Comprehensive Review on the Distribution, Environmental Fate, and Risks of Antibiotic Resistance Genes in Rivers and Lakes of China

Abstract

Antibiotic resistance genes (ARGs) have emerged as globally concerning environmental contaminants, posing serious threats to ecosystem health and public safety. This systematic review summarizes global research trends on ARGs across three key aspects: (i) identification and distribution in river and lake ecosystems, (ii) sources and environmental behaviors, and (iii) ecological and human health risks. Concentration data of ARGs in various rivers and lakes across China were compiled to reveal their spatial distribution patterns. The analysis of ARGs sources and environmental behaviors provides essential insights for designing effective mitigation strategies. Furthermore, this review highlights the potential ecological and human health hazards of ARGs and discusses limitations and improvement directions of current risk assessment methodologies. The main findings indicate that ARGs are widely present in rivers and lakes across China; higher abundances occur in eastern and southern regions compared with central–western and northern areas, such as 4.93 × 10²–8.10 × 10³ copies/mL in Qinghai Lake and 6.7 × 10⁷–1.76 × 10⁸ copies/mL in Taihu Lake. The environmental behaviors of ARGs are highly complex, involving multiple mechanisms and influenced by climatic conditions, nutrient levels, and additional environmental factors. Based on these findings, future efforts should prioritize long-term site-specific monitoring, evaluate their prolonged impacts on aquatic ecosystems, and develop integrated risk assessment models to support evidence-based environmental management.

1. Introduction

Antibiotics are a fundamental component of modern medicine, and China is one of the largest producers and consumers of antibiotics worldwide [1]. The misuse and overuse of antibiotics have resulted in their continuous discharge into aquatic environments [2], with median concentrations in water samples from major rivers and coastal seas in China ranging from 1.30 to 176 ng/L [3]. This directly contributes to the proliferation of ARGs in aquatic ecosystems. ARGs exhibit rapid dissemination and environmental persistence, thereby accelerating the evolution of bacterial resistance [4]. This poses substantial challenges to human health and public health systems [5]. By 2050, it is projected that 10 million people will die annually from infections caused by resistant bacteria, with cumulative costs from healthcare expenditures and productivity losses reaching an estimated USD 100 trillion [6,7]. The World Health Organization (WHO) has classified the emergence of resistant bacteria and ARGs as one of the most severe public health threats of the 21st century [8].

Rivers and lakes are vital freshwater resources, and their pollution status directly affects ecosystem health and drinking water safety. Rivers, due to their high water exchange capacity, play a crucial role in the migration and dispersion of pollutants, whereas lakes function as reservoirs and regulators, with limited hydrodynamics and extended water retention times that promote the accumulation of antibiotics and ARGs [9]. ARGs can migrate spatially with water flow and accumulate in specific regions [10]. In major rivers, lakes, and reservoirs across China, sulfonamide and tetracycline resistance genes are the most frequently detected ARGs, with concentrations in surface waters typically ranging from 10⁷ to 10¹¹ copies/L [11]. However, baseline data on the distribution of ARGs in natural waters remain poorly compiled and summarized. Through horizontal gene transfer (HGT), these elements spread among microbial populations, leading to greater abundance and diversity of antibiotic-resistant bacteria (ARB) in environmental settings. Such alterations disrupt the ecological equilibrium of aquatic systems and present a substantial threat to their microbial stability [10]. Furthermore, environmental factors such as temperature, pH, antibiotic concentration, and light intensity can significantly influence the behavior and fate of ARGs in aquatic environments [12,13].

To mitigate the impacts of ARGs on plants and animals, it is essential to conduct investigations, monitoring, and environmental risk assessments (ERA) of ARGs. These efforts are critical for establishing and strengthening ERA systems. Such measures not only improve the regulation of pollution sources but also facilitate the formulation of controlled antibiotic lists and the implementation of risk control measures under frameworks of prohibition, restriction, and discharge limitation. In recent years, China has made significant progress in regulating antibiotic pollution through a series of national policies and regulatory frameworks [14]. In particular, the Biosecurity Law of the People’s Republic of China, the National Action Plan to Contain Antimicrobial Resistance (2022–2025), and the List of Key Controlled Emerging Pollutants (2023 Edition) highlight the importance of controlling antibiotic emissions and residues, while also addressing the environmental dissemination pathways of ARGs [15]. The management of ARGs is closely related to societal security, and without stricter control, ARGs may hinder industrial upgrading and impede sustainable development. Therefore, identifying the migration, transfer, and dissemination mechanisms of ARGs, and assessing their associated risk levels can promote green transformation of industries and reduce potential threats to human health.

