Rebound of Antibiotic Resistance Genes in Composting: Mechanisms, Challenges, and Control Strategies

Abstract

The proliferation of antibiotic resistance genes (ARGs) in livestock manure has raised growing environmental and public health concerns. Composting is widely recognized as an effective method to mitigate ARG dissemination; however, recent studies have increasingly reported a rebound in ARG abundance during the curing stage of composting, undermining its long-term effectiveness. Here, “rebound” refers to a renewed increase in ARG abundance—either in absolute terms or relative to the 16S rRNA gene—following its decline to a minimum during the thermophilic phase. This review systematically summarizes the dynamic changes in ARGs throughout the composting process, with a particular focus on the mechanisms and drivers underlying ARG rebound. Vertical and horizontal gene transfer, along with microbial succession, are discussed as key contributors to this phenomenon. Current strategies to suppress ARG rebound, including microbial community manipulation, hyperthermophilic composting, and exogenous amendments, are evaluated. Furthermore, the roles of heavy metals and extracellular polymeric substances in promoting ARG persistence are examined, highlighting their potential involvement in ARG rebound. This review aims to provide a comprehensive understanding of ARG rebound in composting and to inform the development of more effective, integrated mitigation strategies.

1. Introduction

Since the discovery of penicillin by Alexander Fleming in 1928, antibiotics have revolutionized human health by effectively treating infectious diseases and saving millions of lives [1]. Following the widespread clinical use and commercialization of antibiotics, resistance emerged rapidly, with penicillin-resistant Staphylococcus reported as early as the 1950s [2,3]. Initially confined to clinical settings, antibiotic resistance became a broader concern by the 1980s, prompting surveillance efforts by the WHO and health authorities in developed countries. In the 21st century, globalization and the extensive use of antibiotics in agriculture have accelerated the spread of resistance, transforming it into a global public health crisis. In response, efforts have shifted from hospital-based monitoring to integrated, multidisciplinary strategies, with growing participation from developing countries and increased international collaboration [4].

The spread of antibiotic resistance genes (ARGs) has significantly compromised the effectiveness of conventional antibiotics, leading to longer treatment durations, higher healthcare costs, and increased mortality. In the United States, antibiotic-resistant infections affect an estimated 3 million people and cause approximately 36,000 deaths annually [5]. Globally, such infections were directly responsible for 1.27 million deaths in 2019, with another 4.95 million deaths associated with resistance-related complications [6]. Without urgent action, antimicrobial resistance could claim 10 million lives each year by 2050 and result in a cumulative economic loss of up to USD 100 trillion [7]. In agriculture and animal husbandry, antibiotic resistance complicates the treatment of infections, increases the difficulty of disease prevention, and raises veterinary costs. It may also impair animal growth and productivity, leading to substantial economic losses for the livestock industry [8]. Moreover, the presence of ARGs can alter the structure and function of environmental microbial communities, potentially disrupting fundamental ecosystem processes such as nutrient cycling and organic matter decomposition, which are essential for maintaining ecological balance.

The widespread use of antibiotics in livestock production has become a major driver of environmental and public health risks associated with antimicrobial resistance. Since their discovery, antibiotics have been extensively applied in animal farming to promote growth and prevent disease. At present, antibiotic use in livestock far exceeds that in human medicine. In 2013, global antibiotic consumption reached 130,000 tons, and is projected to rise to 200,000 tons by 2030, with usage in animal agriculture estimated to be four times that of human consumption [9]. Driven by growing global demand for meat, eggs, and dairy products, the livestock industry has rapidly expanded toward intensified and large-scale production. However, such high-density rearing systems increase the risk of disease outbreaks, often leading to the excessive use of antibiotics.

Once administered, antibiotics exert strong selective pressure on the gut microbiota, promoting the enrichment of ARGs. Due to limited metabolic capacity in animals, 25–75% of administered antibiotics are excreted unmetabolized via feces and urine [10,11,12], entering the environment and sustaining selection pressure for resistant strains. As a result, livestock manure has become a major reservoir of antibiotics, ARGs, and antibiotic-resistant bacteria (ARB), containing resistance genes against tetracyclines, sulfonamides, aminoglycosides, β-lactams, and quinolones [13,14,15].

There is evidence that animal husbandry is a dominant source of ARG accumulation in agricultural soils [16], and long-term application of manure as organic fertilizer significantly elevates soil ARG levels. Within the One Health framework, ARGs originating from livestock waste can disseminate through various human activities—such as manure application—and natural pathways, eventually entering the food chain and posing threats to human health [17].

Given the large volume of livestock manure generated and the need for sustainable waste management, composting has become a widely adopted strategy for stabilizing and recycling animal waste. As a biological treatment process, composting not only reduces the volume and pathogenicity of manure but also facilitates its reuse as an organic fertilizer [18,19]. Importantly, the thermophilic conditions during composting have been shown to reduce the abundance of certain ARGs and ARB, making it a promising tool for mitigating resistance risks in agricultural systems [20,21].

However, emerging evidence indicates that the efficacy of composting in ARG removal may be limited. While significant reductions in ARG abundance are often observed during the thermophilic phase, a rebound in ARG levels frequently occurs during the cooling and maturation stages [22,23,24], referring to a renewed increase in ARG abundance after its thermophilic minimum, expressed in either absolute or relative terms. This rebound effect has been attributed to several factors, including microbial regrowth, horizontal gene transfer, and the enrichment of resistance-carrying microbial populations. Consequently, composted manure remains a potential reservoir and vector for ARGs when applied to agricultural soils and crops, raising concerns about its environmental safety and long-term impacts on the dissemination of antibiotic resistance [25,26].

