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Article

SARS-CoV-2, Noroviruses, Adenoviruses, and Antibiotic-Resistant Coliforms Within Chilean Rural Wastewater Treatment Plants

by
Angela Plaza-Garrido
1,
Cristina A. Villamar-Ayala
1,2,*,
Manuel Ampuero
3 and
Aldo Gaggero
3
1
Departamento de Ingeniería en Obras Civiles, Facultad de Ingeniería, Universidad de Santiago de Chile (USACH), Av. Víctor Jara 3659, Estación Central, Santiago 9170022, Chile
2
Programa para el Desarrollo de Sistemas Productivos Sostenibles, Facultad de Ingeniería, Universidad de Santiago de Chile (USACH), Av. Víctor Jara 3769, Estación Central, Santiago 9170022, Chile
3
Laboratorio de Virología Ambiental, Núcleo Interdisciplinario de Microbiología, Instituto de Ciencias Biomédicas (ICBM), Facultad de Medicina, Universidad de Chile, Santiago 8380453, Chile
*
Author to whom correspondence should be addressed.
Water 2025, 17(22), 3197; https://doi.org/10.3390/w17223197
Submission received: 12 September 2025 / Revised: 22 October 2025 / Accepted: 5 November 2025 / Published: 8 November 2025
(This article belongs to the Section Wastewater Treatment and Reuse)
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Abstract

Wastewater-based epidemiology (WBE) is an effective tool for assessing health risks in rural areas with limited access to health care. Wastewater treatment plants (WWTPs) allow for the monitoring of pathogenic microorganisms, which is key to detecting viral integrity and bacterial viability to assess health risks. This study evaluated five rural WWTPs in Chile during 2022 in two seasons (autumn–winter and spring–summer). SARS-CoV-2, norovirus GI/GII, and HAdV-F40/41 was analyzed, along with antibiotic-resistant coliforms. Influent and effluent samples were used, with viral integrity analysis by propidium monoazide and culture methods to assess bacterial resistance. Despite the low number of clinical cases, SARS-CoV-2 was detected in all influent samples. Intact viral particles of NoV GI (78%), NoV GII (72%), and HAdV-F40/41 (65%) were found. This suggests that they may still be infectious. Viral removal ranged from 74% to 100%, although intact HAdV was detected in effluent (6.2%). Coliforms resistant to various antibiotics were detected and partially removed (22–100%). Removal efficiency depends on the type of treatment and the season of the year. WWTPs act as temporary reservoirs of infectious agents. This study reinforces the usefulness of WBE in rural contexts and WWTPs as barriers or not to these contaminants to the environment.

1. Introduction

One of the most significant global concerns today is the emergence of new pandemics, the proliferation of antibiotic-resistant bacteria, and the increasing scarcity of freshwater resources [1]. These three challenges converge in wastewater, as numerous sectors are actively exploring the reclamation of treated/not treated wastewater as a partial solution to water scarcity [2]. Consequently, effective wastewater management has become a key component in ensuring the sustainable use and protection of water resources [3]. Untreated or poorly treated wastewater that is discharged can cause serious environmental and health problems. For this reason, it is vitally important to study and determine the microbiological aspects of wastewater to understand its composition and the real impact it can have on the environment and health.
Wastewater contains various microorganisms, including viruses and pathogenic bacteria, which pose a potential risk to public health [4]. Some of the viruses that can be found in wastewater and could pose a danger are SARS-CoV-2, Norovirus genogroups I and II (NoV GI/GII), Adenovirus F 40/41 (HAdV-F 40/41), Rotavirus, Hepatitis A Virus, and Enterovirus, many of which are associated with gastrointestinal and systemic diseases [5,6]. In the case of bacteria present in wastewater, they can vary depending on the geographical region and the characteristics of the wastewater entering the different wastewater treatment plants (WWPTs). For instance, in Berlin, Germany, Firmicutes and Proteobacteria dominate the microbial communities [7]. Among the Proteobacteria, several genera, including Escherichia, Klebsiella, and Enterobacter, are considered coliforms and are used as indicators of fecal contamination [8].
In the case of viruses, the wastewater approach was enhanced during the COVID-19 pandemic, where SARS-CoV-2 was detected in wastewater in many countries, with concentrations reaching up to 107 GC/L [9,10]. In Chile, values between <5 and 462 GC/mL have been reported in the influent during the pandemic period in 2021 [11]. Although the virus is enveloped and less stable in the environment, its detection in wastewater is a valuable tool for population-level surveillance through wastewater-based epidemiology (WBE) [12]. However, other viruses can be detected in wastewater and may pose a potential risk to public health. Notably, NoV GI/GII are among the most frequently detected enteric viruses in wastewater and represent a major cause of gastroenteritis worldwide [13]. The presence of Nov GI and GII in Chilean rural areas varies between 10 and 28 and 0 and 75 GC/mL, respectively [11], although limited data exist on viral integrity and potential infectivity. Recent studies using PMA (propidium monoazide) have shown that a significant fraction (34%) of detected NoV may remain infectious [14].
Similarly, HAdVs, especially types F40 and 41, are frequently found in wastewater and have been linked to both gastroenteritis and recent cases of hepatitis in children [15]. A high prevalence of HAdVs has been observed in treated wastewater (76.96%), which may be due to HAdV resistance to a particular type of wastewater treatment [16,17].
Studies have reported that free chlorine and monochloramines are selective in the viral removal of enteroviruses and adenoviruses, respectively, requiring adjustments in contact time and doses [18]. On the other hand, coliform removal (<1000 MPN/100 mL) requires between 4 and 40 mg/L of chlorine (NaClO or Ca(ClO)2) by 30 min [19]. In rural contexts, these requirements are often not met, which increases the risk of coliform and organochlorine compound discharge [20,21]. Therefore, there is a potential risk of coliform viruses and bacteria being discharged into the environment or entering the food chain from wastewater reclamation in irrigation. Moreover, WWTPs are also recognized as reservoirs for antibiotic-resistant bacteria (ARB) due to the co-occurrence of coliforms and antibiotics in wastewater [22]. In domestic wastewater aerosols, there is evidence of ARB to amoxicillin, ciprofloxacin, and tetracycline [23]. ARBs, such as E. coli and Enterococcus spp., have been found in effluents with resistance rates as high as 100% to certain antibiotics [24,25]. Despite reported removal efficiencies of up to 99%, resistant strains are still discharged into the environment [26].
A study determined the presence of E. coli and Enterococci in wastewater, where the effluent presented 104 CFU/100 mL, of which 90% are resistant to antibiotics [25]. These results indicate that they should consider resistant isolates that are disposed of to the environment via effluents. Therefore, the presence of ARB is important not only for public health, but also for ecological aspects, as ARB can be transferred to the environment [27].
Although microbiological wastewater (viruses and bacteria) is relevant due to its impact on the environment and public health, this issue is even more relevant in rural areas where water can be a scarce commodity. Rural areas are trying to reclaim treated wastewater, even for irrigation in recreational areas. Approximately 12% of Chile’s population resides in rural areas, but only 20% have access to wastewater treatment systems [28].
These areas usually have small WWTPs, which are based on technologies, such as activated sludge, artificial wetlands, or vermifilters [29].
This study aimed to detect and quantify SARS-CoV-2, HAdV-F 40/41, and NoV GI/GII, including assessing viral integrity, in domestic wastewater from rural Chilean areas. In addition, total and fecal coliforms resistant to selected antibiotics (amoxicillin, ciprofloxacin, cefadroxil, and azithromycin) were quantified to evaluate the microbiological quality of wastewater effluents and their potential risk to public health. Moreover, determine current microbiological components and their potential risk to the environment and public health.