Examining ARGs within river—lake ecosystems is essential because of their potential ecological and human health implications. Such investigations enhance our understanding of ARGs’ distribution patterns and environmental behavior, while also quantifying their potential risks to natural systems and public health. The resulting knowledge provides a scientific foundation for designing and implementing effective strategies for risk assessment and management. It also facilitates the identification of polluted regions and sources, and elucidates the mechanisms of ARGs’ environmental migration, dispersion, and fate. These findings provide valuable theoretical support for the screening, treatment, and control of ARGs, enabling the formulation of more effective strategies to safeguard the environment and human health.

2. Distribution of ARGs in Lakes and Rivers

In China’s freshwater ecosystems, the geographical distribution of ARGs exhibits pronounced spatial heterogeneity, a pattern shaped by regional differences in human activities, aquaculture practices, eutrophication levels, and climatic factors such as temperature and precipitation. Numerous studies have demonstrated that rivers and lakes in eastern and southern China harbor higher abundances of ARGs compared with those in central, western, and northern regions [16]. The specific abundances of ARGs are presented in

Table 1. Abundance of ARGs in different lakes and rivers.

Lake/River	Abundance
Taihu Lake [17]	6.7 × 107–1.76 × 108 copies/mL
Chaohu Lake [18]	3.5 × 103–2.63 × 105 copies/mL
Pearl River Basin [19]	10−1–105 copies/mL
Poyang Lake [20]	2.26 × 101–1.50 × 108 copies/mL
Yellow River [21]	0–3.6 × 106 copies/mL
Qinghai Lake [22]	4.93 × 102–8.10 × 103 copies/mL
Luoma Lake [10]	2.87 × 103–2.94 × 104 copies/mL
East Dongting Lake [23]	0–4.88 × 104 copies/mL
Nanhu Lake [24]	9.4  ×  107 copies/mL
Xuanwu Lake [24]	2.37  ×  108 copies/mL
Zixia Lake [24]	4.42  ×  107 copies/mL

, and their quantitative spatial distribution is illustrated in Figure 1.

In eastern China, particularly the Yangtze River Delta region—including Taihu Lake, Chaohu Lake, and their adjacent river systems—has consistently been recognized as a hotspot for ARGs. This pattern is attributed to antibiotic inputs from wastewater discharge, high population density, intensive aquaculture activities, and elevated levels of eutrophication and organic pollution. Zhao et al. reported that the relative abundance of ARGs in the surface water of Lake Taihu ranged from 6.7 × 10⁷ to 1.76 × 10⁸ copies/mL [17], with a higher proportion associated with resistance genes against β-lactams, tetracyclines, sulfonamides, and macrolides. Jiang and colleagues reported that Chaohu Lake contained comparatively elevated levels of ARGs, with multidrug, bacitracin, and polymyxin resistance genes representing the predominant categories [18]. Furthermore, multiple investigations have revealed that both ARGs and antibiotic concentrations in the downstream sections of the Yangtze River are typically greater than those observed in upstream regions [25]. In southern China, characterized by a subtropical climate and intensive aquaculture, elevated ARG levels have also been observed [26]. Guo et al. quantitatively analyzed intracellular ARGs (iARGs) and extracellular ARGs (eARGs) in river water from the Pearl River Basin, reporting concentration ranges of 10⁻¹ to 10⁴ copies/mL and 10⁻¹ to 10⁵ copies/mL, respectively, with resistance genes such as sul, tet, and erm constituting a considerable proportion [19]. Moreover, due to the absence of cellular protection, eARGs are more susceptible to HGT, posing potential risks for the dissemination of environmental resistance genes and public health. Wang et al. investigated Poyang Lake, which is influenced by agricultural activities, livestock farming, and urban wastewater discharge, and reported absolute abundances ranging from 2.26 × 10¹ to 1.50 × 10⁸ copies/mL, with sul1, sul2, tetA, and tetM identified as the dominant genes [20].

In the rivers and lakes of central and northern China, ARG contamination also exists, although its concentrations are generally lower [20]. In the Haihe River Basin of northern China, resistance genes such as sul1 and sul2 have been detected, while tetracycline resistance genes were absent [27]. Studies have shown that rising temperatures are positively correlated with ARG abundance [28]; thus, lower temperatures and reduced livestock intensity may contribute to the overall lower ARG levels in these regions compared with eastern and southern China. Zhang et al. investigated the occurrence of ARGs in the Yellow River and observed absolute abundances varying between 0 and 3.6 × 10⁶ copies/mL [21]. Compared with the lower reaches of the Yangtze River, relative abundances were generally lower, while seasonal variation showed higher overall levels during summer than in winter [21]. Due to limited industrial activity and low population density, the occurrence of ARGs in western China was once considered negligible; however, recent studies have revealed that even remote areas may be affected by ARG contamination via long-distance waterborne transport, migratory birds, or antibiotic pollution associated with imported feed and animal products [29]. These regions urgently require more comprehensive monitoring. Jia et al. investigated ARG abundances in typical rivers within the Qinghai Lake Basin and reported that tetracycline genes ranged from 4.93 × 10² to 8.10 × 10³ copies/mL, while genes such as intI1 and sul1 exhibited relatively low abundances [22].

