Influence of Seasonal Variation in Antibiotic Concentration on the Fate and Transport of Antibiotics Within an Artificial Pond System

Abstract

Seasonal variability significantly influences the fate and transport of antibiotics (Abs) in wastewater stabilization ponds by affecting their concentration, degradation kinetics, sorption behavior, and ecological interactions. This study investigated the influence of seasonal variability for a large number of Ab classes—eleven sulfonamides (SAs), eight fluoroquinolones (FQs), five macrolides (MLs), one diaminopyrimidine (DIA), two tetracyclines (TETs), two lincosamides (LICs), and three phenicols (Phens)—on their fate and transport in an artificial stabilization pond system (SPS) receiving treated WWTP effluent. Two sampling campaigns were conducted during China’s long-lasting seasons (summer and winter). The detection frequency for sulfamethoxazole (SMX), sulfapyridine (SPY), and ofloxacin (OFX) was 100%, for sulfamethazine (SMZ) 63.3%, and for clindamycin (CLN) 83.3% in both seasons. The detection frequency for the other Abs was equal or below 50% in both seasons. In addition, the maximum concentration of SMX, SMZ, SPY, OFX, and CLN in summer was 10.51, 19.37, 6.93, 22, and 4.04 ng/L, respectively, and 4.27, 0.14, 3.15, 9.29, and 8.78, respectively, in winter). The rest of the Abs were either detected in summer or winter. It was observed that environmental fluctuations (such as temperature, precipitation, SPS flow patterns, light intensity), differences in antibiotic use and consumption between seasons, and differences in physicochemical properties of the Abs were the main factors influencing their fate and transport within the SPS. The potential environmental risks of Abs detected in the SPS were assessed using the risk quotient (RQ) approach. Typically, RQs in summer were remarkably higher than in winter. Norfloxacin and chlortetracycline posed a medium risk in summer; however, ofloxacin posed a medium risk in winter and a high risk in summer. Therefore, management strategies should consider the dynamic nature of antibiotic contamination, accounting for seasonal influences on fate and transport within the studied SPS and maybe for other wastewater stabilization ponds by adjusting operational practices, optimizing treatment processes, and implementing source control measures to mitigate the environmental impacts of seasonal antibiotic variability.

1. Introduction

Seasonal variability in antibiotic concentration in aquatic environments refers to the fluctuations in the levels of antibiotics present in a water body for a year. This phenomenon is of significant concern due to its potential impact on aquatic ecosystems and public health. Antibiotics are usually detected in aquatic environments, with concentrations ranging from nanograms to micrograms per liter [1]. Antibiotics in aquatic environments can result from various sources, including agricultural runoff, wastewater discharge, and improper disposal of pharmaceuticals [2,3]. Seasonal variability in antibiotic concentration in aquatic environments has been the subject of extensive environmental science and public health research. Recent reports revealed that the levels of antibiotics in water bodies can vary throughout the year, with certain seasons exhibiting higher concentrations than others [3,4,5]. For example, it was reported that the levels of antibiotics in marine environments during colder seasons surpassed those observed in warmer seasons [6]. Previous studies have highlighted the dependency of antibiotic concentrations in sediments on factors such as seasonal variations, water flow, sediment characteristics, and the quantities of antibiotics used in the region [7]. These studies also demonstrated that microbial activity in sediments can contribute to a gradual reduction in antibiotic concentrations over time. The escalation of pollutants, including antibiotics, in rivers traversing urban and rural areas was attributed to sewage discharge into these water bodies.

Furthermore, it has been identified that downstream factors such as higher population density, dehydration, and inadequate continuous flow during seasonal runoff contribute significantly to elevated antibiotic concentrations in the lower reaches of rivers [8]. Tang et al. similarly discovered that contamination levels of most antibiotics in Chaohu Lake were more pronounced in winter compared to other seasons, but this was observed under conditions where the lake was not covered by ice [9]. Penicillin V was found to be higher in winter (216.43 ng/L) than in summer (89.91 ng/L) in a study conducted on Qingcaosha Reservoir, which is an important drinking water source for Shanghai residents [3]. This variability can be attributed to factors such as weather patterns, agricultural practices, and human activities.