Therefore, this review focuses on the occurrence, fate, and dynamics of ARGs during the composting of livestock and poultry manure. Special emphasis is placed on the phenomenon of ARG rebound during composting. It aims to summarize current understanding of ARG behavior during composting, elucidate the underlying mechanisms driving their enrichment and rebound, and propose effective mitigation strategies. Relevant studies were primarily collected from the Web of Science Core Collection using combinations of the keywords “compost,” “antibiotic resistance genes,” “rebound,” “microbial community,” “heavy metal,” and “extracellular polymeric substances.” The search covered publications up to June 2025. Priority was given to peer-reviewed articles that provided mechanistic insights or experimental evidence related to ARG fate during composting. Additional references from wastewater treatment or sludge systems were included when compost-specific studies were limited but conceptually relevant.

2. ARG Dynamics During Composting

2.1. The Composting Process

The composting process can be broadly divided into four sequential stages: the mesophilic phase, thermophilic phase, cooling phase, and maturation phase. Initially, composting enters the mesophilic phase, which lasts from a few hours to several days. During this stage, moderate-temperature microorganisms decompose easily degradable organic matter, leading to a rapid increase in temperature. This is followed by the thermophilic phase, which typically lasts from several days to a few weeks. It is characterized by intense microbial activity and elevated temperatures (40–70 °C), which accelerate the breakdown of complex compounds and contribute to the inactivation of pathogens and weed seeds. As the easily degradable substrates are depleted, the temperature gradually decreases, marking the transition to the cooling phase, which may span several weeks. During this phase, mesophilic microbes re-establish and continue decomposing more resistant organic materials. Finally, the process enters the maturation phase, which can last from several weeks to a few months, during which microbial activity declines, and the compost stabilizes into a humus-rich, mature product suitable for agricultural application [27,28].

2.2. Dynamic Changes in ARGs During Livestock Manure Composting

During the composting process of livestock and poultry manure, the dynamic changes in ARGs show a certain regularity. In the initial mesophilic stage, the composting temperature is relatively low, and ARGs are relatively close to the original ARG state in manure. The activities and metabolic activities of microorganisms are in the initiation stage, and the types and abundances of ARGs change relatively little. With the increase in composting temperature, the abundance of many ARGs decreases significantly. Yang et al. [29] evaluated the profile of ARGs during composting of swine manure, food waste, and cornstalk using high-throughput quantitative PCR. A total of 33 ARG subtypes conferring resistance to tetracyclines, sulfonamides, chloramphenicol, macrolide-lincosamide–streptogramin (MLS), quinolones, aminoglycosides, and β-lactams were detected. With the increase in temperature, the abundances of ARGs in both treatments showed a downward trend, with a decrease in target ARGs from 1.19 × 10⁹ copies/g DW (dry weight) of the initial material to 1.30 × 10⁷ copies/g DW on the 10th day. Previous studies suggest that this might be because high temperature is not conducive to the survival of some microorganisms carrying ARGs [30,31]. Meanwhile, under high-temperature conditions, the metabolic activities of microorganisms accelerate, which may lead to the degradation or inactivation of ARGs.

Yang et al. [29] also found that ARGs rebounded during the cooling and maturation stages; the abundance of ARGs in their conventional composting had rebounded to 5.05 × 10⁸ copies/g DW on the 40th day. In other composting studies, similar rebounds were also observed when the composting temperature began to drop. Hao et al. [32] conducted a 35-day composting experiment using pig manure and straw and found that the total abundance of ARGs increased by 729.23% from day 14 to the end of the composting process. The main contributors to this rebound were genes such as tetX, tetG-01, tetG-02, sul1, sul2, strB, aadA9-01, and intI-1. Wen et al. [33] found that in the compost made from pig manure and sawn wood as raw materials, multidrug and sulfonamide subtypes were the main ARGs after composting. Although the total relative abundance of residual multidrug resistance and sulfonamide ARG subtypes decreased significantly during the thermophilic period of composting, at the end of composting, the relative abundances of each group increased significantly and were even significantly higher than those before composting (p < 0.05). Tong et al. [34] conducted a 50-day continuous composting of pig manure and observed the phenomenon that the absolute abundance of ARGs was enriched during the composting process. Guo et al. [35] quantified sixteen ARGs frequently detected in composting systems of chicken manure, including those conferring resistance to tetracyclines (tetM, tetW, tetG, tetC, tetX), macrolides (ermQ, ermF, ermX), sulfonamides (sul1, sul2, dfrA7), quinolones (qnrA, gyrA, aac(6′)-Ib-cr), and β-lactams (blaCTX-M, blaVIM). They reported that the total ARG copy number significantly decreased after the thermophilic stage and rebounded after 17 days. These findings indicate that conventional compost is still insufficient to control antibiotic resistance (Figure 1).