2. Materials and Methods

2.1. Study Area and Monitoring

The wastewater sampling sites selected for this study were rural zones from the two Chilean regions (Metropolitana and Valparaiso), and samples were collected from five WWTPs (Figure 1). Monitoring was conducted twice during the autumn–winter season (June–August 2022) and twice during the spring–summer season (November–December 2022), except for the constructed wetland (CW), which was monitored only once per season due to technical problems encountered by the WWTPs. Samples (semi-composite) were taken between 10:00 am and 2:30 pm, with 1000 mL of wastewater collected every 45 min [11]. Samples were stored in propylene bottles and transported at 4 °C in dark conditions. Analyses were performed within 24 h [11]. Samples were taken at the WWTP entry point (influent) and after the disinfection process (effluent). The five WWTPs visited are in residential areas (only housing) in rural areas; there are no hospitals or industrial areas connected to the WWTPs.
Table 1 summarizes the physicochemical characteristics of the WWTPs studied; the flow rate (Q) was estimated based on the number of inhabitants of each WWTP, considering 150 L/inhabitant/day as wastewater discharge in Chile [28].

2.2. Analytical Methods and Instrumentation

2.2.1. Wastewater Viral Detection

Viral detection was conducted by concentrating 42 mL of each wastewater sample using ultracentrifugation. The final pellet was resuspended in 200 uL of sterile phosphate-buffered saline (PBS) at pH 7.4 and stored at −80 °C [11,30]. This concentrate was used to extract genetic material for the detection of Hepatitis A, Norovirus GI, Norovirus GII, SARS-CoV-2, Adenovirus F40/41, and JC polyomavirus. This last virus was employed as an internal control for the concentration and detection process of the virus [31]. Total viral nucleic acid was extracted from concentrated wastewater using the QIAamp®Viral RNA Mini kit (QIAGEN, Valencia, CA, USA) as described by the manufacturer. All samples were analyzed directly and diluted 1:10 to discard the eventual effects of inhibitors. The detection process of SARS-CoV-2 and Norovirus Genogroups I and II was previously described in [11]. Adenovirus 40/41 primers and probes synthesized at Macrogen Inc., Republic of Korea, with the following sequences were used: Adeno.fwd: 5′TTCCAGCATAATAACTCWGGCTTTG’3, Adeno.rev: 5′AATTTTTTCTGWGTCAGGCTTGG’3 and Adeno.probe 5′FAM-CCWTACCCCCTTATTGG-BHQ1′3. As a positive control, a plasmid was synthesized with the insert to be amplified, which was quantified. To carry out the amplification, the TaqPathTM1-step Multiplex Master Mix kit, Applied BiosystemsTM, was used with the conditions indicated by the manufacturer. The removal efficiency was determined using the viral load of influent and effluent in GC/mL (genome copies per mL of wastewater).

2.2.2. Propidium Monoazide (PMA)

For PMA assays, a 200 μm stock solution was prepared, from which 25 μL of PMAxx was taken and mixed with 50 μL of viral concentrate and 25 μL of nuclease-free water (the mixture was homogenized using a vortex (VWR Model G-560 Vortex Genie 3, USA)). The mixture was centrifuged for 10 min at 300 rpm (20 °C), and then the photoactivation process was performed using blue light equipment with a wavelength of 460 nm for 20 min. After this process, the genetic material (DNA/RNA) was extracted.

2.2.3. Detection of Antibiotic-Resistant Coliform Bacteria

Influent wastewater samples were diluted in physiological saline, and 20 μL of appropriate dilutions were spread on Violet Red Bile Agar (VRB Agar), a selective medium for the detection and enumeration of coliform organisms. The plate with VRB agar was incubated at 35 °C for 24 h. On the other hand, 10 mL of effluent wastewater samples were filtered through a membrane filter with a pore size of about 0.45 μm. The filter was cultured in the plate with mFC medium, a selective and differential medium to detect and enumerate fecal coliform. The plate with mFC agar was incubated at 44 °C for 24 h (membrane filtration: 8074 methods; HACH). For the detection of antibiotic-resistant coliforms, the same protocol for coliforms was used with the difference that the VRB agar and mFC media plates were supplemented with amoxicillin (25 μg/mL), cefadroxil (5 μg/mL), ciprofloxacin (5 μg/mL), azithromycin (15 μg/mL), and doxycycline (30 μg/mL). After the incubation period, a colony count was performed to determine CFU/mL. Removal efficiencies were calculated based on Log10 (influent/effluent), considering LOD 0.05 CFU/mL for fecal coliforms and LOD 25 CFU/mL for total coliforms.