3. Environmental Fate of ARGs in Lakes and Rivers

3.1. Source of ARGs

The widespread application of antibiotics across healthcare, agricultural practices, aquaculture, and livestock farming has been shown to directly promote the development of microbial resistance by targeting bacterial infections for prevention and therapy [30,31]. ARGs enter freshwater systems through multiple pathways, including treated and untreated wastewater discharges, agricultural runoff, and the release of antibiotics and their by-products (Figure 2).

Studies have demonstrated that antibiotic resistance is a natural phenomenon that predates the selective pressures imposed by modern clinical antibiotic use. Analyses of ancient DNA have revealed the presence of ARGs in permafrost sediments dating back approximately 30,000 years, including genes conferring resistance to β-lactams, tetracyclines, and glycopeptides [32]. In contemporary environments, urban wastewater treatment plants (UWTPs) are considered the primary anthropogenic sources of antibiotics, ARGs, and ARB entering ecosystems. The biological treatment processes within UWTPs create favorable conditions for resistance selection and dissemination, as microbial communities are continuously exposed to sub-inhibitory concentrations of antibiotics [33]. In livestock and poultry farming, antibiotics are routinely used to promote growth, prevent diseases, and treat infections. For instance, tetracyclines are frequently administered; however, their low absorption efficiency in animals results in the accumulation of active residues in tissues, urine, and feces, whereby some ARGs enter the food web while others are discharged into river and lake systems through fecal excretion [34]. In agricultural practices, wastewater is often used for irrigation, and animal manure is widely applied as fertilizer. These practices introduce antibiotics, ARGs, as well as various inorganic pollutants, organic contaminants, and pathogens into soils. Consequently, antibiotics may enter the environment through the application of manure and liquid fertilizers, direct deposition by grazing livestock, or discharges of untreated wastewater [35]. Aquaculture represents another significant pathway through which antibiotics are introduced into aquatic environments. Globally, antibiotics are widely employed as therapeutic agents and feed additives to control fish diseases. Driven by strong market demand, the quantities of antibiotics used in aquaculture often exceed those applied in human medicine. Such practices promote the emergence and dissemination of ARGs among aquaculture-associated pathogens. The prolonged and excessive use of antibiotics, coupled with their environmental persistence, inevitably results in elevated levels of antibiotic residues in aquaculture systems. These residues can be further redistributed between water and sediments, forming long-term reservoirs of resistance determinants and posing risks to ecological integrity and public health [36].

3.2. Migration and Dissemination of ARGs

As antibiotics progressively accumulate in environmental settings, they impose selective stress on microbial communities, driving the proliferation of ARB and ARGs. These genetic determinants are propagated through both HGT and vertical gene transfer (VGT), thereby amplifying the environmental threat posed by antibiotic resistance [37].

3.2.1. Horizontal Gene Transfer

HGT is a central process driving the dissemination, amplification, and recombination of ARGs among different bacterial species. Bacteria achieve HGT through multiple mechanisms, including transformation, transduction, and conjugation [38], while mobile genetic elements (MGEs), acting as carriers and facilitators, accelerate rapid evolution and adaptation. MGEs include conjugative components such as plasmids and integrative conjugative elements (ICEs), as well as transposable elements like transposons and integrons, in addition to bacteriophages [39]. These elements carry a diverse repertoire of functional genes that contribute to antimicrobial and heavy metal resistance, pathogenicity, and metabolic processes.

