9, 19]. Briefly, for the ALB treatment, a stock solution was prepared using acetone and sprayed onto 500 g of soil. After acetone evaporation, 1.5 kg of additional soil and sterile deionized water were incorporated into the mixture, which was then homogenized by passing it through a 2-mm sieve. For the ZnO treatment, ZnO NP powder was directly mixed into 500 g of soil supplemented with 1.5 kg of additional soil and sterile deionized water and homogenized identically through a 2-mm sieve. Subsequently, the prepared soils were transferred into plastic pots (height: 65 mm, upper diameter: 95 mm, bottom diameter: 70 mm). The control pots (CK) received acetone-only treatments equivalent in volume to the experimental groups. Each pot was inoculated with twenty E. fetida (300–400 mg) and sealed with perforated aluminum foil (1-mm pore size). All pots were incubated in an artificial climate chamber (20 ± 1 °C, 75% relative humidity, 12:12 h light/dark cycle) for 28 d. Soil moisture was maintained at 60% water-holding capacity by replenishing water loss every two days. All treatments were performed in triplicate. Soil and earthworm samples were collected at 0, 1, 3, 7, 14, 21, and 28 d for ALB residue analysis and total deoxyribonucleic acid (DNA) extraction.

Determination of ALB residues in soil and earthworms

The determination of ALB in soil and earthworms was conducted following a modified method based on a previous study [20]. Briefly, earthworm gut contents were emptied overnight before extraction. Approximately 5 g of soil or 3 earthworms were weighed, and water and acetonitrile were added at a 1:1 (v/w), followed by homogenization using a tissue crusher. The mixture was vortexed for 2 min, after which 1 g of NaCl and 2 g of Na_2_SO_4_ were added, followed by an additional 2 min of vortexing. The sample was then centrifuged at 8000 rpm for 5 min. For the cleanup phase, 50 mg of primary secondary amine (PSA) was added to 1 mL of the supernatant, vortexed for 2 min, and centrifuged at 5000 rpm for 5 min. Finally, the supernatant was filtered through a 0.22-µm organic filter before high-performance liquid chromatography (HPLC) analysis. Three concentrations (0.3, 1, and 3 mg/kg)) were set to validate the analytical method. The analysis of albendazole was performed by employing an Agilent XDB-C18 column (4.6 mm × 150 mm, 5 μm particle size) under the following conditions: mobile phase, water:acetonitrile (60:40, v/v); flow rate, 1.0 mL/min; detection wavelength, 292 nm; injection volume, 20 μL; and temperature, 30 °C.

DNA extraction and metagenomic sequencing

All earthworms were dissected after 28 d of incubation following the method described in a previous study [21]. Briefly, the earthworms were rinsed five times with sterile water and then placed on ice for 10 min. After immersion in 75% ethanol, they were washed five times with sterile water. Using sterile scissors, dissecting needles, and forceps, the body tissues were carefully dissected under aseptic conditions, and the gut posterior to the gizzard was collected. The gut samples were resuspended in 0.1 mol/L phosphate buffer and stored at −80 °C.

Total DNA from both earthworm and soil samples was extracted using the Fast DNA™ Spin Kit for soil (Qbiogene, CA, USA) following the manufacturer's instructions. Each treatment was performed in triplicate, and all DNA samples were stored at −80 °C before metagenomic sequencing.

Database construction

The BDG protein database and ALB Degradation Gene (ADG) protein database were constructed following the method described by Fang, Cai, Yang, Ju, Li, Yu and Zhang [22]. The BDG protein database was obtained from the FunGene protein database (http://fungene.cme.msu.edu/) and the NCBI database, which consists of 28 sub-databases (ppo, P450, xylA, benA, lip, cutC, carA, alkB, mnp, xenB, bphA1, mmoX, glx, diaA, xenA, bphA2, dbfA1, dxnA-dbfA1, ppah, npah, bph, cntA, BSH, PSA, HSDH, nfsI, etnC, dxnA). The ADG protein database included three sub-databases (ami, hmr, and mno; see Table S3). All databases were subjected to redundancy removal processing before analysis.