Seasonal changes in precipitation and temperature can influence the transport and fate of antibiotics in aquatic environments [5]. For instance, heavy rainfall can lead to increased runoff from agricultural fields, carrying antibiotics into nearby water bodies. Similarly, higher temperatures can enhance the degradation of antibiotics in water, affecting their persistence and concentration levels [5].

Furthermore, seasonal variations in agricultural practices, such as planting and harvesting seasons, can impact the use of antibiotics in livestock farming. The application of antibiotics in animal husbandry can contribute to the presence of these compounds in surface water and groundwater [5,10,11]. Additionally, variations in human activities, such as tourism and recreational events during certain seasons, can also influence the input of antibiotics into aquatic ecosystems through wastewater discharges [8].

Wastewater stabilization ponds are crucial components of wastewater treatment systems, playing a pivotal role in removing contaminants and pathogens from municipal and industrial effluents. They are expansive shallow reservoirs that leverage physical and biological processes to eliminate organic substances, contaminants, and pathogens from untreated wastewater. This approach stands as one of the uncomplicated techniques for wastewater treatment. It is widely implemented globally, particularly in developing nations where ample land is typically accessible. The climate is conducive (characterized by high temperatures and abundant sunlight) for efficient operation [12]. Many countries favor SPS as the preferred method, employing it for the secondary treatment of municipal sewage due to its relatively modest investment requirements. Additionally, it is considered advantageous for being simple and cost-effective to operate and maintain, using locally available personnel [13].

Traditional wastewater stabilization ponds, commonly known as lagoons, typically incorporate a mix of three distinct pond types: anaerobic ponds (APs), facultative ponds (FPs), and maturation ponds (MPs) [12]. Anaerobic ponds, or APs, are characterized by greater depth and a lower hydraulic retention time. They function primarily without dissolved oxygen, operating under methanogenic conditions. Facultative ponds, or FPs, represent a blend of aerobic and anaerobic environments, supporting both aerobic and anaerobic conditions. Maturation ponds, analogous to aerobic ponds, bear a minimal organic waste load. Their primary purpose is to augment the quality of treated effluents through secondary treatment processes [14].

Generally, the effectiveness of wastewater stabilization ponds in eliminating organic micropollutants exhibits significant variability, influenced by multiple factors. These factors encompass the pond type and configuration, operational parameters, wastewater quality, ambient circumstances (such as sunlight, temperature, redox conditions, and pH), and the unique characteristics of the pollutant. The elimination procedures for organic micropollutants in SPS are primarily ascribed to biodegradation, photodegradation, and sorption, predominantly taking place during the first treatment phases, such as in anaerobic or facultative ponds [5,12].

Although stabilization ponds are widely used worldwide due to their various advantages, the seasonal detection and occurrence of emerging pollutants, such as antibiotics, have rarely been reported in artificial stabilization pond system (SPS). Recent reports have primarily focused on seasonal detection and occurrence in wastewater treatment plants [15,16,17,18,19], rivers [20,21,22,23,24], lakes [25,26], reservoirs [3,27], and tap and well water [28]. Additionally, the influence of seasonal variation on the fate and transport of antibiotics, especially within artificial stabilization pond systems (SPSs), has not yet been well documented. However, the seasonal variation of antibiotics present in the wastewater, seasonal changes, such as temperature fluctuations, precipitation patterns, and agricultural practices, and variations in microbial activity, can significantly affect the degradation, transformation, and transport of antibiotics within these treatment systems. Therefore, the impact of seasonal variations on the fate and behavior of antibiotics in artificial stabilization pond systems (SPSs) is a critical aspect that requires comprehensive understanding. Understanding how these variations influence the persistence and removal of antibiotics is essential for optimizing wastewater treatment processes and minimizing the environmental impact of antibiotic residues. Moreover, understanding the seasonal dynamics of antibiotics in wastewater stabilization ponds is essential for implementing effective management strategies and optimizing treatment processes.