Bacteriophages consist of nucleic acids enclosed within a protein capsid, capable of infecting bacteria or archaea and replicating within them [40]. During integration and excision within host genomes, phage DNA can exchange with host chromosomal sequences and potentially excise adjacent bacterial genes, thereby mediating HGT and highlighting the pivotal role of bacteriophages in driving microbial evolution [41,42]. Plasmids are extrachromosomal replicons, either circular or linear, that can be disseminated through conjugation [43]. Notably, diverse genes can be efficiently transferred among bacteria via conjugation, making this process a major driving force for rapid bacterial evolution and adaptation. Multiple types of chromosomally integrated mobile genetic elements (ciMGEs) have been characterized, encompassing integrative conjugative elements (ICEs), integrative mobilizable elements (IMEs), and cis-mobilizable elements (CIMEs) [44]. ICEs, also referred to as conjugative transposons, are modular MGEs that integrate into host genomes and proliferate passively through chromosomal replication and cell division. Induction of ICE gene expression triggers excision from the chromosome and assembly of a conserved conjugation machinery (type IV secretion system), enabling DNA transfer into suitable recipient cells [45]. IMEs resemble ICEs but cannot complete transfer independently. IMEs are capable of inserting into host chromosomes but lack the essential genes required for self-mediated conjugation. As a result, their mobilization and transfer to other bacterial cells rely on helper conjugative elements, such as plasmids or ICEs. CIMEs are flanked by recombination sites, attL and attR, which can be recognized by integrases encoded by related ICEs or IMEs; nevertheless, CIMEs themselves do not encode the machinery necessary for conjugation or recombination [46]. Transposons, in contrast, are mobile genetic elements that relocate between replicons via transposase activity. Although they cannot move directly between cells, their incorporation into plasmids or ICEs allows them to be transmitted to other bacterial hosts [47]. These diverse mechanisms of gene transfer not only facilitate the rapid dissemination of ARGs but also enhance bacterial adaptability to environmental changes, thereby posing potential threats to public health and ecosystem integrity.

3.2.2. Vertical Gene Transfer

VGT complements HGT to some extent, jointly facilitating the dissemination of ARGs. Unlike HGT, which involves the transfer of genetic material between different bacterial species, VGT occurs during microbial reproduction, transmitting resistance genes from mother cells to daughter cells through cell division [48]. This process is essential for the transmission of resistance genes within a single species, particularly in environmental microbial communities. VGT ensures the long-term inheritance of resistance genes within microbial populations, leading to the gradual accumulation of antibiotic resistance in these communities. This mechanism contributes to the stability of microbial populations, enabling some environmental microorganisms to retain resistance traits even in the absence of antibiotic selective pressure, thereby ensuring their survival under adverse conditions [49].

3.2.3. Impact of Pollutants on the Migration of ARGs

Antibiotics play a pivotal role in shaping the environmental propagation of ARGs. One major pathway involves antibiotic-induced oxidative stress, which has been widely recognized as a central factor in stimulating ARGs mobilization [50]. For example, exposure to tetracycline increases the production of intracellular reactive oxygen species (ROS), disrupts bacterial membrane integrity, and facilitates plasmid translocation across cell membranes. Empirical evidence further indicates that when antibiotics are present at sub-inhibitory levels—for instance, tetracycline ranging from 3.9 to 250 ng/mL—they can not only accelerate conjugative transfer but also boost plasmid dissemination by activating conjugation-associated genes, including trbBP and trfAP [51]. Similarly, even minimal concentrations of sulfonamides have been shown to enhance plasmid exchange from Escherichia coli to genera such as Pseudomonas and Salmonella, thereby amplifying the environmental spread of resistance determinants [52].

Agricultural contaminants substantially contribute to the spread of ARGs. According to Malagón-Rojas and colleagues, pesticides can intensify bacterial resistance through several pathways, such as stimulating efflux pump activity, reducing the permeability of outer membrane porins to antimicrobial agents, and triggering genetic mutations [53]. Evidence also indicates that the widespread application of plant growth regulators (PGRs) leads to measurable residues in edible crops, with concentrations in fruits and vegetables reported between 21.18 and 51.20 mg/kg. Moreover, sub-environmental levels of PGRs, typically ranging from 0.5 to 10 μg/L, including indole-3-acetic acid (IAA) and ethephon (ETH), have been shown to enhance horizontal plasmid transfer between Escherichia coli and Enterococcus faecalis. Mechanistically, IAA facilitates this process by inducing oxidative stress and disrupting bacterial membrane stability, while ETH augments conjugative plasmid exchange by intensifying cellular stress responses [54].

Environmental contamination with heavy metals imposes selective pressures that foster both the expression and propagation of ARGs within microbial populations. Several metals share common resistance pathways with antimicrobial agents; for example, bacteria can mitigate metal toxicity by activating specialized efflux systems or by modifying the permeability of their cell envelopes, strategies that closely parallel mechanisms of drug resistance [55]. As a result of these overlaps, the simultaneous presence of metals and antibiotics can intensify ARG enrichment and distribution, ultimately strengthening microbial adaptability to a wide spectrum of pollutants and pharmaceuticals [56]. This interaction also promotes HGT, further amplifying the environmental spread of resistance determinants [57]. Collectively, these insights highlight the necessity of coordinated strategies addressing both antibiotic use and pollutant control to mitigate ARG proliferation in natural ecosystems.