3. Results and Discussion

3.1. Seasonal Behavior of Viruses in Rural WWTPs

Virus detection can contribute to scientific, sanitary, and environmental levels based on several aspects; one of them is epidemiological surveillance, as in the case of SARS-CoV-2, which allows for the detection of the circulation of viruses in a population. Viral variants can also be monitored as viruses with potential health impacts, such as NoV GI/GII and HAdV F40/41. It can also help to evaluate the efficiency of different wastewater treatments.

3.1.1. SARS-CoV-2

Figure 2 describes the quantification and detection of SARS-CoV-2 in influents from the studied WWTPs. Specifically, SARS-CoV-2 was detected in all influent samples analyzed (n = 9) during the autumn–winter and spring–summer seasons (Figure 2). Concentrations ranged from 26 to 822 GC/mL across different technologies. In contrast, detection in effluent samples was lower: the virus was present in 5 of 9 samples in AW (7–47 GC/mL) and 6 of 9 in spring–summer (<5–81 GC/mL). The AS_1 consistently exhibited the highest influent concentrations (819 and 822 GC/mL in autumn–winter and spring–summer, respectively), whereas VF reported the lowest (26 and 16 GC/mL in autumn–winter and spring–summer, respectively). Notably, these viral loads did not directly correlate with the number of clinically confirmed SARS-CoV-2 cases in each locality, likely due to reduced diagnostic testing [32]. These discrepancies highlight the limitations of clinical case data during the post-pandemic phase and underscore the value of wastewater-based surveillance.
Globally, SARS-CoV-2 test positivity rates have decreased significantly, 44.6% (2020), 26.4% (2021), 51.2% (2022), 9.5% (2023), and 11.5% (2024) [32], illustrating the decline in clinical testing intensity.
Despite this tendency, wastewater-based epidemiology (WBE) remains a robust tool for monitoring community-level viral circulation. During the pandemic (2021), average influent concentrations in rural Chilean areas were around 462 GC/mL [11].
In the present study, concentrations reached up to 842 GC/mL, suggesting a potential resurgence in viral circulation, possibly due to increased population mobility and reduced confinement measures. Variant-specific detection revealed the presence of Omicron subvariants BA.2 and BA.4/BA.5 during the AW season, while only BA.4/BA.5 was found in spring–summer. No detections were observed for Mu, Lambda, Gamma, or Delta variants. BA.2, first identified in Chile in February 2022, has been reported to be ~1.5 times more transmissible than earlier Omicron subvariants [33]. The emergence of BA.4/BA.5 between April and June 2022 coincides with its initial detection in wastewater (July, vermifilters), and its dominance during spring–summer aligns with reported increases in transmissibility [34].
These findings reaffirm the value of WBE for the real-time monitoring of variant dynamics. WBE provides a cost-effective and non-invasive alternative to individual testing, offering early warning signals of increased transmission and variant emergence [35,36,37,38]. To assess removal efficiency, WWTPs were grouped by biological treatment type: nature-based solutions (NBSs)—including constructed wetlands (CWs) and vermifilters (VFs)—and activated sludge-based (ASB) systems (AS_1, AS_2, AS_BD). Among NBS, CW achieved average SARS-CoV-2 removal rates of 0.60 log10 (autumn–winter) and 0.59 log10 (spring–summer). In comparison, VF improved from 0.71 log10 (autumn–winter) to 0.73 log10 (spring–summer), likely due to higher temperatures enhancing earthworm activity and enzymatic processes [39]. In both constructed wetlands and vermicomposting, filtration processes can contribute to virus removal, such as SARS-CoV-2. In some cases, removal times can affect removal efficiency. Despite generally lower hydraulic retention times (HRTs)—averaging only 0.8 days in 2021—NBS systems demonstrated moderate removal efficiency.
The literature suggests that optimal HRTs of 3–6 days are required for effective viral removal [40,41]. ASB systems exhibited superior removal performance. AS_1 achieved 2.2 log10 removal in both seasons. AS_2 also reached 1.18 log10 in the autumn–winter, with a slight reduction to 1.06 log10 in the spring–summer season. AS_BD removed 1.42 log10 in autumn–winter and 0.88 log10 in spring–summer. These differences may be linked to operational parameters, such as sludge age and HRT. Longer sludge ages—reported as 30–76 days (AS_1), 46–56 days (AS_BD), and 31–53 days (AS_2)—enhance enzymatic degradation and viral adsorption onto microbial biomass [42].
The high removal efficiencies observed in ASB systems underscore their operational robustness for enveloped viruses, such as SARS-CoV-2. The virus’s lipid envelope facilitates inactivation through enzymatic degradation, microbial activity, and adsorption to suspended solids [11,43]. These findings underscore the importance of biological wastewater treatment in mitigating pathogens.