This study’s aims were: (a) to describe and cover the knowledge gap on seasonal occurrence, detection frequency, and the environmental risk assessment of a wide number of antibiotics belonging to different antibiotic classes (sulfonamides, diaminopyrimidine, fluoroquinolones, macrolides, tetracyclines, phenicols, and lincosamides) within an SPS for the first time; (b) to describe and investigate the specific mechanisms through which seasonal variations (summer and winter) influence the fate and transport of these antibiotics within an artificial stabilization pond system; and (c) to describe the treatment technology of an artificial stabilization pond system (SPS) (configuration of an artificial stabilization pond system) in order to provide information about the powerful linkage for algae bacteria symbiosis in degrading antibiotics, ultimately contributing to sustainable and efficient wastewater treatment practices.

Therefore, field investigations, sampling, and analysis were conducted regularly throughout the year for the two long-lasting seasons (summer and winter) in China to understand the seasonal variation in antibiotic concentration in an artificial stabilization pond (SPS). Sampling at different depths and locations within the pond was performed to provide insights into spatial variations.

2. Materials and Methods

2.1. Chemicals and Standards

Sulfonamides (SAs): sulfamethoxazole (SMX), sulfamethazine (SMZ), sulfapyridine (SPY), sulfadiazine (SDZ), sulfachloropyridazine (SPZ), sulfacetamide (SCD), trimethoprim (TMP). Fluoroquinolones (FQs): ofloxacin (OFX), norfloxacin (NFX), enrofloxacin (EFX), difloxacin (DFX). Macrolides (MLs): clindamycin (CLN), lincomycin (LCN), roxithromycin (ROX), erythromycin (ERY), azithromycin (AZN). Tetracyclines (TETs): oxytetracycline (OXY), chlortetracycline (CLT). Phenicols (PHENs): florfenicol (FFN). These antibiotic compounds, together with their labeled compounds, include ciprofloxacin D8 (CFX-D8), norfloxacin D5 (NFXD5), sulfadiazine D4 (SDZD4), sulfamethazine d4 (SMZd4), sulfamethoxazole d4 (SMX-d4), sulfapyridine d4 (SPD-d4), sulfadimethoxine d6 (SDM-d6), ciprofloxacin d8 (CFC-d8), ofloxacin d3 (OFC-d3), enrofloxacin d5 (EFC-d5, norfloxacin d5 (NFC-d5), clarithromycin d3 (CTM-d3), erythromycin d3 (ETM-d3), roxithromycin d7 (RTMd7), demeclocycline (DMC), doxycycline d3 (DCX-d3), clindamycin, and trimethoprim d3 (TMP-d3), and were bought from China’s Sinopharm Chemical Reagent Co., Ltd. (Tianjin) with purity >98%. Each compound was prepared by diluting the stock solution with methanol at 1000 mg/L and a mixture of working standards containing each compound at 10 mg/L. Other chemicals of analytical grade such as formic acid and disodium ethylenediamine tetraacetate (Na₂EDTA) and hydrochloric acid, HPLC grade acetonitrile, and methanol were also supplied by China’s Sinopharm Chemical Reagent Co., Ltd. (Tianjin, China). Ultrapure water (MQ) was obtained from a Milli-Q water purification system (Millipore, Billerica, MA, USA). Antibiotic properties in this study are summarized in

Table 1. The properties of the target antibiotic compounds.