Metagenomic sequencing and identification of BDGs and MGEs

The metagenomic sequencing libraries were constructed using the extracted DNA and sequenced on Novogene's Illumina MiSeq platform (China). All raw reads were filtered using fastp at default settings [23]. For gut samples, host contamination was removed by aligning reads to the reference genomes of E. fetida (GenBank: GCA_003999395.1 and GCA_900000155.1). A summary of raw and clean data is presented in Table S2. BDGs and MGEs were identified using BLSATP (e-value ≤ 1e-5, identity ≥ 80%) based on previous studies [12, 24]. Clean reads were assembled into contigs (> 500 bp) using MEGAHIT software [25], and assembly quality was evaluated with QUAST [26]. A total of 170,779–489,801 contigs were generated across all the samples (Table S2). Open reading frames (ORFs) were predicted using prodigal [27]. BDG-carrying contigs (BCCs), ADG-carrying contigs (ADCs), and MGE-carrying contigs (MCCs) were identified via BLASTP (e-value ≤ 1e-5, identity ≥ 80%). The co-occurrence arrangements of BCCs, ADCs, and MCCs were selected if they were concurrently found on the same contigs. For BCCs, Plasflow V1.133 was used to locate them on plasmids or chromosomes [28].

The abundance of BDGs, ADGs, and MGEs was calculated by the following formula:

\documentclass[12pt]{minimal} \usepackage{amsmath} \usepackage{wasysym} \usepackage{amsfonts} \usepackage{amssymb} \usepackage{amsbsy} \usepackage{mathrsfs} \usepackage{upgreek} \setlength{\oddsidemargin}{-69pt} \begin{document}$$Abundance = \frac{{N}_{mapped itags}}{{N}_{total itags}}$$\end{document}

where N_mapped itags_ is the number of the itags mapped to target BDGs or MGEs, and N_total itags_ is the number of itags obtained by clean data.

Microbiome analysis and functional annotation

Based on the clean data, the soil microbiome was analyzed using Kraken2 and Bracken software to determine the relative abundance of the taxa in each sample [29, 30]. The clean data were annotated with KofamScan [31] (V 1.3.0), and Kyoto Encyclopedia of Genes and Genomes (KEGG) pathways were hierarchically annotated.

Metagenomic binning

The clean reads were mapped to the assembled contigs (> 1000 bp) using Bowtie2, and the metagenome-assembled genomes (MAGs) were reconstructed using MetaBAT2 (V2.15) [32], Maxbin2 V2.0 [33], and Concoct V1.1 [34]. Only MAGs with completeness > 50% and contamination < 10% that contained BDGs or ADGs were retained for downstream analysis. Taxonomy classification of the MAGs was performed using GTDB-Tk [35]. The abundance of contigs carrying MGEs and MAGs was calculated using the following formula:

\documentclass[12pt]{minimal} \usepackage{amsmath} \usepackage{wasysym} \usepackage{amsfonts} \usepackage{amssymb} \usepackage{amsbsy} \usepackage{mathrsfs} \usepackage{upgreek} \setlength{\oddsidemargin}{-69pt} \begin{document}$$Aubundance (coverage, \times /Gb) = {\sum }_{1}^{n}\frac{{N}_{mapped reads}\times {L}_{reads}/{L}_{ORF}}{S}$$\end{document}

where N_mapped reads_ is the number of reads mapped to target ORFs, L_reads_ is the sequence length of the Illumina reads, L_ORF_ is the sequence length of target ORFs, n is the number of target ORFs belonging to the same type, and S (Gb) is the size of the metagenomic data.

Data analysis and visualization

All data are presented as the mean of three replicates and were analyzed using one-way analysis of variance (ANOVA) followed by Tukey's honestly significant difference (HSD) test using SPSS. Differences were considered statistically significant at p < 0.05 and marked with an asterisk (*). All figures were generated using R (V 4.2.3), Gephi (V 0.10.1), and online platforms [36, 37].

Results

Dissipation and bioaccumulation of ALB in the soil-earthworm system

The average fortified recoveries of ALB in soil and earthworms were 84.36%–99.27% and 86.89%–102.31%, respectively, with corresponding relative standard deviations (RSDs) ranging from 3.24%–6.43% and 1.61%–6.33%. The limit of detection (LOD) and limit of quantification (LOQ) in soil and earthworms were 0.01 mg/kg and 0.05 mg/kg, respectively. The above results indicated that the analytical methods of ALB in soil and earthworms meet the analysis requirements for pesticide residues.