Moreover, during the storage period after composting, a relatively greater rebound of ARGs was observed in Liao et al. [36]. Such a rebound poses a considerable risk to soil–crop systems. The application of composted swine manure has been reported to significantly increase the abundance of ARGs in agricultural soils and plants, thereby facilitating the spread of ARGs within farmland ecosystems [37]. Similarly, the use of chicken manure fertilizer not only enhanced lettuce yield but also markedly elevated the ARG abundance in the soil–crop continuum [38]. In addition, manure supplementation can lead to increased concentrations of heavy metals and integrase genes in soils, as well as their accumulation in crops, accelerating the environmental dissemination of antibiotic resistance [26]. Therefore, it is essential to elucidate the mechanisms underlying ARG rebound during composting and to develop effective strategies to mitigate this risk.

3. Drivers of ARG Rebound in the Curing Stage of Composting

3.1. Inheritance and Dissemination of ARGs: Roles of HGT and VGT

To better understand the mechanisms underlying the rebound of ARGs during composting, it is essential to first examine how ARGs disseminate in the environment. The environmental dissemination of ARGs occurs predominantly via vertical and horizontal gene transfer (VGT and HGT). Each mechanism operates through distinct biological processes that enhance the spread of resistance traits, thereby exacerbating the risk of antibiotic resistance proliferation and threatening both public health and ecological integrity (Figure 2).

Under selective pressure from external factors such as antibiotics, certain bacteria acquire ARGs through genetic mutations and subsequently become resistant strains. During reproduction, these resistant bacteria can transmit ARGs from the mother cell to the daughter cells [39]. In other words, resistance genes are stably inherited by progeny during bacterial cell division. Although the rate of VGT is relatively slow, its cumulative effect in specific environments should not be overlooked. In environments with frequent antibiotic use—such as livestock farms—ARGs can become stably maintained and gradually accumulate within bacterial populations through VGT, leading to the persistent presence of ARGs in the environment [39]. However, the VGT process is generally restricted to bacteria of the same species or closely related taxa.

ARB can transfer ARGs to other non-resistant bacteria and even different species through HGT [40]. HGT significantly broadens the dissemination potential of ARGs and increases the risk of their transfer to human pathogens [41]. The three main mechanisms of HGT include conjugation, transduction, and transformation. Transformation involves the uptake of free extracellular DNA from the environment, including fragments released by lysed bacteria. These exogenous genetic materials can be incorporated into the genome of competent recipient cells, potentially conferring antibiotic resistance [42]. Conjugation requires the establishment of physical contact between donor and recipient bacteria, typically mediated by a long, thin bacterial appendage known as the sex pilus. The sex pilus is produced by the donor cell and extends to attach to the recipient, forming a conduit for genetic material transfer. During conjugation, mobile genetic elements (MGEs) such as plasmids, transposons, or integrons carrying ARGs are transferred from the donor to the recipient bacterium [42]. The dynamics of MGEs varied with composting material, treatments, and other environmental factors. Xie et al. [43] investigated 10 MGEs (2 integrase and 8 transposase genes) during composting and observed distinct patterns between manure types: MGE abundances decreased in cattle manure compost but rebounded after the thermophilic phase in poultry manure compost. During the subsequent storage period, integrases (intI1, intI2), transposases (Tn916/1545), IS91, and tniB all increased, accompanied by a corresponding rise in ARG levels [36]. This increase in MGEs may enhance the potential of ARG transmission.

In transduction, bacteriophages act as vectors that transfer ARGs between bacterial hosts, promoting the spread of resistance across microbial communities. Recent studies further suggest that such phage-mediated ARG dissemination also occurs in organic waste treatment systems. In activated sludge from pig farms employing A/O wastewater treatment, phages—particularly those belonging to the Siphoviridae, Myoviridae, and Podoviridae families—have been identified as potential hosts of ARGs, suggesting that phage-mediated HGT may contribute to ARGs spread within microbial communities [44]. Evidence from soil fertilized with dairy manure or biosolids further supports this view, showing that bacteriophages can harbor ARGs and mediate their transfer under selective pressure [45]. During composting, environmental stresses can activate prophages integrated in bacterial genomes, releasing phage particles that mediate ARG transfer and potentially contribute to ARG persistence and rebound [46]. However, direct evidence from composting systems remains scarce, and the ecological role of phage-mediated transduction in ARG dynamics requires further investigation.

Vertical and horizontal gene transfer are not isolated processes; rather, they often act synergistically to expand the dissemination of ARGs. While horizontal gene transfer enables the rapid movement of resistance genes across different bacterial populations, vertical gene transfer ensures their stable inheritance within a population over time [39]. Together, these mechanisms facilitate the persistence and proliferation of antibiotic resistance within microbial communities. The rebound of ARGs in composting is associated with the reactivation of ARG-carrying microorganisms that were suppressed during the thermophilic phase, as well as the continuous enrichment of ARGs within newly established microbial communities through HGT.

3.2. Phylogenetic Shifts in Bacterial Hosts Influencing ARG Dynamics

The profile of ARGs is strongly associated with the composition of bacterial hosts, as most ARGs exhibit host-specific characteristics. A metagenomic analysis of ARGs and their bacterial hosts in pig farm wastewater treatment systems and adjacent farm soils revealed that the abundance and distribution of ARGs varied significantly among different samples, such as feces, wastewater, and soil [47]. This variation was primarily attributed to the differences in the composition of bacterial hosts, with specific bacterial families being identified as major ARG carriers in each environment [48]. Composting systems host diverse microbial communities that undergo temperature-driven succession throughout the process. Such phylogenetic shifts may partly account for the dynamic variations and rebound of ARGs.