Antibiotic Class	Antibiotic Compounds	CAS Number	Molecular Formula	Log Kow	pKa	Usage
11 Sulfonamides	Sulfamethoxazole (SMX)	723–46–6	253.28	0.89 a	1.85, 5.60 b	Veterinary/Human
	Sulfamethazine (SMZ)	57–68–1	278.32	0.89 a	2.65, 7.65 b	Veterinary
	Sulfapyridine (SPY)	144–83–2	249.27	n.a	1.98, 5.96 b	Human
	Sulfamethizole (SML)	144–82–1	260.29	n.a	n.a	Human
	Sulfathiazole (STL)	72–14–0	255.28	n.a	n.a	Veterinary/Human
	Sulfacetamide (STD)	144–80–9	254.26	n.a	n.a b	Veterinary/Human
	Sulfadiazine (SDZ)	68–35–9	250.27	−0.09 a	2.00, 6.48 b	Veterinary/Human
	Sulfamonomethoxine (SMM)	1220–83–3	311.33	0.7 a	n.a	Veterinary
	Sulfadimethoxine (SDM)	1220–87–7	310.33	n.a	n.a	Veterinary
	Sulfaquinoxaline (SQL)	59–40–5	314.32	n.a	n.a	Veterinary
1 Diaminopyrimidine	Trimethoprim (TMP)	738–70–5	290.31	0.91 a	3.23, 6.76 b	Veterinary/Human
8 Fluoroquinolones	Ofloxacin (OFX)	82419–36–1	361.38	−0.39 a	5.97, 8.28 b	Veterinary/Human
	Norfloxacin (NFX)	70458–96–7	319.34	−1.03 a	3.11, 6.10, 8.60, 10.5 b	Veterinary/Human
	Enrofloxacin (EFX)	93106–60–6	359.39	0.7 a	3.86, 6.19, 7.59, 9.86 b	Veterinary
	Ciprofloxacin (CPX)	85721–33–1	331.34	0.28 a	3.10, 6.14, 8.70, 10.5b	Veterinary/Human
	Fleroxacin (FLX)	79583–68–1	346.33	1.12 a	n.a	Veterinary
	Difloxacin (DFX)	98105–99–8	349.35	2.36 a	n.a	Human
	Lomefloxacin (LFLX)	98079–51–7	362.36	-0.3 a	5.00, 5.87, 9.23 b	Human
	Sarafloxacin (SFX)	98105–99–8	362.36	n.a	n.a	Human
5 Macrolides	Roxithromycin (ROX)	80214–83–1	837.94	2.75 a	9.17 b	Veterinary/Human
	Erythromycin (ERY)	114–07–8	733–749	3.06 a	8.9 b	Veterinary/Human
	Azithromycin (AZT)	83905–01–5	749–767	4.02 a	8.74, 9.45 b	Veterinary/Human
	Rifampicin (RFN)	13292–46–1	822–823	n.a	n.a	Veterinary/Human
	Tylosin (TYL)	n.a	916–918	1.63 a	n.a	Veterinary
2 Tetracyclines	Oxytetracycline (OXY)	n.a	460–475	−0.9 a	3.27, 7.32, 9.11 b	Veterinary/Human
	Chlortetracycline (CLT)	n.a	479–480	−0.62 a	3.30, 7.55, 9.15 b	Veterinary/Human
others	Lincomycin (LIN)	154–21–2	406–407	0.56 a	7.60 b	Veterinary/Human
	Clindamycin (CLN)	18,323–44–9	714–716	n.a	n.a	Veterinary/Human
	Chloramphenicol (CPL)	n.a	323–325	1.14 a	5.5 b	Veterinary/Human
	Thiamphenicol (TPL)	n.a	356–358	n.a	n.a	Veterinary
	Florfenicol (FFL)	n.a	358–359	n.a	n.a	Veterinary

n.a: not available; ^a log Kow (octanol–water coefficient) from ChemIDPlus Advanced (https://chem.nlm.nih.gov/chemidplus/) and PubChem (https://pubchem.ncbi.nlm.nih.gov/), U.S. National Library of Medicine, last accessed 15 August 2022; ^b TOXNET, Toxicology Data Network (https://www.nlm.nih.gov/toxnet/index.html), last accessed 15 August 2022, U.S. National Library of Medicine.

.

2.2. Study Site and Sample Collection

Triplicate samples were collected in two different periods (summer and winter). Summer water samples were collected in July 2023, and winter water samples were collected in December 2023. Sampling was undertaken at five sampling point areas (sites) of the artificial stabilization pond system (SPS), denoted A, B, C, D, and E. These sites were chosen according to the flow path of the SPS and its partitions. For the flow path, A is the inlet of the SPS and E is the outlet of the SPS. For the partitions, A is the ultramicro aeration pond, BCD is the subsurface flow pond, and E is the depth conservation pond. The sites are situated at 31°10′47” latitude north and 120°54′0” longitude north of the SPS. The sites of the stabilization pond system are shown in Figure 1, which shows the sampling location and schematic overview of the SPS located in Kunshan City, Jiangsu Province, China. Composite sampling was performed across different times of the day (morning, afternoon, and late evening). At each of the sampling points, 6 L of pond water was collected within three newly purchased plastic bottles (with a capacity of 2 L each). Before sampling, each of the plastic bottles was first washed with tap water and rinsed three times each with distilled water and pond water to prevent sample contamination. To avoid degradation during on-site sampling and transport, 150 mg/L of ascorbic acid and 0.25 g/L of disodium EDTA were added to each liter of the water sample vial. The addition of EDTA was to chelate the interfering divalent cations such as Ca²⁺ and Mg²⁺ and other heavy metals. During sampling, the sample bottle had to be filled with the sample, leaving no space for the liquid. Immediately after sample collection, hydrochloric acid solution was added to adjust the sample pH to 2 to cease microbial activities. The pH of the sample was adjusted on-site to 2 using hydrochloric acid (0.1 M HCl), (0.1 M HCl) and measured with a pH meter. The samples were carefully placed in a cooling box and subsequently carried to the laboratory, where they were maintained at 4 °C before analysis within 2 days.