At the beginning of the experiment, 20 earthworms were added to each plastic pot. After 28 d of exposure, the mortality rate of earthworms in all treatments remained below 10%, and no significant differences in the mortality rate of earthworms were observed among the different treatments. As illustrated in Fig. 1a, the accumulation concentration of ALB in the earthworm followed a dynamic trend of initial decrease, subsequent increase, and final decrease. Compared to ALB treatment alone, ALB accumulation was significantly reduced in the ALB-ZnO treatment samples (p < 0.05). After 28 d, the residual ALB concentrations in earthworms were 0.061 mg/kg (ALB treatment) and 0.031 mg/kg (ALB-ZnO treatment). Additionally, the bioaccumulation factor (BAF) of ALB in earthworms increased over time (Fig. 1b), with ZnO NPs significantly lowering the BAF (p < 0.05). Meanwhile, the dissipation of ALB in soil followed first-order kinetics (0.9918 < r < 0.9930). The half-lives of ALB in soil were 4.70 d (ALB treatment) and 4.15 d (ALB-ZnO treatment), respectively, and the corresponding final soil residual levels were 0.20 mg/kg and 0.11 mg/kg, respectively. Collectively, the accelerated dissipation of ALB in soil and its reduced accumulation in earthworms suggest that ZnO NPs decrease both the environmental persistence and biological exposure of ALB, thereby limiting its transfer to soil biota and reducing internal exposure and potential long-term toxicity risks in soil organisms.Fig. 1. Bioaccumulation concentration a and bioaccumulation factor b of albendazole in the earthworm gut

Effects of ALB on related degradation genes in the soil-earthworm system

As shown in Fig. 2a, both BDGs and ADGs were detected in all earthworm samples, and their total abundance varied from 0.085 to 0.53. Compared to the control, ADG abundance in the earthworm gut significantly decreased by 17.75% and 7.78% under the E-ZnO and E-ALB-ZnO treatments, respectively, while no significant changes were observed in the E-ALB treatment. In contrast, BDG abundance significantly declined (p < 0.05) across all treatments, with reductions of 11.70%, 19.01%, and 11.91% in the E-ALB, E-ZnO, and E-ALB-ZnO treatments, respectively.Fig. 2. Biodegradation gene profiles in the earthworm gut. a Relative abundance of BDGs and ADGs in different treatments. b Heatmap of the BDGs and ADGs based on common logarithmic transformed abundance. Abbreviations: BDGs, biodegradation genes; ADGs, albendazole degradation genes

The normalized and log-transformed heatmap further illustrated the abundance variations of specific ADG and BDG gene clusters under different treatments (Fig. 2b). ALB exposure notably increased the abundance of ADGs (hmr and ami) in the earthworm gut by 20.26% and 29.22%, respectively, while reducing dominant BDGs (benA, p450, lip, and mmoX) by 11.91%–68.51%. Strikingly, ALB also triggered a substantial upregulation (19.26%–5264.29%) in the ppo, xylA, cutC, and nfsl genes. These findings imply that ALB may promote earthworm metabolic adaptability to pollutants by selectively activating biodegradation-related gene clusters.

In contrast to the earthworm gut, soil samples showed no significant changes in the total abundance of ADGs and BDGs (Fig. S2). Notably, ALB treatment resulted in a 460.32% increase in nfsl gene abundance compared to the control, which was consistent with previous observations in the earthworm gut.

Effects of ALB on earworm gut microbial metabolic function

Based on the KEGG database analysis, significant differences were observed in microbial metabolic pathway abundances between E-ALB and control samples. At level 1 of the KEGG classification, three major functional categories showed significantly higher relative abundance in E-ALB samples: environmental information processing, genetic information processing, and cellular processes (Fig. 3a). Further analysis of level 2 KEGG pathways revealed that 21 out of 36 microbial metabolic pathways exhibited increased relative abundance in E-ALB samples, with enhancement percentages ranging from 0.126% to 36.30% (Fig. 3b). Additionally, functional annotation identified 3,915 enzymes, among which 2,693 enzyme-coding genes displayed higher relative abundance in E-ALB samples compared to the control (Fig. S3). The top 10 enzymes with the most significant increases included tryptophan 6-halogenase (EC 1.14.19.59), galactinol-sucrose galactosyltransferase (EC 2.4.1.82), sorbose reductase (EC 1.1.1.289), long-chain fatty acid adenylyltransferase FadD28 (EC 6.2.1.49), ent-pimara-9(11),15-diene synthase (EC 4.2.3.31), sortase B (EC 3.4.22.71), (R)−3-((carboxymethyl)amino)fatty acid dioxygenase (EC 1.14.11.78), prosolanapyrone-II oxidase (EC 1.1.3.42), prosolanapyrone-III cycloisomerase (EC 5.5.1.20), and NADH peroxidase (EC 1.11.1.1).Fig. 3. Changes in microbial metabolic functions in the earthworm gut.** a** Relative abundance of microbial metabolic pathways at level 1 among different treatments. b Heatmap of the microbial metabolic pathways at level 2 and the corresponding relative abundance based on common logarithmic transformed abundance among different treatments