3.2.4. Impact of Environmental Factors on the Migration of ARGs

Numerous studies have observed that essential metabolic and behavioral genes, including ARGs, undergo positive selection to adapt to fluctuating environments [58,59]. In addition to positive selection, environmental stress can facilitate HGT, which serves as a critical pathway for the evolution of novel resistant strains [60]. Climate change exerts broad impacts on human health, including heat-related mortality, food insecurity, and reduced crop yields, increased transmission of infectious diseases, and diverse health outcomes triggered by extreme weather events such as floods and droughts [61]. Moreover, the influence of nutrients on ARGs has emerged as a research hotspot in recent years. This study focuses on the effects of climate change (temperature and precipitation) and nutrients on the behavior of ARGs.

Elevated temperatures accelerate microbial metabolic rates, thereby increasing the frequency of genetic mutations and HGT. Under conditions of heightened metabolic activity, microorganisms more readily acquire ARGs, enhancing the potential for resistance dissemination within microbial communities [62]. MacFadden et al. demonstrated that the rise in antibiotic resistance is associated with increases in average minimum temperatures [63], which are continuously elevated by climate change. With the growing incidence of infections and the spread of resistant pathogens, the intensification of climate change will inevitably result in a further rise in the prevalence of resistant pathogens. Moreover, elevated temperatures facilitate the process of HGT. For instance, one study reported that conjugative transfer induction at 37 °C was 4.80 times higher than at 17 °C [62]. This temperature-dependent mechanism of gene transfer is particularly pronounced in aquatic ecosystems, where ARG levels in summer are markedly higher than in winter [64]. Temperature not only regulates microbial activity and HGT but also substantially influences the spatial distribution of ARGs. Studies conducted in Qinghai Lake and Poyang Lake revealed that ARGs abundance in Qinghai Lake is considerably lower than in Poyang Lake, likely due to its location on the Qinghai–Tibet Plateau, where high altitude, cold climate, and low water temperatures restrict human activity, microbial metabolism, and HGT rates [22,26].

With global warming, the water-holding capacity of the atmosphere increases exponentially, implying that storms will become more intense and accompanied by heavier precipitation. Increased precipitation may result in flooding, flood-related infections, population displacement, refugee crises, and overcrowding [65]. Seasonal runoff facilitates the long-distance migration of ARGs, while flooding may exacerbate the spread of waterborne infections through contaminated sewage overflows or pollution associated with livestock farming. Studies have shown that organic compounds can enhance antibiotic resistance [66]; therefore, during severe flood events triggered by climate change, the co-migration of organic compounds and ARGs from soils into aquatic systems may accelerate their dissemination in aquatic environments. Flood-induced eutrophication may be intensified by increased precipitation, thereby enhancing antibiotic resistance and potentially facilitating the spread of resistant pathogens and ARGs [67]. Extreme weather events that trigger flooding may severely damage fragile sanitation infrastructure, exacerbate overcrowding in densely populated areas, and promote the spread of antibiotic resistance through sewage, which is a known reservoir of ARGs [68]. In addition to flooding, extreme weather events may also induce drought in certain regions. During drought periods, elevated concentrations of waterborne contaminants may promote the aggregation of resistant microorganisms. Water scarcity can undermine sanitation facilities and force larger populations to share limited water resources, thereby creating conditions conducive to waterborne infectious disease outbreaks. Water shortages are often accompanied by food insecurity, which may further exacerbate malnutrition and increase the incidence of diarrheal diseases [69].

Nutrients, including nitrogen, phosphorus, and carbon sources, are considered essential for conjugative transfer, as both the transfer of ARGs from donors and their acquisition by recipients require energy; appropriate concentrations of N, P, and C sources can enhance conjugation by providing the necessary transfer energy. High nutrient concentrations can promote the growth of microalgae and bacteria. It has been reported that increased carbon fixation was observed in algal–bacterial systems containing exogenous ARG plasmids carrying sul1, as the adsorption and capture of exogenous plasmids require additional energy, thereby providing strong support for the biological process of plasmid sequestration jointly mediated by algae and bacteria [70]. Guo et al. examined how sulfonamide resistance genes (sul1, sul2) and microbial assemblages responded to different concentrations of sulfamethoxazole (SMX) under varying wastewater chemical oxygen demand (COD) [71]. Their results demonstrated that COD availability influenced the impact of SMX on resistance genes; when COD was inadequate, bacterial growth was suppressed, which restricted the expansion of sul1 and sul2. One possible explanation is that compounds contributing to COD serve as nutrients for bacteria, and even under the selective pressure of SMX, organisms cannot grow normally without an adequate nutrient supply. When nutrient levels exceed bacterial requirements, growth may restrict the promotion of ARGs.