2.3. Sample Preparation and Extraction

As the study scope consisted of a wide number of antibiotics, the method of determining multiple antibiotics in a solution was used and consisted of three steps: (1) pretreatment process, (2) calibration configuration, and (3) water sample preparation.

(1) Pretreatment process

A standard configuration of 10 mg of antibiotic standard was accurately weighed, and 10 mL chromatographically pure methanol or acetonitrile was used to dissolve and prepare an antibiotic stock solution with a final concentration of 1 g/L.

(2) Calibration configuration

The antibiotic stock solution was gradually diluted with the initial mobile phase, configured as a series of mixed antibiotic standard samples, and a certain amount of antibiotic internal standard was added to be tested.

(3) Water sample preparation

One liter of water was filtered and extracted using a cartridge (200 mg, 6 mL) of Oasis hydrophilic–lipophilic balance solid-phase extraction (Oasis HLB SPE) (Anpel, Shanghai, China). The HLB cartridge was preconditioned with 10 ml of methanol, 10 ml of pure water, and 10 mL of pure water, then flushed at a flow rate of 1 mL/min. Before extraction, the samples were acidified to a pH of 3–4 using 300 μL of hydrochloric acid. Then, a chelating agent of 0.2 g of Na₂EDTA was added, and subsequently 25 ng of internal standards was spiked. The extraction rate of the SPE was 5 mL/min. After loading, the HLB column was soaked in 10 mL of pure water, flushed, dried for 30 minutes under a stream of mild nitrogen N₂, dissolved three times with 6 ml of methanol, and redissolved in the first mobile phase (0.1% formic acid–ammonium formate aqueous solution/acetonitrile) for analysis.

2.4. Quantification of Antibiotics

Antibiotics were analyzed using an Agilent 6410B triple-quadrupole LC-MS/MS system (Thermo Fisher Scientific, Waltham, MA, USA). The separation was performed with an Agilent C18 column (100 mm × 2.1 mm, 3.5 μm). The column temperature was maintained at 25 °C, the flow rate was 0.25 mL/min, and the injection volume was 200 μL. The mobile phase was composed of phase A (0.1% formic acid–ammonium formate) and phase B (acetonitrile). The linear gradient program was structured as: 0 min, 5% B; 0.1 to 10 min, 10% to 60% B; 10 to 12 min, 60% B; and 12.1 to 22 min, 10% B. The operating parameters for mass spectrometry (MS) were gas temperature at 350 °C, gas flow rate at 8 ml, atomizer pressure at 25 psi, and capillary voltage at 4000 V. The mass spectrometric analysis was realized using a positive electrospray ionization source (ESI+) in multiple reaction monitoring (MRM) modes. The mass spectrometric characteristics of target antibiotics are summarized in Table S1. The experimental workflow of the study is shown in Figure 2.

2.5. Risk Assessment

To evaluate the environmental risk associated with specific antibiotics in the SPS, the risk quotient (RQ) was used and calculated as the ratio of the measured environmental concentration (MEC) to the predicted no-effect concentration (PNEC). The PNEC value was calculated from acute toxicity data (CE50 or CL50) provided by the Environmental Protection Agency (EPA) and then divided by a safety factor of 1000. The values are detailed below [3,29]:

$$\mathrm{PNEC}={\displaystyle \frac{{\mathrm{EC}}_{50}{ \mathrm{or} \mathrm{LC}}_{50}}{1000}}$$

(1)

and:

$$\mathrm{RQ}={\displaystyle \frac{\mathrm{MEC}}{\mathrm{PNEC}}}$$

(2)

In Equation (1), the lowest PNEC of each trophic level was used, and in Equation (2), the highest MEC of all sampling sites was used.