During composting, the microbial community undergoes a characteristic succession at the phylum level. In the initial phase of composting, the bacterial community is often dominated by phyla such as Firmicutes, Bacteroidetes, and Proteobacteria [48,49]. As the temperature rises to the thermophilic phase, Firmicutes (particularly genera such as Bacillus) become predominant [49], reflecting their strong tolerance to heat and ability to decompose complex macromolecules, including cellulose and lignin. With the decline in temperature in the cooling and maturation stages, Actinobacteria typically increase in abundance, contributing to the humification and stabilization of organic matter. In sheep manure composting, Actinobacteria accounted for 52.22% of the bacterial community in the maturation phase [48]. Similarly, in cow manure and corn straw composting, Actinobacteria gradually increased from the thermophilic phase to the end of composting [49].

Proteobacteria are widely recognized as one of the primary reservoirs of antibiotic resistance genes (ARGs) across diverse environments due to their high genomic plasticity and frequent association with mobile genetic elements. Wang et al. [50] examined the microbial communities and ARG profiles in a pig farm and reported that Proteobacteria exhibited the greatest potential for ARG carriage—approximately 9 to 20 times higher than that of other bacterial groups. Similarly, Korry and Belenky [51] found a positive correlation between the abundance of Proteobacteria and the overall ARGs within microbial communities. These bacteria, which are predominantly abundant in animal feces and feed, are key hosts of resistance determinants such as sul, tet, and β-lactam resistance genes [52,53]. During the thermophilic phase of composting, the reduction in Proteobacteria abundance may contribute to the overall decline in ARG levels.

During maturation, increased humification and stabilization of organic matter create microhabitats that can favor the persistence and regrowth of microorganisms carrying ARGs, thereby facilitating ARG retention in the compost matrix [54]. The resurgence of Proteobacteria is believed to contribute to the rebound of ARGs during the cooling and maturation phases [55]. However, this resurgence alone cannot fully account for the observed increase; rather, it reflects the combined effects of microbial succession and the persistence of resistant populations adapted to the changing compost environment. Consistent with this view, metagenomic host-assignment and MGE analyses from composting studies indicate that ARG dynamics are shaped by the interplay between host shifts, vertical inheritance, and MGE-mediated horizontal transfer [56].

3.3. Selective Pressure of Heavy Metals on ARG Persistence

Although antibiotics, as the primary selective pressure inducing the generation of ARGs, are easily degraded during the high-temperature phase of composting, other inducing factors, such as heavy metals, can persist throughout the composting process and play a significant role in the enrichment and dissemination of ARGs. In intensive animal husbandry, heavy metal additives are widely used in feed to promote animal growth and prevent diseases [57]. However, these heavy metals accumulate in animal manure and subsequently enter the environment through the application of manure as fertilizer, posing significant environmental and health risks. Moreover, due to the presence of co-selection pressures, heavy metals in the environment further facilitate the spread and dissemination of ARGs.

During livestock and poultry breeding, heavy metals such as copper (Cu), zinc (Zn), and chromium (Cr) are commonly used as feed additives, primarily for improving feed utilization, promoting animal growth, and preventing diseases [58]. Although some countries have regulated the addition of heavy metals and banned the use of certain toxic heavy metals, the sources of feed raw materials and the production processes may still introduce additional heavy metals and toxic heavy metals. A study collected 104 livestock feed samples and 118 animal manure samples from farms of different scales in Northeast China and measured their heavy metal contents: All feed samples contained the heavy metals Cu, Zn, and arsenic (As), indicating the widespread use of these heavy metal additives in animal production in the region. In cattle feed, the content of Cr was generally below the detection limit, while Cr was detected in more than 13% of pig feed and over 20% of poultry feed. Although cadmium (Cd) is not an essential element for animal growth, it was detected in more than 60.6% of the feed samples surveyed [59].

Due to the limited metabolic capacity of animals, most of the ingested heavy metals are excreted through feces and urine [11], thereby entering the environment. These heavy metals exhibit significant enrichment in livestock and poultry manure, with concentrations typically around four times higher than those in feed [59]. Additionally, the heavy metal content in different types of livestock and poultry manure varies considerably. In a study, researchers collected data on heavy metal concentrations in livestock and poultry manure from all over China and analyzed the differences in eight heavy metals (Zn, Cu, Pb, Cd, Cr, Hg, As, and Ni) [60]. In terms of average concentrations, Zn, Cu, Cd, and As in pig manure were significantly higher than those in poultry manure (such as chicken manure) and much higher than those in cattle and sheep manure. In contrast, the average concentrations of Pb, Cr, and Ni were relatively higher in poultry manure [60]. A recent study established a dataset comprising 4082 records of eight heavy metals in animal manure from various regions of China. The analysis revealed that the average concentration of Zn in manure was the highest (527.55 mg/kg), followed by Cu (247.36 mg/kg). The concentrations of the remaining heavy metals in descending order were Cr (33.88 mg/kg), Pb (14.04 mg/kg), Ni (13.07 mg/kg), As (5.62 mg/kg), Cd (2.72 mg/kg), and Hg (0.16 mg/kg) [42]. The maximum concentrations of these eight heavy metals in manure were 16,928 mg/kg for Zn, 4187 mg/kg for Cu, 4304 mg/kg for Cr, 352.7 mg/kg for Pb, 51.7 mg/kg for Ni, 685.08 mg/kg for As, 203.4 mg/kg for Cd, and 21 mg/kg for Hg [61].