In summary, climate change exacerbates the migration and dissemination of ARGs through multiple pathways. Rising temperatures directly enhance bacterial growth and the rate of HGT, while extreme weather events accelerate the dispersal of ARGs through the mobilization and mixing of pollutants. Moreover, climate change may indirectly influence the distribution patterns and transmission pathways of ARGs by altering microbial community composition and ecosystem functions. The concentration and type of nutrients also exert significant effects on the behavior of ARGs; appropriate nutrient levels can enhance HGT, whereas nutrient deficiency or excess may suppress ARG dissemination.

3.3. Persistence of ARGs

The final fate of ARGs is central to understanding their ecological risks and formulating effective management strategies. In general, the final fate of ARGs can be categorized into three major outcomes: degradation and elimination, stabilization within environmental matrices, and integration and dissemination. These processes collectively determine the persistence, bioavailability, and dissemination potential of ARGs in natural ecosystems (Figure 3).

Extracellular DNA (eDNA), originating from microbial cell lysis or active secretion, represents a major form of ARGs in the environment [72]. ARGs associated with eDNA can be degraded by nuclease activity, microbial metabolism, and physicochemical factors, thereby reducing their abundance and ecological activity. Environmental nucleases, such as deoxyribonuclease I (DNase I), fragment DNA molecules, thereby impairing the structural integrity and functional potential of ARGs [73]. Certain microorganisms utilize eDNA as a source of carbon and nitrogen, metabolically degrading ARG-containing fragments and further diminishing their stability. Ultraviolet radiation, oxidants (e.g., hydrogen peroxide), and other environmental stressors accelerate DNA damage and degradation, thereby enhancing the breakdown of ARGs within eDNA [74]. Collectively, these processes act as natural attenuation mechanisms that degrade and eliminate ARGs in the environment.

Despite the existence of degradation pathways, ARGs can remain stabilized within environmental matrices, thereby prolonging their ecological half-life and sustaining a potential reservoir of resistance genes. Stabilization can occur when eDNA-associated ARGs adsorb onto clay minerals, humic substances, and organic colloids, which reduces nuclease accessibility and protects ARGs from rapid decay [75,76]. Within biofilms, extracellular polymeric substances (EPS) form physical barriers that shield ARGs from environmental stress while simultaneously promoting HGT among resident microorganisms. In addition, other persistent pollutants impose long-term selective pressures that further facilitate the retention of ARGs. These stabilization processes extend the persistence of ARGs, maintaining a reservoir that can be reactivated under favorable conditions or selective pressures, thereby posing sustained ecological and public health risks.

In addition to degradation and stabilization, ARGs may achieve long-term persistence by integrating into microbial genomes and disseminating within microbial communities. As noted previously, ARGs can integrate into bacterial chromosomes through recombination or mobile genetic element–mediated HGT; once integrated, they are stably inherited through VGT during host cell division. ARGs can also rely on conjugative plasmids to spread rapidly across different taxonomic groups, thereby expanding their host range [43]. Integrons, transposons, and integrative ICEs are capable of capturing, rearranging, and disseminating ARGs within microbial communities [45]. Meanwhile, bacteriophages facilitate the transfer of ARGs across species and environmental boundaries through transduction [41]. The integration and dissemination of ARGs ensure that they are not merely transient pollutants but persistent features of microbial ecosystems, thereby sustaining environmental resistomes and increasing the risk of resistance gene transfer to pathogens, which exacerbates public health challenges.

4. Potential Risks of ARGs in Lakes and Rivers

4.1. Ecological Risks of ARGs

With the escalating issue of antibiotic pollution, elucidating its potential impacts on ecosystems and accurately assessing ecological risk levels have become pressing challenges in environmental science.

Contemporary risk assessment frameworks, such as probability density functions, the Nemerow multi-factor index, and the potential ecological risk model, have been applied in monitoring antimicrobial resistance [77]. However, due to the absence of baseline inventories of resistance genes, these paradigms exhibit limited predictive validity in quantifying ARG-related risks [77]. This highlights the urgent need to establish a comprehensive and accurate ARGs database, as defining baseline levels is essential to provide reliable benchmarks for risk assessment and to more precisely quantify the ecological risks of ARGs. Within this methodological framework, multiple ecological risk assessment approaches have been proposed (

Table 2. Comparison of commonly used risk assessment methods for ARGs.