3.1.2. Norovirus GI and GII

During the autumn–winter season, NoV GI was detected in one of nine influent samples, specifically in the VF system, where a 30 GC/mL concentration was recorded. Additionally, <5 GC/mL was attributed to intact viral particles (as determined by PMA treatment). Effluent samples yielded only one positive detection for NoV GI (VF), with levels below the quantification threshold (<5 GC/mL); NoV GI was detected in 2 of 9 influent samples in the spring–summer season. The AS_BD system reported 18 GC/mL, with 77.8% (14 GC/mL) of the load corresponding to intact viral particles. A second sampling at AS_BD recorded levels below the detection limit (<5 GC/mL). No NoV GI was detected in any effluent sample during the spring–summer season. Norovirus GII (NoV GII) was detected in five of nine influent samples during the autumn–winter season. The CW system showed the highest viral load (248 GC/mL), with only 7.8% (31 GC/mL) of particles identified as intact. The AS_BD system reported 40 GC/mL, of which 72.5% (29 GC/mL) were intact. Other influents (VF, AS_1, AS_2) showed <10 GC/mL. Only CW yielded detectable levels (<10 GC/mL) in the effluent, but integrity assessment was not feasible due to concentrations near the detection limit.
During the spring–summer season, NoV GII was detected in seven of nine influent samples. CW influents recorded 18 GC/mL, with 55.6% (10 GC/mL) intact. AS_BD showed 19 GC/mL, with 68.4% (13 GC/mL) intact. VF, AS_1, and AS_2 had concentrations ranging from 12 to 16 GC/mL; however, PMA quantification was impossible due to detection limits. NoV GII was detected in only 1 of 10 samples, corresponding to effluent VF (<10 GC/mL). These findings support a seasonal pattern, with higher norovirus prevalence and viral loads in the autumn–winter season. Similar trends were observed in 2021 in Chilean rural WWTPs, where NoV GII concentrations reached 1875 GC/mL [11], and in the U.S., where peak levels between 3012 and 3215 GC/mL were reported from October to May [44]. Norovirus is primarily transmitted through contact, contaminated food and water, and aerosols [45]. Aerosolized particles from vomiting events are exceptionally infectious, requiring <100 particles to initiate infection. In wastewater-associated aerosols, concentrations of 320 and 150 GC/m3-air for NoV GI and GII were reported in Kanpur, India. In Bolivia, levels ranged from 13 to 2.4 GC/m3-air [46]. In our study, the proportion of intact particles in influents reached up to 77.8% for NoV GI (AS_BD) and varied from 7.8% (CW) to 72.5% (AS_BD) for NoV GII. These observations highlight a potential occupational risk for WWTP workers [14,47]. While wastewater aerosols may contribute to NoV transmission, further studies are needed to assess aerosol infectivity and viral integrity directly.
Regarding removal efficiency, VF showed a 1.84 log10 reduction in autumn–winter and 0.08 el removal during spring–summer. CW achieved 1.6 log10 removal in autumn–winter. These seasonal differences may reflect variations in operating conditions, viral characteristics, and pathogen removal mechanisms inherent to each technology.
Noroviruses are non-enveloped viruses, and their protein capsid confers greater resistance to environmental stressors. As such, NoV can persist in wastewater and resist disinfectants such as UV, alcohol, and quaternary ammonium compounds ([48,49,50]). Viral removal may depend on viral concentration and adsorption to solids [11]. Interactions with bacterial communities may also influence removal dynamics [51]. Virus–surface interactions contribute to through adsorption, filtration, and sedimentation processes, and these interactions are modulated by factors such as pH, viral surface charge, and the type of adsorbent [52]. Norovirus particles can adhere to suspended solids or sludge, facilitating their physical removal depending on the treatment configuration [53].