BDGs were highly correlated with MGEs

As shown in Fig. 4a, the total relative abundance of MGEs in all earthworm gut treatment samples was lower than that in the control. Specifically, the relative abundance of plasmids, transposons, and integrons in E-ALB samples decreased by 26.7%, 36.03%, and 26.94%, respectively. Linear regression analysis revealed a strong correlation between BDGs and the three MGEs, which exhibited statistically significant coefficients ranging from R^2^ = 0.7357 to 0.7888 (p < 0.001, Fig. 4b). In contrast, the correlation between ADGs and MGEs was weak (Fig. S4). Further analysis of contigs after metagenomic assembly confirmed the co-occurrence of MGEs and BDGs. As shown in Fig. 4c, the proportion of contigs containing both MGEs and BDGs was significantly higher in the control than in the E-ALB, E-ZnO, and E-ALB-ZnO treatments, consistent with previous findings.Fig. 4. Links between BDGs and MGEs in the earthworm gut.** a** Total relative abundance of MGEs among different treatments. b Linear regression analysis between the relative abundance of BDGs and MGEs. c Proportions of BCCs with MGEs among different treatments. d The mapping relationship between BDGs and MGEs in the E-ALB samples. The bars on the inner ring represents the co-occurrence of BDGs and MGEs on BCCs. Abbreviations: BDGs, biodegradation genes. MGEs, mediated by mobile genetic elements. BDGs, biodegradation genes. BCCs, BDG-carrying contigs. E-ALB, earthworm samples treated with albendazole. E-ZnO, earthworm samples treated with zinc oxide nanoparticles. E-ALB-ZnO, earthworm samples treated with albendazole and zinc oxide nanoparticles and E-CK, untreated earthworm samples

Significant differences in the BDG carriers among all samples were observed (Fig. S5). Chromosome-borne and plasmid-borne MGEs accounted for 36.97% and 23.14% of these carriers, respectively. Although plasmid-borne MGEs were less abundant than chromosome-borne MGEs, plasmids are widely recognized as highly mobile genetic elements capable of mediating efficient horizontal gene transfer across diverse bacterial taxa. Furthermore, the co-occurrence of BDGs and MGEs provided stronger evidence for the mobility of BDGs. As shown in Fig. 4d, a distinct selective pattern of co-existence of MGEs and BDGs was observed in the earthworm gut under ALB treatment. Among the 16 BDGs detected across all samples, most exhibited extensive colocalization with plasmids and transposases, while others (e.g., mmoX, bphA2, cntA, dbfA1, alkb, bph) were linked with specific MGEs. For instance, in the E-ALB samples, bph frequently co-localized with plasmids, whereas in the E-CK samples, it was predominantly associated with integrases (Fig. S6). Additionally, certain contigs contained multiple MGEs linked to specific BDGs, such as ppo and nfsl, which were associated with plasmids, integrases, resolvases, and recombinases (Fig. S7).

ALB induced shifts in the microbiome, ADG, and BDG hosts of the earthworm gut

As shown in Fig. 5a, the dominant bacterial phyla in the earthworm gut microbiome were Pseudomonadota, Actinomycetota, and Bacillota, collectively accounting for over 95% of the total abundance. Exposure to ALB and ZnO NPs significantly altered the bacterial community structure in the earthworm gut but had no notable effect on soil bacterial communities (Fig. S8). Compared to the control, the relative abundance of Pseudomonadota decreased by 22.77% (E-ALB) and 24.50% (E-ALB-ZnO), whereas that of Actinomycetota and Bacillota increased by 36.18% − 36.33% and 23.87% − 40.28%, respectively. The normalized and log-transformed heatmap (Fig. S9) further revealed shifts in dominant gut bacterial genera, including Bradyrhizobium, Brucella, Microbacterium, Pseudomonas, Micromonospora, Agromyces, Methylobacterium, and Microvirga. Alpha diversity analysis (Fig. S10) revealed a Shannon index range of 3.72–4.99 in the earthworm gut, and ALB significantly increased the diversity of earthworm gut bacteria (p < 0.05). Principal component analysis (PCA, Fig. S11) demonstrated that the E-ALB, E-ZnO, and E-ALB-ZnO treatments were significantly separated from the control at the genus level (p < 0.05, ANOSIM), although no significant differences were observed among the treatment groups.Fig. 5. Shifts in the microbiome, BDG and ADG hosts in the earthworm gut. a The structural composition of the soil microbial community among the different treatments. The co-occurrence network of BDGs b and ADGs c with genera (top 100) in the earthworm gut. Data in all samples with significant correlation (|ρ|> 0.7) and significance (p < 0.05) were screened based on Spearman analysis. The red line represents negative correlation, and the green line represents positive correlation. d Visualization based on the relative abundance and taxonomic assignment of the recovered MAGs carrying BDGs. e Relative abundance of BDG-carrying bacterial genera based on the recovered MAGs in the E-ALB samples. f The relative abundance of relevant hosts carrying both BDGs and ADG. Abbreviations: BDG, biodegradation gene; ADG, ALB Degradation Gene; MAGs, metagenome-assembled genomes