The heavy metals enriched in manure not only exert selective pressure on microorganisms to acquire metal resistance genes (MRGs) but also promote the selection of ARGs through co-selection mechanisms. The co-selection mechanisms of heavy metals mainly include three types: co-resistance, cross-resistance, and co-regulated resistance. Co-resistance occurs when antibiotic and heavy metal resistance genes are located on the same genetic element, such as a plasmid, transposon, or integron. When heavy metals exert selective pressure to select for MRGs, all ARGs located on the same genetic element are co-selected. Hasman and Aarestrup [62] demonstrated a correlation between copper resistance and resistance to macrolides and glycopeptides in copper-exposed Enterococcus faecium. They also observed the co-transfer of copper and macrolide resistance phenotypes to E. faecium [63]. This indicates a physical linkage between copper resistance (acquired via the tcrB gene) and the vanA gene cluster and the ermB gene (macrolide resistance) on a single transferable plasmid. Fang et al. [64] found that the IncHI2 plasmid harbors both the pco and sil operons (heavy metal resistance) and the oqxAB/blaCTX-M (antibiotic resistance).

Cross-resistance occurs when a single resistance mechanism in microorganisms confers tolerance to both heavy metals and antibiotics. Some efflux pumps are capable of expelling both heavy metals and antibiotics from bacterial cells. For example, the presence of the MexGHI–OpmD efflux pump system in Pseudomonas aeruginosa increases bacterial resistance to both metals (vanadium) and antibiotics (ticarcillin) [65]. The third mechanism of co-selection—co-regulation—involves shared regulatory pathways that can respond simultaneously to both metals and antibiotics. When bacteria are exposed to metals or antibiotics, regulatory responses are induced, thereby upregulating genes that confer resistance to both metals and antibiotics. For instance, in Escherichia coli, the mdtABC efflux pump operon is upregulated in the presence of zinc, thereby conferring resistance to certain antibiotics.

Heavy metals, unlike organic pollutants, cannot be decomposed or removed through biological degradation processes and exhibit high persistence in the environment. During composting, organic matter is gradually decomposed by microorganisms into carbon dioxide, water, and simple inorganic compounds, leading to a relative increase in the concentration of heavy metals. This enrichment phenomenon results in significantly higher concentrations of heavy metals in the final compost product compared to the original manure. The elevated levels of heavy metals exert stronger and more prolonged selective pressures on microbial communities [66], thereby promoting the enrichment of ARGs through co-selection mechanisms. Numerous studies have demonstrated that the presence of heavy metals in composting facilitates the spread of ARGs. Yin et al. [67] found that the addition of a certain amount of Cu to compost increased selective pressure, which was not only manifested by the increased relative abundance of copper resistance genes (tcrB, cusA, pcoA, and copA) but also by the increased relative abundance of macrolide resistance genes [erm(A) and erm(B)] and the integron intI1. Wang et al. [68] identified a strong correlation between heavy metals (Zn and Cu) and ARGs and MGEs. Deng et al. [69] showed that heavy metals are one of the factors influencing the persistence of ARGs (sul1, aac(60)-Ib-cr, ermB, blaCTX-M, and tetM) during poultry manure composting. However, most current studies focus on heavy metals with relatively high concentrations in manure, neglecting those with lower concentrations but higher bio-toxicity, such as As and Cr, and their impacts on ARGs during composting. Moreover, most studies typically focus on a limited number of genes, failing to comprehensively reflect the dynamic changes in ARGs during composting. This limitation prevents the establishment of a clear relationship between specific heavy metals and certain resistance genes.

Moreover, the selective pressure exerted by heavy metals cannot be solely represented by the total amount of a specific heavy metal, as not all forms of heavy metals exhibit bio-toxicity. In the 1980s, Tessier et al. [70] proposed a five-step sequential extraction method, which classifies heavy metals into five forms: exchangeable, carbonate-bound, iron/manganese oxide-bound, organic matter and sulfide-bound, and residual lattice-bound. During the same period, other single-step or sequential extraction methods emerged to assess the “bioavailable” forms of metals in soils and sediments. However, the lack of standardization in these extraction procedures led to data that could not be compared globally. Subsequently, the European Commission, through the BCR program, established a three-step sequential extraction method to enhance data comparability, although inter-laboratory reproducibility remained suboptimal. Consequently, Rauret et al. [71] proposed an improved BCR extraction scheme, which has been widely adopted to date. According to the improved BCR sequential extraction method, heavy metals are categorized into four forms: exchangeable, reducible, oxidizable, and residual. The exchangeable and reducible forms are considered bioavailable, while the oxidizable and residual forms are considered non-bioavailable. Therefore, when considering the reduction in co-selection pressure on ARGs exerted by heavy metals, it is essential to focus on decreasing the bioavailability of heavy metals.

3.4. Extracellular Polymeric Substance (EPS)-Mediated Protection and Dissemination of ARGs

ARGs in the environment exist in two forms: intracellular ARGs and extracellular ARGs, with distinct distribution and dissemination characteristics. Intracellular ARGs, as the name suggests, reside within cells, while extracellular ARGs are released into the environment through active secretion by living cells or lysis of dead cells [72]. Intracellular ARGs are protected by the cell membrane, shielding them from harsh environmental conditions. In contrast, extracellular ARGs, despite being outside the cell, exhibit greater mobility and can bind to other particles, thereby avoiding degradation by nucleases and persisting in the environment [73]. Studies have reported that extracellular free DNA can remain stable even at 70 °C [74]. The highly resilient extracellular ARGs warrant more attention, as they may play a significant role in the spread and dissemination of ARGs. Horizontal gene transfer of intracellular ARGs primarily occurs through conjugation and transduction, while horizontal gene transfer of extracellular ARGs takes place via transformation [73].