3.1.3. Adenovirus F40/41

Figure 3 summarizes the detection, quantification, and occurrence of intact viral particles of HAdV-F40/41 from WWTPs. HAdV F40/41 was detected in nine of the nine influent samples during the autumn–winter season. The highest concentration detected was 5218 GC/mL (AS_BD) and the lowest was 161 GC/mL (AS_2). The total influent with HAdV-F40/41, all present intact viral particles (with PMA), had percentages varying from 7.0 to 62.6%. When analyzing the intact HAdV-F40/41 viral particles, it can be observed that the AS_BD influent showed 62.6% (159 GC/mL) of intact viral particles. Meanwhile, effluent during the autumn–winter season, HAdV-F40/41, was detected in nine of the nine samples analyzed. The wastewater sample with the highest viral load was 105 GC/mL (AS_BD), and the lowest was 12 GC/mL (CW). Although all samples have HAdV-F40/41, AS_1 is interesting since its effluent had a viral load of 61 GC/mL, with 5 GC/mL (6.2%) integrated viral particles. In contrast, for the rest of the WWTPs, HAdV-F40/41 levels were within the effluent detection limit (<5 GC/mL). It is impossible to indicate whether these particles are potentially infectious in environments where these wastewaters are discharged.
HAdV-F40/41 during the spring–summer season was detected in nine of the nine samples analyzed. The highest concentration detected during this season was 3797 GC/mL in CW, and the lowest was 21 GC/mL in VF. The total influent with HAdV-F40/41, all present integral viral particles, had percentages from 6.1 to 65.3%. Meanwhile, in AS_BD influent, 65.3% (158 GC/mL) are integral viral particles. The effluent during the spring–summer season showed that HAdV-F40/41 was detected in eight of the nine samples analyzed. The effluent with the highest viral load was 82 GC/mL (AS_BD); the lowest was 10 GC/mL (CW). Although all samples contain HAdV F40/41, VF is notable because it effectively removes the total HAdV-F40/41 from the wastewater, which was previously undetected. However, for the other effluents, although they show a viral load, the particles cannot be determined to be intact or not, since the analysis with PMA was detected at most <5 GC/mL (detection limit) in the effluents of AS_1, AS_2, and AS_BD. Inside CW and VF, no HAdV-F40/41 levels were detected with PMA.
The data highlights a low percentage (6.2%) of integral viral particles. HAdV F40/41 can be transmitted by fecal–oral contact or contaminated surfaces, food, or water [54]. Therefore, whole viral particles could reach people and generate infectious conditions associated with gastrointestinal cases. When observing the seasonality of HAdV inside wastewater, there is a tendency for a higher viral load in the cold period (autumn–winter) with an average load of 1287 GC/mL, while a lower viral load is observed in the warm period (spring–summer) with an average viral load of 543 GC/mL. In USA studies, something similar is observed, since it is during the winter period that they observed a higher detection rate, with 42.3% of the samples positive for HAdV-F40/41 during the autumn–winter, with detection ranges between 1.7 × 104 and 7.5 × 107 GC/mL [55]. HAdV is a year-round virus, but its peak is during the cold period, specifically winter.
Another approach to use HAdV-F40/41 measurements is for epidemiological studies based on wastewater or EBW. Northern Ireland has used wastewater-based epidemiology to monitor the levels of circulating HAdV-F40/41, where they have observed an increase in cases of acute hepatitis associated with HAdV-F40/41. The study allowed for the detection of HAdV F40/41 from wastewater, with a sequence like that of HAdV-F40/41 in clinical pictures with acute hepatitis [15]. These studies support the potential to inform the community surveillance of viruses, such as HAdV-F40/41, as they can contribute to their detection and generate the corresponding health alerts. The study of HAdV from wastewater involves assessing the potential health risk of viral diseases transmitted via the fecal–oral pathway. When observing the relationship between seasonality and HAdV-F40/41 levels, the highest levels were observed in the autumn–winter period, with values ranging from 5218 to 161 GC/mL. In contrast, the levels were 21—3797 GC/mL in spring–summer. Therefore, in this study, HAdV levels presented a higher prevalence during the autumn–winter, contrary to what was reported in Saudi Arabia, where the prevalence is higher in late summer and autumn at high temperatures (32–43 °C) [56]. It may be that the prevalence of HAdV is associated with the number of infected persons rather than seasonality.
When comparing the HAdV-F40/41 removal efficiencies, the nature-based solutions (CW and VF) remove between 1.0 and 2.5 log10 of the viral load. Meanwhile, activated sludge technologies (AS_1 and AS_2) remove between 0.02 log10 and 0.98 log10 of the viral load. Hybrid technologies like AS_BD report viral removal in the 0.29 to 1.69 log10 range. This result is interesting because ASB technologies may not completely remove the viral load, and even some of these particles may remain intact, which could generate infectious symptoms in the population that meets this type of wastewater. However, the effluents presented <5 GC/mL (detection limit) of potential intact viral particles. Adenovirus is a non-enveloped, double-stranded DNA virus with an icosahedral capsid, which could favor their adsorption and NBS regarding ASB technologies. Remove the viral load, and even some of these particles may remain intact, which could generate infectious symptoms in the population that meets this type of wastewater.
Regarding the removal of HAdV F40/41, this can occur through various processes, including sedimentation and filtration [57], which can be combined with absorption processes in elements such as biofilms [37]. All these processes are observed in nature-based treatments, where filtration, sedimentation, and biofilm formation occur in gravel in CW, and something similar happens in vermifilters, where the filtration process and the presence of biofilms in support material contribute to the removal of HAdV F40/41. This is reflected in the removal percentages, where NBS (CW and VF) have higher removal rates than “conventional” systems. This may be due to the diversity of processes involved in pathogen removal.

3.1.4. Total and Fecal Coliforms in Rural WWTPs

Table 2 summarizes the quantification and removal efficiency of total coliforms (TCs) and fecal coliforms (FCs) across the evaluated WWTPs. During the autumn–winter season, TC was detected in all influent samples (9/9), with concentrations ranging from 3.2 × 106 CFU/100 mL (AS_1) to 7.6 × 109 CFU/100 mL (AS_BD). In effluent samples, TC was detected in seven of nine cases, with concentrations ranging between 7.5 × 103 and 3.0 × 109 CFU/100 mL. TC was again detected in all influent samples during the spring–summer season. Concentrations ranged from 1.8 × 107 CFU/100 mL (AS_1) to 5.2 × 109 CFU/100 mL (AS_BD). TC was found in eight of nine effluent samples, with values between 2.6 × 105 CFU/100 mL (AS_1) and 6.0 × 108 CFU/100 mL (AS_BD).
Similarly, FCs were detected in all influent samples (9/9) during autumn–winter, ranging from 5.0 × 103 CFU/100 mL (AS_1) to 4.7 × 107 CFU/100 mL (VF). Effluent FCs were detected in seven of nine samples, with concentrations between 1.6 × 102 CFU/100 mL (AS_2) and 3.3 × 106 CFU/100 mL (CW). In spring–summer, FCs were detected in all influent samples, ranging from 6.67 × 103 to 9.00 × 106 CFU/100 mL (AS_1). FCs were found in six of nine effluent samples, ranging from 3.89 × 102 CFU/100 mL (AS_1) to 1.39 × 106 CFU/100 mL (AS_BD). Seasonal comparison showed higher average effluent loads in autumn–winter for both TC (1.60 × 109 CFU/100 mL) and FC (1.75 × 107 CFU/100 mL) compared to spring–summer (1.25 × 109 CFU/100 mL and 4.48 × 106 CFU/100 mL, respectively).
In terms of removal efficiency, autumn–winter removal of TC ranged from 5.46 log10 (AS_1) to 0.13 log10 (CW), while in spring–summer, it varied from 6.29 log10 (AS_2) to 0.42 log10 (VF). For FC, autumn–winter removal ranged from 6.87 log10 (AS_2) to 0.65 log10 (CW), and in spring–summer, it ranged from 6.87 log10 (AS_1 and AS_2) to 0.31 log10 (VF). Overall, nature-based solutions (NBSs), such as constructed wetlands (CWs) and verifiers (VF), showed lower average removal rates compared to activated sludge-based systems (ASB). Although NBS involves multiple mechanisms adsorption, filtration, sedimentation, and microbial degradation—its effectiveness may depend on design parameters and operational conditions. For instance, CW performance is influenced by substrate grain size and water depth, which affect contact with plant root systems. Systems with lower depths (e.g., 0.27 m) and smaller substrates (e.g., 3.5 mm) have been associated with improved FC removal [58,59]. Root-associated microbial communities may also secrete antibacterial compounds, contributing to pathogen reduction [60]. However, external factors such as organic load and temperature can reduce NBS effectiveness; for example, Brazilian CWs have reported lower FC removal under increased organic loads and lower temperatures [61].
Seasonal differences were evident in VF systems. In autumn–winter, VF achieved 1.11 log10 TC and 2.06 log10 FC removal, while in spring–summer, efficiency decreased to 0.42 log10 for TC and 0.83 log10 for FC. This trend may be explained by microbial activity: warmer temperatures in spring–summer enhance microbial metabolism and earthworm-associated bioprocesses [60], while in colder seasons, activity may be suppressed. VF systems rely on the synergistic action of earthworms and microorganisms. Earthworms ingest organic matter and associated microbes, enhancing degradation and promoting antibacterial activity via their gut microbiota [39]. While these systems demonstrate potential, their performance is susceptible to environmental and operational factors such as HRT and support media.
In Chile, fecal coliform levels in treated wastewater are regulated by Supreme Decree 90/2001, which states that the maximum value for discharge into surface waters is 1000 NMP/100 mL [62]. In terms of international regulations, the WHO has established values for the reuse of gray water that depend on the reuse of treated wastewater, ranging from 1000 NMP/100 mL (raw crop irrigation) to 10,000 NMP/100 mL (non-edible crop irrigation) [63].