The co-occurrence network based on Spearman analysis identified 83 potential hosts of BDGs (Fig. 5b). Among them, Microvirga was the most prevalent host, associated with 9 BDGs, followed by Methylorubrum, Methylocystis, Methylobacterium, Cohaesibacter, Chelatococcus, and Azospirillum, each linked to 8 BDGs. Additionally, six ADG hosts were detected (Fig. 5c). Notably, Sphaerobacter, Saccharothrix, Actinomadura, and Nocardia were identified as dual hosts for both BDGs and ADGs. These bacterial genera harboring BDGs and ADGs were further confirmed through the binning method (Fig. 5d and Fig. S12). Compared to the control, ALB significantly enhanced the diversity of BDG hosts (ppo, xylA, cutC, and nfsl) and enriched ADG hosts (Fig. 5e, Fig. S13a, and Fig. S13b). As shown in Fig. 5f, ALB exposure markedly increased the relative abundance of Actinomadura, Sphaerobacter, and Saccharothrix by 88.23% to 132.63%, suggesting that ALB promoted the proliferation of microbial hosts carrying dual functional degradation genes.

Discussion

Nanomaterials such as ZnO NPs have attracted significant attention due to their unique physicochemical properties and potential agricultural applications, including their use in combination with pesticides to enhance efficacy [38]. However, their environmental distribution and potential toxicity remain controversial [39, 40]. During the 28-d exposure experiment in this study, 100 mg/kg ZnO NPs showed no significant toxicity to earthworms and did not significantly disrupt the ecological balance of soil bacterial communities, which is similar to the results reported in previous studies [41, 42]. Daqa et al. [43] reported that ZnO NPs could promote pesticide degradation in soil, and our study also observed an enhanced degradation of albendazole (ALB), which may be due to its adsorption on ZnO NPs. Furthermore, ZnO NPs could significantly reduce the bioaccumulation of ALB in earthworms, which may be associated with the stimulation of ALB detoxification pathways [44].

Microbially driven enzymatic reactions are the cornerstone of exogenous pollutant biodegradation [45], yet studies on functional genes of ALB-degrading enzymes remain limited. In this study, both ZnO NPs and ALB broadly inhibited functional genes associated with xenobiotic degradation in the soil-earthworm system, potentially impairing the detoxification capacity of earthworms. However, under ALB exposure, the relative abundance of specific BDGs—including polyphenol oxidase gene (ppo), xylose isomerase gene (xylA), choline trimethylamine lyase gene (cutC), and nitroreductase gene (nfsl)—significantly increased in the earthworm gut. This selective enrichment of BDG may reflect adaptive evolution of ALB metabolism by gut microbiota, enhancing pollutant removal through substrate-specific pathways such as polyphenol oxidase-mediated aromatic ring cleavage and xylose isomerase-associated cofactor regeneration [46–49]. Additionally, the abundance of ADGs such as aminotransferase gene (ami) and hemoprotein reductase gene (hmr) was significantly increased, suggesting that the gut microbiota preferentially activated two key metabolic pathways: (i) sulfur reduction and amination of ALB sulfone, yielding low-toxicity metabolites via aminotransferases [50]; and (ii) hydroxylation of ALB to form hydroxyalbendazole through side-chain oxidative modification [51]. KEGG database analysis further revealed that ALB enhanced four metabolic pathways (metabolism, environmental information processing, genetic information processing, and cellular processes) and upregulated potential functional enzymes (e.g., CYP121, EC 1.14.19.70), indicating that earthworm gut microbiota achieved efficient ALB degradation via metabolic remodeling.