3. Results

3.1. Occurrence of Antibiotics in the Artificial Stabilization Pond System (SPS)

Of the 32 antibiotic compounds from six groups, twelve antibiotics—sulfamethizole (SML), sulfathiazole (STL), sulfamonomethoxine (SMM), sulfadimethoxine (SDM), sulfaquinoxaline (SQN), ciprofloxacin (CIP), fleroxacin (FLX), sarafloxacin (SFX), rifampicin, (RFN) tylosin (TLN), chloramphenicol (CPL), and thiamphenicol (TPL)—were not detected at any sampling point area of the SPS in both seasons. Thirteen antibiotics—SPZ, sulfadiazine (SDZ), roxithromycin (ROX), oxytetracycline (OXY), lomefloxacin (LFLX), lincomycin (LCN), erythromycin (ERY), difloxacin (DFX), enrofloxacin (EFX), chlortetracycline (CLT), azithromycin (AZN), Florfenicol (FFL), and trimethoprim (TMP)—were detected at least in one season. The remaining seven antibiotics—sulfapyridine (SPY), sulfamethazine (SMZ), sulfamethoxazole (SMX), sulfacetamide (STD), ofloxacin (OFX), norfloxacin (NFX), and clindamycin (CLN)—were detected in both seasons (Figure 3). Three antibiotics—sulfamethoxazole (SMX), sulfapyridine (SPY), and ofloxacin (OFX)—were the most frequently detected compounds with the highest detection rate (100%) in both seasons. Clindamycin was also detected with a relatively higher detection rate (80%), while the detection frequency of the rest of the antibiotics in both seasons was below 80%. The detection frequencies and the concentration of antibiotics in the SPS are shown in Figure 3 and Table S2.

As shown above, concerning the sulfonamide (SA) antibiotic group, SMX and SPY showed 100% detection frequency, while SMZ, SDZ, STD, and SPD showed a detection frequency of 63.34%, 36.66%, 16.66%, and 10%, respectively (Figure 3). In addition, more than half of all the investigated SAs were detected, suggesting that sulfonamides are likely to be detected in the aquatic environment especially due to their inherent physicochemical properties and their resistance to being degraded. Previous studies revealed that due to the characteristics of recalcitrance and hydrophilicity contribute to the widespread presence of SAs in surface water [30,31]. The frequent detection of sulfonamides was also recently reported in fishponds [32]. Another recent study also observed high detection of sulfonamides in aquaculture pond and revealed that the high detection of SAs was attributable to the SAs’ recalcitrance and inability of SAs to be absorbed into sludge [33].

In the group of fluoroquinolones (FQs), OFX and NFX showed the highest detection frequency of 100% and 50%, respectively, while the other FQs, EFX, DFX, and LFX, showed a lower detection frequency of 3.33%. The mean concentrations of antibiotics in this group followed the rank order: OFX (10.74 ng/L) > NFX (3.65 ng/L) > EFX (1.94) > DFX (0.094). A recent study on karst groundwater in northern and southern China, Guizhou Province, revealed the highest detection of ofloxacin with a maximum concentration of 1200 ng/L [34]. Ofloxacin was also reported to be the highest detected in sewage sludge of eight different WWTPs, with maximum concentration of 2100 and average concentration of 1100 μg kg^–1 [35]. This was mainly due to the property of adsorption that is observed for the FQs.

In the group of macrolides, ERY had the highest detection frequency (66.6%) among macrolides; however, ROX and AZN had equal detection frequency (50%) (Figure 3). The mean concentrations of antibiotics in this group were ranked as follows: ROX (10.18 ng/L), AZN (2.53 ng/L), and ERY (1.44 ng/L) (Table S2). The analysis revealed that the macrolides (MLs), specifically ROX, ERY, and AZT, were present at low mean concentrations, with ranges of not detected (n.d.)–10.18 ng/L, n.d.–1.44 ng/L, and n.d.–2.53 ng/L, respectively. These concentrations were significantly lower than those found in Australian urban water (0–60 ng/L) [31], but higher than those reported in the southern Yellow Sea (n.d.–1.7 ng/L) [36].