Method	Principle	Advantages	Limitations
RQ [81]	Compares measured environmental concentration with a PNEC.	Simple and data-efficient, suitable for rapid large-scale screening.	Does not directly reflect human health risks; high uncertainty in the PNEC for ARGs.
SSD [82]	Uses dose–response data from multiple species to derive hazard thresholds.	Considers species differences; provides probabilistic risk estimates.	Lack of standardized toxicity data for ARGs limits applicability.
PERA [84]	Evaluates ARG mobility, host pathogenicity, and transfer potential.	Integrates biological behavior and environmental factors.	Data-intensive, model-sensitive, and computationally complex.
QMRA [85]	Predicts infection risk based on exposure pathways and dose–response models.	Directly linked to human exposure and disease burden.	Data-deficient, assumption-dependent, and uncertainty-prone.

). Risk quotient (RQ) analysis serves as a straightforward metric that relies solely on ecotoxicological datasets to allocate preliminary ecological risks, which is particularly valuable under current constraints in toxicokinetics and environmental baselines of ARGs [78]. Its principle involves comparing the measured environmental concentration (MEC) of antibiotics with the predicted no-effect concentration (PNEC), thereby enabling rapid determination of risk levels through simple numerical calculations and providing a direct basis for preliminary ecological risk judgments. For instance, Chen et al. applied the RQ method to assess the environmental risks of antibiotics in the Huangpu River and found that only SMX had an RQ > 0.1, suggesting a moderate risk to daphnia [79]; Zhou et al. reported that SMX exhibited RQ > 1 in Taihu Lake, posing severe risks to aquatic ecosystems, while tetracyclines and quinolones such as chlortetracycline, doxycycline, oxytetracycline, ciprofloxacin, enrofloxacin, ofloxacin, and norfloxacin were also reported with RQ > 1, indicating high environmental risks [80]. Zhou et al. further surveyed shallow lakes in the middle and lower reaches of the Yangtze River and found that sulfamethoxazole, erythromycin, and ofloxacin in surface waters posed moderate risks to algae or bacteria in aquatic ecosystems. In summary, the RQ approach is advantageous for its simplicity but exhibits limitations when assessing the combined effects of multiple antibiotics [81]. In contrast, the species sensitivity distribution (SSD) method integrates toxicity response data of multiple organisms to specific antibiotics and constructs sensitivity distribution curves to reflect the susceptibility of the entire biological community [82], thereby yielding a more comprehensive and scientific assessment. A previous study [83] that combined toxicity data of ETM, TC, NOR, and SMX from Chinese coastal estuaries found significant differences in antibiotic toxicity across aquatic taxa, with ETM and NOR being more toxic while TC and SMX exhibited relatively lower toxicity. Nevertheless, the implementation of SSD also faces challenges, as it requires extensive high-quality toxicity data; when data are insufficient or inconsistent, the uncertainty increases substantially, thereby undermining the accuracy and reliability of the assessment. Therefore, when applying the SSD method, it is essential to emphasize data collection, curation, and quality control to ensure the scientific validity and credibility of the assessment outcomes.

Overall, research on the ecological impacts of combined antibiotic pollution remains in the exploratory stage, necessitating multidisciplinary integration to advance ERA and management strategies, thereby enhancing source identification, risk prediction, and mitigation capacity. In their investigation of the ecological impacts of combined antibiotic pollution in the Haihe River system, Chen et al. employed the concentration addition prediction model proposed by Altenburger et al. and found that the joint exposure of TC and ENR posed synergistic risk effects to aquatic organisms [86]. These findings not only reveal the potential ecological hazards of combined antibiotic pollution but also provide new perspectives and methodologies for composite pollution risk assessment. Furthermore, the probabilistic ecological risk assessment (PERA) method offers an effective analytical framework for assessing ecological risks arising from the interactions of multiple pollutants in complex environments. This approach, grounded in probabilistic statistics, models the exposure concentrations of different pollutants and their corresponding ecological effect data as independent observations, incorporating statistical distribution features and significance analysis into risk inference. The advantage of PERA lies in its ability to provide quantitative risk evaluations along with uncertainty intervals for risk levels; although it demands high data quality and large sample sizes and involves relatively complex analyses, its outcomes are more scientifically robust and explanatory [84]. With the continuous advancement of relevant technologies and the growing availability of data resources, PERA is expected to play an increasingly important role in the ecological risk assessment of combined antibiotic pollution.

4.2. Human Health Risks of ARGs

Humans may be exposed to ARGs through multiple pathways, including drinking contaminated water, bathing, engaging in water-related recreational activities, occupational exposure during agricultural irrigation, and consuming crops irrigated with reclaimed water. Hooban et al. demonstrated in their study on drinking water that the transfer of ARGs from freshwater environments to clinical settings poses significant health risks, as clinically relevant resistance genes were detected in drinking water samples [87]. The health risks posed by ARGs vary depending on multiple factors, including host pathogenicity, genetic background, and the likelihood of transfer to human pathogens. The environmental potential pathways of ARGs and their associated health impacts are illustrated in Figure 4.