Early studies on ARGs in the environment did not distinguish between extracellular and intracellular ARGs. However, as research has progressed, the significance of extracellular ARGs has gradually emerged, and an increasing number of studies have focused on extracellular ARGs [75,76,77,78,79,80]. Dong et al. [76] evaluated the abundance of 10 ARGs in both intracellular and extracellular compartments in samples from hospitals, pharmaceutical factories, wastewater treatment plants, pig manure sludge, and urban lake sediments. They found that the relative abundance of extracellular ARGs was higher than that of intracellular ARGs in wastewater treatment plant sludge and pharmaceutical sludge. Additionally, they tested the transformability of extracellular ARGs and found that adsorbed extracellular ARGs were more likely to bind with competent cells than free extracellular ARGs. He et al. [78] quantified the abundance of extracellular ARGs in atmospheric fine particulate matter (PM_2.5) and discovered that a significant proportion of ARGs in PM_2.5 were extracellular. Their analysis also revealed that in some cities, extracellular ARGs could account for up to 60% of the inhaled and ingested ARGs. The review by Zarei-Baygi and Smith [80] indicated that extracellular ARGs dominate in natural water bodies receiving wastewater discharge or other pollutants, such as rivers, lakes, estuaries, and oceans. In particular, extracellular ARGs accounted for up to 78% of ARGs in riverine and marine sediments (e.g., Hai River and Yangtze River Estuary). These studies have emphasized the critical role of extracellular ARGs in the dissemination of antibiotic resistance in the environment. However, few studies have quantified extracellular ARGs during the composting of manure. Therefore, it is essential to investigate the distribution characteristics, dynamic transformation patterns, and potential environmental risks of extracellular ARGs during manure composting. This will allow for a systematic assessment of the effectiveness of composting in reducing extracellular ARGs and for the optimization of composting process parameters to minimize the spread of antibiotic resistance through the extracellular gene pathway into agricultural environments and food chains.

Extracellular ARGs primarily exist in the environment in an adsorbed state, which has significant implications for their environmental behavior and dissemination mechanisms [81]. These ARGs can enhance their stability and persistence in the environment by adsorbing onto the surfaces of particulate matter, such as soil minerals, biochar, and microplastics. Adsorbed extracellular ARGs are less susceptible to degradation by nucleases compared to their free counterparts, thus conferring a survival advantage under environmental stress [81]. In microbially rich environmental systems, EPS serve as an important reservoir for extracellular ARGs, enhancing their stability and persistence. Moreover, EPS also acts as a critical medium for the dissemination of extracellular ARGs in the environment. It should be noted that most current evidence on EPS-mediated adsorption and protection of extracellular ARGs is derived from sludge treatment and wastewater systems, while compost-specific data remain limited and warrant further validation.

EPS are high molecular weight polymers secreted by microorganisms (primarily bacteria) under specific environmental conditions. The main components of EPS include polysaccharides and proteins, as well as other organic substances such as nucleic acids, lipids, humic substances, and sugar aldonates [82,83]. EPS can enhance microbial aggregation by adhering to individual bacterial cells and help form granules, thereby influencing microbial flocculation [84]. The abundant functional groups on EPS, such as carboxyl, phosphoryl, sulfydryl, phenolic, and hydroxyl groups, provide numerous adsorption sites for extracellular DNA [85], thereby enabling the adsorption of a higher amount of ARGs onto EPS. Studies have shown that the abundance of EPS-associated ARGs in sludge can reach 1.49 × 10⁷–4.45 × 10⁹ copies g⁻¹-VSS, which is 2–3 orders of magnitude higher than that of cell-free ARGs [82]. Current research on EPS-associated ARGs is mainly focused on sludge treatment systems and microalgae–bacteria co-culture systems, with relatively few studies in composting systems. However, the high concentration of organic matter and active microorganisms in composting may promote the extensive secretion of EPS, forming a dense biological barrier that adsorbs or encapsulates ARG-hosting bacteria and extracellular ARGs. Therefore, it is suggested that the dynamic changes in EPS during composting are essential for regulating the dissemination of ARGs.

Although existing studies have consistently shown that EPS can adsorb extracellular free ARGs, the mechanism by which EPS acts as a medium for the horizontal transfer of ARGs remains controversial. Wang et al. [82] conducted transformation experiments using EPS-associated ARGs and cell-free ARGs isolated from sludge, and the results showed that the former had a transformation efficiency 3.55–4.65 log units higher than the latter. In contrast, another transformation experiment indicated that the binding of plasmids to EPS hindered the horizontal transfer of ARGs carried by the plasmids to E. coli [86]. Li et al. [87] speculated that a portion of EPS may primarily act as a buffer in the dissemination of ARGs, thereby reducing horizontal gene transfer. Currently, no studies have investigated the impact of EPS on the horizontal transfer of ARGs during composting. Composting EPS may act as a “double-edged sword” through a “protection-release” mechanism, potentially serving as a “transfer station” or “buffer zone” for the environmental dissemination of ARGs. Conducting relevant research is not only a key piece in completing the theoretical framework of the environmental behavior of antibiotic resistance genes, but also provides a theoretical basis for ensuring the safe utilization of organic solid waste and blocking the chain of antibiotic resistance dissemination.