3.1.5. Antibiotic-Resistant Total and Fecal Coliforms in Rural WWTPs

Antibiotic-resistant coliforms in treated wastewater are of growing concern, particularly in the context of potential reuse for irrigation or environmental discharge. In this study, the occurrence and removal of total and fecal coliforms resistant to four commonly used antibiotics in Chile—amoxicillin, cefadroxil, ciprofloxacin, and azithromycin [64]—were evaluated and are summarized in Table 3.
During the autumn–winter season, total coliforms resistant to amoxicillin were effectively removed by AS_1 and AS_2 systems (2.72 log10 removal), while VF achieved 0.53 log10. In the spring–summer season, removal efficiencies ranged from 2.76 log10 (AS_1 and AS_2) to 0.63 log10 (VF). For fecal coliforms, autumn–winter removal ranged from 3.12 log10 AS_1 and 1.8 log10 AS_2 to 0.95 log10 (CW), while spring–summer removal ranged from ~3.00 log10 (AS_1 and AS_2) to 2.72 log10 (VF).
In the case of cefadroxil resistance, total coliforms were removed with efficiencies from 2.59 log10 (AS_2) to 1.50 log10 (VF) during autumn–winter and from 2.68 to 2.12 log10 (AS_1 and AS_2) to 0.78 (VF) in spring–summer. For fecal coliforms, autumn–winter removal ranged from 1.00 to 5.2 log10 (AS_1 and AS_2) to 0.64 log10 (AS_BD), while in spring–summer, the range was from 4.45 to 5.29 log10 (AS_1, AS_2, AS_BD) to 1.98 log10 (AS_BD).
Total coliforms resistant to ciprofloxacin were effectively removed ~1 log10 by AS_1 and AS_2 in both seasons, while CW had the lowest removal: 0.11 log10 in autumn–winter and 0.18 log10 in spring–summer. For fecal coliforms, autumn–winter removal ranged from 3.5 log10 (AS_1) to 0.24 log10 (AS_BD), and in spring–summer, from 3.8 log10 (AS_1) with no detection in AS_2. During spring–summer, total coliforms resistant to azithromycin were removed between 1.85 log10 (AS_2) and 0.21 (VF). In autumn–winter, removal ranged from 1.54 to 2.25 log10 (AS_1 and AS_2) to 0.49 log10 (VF). Autumn–winter removal ranged from 1.65 (AS_1) to only 0.18 log10 (CW) for fecal coliforms. Interestingly, no resistant fecal coliforms were detected in AS_BD in autumn–winter, and none were detected in any system during spring–summer (complete removal or absence).
WWTPs are designed to reduce microbial loads, particularly coliforms, which serve as key indicators of fecal contamination. However, the persistence or increase in antibiotic-resistant coliforms (ARCs) during wastewater treatment is an emerging public health concern. Our results confirm the presence of ARCs across all evaluated systems, with removal efficiencies ranging widely depending on the treatment technology and antibiotic. Some technologies, particularly activated sludge systems (AS_1 and AS_2), consistently achieved maximum removal of total and fecal coliforms resistant to various antibiotics. In contrast, nature-based systems (CW and VF) exhibited lower performance, especially against ciprofloxacin- and azithromycin-resistant strains. Previous studies have reported increased ARC counts after biological treatment, potentially due to selective pressure and gene exchange among microbial populations. For instance, studies have found that aerated lagoon systems with long hydraulic retention times (20–30 days) can enhance the development of ARCs [64,65]. Similarly, another study observed that although total coliforms were effectively removed, the proportion of resistant strains increased after treatment—suggesting horizontal gene transfer within the microbial community [66].
These findings underscore the importance of monitoring ARCs in wastewater as part of risk assessment frameworks for water reuse and environmental protection. Wastewater treatment technologies should be evaluated for their capacity to reduce microbial loads and limit the spread of antibiotic resistance in environmental matrices.
One of the key findings regarding resistant coliforms is that the effluents from various treatment plants contain several coliforms resistant to amoxicillin, ciprofloxacin, cefadroxil, and azithromycin. This suggests that the most significant contribution of resistant coliforms comes from the population, rather than from wastewater treatment. However, the treatments are not entirely efficient, and further studies are needed to determine the type of resistant bacteria that enter the treatment plant versus those that leave it.
Although there are some WWTPs that allow for the removal of almost 100% (AS_1 and AS_2) of resistant fecal coliforms, work must be conducted on the operating parameters of each plant, evaluating parameters such as secondary treatment efficiency, solids removal, and hydraulic retention times. By evaluating these types of parameters, the aim is to achieve proper secondary treatment, which can contribute to better disinfection processes, as this could contribute to the elimination or reduction in resistant coliforms in treated wastewater. Another way is to evaluate other types of technologies that are currently under study, but the initial cost is usually higher, which complicates the implementation of these types of technologies in more rural areas. Other alternatives could be to complement technologies in both the secondary process and the disinfection process [67], which would allow for better effluent with respect to the removal of fecal coliforms. Even supplementing higher concentrations of chlorine and UV in disinfection processes can remove resistant coliforms [68]. Therefore, it is important to conduct a detailed analysis of the critical points in the operation of the different WWTPs to ensure better treatment.
While optimization processes could be carried out on the different technologies, other measures would correspond to public policies on access to and administration of antibiotics, since the main contributor to resistant coliforms would be the population, and therefore it would be a measure beyond wastewater treatment plants.
As for new methods with ultrafiltration [69] and advanced oxidation processes [70] that help control wastewater more effectively and, in less time, other research groups are developing new technologies that are more efficient and faster, but they tend to be more expensive to implement, which makes them even more complex. Therefore, the first step would be to review in detail the operating parameters of each system to optimize each treatment and make the removal of these pathogens more efficient.