The strong correlation between BDGs and MGEs underscores the crucial role of HGT in the transmission of functional genes such as BDGs across bacterial communities [22, 24]. This is further supported by the finding that BCCs in the earthworm gut are predominantly associated with plasmids and transposons. The observed reduction in BDGs under ALB exposure may result from the regulation of related mobilomes, leading to decreased potential mobility of these genes [13]. Co-occurrence analysis revealed that key BDGs—such as mmoX, bphA2, cntA, dbfA1, alkb, and bph—were frequently linked to specific MGEs, suggesting their capacity for horizontal transfer. However, this transfer potential may be suppressed in the earthworm gut under ALB exposure. Notably, bph, a key BDG involved in aromatic hydrocarbon degradation, shifted from its original co-localization with integrase to plasmids, possibly due to ALB-induced alterations in gene transfer regulatory mechanisms [14, 52]. Additionally, the difference in mobility potential between BDGs and ADGs can be attributed to the distinct characteristics of these two gene types. The distribution of ADGs specifically depredating ALB within microbial communities may be more restricted compared to BDGs [53]. Their horizontal transfer potential is likely constrained by their specificity to ALB exposure and the environmental conditions that select for these genes.

The co-occurrence network identified Microvirga, Methylobacterium, and related genera as keystone BDG hosts, all belonging to the phylum Pseudomonadota, highlighting their critical role in complex organic matter degradation [12, 54]. Previous studies have demonstrated that Microvirga and Methylobacterium participate in the biodegradation of various pollutants, suggesting their potential application in environmental pollution remediation [55–57], which aligns with our findings. Interestingly, certain genera (e.g., Sphaerobacter, Saccharothrix, Actinomadura, and Nocardia) exhibited dual functionality by hosting both BDGs and ADGs, implying a synergistic detoxification mechanism where ADGs may preprocess substrates for subsequent BDG-mediated degradation of ALB. Notably, ALB not only enriched ADG hosts but also increased the diversity of BDG hosts (e.g., xylA, cutC), indicating a compensatory response to enhance metabolic redundancy [58, 59]. This redundancy improves microbial community resilience and adaptability in fluctuating pollutant conditions—a key survival trait in polluted ecosystems. By occupying pivotal ecological niches, these genera facilitate pollutant breakdown and detoxification [59].

Consistent with prior reports, single exposure to ALB induced significant structural shifts in the earthworm gut microbiome [53]. In contrast, co-exposure to ALB and ZnO NPs did not induce further significant structural alterations or detectable adverse effects in the microbiota compared to single ZnO NPs treatment. Moreover, the reduction in BDG abundance under ALB exposure was partly attributed to a notable decrease in Pseudomonadota. However, the significant increase in Actinomadura, Sphaerobacter, and Saccharothrix indicated that ALB reshaped the gut microbiota into a functionally specialized community, driving the enrichment of specific BDGs such as ppo, xylA, cutC, and nfsl. The co-enrichment of both ADGs and BDGs in the earthworm gut likely served to mitigate xenobiotic toxicity and enhance adaptability in polluted environments. Concurrently, the increased relative abundance of Actinomycetota and Bacillota may aid earthworms in decomposing complex organic matter, improving food processing, digestion, nutrient absorption, and immune function [60, 61]. Metagenomic analysis showed that microbial populations carrying different degradation genes were involved in the corresponding degradation pathways of ALB in soil, but the specific metabolic pathways of ALB still need further verification. Future studies are essential to identify and track the dynamics of key ALB metabolites (e.g., albendazole sulfoxide, albendazole sulfone, and further hydrolyzed or ring-opened products) in both soil and earthworms by employing high-resolution HPLC–MS/MS.

Conclusion

The results obtained in this study showed that earthworms selectively enriched specific BDGs involved in detoxification, metabolic processes, and antioxidant responses, as well as ADGs related to ami and hmr, to synergistically degrade ALB. Bacterial genera such as Sphaerobacter, Saccharothrix, Actinomadura, and Nocardia were identified as potential hosts for both BDGs and ADGs. Additionally, ALB inhibited the spread of BDGs mediated by MGEs, thereby reducing BDG abundance in the earthworm gut. Notably, ZnO NPs significantly reduced the bioaccumulation of ALB in the earthworm gut and enhanced its degradation in soil. These insights enhance our understanding of pollutant-microbiome interactions and provide novel perspectives for bioremediation strategies in contaminated environments. Future research will incorporate metatranscriptome and qPCR analyses to validate the expression activity of functional genes.

Supplementary Information

Supplementary Material 1.