4. Control Strategies to Mitigate ARG Rebound in Composting

4.1. Conventional Approaches for ARG Suppression

Currently, it has been recognized that the basis for the rebound of ARGs in the curing stages of composting is the proliferation and succession of microorganisms following the cooling of the compost. The key factors are the vertical and horizontal transfer pathways of ARGs. Figure 2 summarizes conventional control strategies to suppress ARG rebound in compost. As microbial composition and succession are considered key factors influencing the dynamics of ARGs during composting [88], microbial community manipulation has been proposed as a potential strategy to mitigate the rebound of ARGs. Hao et al. [32] attempted to occupy ecological niches by inoculating compost with bacterial communities carrying low levels of ARGs, thereby suppressing the proliferation of ARG-harboring bacteria. Specifically, on day 14 of pig manure composting, they introduced humus-rich high-altitude soil at a concentration of 20% (w/w, dry weight). This intervention effectively reduced ARG abundance in the curing stages of composting, potentially by enhancing competition among non-host microorganisms and altering the dynamics of ARG-hosting bacteria [32]. Other studies have also demonstrated that the use of microbial inoculants to modulate microbial community structure can effectively mitigate the rebound of ARGs. For instance, Guo et al. [35] and Liu et al. [89] have shown the efficacy of microbial inoculants such as Streptomyces avermitilis, Aspergillus niger, and Bacillus licheniformis, or 5% Bacillus megaterium in suppressing the rebound of ARGs.

High temperature is considered a key factor in the reduction in ARGs during the composting process, as it has the potential to kill a majority of bacterial species, which includes the host microorganisms carrying ARGs [90]. Some studies have attempted to enhance the removal of ARGs by increasing temperature and prolonging the duration of high temperatures. Liao et al. [91] found that hyperthermophilic composting was more effective in removing ARGs and MGEs than conventional composting (with removal rates of 89% and 49%, respectively). They also demonstrated that this was associated with a greater reduction in the abundance and diversity of potential ARG-hosting bacteria. However, a rebound of ARGs was observed after day 21 of composting. During the storage process of hyperthermophilic composting, Liao et al. [36] observed a significant rebound of ARGs, which exceeded the initial levels in untreated sewage sludge. Their research indicated that this rebound effect was primarily driven by the regrowth of indigenous ARB that had survived the composting process. Yang et al. [29] extended the duration of high temperatures through external heating to achieve rapid composting, which improved the removal rates of ARGs and MGEs and suppressed the rebound of ARGs. However, external heating may lead to increased energy consumption and costs, while the shortened composting time may affect the maturity of the compost. In a composting environment that is rich in microorganisms or where microorganisms are the primary functional players, the complete removal of ARGs seems paradoxical and may be unattainable.

In addition to regulating temperature and microbial communities, researchers have focused on optimizing composting techniques and additives to enhance the reduction in ARGs during composting, in order to address the challenge of ARG rebound. Although some progress has been made, there is still a gap between the current results and the ideal outcome of complete removal. Fu et al. [92] achieved a high removal rate of 84.8% by simultaneously adding hyperthermophiles and biochar. Cheng et al. [93] investigated the effects of biochar produced at different temperatures from corn straw on the reduction in ARGs in composting and found that biochar produced at 400 °C was the most effective in removing ARGs, yet it could not prevent the rebound of ARGs in the curing stages of composting. Tong et al. [34] added biochar to pig manure composting, which, to some extent, inhibited the increase in the absolute abundance of ARGs, but the abundance of ARGs after composting was still higher than before composting. Zhang et al. [94] added two mineral amendments (diatomite and zeolite), two chemical amendments (sodium selenite and Tween 80), and one plant-based amendment (humic acid) to pig manure and mushroom composting, respectively. The results showed that at the end of composting on day 30, the levels of ARGs in all treatments were higher than those in the control.

These studies have revealed the limitations and instability of ARG removal through additives and have also highlighted the need and urgency for seeking other, more effective ways to suppress the rebound of ARGs. Based on our understanding of composting and a review of the literature, we propose that heavy metals and EPS play significant roles in the rebound of ARGs. The specific reasons are detailed in the following sections, and we also present a potential approach to suppress the rebound of ARGs in composting based on these insights. Based on literature reports and our understanding of composting, it is believed that heavy metals and EPS play significant roles in the rebound of ARGs. The specific impacts are detailed in the following sections, and we also present a potential approach to suppress the rebound of ARGs in composting based on these insights.

4.2. Electric Field-Assisted Composting—A Potential Strategy Targeting Heavy Metals and EPS

In recent years, electric field technology has been introduced and applied to solid composting systems and has been proven to be highly effective, with a promising application prospect. Electric field-assisted composting applies an external electric field to the composting matrix to regulate physicochemical conditions and microbial activities; schematic representations of several typical configurations are provided in Figure 3. The core configuration of electric field-assisted composting involves the installation of electrodes connected to a direct current (DC) or alternating current (AC) power supply. This configuration can, in principle, be applied to any composting system; however, most current studies have focused on its use in closed composting reactors, where electric fields can be precisely controlled [95,96,97].