4. Conclusions

This study confirmed the presence and variability of SARS-CoV-2, NoV GI, NoV GII, HAdV F40/41, and antibiotic-resistant coliforms in Chilean rural domestic wastewater, highlighting the epidemiological potential of wastewater surveillance. High SARS-CoV-2 (822 GC/mL) viral loads were detected in influents across several sectors, supporting the utility of wastewater-based epidemiology as a rapid, cost-effective, and non-invasive tool for tracking viral circulation and detecting emerging variants at the community level. Regarding enteric viruses, the monitoring of NoV GI and GII demonstrated a consistently higher prevalence of NoV GII in both seasons (seven of nine influents, with a maximum of 399 GC/mL in spring–summer). Although most WWTPs effectively removed NoV, HAdV F40/41 was detected at high concentrations (5218 GC/mL) with lower removal efficiencies and the presence of potentially infectious intact particles. Quantifying total and fecal coliforms in influents (5 × 103–7 × 109 CFU/100 mL) and effluents (1 × 102–3 × 109 CFU/100 mL) revealed variable microbial loads, with evidence of persistence in effluents. In the case of resistant coliforms, some of the technologies have low efficiency in removing coliforms.

Author Contributions

A.P.-G.: Data Curation, Formal analysis, Investigation, Writing—Original Draft, Resources. C.A.V.-A.: Conceptualization, Writing—Review and Editing, Project Administration. M.A.: Data Curation, Investigation. A.G.: Conceptualization, Methodology, Writing—Review and Editing. All authors have read and agreed to the published version of the manuscript.

Funding

This research and the APC was funded by Fondecyt’s support, grant number 3220570. Cristina A. Villamar-Ayala thanks FONDEF ID23I10184. Aldo Gaggero thanks ANID-Fondecyt’s support, grant number 1181656 and ANID-Anillo, grant number ATE220007.

Data Availability Statement

The original contributions presented in this study are included in the article. Further inquiries can be directed to the corresponding author.

Acknowledgments

Authors thank the communities/municipalities in charge of the rural wastewater treatment plants that were part of this study. Our motivation is always to move towards a more equitable world where access to sanitation is a guaranteed right for all.

Conflicts of Interest

The authors declare no conflicts of interest.