Evaluating the potential health risks posed by ARGs is inherently complex. To address this, Zhang and colleagues developed a risk assessment framework incorporating a decision-tree methodology that emphasizes three main criteria: (i) the degree of enrichment in human-associated environments, (ii) the mobility of the gene, and (iii) whether the host organism is pathogenic [88]. This simplified framework generates four risk categories. ARGs that do not meet the first criterion are assigned to Category IV, indicating human-unrelated ARGs that are least likely to threaten human health. Those meeting the first but not the second criterion are assigned to Category III, representing non-mobile ARGs in human-associated environments that are unlikely to introduce new resistance in pathogens. ARGs that meet the first two but not the third criterion are assigned to Category II, indicating a high risk of emerging resistance, though not yet observed in pathogens. ARGs that satisfy all three criteria are classified as Category I, representing the highest risk due to the emergence of novel or multidrug resistance in pathogens. However, this study is based on genotypic data rather than phenotypic evidence for all ARGs, which may lead to underestimation or misjudgment of their actual threats. Amarasiri et al. [85] evaluated and quantified health risks associated with ARB and ARGs in aquatic ecosystems, proposing the quantitative microbial risk assessment (QMRA) method. Nevertheless, limited information exists regarding ARB and ARG exposure in diverse aquatic environments, and specific dose–response models for ARB infections are yet to be developed. Zhang et al. developed a health risk assessment framework that evaluates each ARG based on human accessibility, mobility, pathogenicity, and clinical relevance. Based on these four indicators, the authors defined a “Risk Index (RI)” to quantify the health risk of each ARG [89]. Results indicated that 23.78% of ARGs had RI > 0, suggesting potential health risks, particularly those associated with multidrug resistance. Zhu et al. evaluated health risks associated with ARGs by considering several factors, including gene abundance, relevance to clinical treatment, potential pathogenicity in humans, human exposure pathways, and the mobility of the genes [90]. However, the study did not specify how these indicators should be weighted. This indicates that although current assessment frameworks have made notable progress in both theory and practice, further refinement is necessary to enhance their scientific rigor and practical applicability.

Although notable progress has been made in assessing the health risks of ARGs, current evaluation methods still exhibit significant limitations. Future research should further refine the assessment frameworks by integrating more comprehensive data and information to enhance the accuracy and reliability of the outcomes. Moreover, additional experimental and monitoring data are required to support the practical application of these evaluation methods, thereby providing a robust scientific basis for developing effective policies and management strategies.

5. Conclusions

ARGs have emerged as critical environmental contaminants with profound implications for ecosystem health and public safety. This review consolidates current insights regarding the distribution, environmental behavior, and associated risks of ARGs. Available evidence demonstrates that ARGs are widely present in aquatic ecosystems, with their prevalence shaped by a combination of factors, including human activities, patterns of land use, and climatic conditions. The environmental behavior of ARGs is complex, involving HGT, VGT, microbial metabolism, and physicochemical processes, which collectively determine their persistence, bioavailability, and dissemination potential. These mechanisms not only enhance the environmental mobility of ARGs but also magnify their ecological and health risks. In terms of risk assessment, existing frameworks such as RQ analysis, SSD, and PERA have made progress in evaluating the ecological risks of ARGs, but limitations remain. These methods may underestimate or misjudge the actual threats posed by ARGs, particularly when assessing the combined effects of multiple antibiotics and their long-term ecological impacts. Therefore, future research should further refine assessment frameworks by integrating more comprehensive data and information to improve the accuracy and reliability of evaluation results. Moreover, current studies often focus on single environmental compartments or short-term dynamics, while the understanding of long-term fate, cross-compartment transport, and the interactive effects of climate change and nutrient loading remains limited.

Future efforts should prioritize (i) assessing the long-term impacts of ARGs on ecosystem structure and function through long-term monitoring and experimental studies, (ii) integrating multidisciplinary data to develop more robust and practical risk assessment models, particularly for the combined effects of multiple pollutants, and (iii) formulating and implementing effective policies and management strategies to reduce the misuse and overuse of antibiotics and control the environmental release of ARGs. In conclusion, the extensive presence and intricate dynamics of ARGs represent a considerable risk to both environmental integrity and human health. Through integrative research and multidisciplinary collaboration, a better understanding and management of ARG-related risks can be achieved, thereby safeguarding ecosystems and human health.