Compared with traditional aerobic composting, the total greenhouse gas emissions during the electric field-assisted composting process are reduced by approximately 70%, with nitrous oxide emissions decreasing by 73% and methane emissions significantly reduced from 1.6 ± 0.024 mol/kg DW to 0.045 ± 0.007 mol/kg DW [95]. The electric field enhances microbial metabolic activity, significantly increasing the composting temperature and prolonging the high-temperature phase, thereby accelerating the decomposition of organic matter and the inactivation of pathogens [98]. Electric field-assisted composting shortens the maturation period by 33%, thereby improving composting efficiency [95]. The electric field also has a certain promoting effect on the humification process of composting. Under the influence of the electric field, the humus content in the compost product increases significantly by 19.1%, the humic acid content increases significantly by 69.0%, and the stability of humus is enhanced [97]. Proteinaceous substances are gradually converted into humic acid and humus during composting, and this humification process is accelerated under the action of the electric field, thereby improving the degree of humification and the quality of the compost product [99].

Humus can passivate heavy metals through adsorption mechanisms such as chemical and physical adsorption, thereby reducing their toxicity to plants and microorganisms [100]. In electric field-assisted composting, the content of heavy metals (Cu, Zn, As, and Cd) complexed with humic acids significantly increased, indicating that the electric field promotes the formation of stable metal–humus complexes. Specifically, the contents of Cu, Zn, As, and Cd complexed with humic acids increased by 68.4%, 44.4%, 30.5%, and 226.5%, respectively [97]. This suggests that the increase in humus under the action of the electric field not only improves the quality of compost but also enhances the immobilization of heavy metals. As discussed in Section 3.3, heavy metals exert co-selection pressure on ARGs. Electric field-assisted composting can mitigate this pressure by altering heavy metal speciation and reducing their bioavailability. However, the exact effects of the electric field on the speciation of heavy metals in compost are not yet clear, and whether the electric field can reduce the overall bioavailability of heavy metals requires further investigation.

For EPS, the application of an electric field could induce changes in its structure. For instance, under the influence of an electric field, components such as proteins and polysaccharides within EPS may undergo rearrangement or partial degradation. Previous research on sewage sludge electro-dewatering has shown that electric field treatment can transform the α-helix and β-sheet structures in proteins into β-turns and random coils, which may weaken the integrity of EPS [101]. Under certain conditions, the electric field could disrupt the colloidal stability of EPS, causing it to transition from a compact to a loose structure. In the presence of a strong electric field, EPS may even rupture. The intermolecular forces among EPS components, which are crucial for the colloidal stability of flocs, are disrupted during the electrolysis process [102]. Free radicals (such as hydroxyl radicals) generated by the electric field could also degrade EPS, leading to its rupture. This is because oxidizing agents can enhance the oxidative degradation of EPS by the electric field [103]. Although direct composting studies are limited, related evidence from sludge or wastewater treatment systems provides mechanistic insights into the potential effects of electric fields on EPS in compost.

The disruptive effect of the electric field on EPS implies that EPS may release substances that were previously tightly adsorbed and encapsulated into the environment. In a study focusing on sludge, the application of an electric field caused the EPS structure to become more loosely fragmented, and the disruption of EPS led to the release of bound water that was otherwise difficult to remove [104]. As previously discussed (Section 3.4), EPS plays a crucial role in stabilizing extracellular ARGs. The electric field may indirectly influence this process by altering EPS conformation and integrity, thereby affecting ARG stability. In addition to potentially affecting the presence of extracellular ARGs by disrupting the structure of EPS, the electric field may also influence the dissemination and spread of ARGs in the composting environment by affecting the transformation of extracellular ARGs. It should be noted that some mechanistic interpretations in this review are inferred from related sludge or wastewater studies due to limited compost-specific evidence. Future research should focus on verifying these mechanisms under real composting conditions.

In summary, heavy metals and EPS are two critical drivers influencing the rebound of ARGs during composting through co-selection and protection mechanisms, respectively. Building upon these mechanisms, electric field-assisted composting could potentially regulate HM bioavailability and EPS integrity, which might contribute to mitigating ARG persistence and dissemination; however, this remains a hypothesis that requires experimental validation (Figure 4). The rebound of ARGs during composting is a multifactorial process, in which heavy metals and EPS represent only two of several interacting drivers, together with antibiotic residues, microbial succession, and physicochemical variations within the composting environment. Future experiments could be designed to test whether electric field-assisted composting affects ARG dynamics by modulating these factors under controlled thermophilic conditions.

5. Conclusions

The rebound of ARGs after the thermophilic phase of composting has been increasingly documented and is now recognized as a significant concern. Current research suggests that the proliferation and succession of microbial communities in the curing stages of composting form the basis for ARG resurgence. To address this challenge, various strategies have been explored, including the addition of exogenous amendments, hyperthermophilic composting, prolongation of the thermophilic phase, and manipulation of microbial communities. However, the effectiveness and stability of these approaches in suppressing ARGs remain limited and inconsistent. Given that heavy metals and EPS can influence the dissemination of ARGs during composting, they may serve as critical drivers in the rebound process. Electric field-assisted composting has emerged as a promising technology that may mitigate ARG proliferation by reducing the bioavailability of heavy metals and disrupting EPS structures. Nonetheless, its practical effectiveness has yet to be fully validated. Future research should focus on elucidating the mechanisms underlying ARG rebound and developing transformative technologies, while also integrating currently available strategies to achieve optimal ARG removal efficiency.