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Water 17 03197 g001
Figure 1. Geographical map of the Chilean regions of Valparaiso and Metropolitan regions, where the different WWTPs analyzed in this study are located. CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk.
Figure 1. Geographical map of the Chilean regions of Valparaiso and Metropolitan regions, where the different WWTPs analyzed in this study are located. CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk.
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Figure 2. Quantification and detection of SARS-CoV-2 and its variants from rural domestic wastewater influent. (A) Autumn–winter. (B) Spring–Summer. Active cases: correspond to the population numbers infected with SARS-CoV-2 within each location where WWTPs are located (MINSAL). % Population: Corresponds to the percentage of the supplied population from studied WWTPs concerning the total location population. CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk.
Figure 2. Quantification and detection of SARS-CoV-2 and its variants from rural domestic wastewater influent. (A) Autumn–winter. (B) Spring–Summer. Active cases: correspond to the population numbers infected with SARS-CoV-2 within each location where WWTPs are located (MINSAL). % Population: Corresponds to the percentage of the supplied population from studied WWTPs concerning the total location population. CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk.
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Figure 3. Quantification and detection of integral viral particles of HAdV from rural domestic wastewater. (A) Influent autumn–winter, (B) effluent autumn–winter, (C) influent spring–summer, (D) effluent spring–summer. 1: First visit; 2: second visit. CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk.
Figure 3. Quantification and detection of integral viral particles of HAdV from rural domestic wastewater. (A) Influent autumn–winter, (B) effluent autumn–winter, (C) influent spring–summer, (D) effluent spring–summer. 1: First visit; 2: second visit. CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk.
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Table 1. General information from rural WWTPs.
Table 1. General information from rural WWTPs.
WWTPsNameLocationPopulation WWTPs TreatmentDate Autum–WinterDate Spring–Summer
Q (m3/day) *Kg BOD/daySecondaryTertiaryFirst
Sample		Second SampleThird SampleFourth
Sample
CWPolpaico33°10′05.8″ S 70°53′12.3″ W26740.0510.21Constructed wetlandCa (OCl)208/06/22nd30/11/22nd
VFRungue33°00′29.7″ S 70°53′22.7″ W37255.8022.38Verm filtersNaClO15/06/2220/07/2216/11/2221/12/22
AS_1Huelquén33°49′59.6″ S 70°38′49.6″ W27941.8513.32Activated sludgeNaClO06/07/2203/08/2202/11/2207/12/22
AS_2Los Loros32°50′44.4″ S 70°56′23.5″ W4850727.5124.49Activated sludgeCa (OCl)213/07/2224/08/2223/11/2228/12/22
AS_BDBicentenario33°44′45.0″ S 70°52′18.0″ W4340651.00118.04Activated sludgeNaClO22/06/2210/08/2209/11/2214/12/22
Notes: CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk. * Theoretical flow rates calculated based on population supplied, considering a per capita consumption of 150 L/capita-day.
Table 2. Detection and quantification of total and fecal coliforms from rural WWTPs (influent and effluent).
Table 2. Detection and quantification of total and fecal coliforms from rural WWTPs (influent and effluent).
WWTPsVisitSampleTotal ColiformsFecal Coliforms
CFU/100 mLRemoval Efficiency (log10)CFU/100 mLRemoval Efficiency (log10)
Autum WinterSpring SummerAutum
Winter		Spring
Summer		Autum
Winter		Spring
Summer		Autum
Winter		Spring
Summer
CW1Influent1.31 × 1081.33 × 1090.131.851.52 × 1071.50 × 1050.656.48
Effluent9.7 × 1071.88 × 107 3.39 × 1060.00 × 100
2Influentn.dn.dn.dn.dn.dn.dn.dn.d
Effluentn.dn.d n.dn.d
VF1Influent7.78 × 1075.50 × 1071.04nr4.77 × 1073.67 × 1051.910.31
Effluent7.08 × 1063.35 × 108 5.85 × 1051.80 × 105
2Influent6.50 × 1071.03 × 1081.170.423.69 × 1078.64 × 1062.211.35
Effluent4.35 × 1063.88 × 107 2.25 × 1053.83 × 105
AS_11Influent7.25 × 1061.83 × 1075.461.855.00 × 1039.89 × 1055.002.73
Effluent0.00 × 1002.60 × 105 0.00 × 1001.83 × 103
2Influent3.25 × 1066.45 × 1075.111.968.33 × 1036.67 × 1031.481.23
Effluent0.00 × 1007.00 × 105 2.78 × 1023.89 × 102
AS_21Influent1.98 × 1074.78 × 1073.420.787.22 × 1033.67 × 1051.646.87
Effluent7.50 × 1038.00 × 106 1.67 × 1020.00 × 100
2Influent2.63 × 1074.83 × 1072.326.293.67 × 1051.17 × 1056.876.37
Effluent1.25 × 1050.00 × 100 0.00 × 1000.00 × 100
AS_BD1Influent7.63 × 1095.25 × 1090.400.943.20 × 1072.07 × 1071.431.17
Effluent3.05 × 1096.00 × 108 1.20 × 1061.39 × 106
2Influent6.45 × 1094.38 × 1090.960.902.49 × 1079.00 × 1061.230.90
Effluent7.00 × 1085.55 × 108 1.46 × 1061.13 × 106
Notes: CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk. nr: Non-removed; n.d: Non-data.
Table 3. Efficiency of removal of antibiotic-resistant coliforms in wastewater from rural areas (influent and effluent).
Table 3. Efficiency of removal of antibiotic-resistant coliforms in wastewater from rural areas (influent and effluent).
WWTPsVisitColiformsAmoxicillinCefadroxilCiprofloxacinAzithromycin
Autum–WinterSpring–SummerAutum–WinterSpring–SummerAutum–WinterSpring–SummerAutum–WinterSpring–Summer
CW1Total1.871.121.000.820.110.180.490.37
Fecal0.600.001.024.540.000.000.183.30
2Totalndndndndndndndnd
Fecalndndndndndndndnd
VF1Total0.580.650.790.810.120.300.220.26
Fecal1.001.601.081.600.000.000.053.00
2Total0.480.600.760.760.150.220.400.15
Fecal0.903.851.381.410.000.000.300.00
AS_11Total2.782.781.882.781.180.781.382.35
Fecal4.654.545.205.293.704.000.000.00
2Total2.672.702.282.581.301.301.692.30
Fecal1.584.664.995.273.303.600.050.00
AS_21Total2.302.782.482.781.381.002.232.04
Fecal0.000.001.000.000.000.003.303.00
2Total2.782.782.701.461.200.482.261.65
Fecal3.600.003.704.450.000.000.003.30
AS_BD1Total1.000.911.200.930.700.780.330.30
Fecal0.000.000.703.480.000.000.000.00
2Total1.011.031.220.900.300.270.370.20
Fecal0.000.000.580.480.483.000.000.00
Notes: CW: Constructed Wetland; VF: Vermifilter; AS: Activated Sludge; AS-BD: Activated Sludge/Bio-disk. nd: Non-data.
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Plaza-Garrido, A.; Villamar-Ayala, C.A.; Ampuero, M.; Gaggero, A. SARS-CoV-2, Noroviruses, Adenoviruses, and Antibiotic-Resistant Coliforms Within Chilean Rural Wastewater Treatment Plants. Water 2025, 17, 3197. https://doi.org/10.3390/w17223197

AMA Style

Plaza-Garrido A, Villamar-Ayala CA, Ampuero M, Gaggero A. SARS-CoV-2, Noroviruses, Adenoviruses, and Antibiotic-Resistant Coliforms Within Chilean Rural Wastewater Treatment Plants. Water. 2025; 17(22):3197. https://doi.org/10.3390/w17223197

Chicago/Turabian Style

Plaza-Garrido, Angela, Cristina A. Villamar-Ayala, Manuel Ampuero, and Aldo Gaggero. 2025. "SARS-CoV-2, Noroviruses, Adenoviruses, and Antibiotic-Resistant Coliforms Within Chilean Rural Wastewater Treatment Plants" Water 17, no. 22: 3197. https://doi.org/10.3390/w17223197

APA Style

Plaza-Garrido, A., Villamar-Ayala, C. A., Ampuero, M., & Gaggero, A. (2025). SARS-CoV-2, Noroviruses, Adenoviruses, and Antibiotic-Resistant Coliforms Within Chilean Rural Wastewater Treatment Plants. Water, 17(22), 3197. https://doi.org/10.3390/w17223197

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