In the tetracycline (TC) group, OXY and CLT were 10% and 6.66% detected, with mean concentrations of 1.83 and 2.54 ng/L, respectively. Tetracyclines are rarely detected in the water phase; however, they are mostly detected in solid matrices (soil, sludge, and sediment) [37]. A previous study reported that tetracyclines are easily broken down and therefore are directly adsorbed into solid matrices such as sediment [38]. This can explain why the detection frequency of OXY and CLT was low in the artificial pond system’s water column.

In the group named “others,” only FFL was detected, at 46.66%, with a mean concentration of 3.36 ng/L. CLN had the highest detection frequency (80%), with a mean concentration of 4.2 ng/L. LCN had a detection frequency of 50% and mean concentration of 0.175 ng/L. Previous studies have reported LCN concentrations ranging from 1 to 50 ng/L in urban water [39] and from 4 to 171 ng/L in surface water in Spain [31], as well as from n.d. to 11.5 ng/L in the Qingcaosha Reservoir [3] compared to the 0.175 ng/L observed in our study (Table S2).

3.2. Seasonal Variation in Antibiotics Within the SPS

Summer and winter are the long-lasting seasons in China, and their longevity may influence the fate and transport of antibiotics within the SPS and other water environments. The seasonal variations of antibiotics from six groups of antibiotics (sulfonamides, fluoroquinolones, macrolides, diaminopyrimidine, tetracyclines, and antibiotics classified as “others”: lincomycin, clindamycin, and florfenicol) were investigated at five sampling sites of the SPS (Figure S1).

Generally, antibiotics were detected in the SPS during both sampling campaigns (summer and winter). Some antibiotics were observed with high concentrations in summer and low concentrations in winter and vice versa. Some other antibiotics were ultimately not observed in summer, but in winter and vice versa, as shown in Figure S1.

3.2.1. Seasonal Variation of Sulfonamides Within the SPS

The mean concentrations of most sulfonamide antibiotics in summer were notably higher than in winter. For example, during summer, the SMX, SMZ, and SPY concentrations were 9.05, 6.19, and 4.85 ng/L, respectively, at sampling point A (inlet of the SPS). In contrast, their concentration in winter was 4.27, 0.14, and 3.15 ng/L, respectively. The concentration of STD was 7.19 ng/L in summer and 0.06 ng/L in winter, respectively, at the inlet of the SPS. The concentration of SPD was n.d.–12.2 ng/L in summer and n.d. in winter at sampling point area B of the SPS. SDZ and TMP were not detected (n.d.) in summer at any sampling point of the SPS; however, they were detected in winter at lower concentrations. For instance, their mean concentration in winter at the inlet of the SPS was 0.42 and 1.27 ng/L (Figure 4). Therefore, the different occurrences of the antibiotics in the SPS for summer and winter may be attributed to different reasons: the physicochemical properties of antibiotics (e.g., antibiotic inherent chemical structure, sorption capability, and acid dissociation constant—pKa, flow conditions of the SPS, consumption patterns) and environmental conditions (e.g., pH, temperature, light intensity).

Sulfonamides exhibit significant resistance to degradation and possess sufficient hydrophilicity to facilitate long-distance transport within aquatic environments. These properties may play a significant role in their elevated detection rates and concentrations within aquatic ecosystems [34,40]. The higher concentration of sulfonamides in summer than in winter may be attributed to their lower adsorption properties into soil and sediments, which can transport them from one medium to another. In addition, high temperatures observed in summer may impact the sorption of SAs by lowering the sorption coefficient, which results in higher mobility in summer than in winter [34].

Summer is a period of abundant rainfall, which may affect the SPS flow conditions. For example, a secondary source of pollution during the summer season, sulfonamides, may be transported to the SPS through runoff, which may increase their concentration in the SPS compared to the period of low rainfall (winter). In addition, water usage is higher in summer than in winter, where high loads of treated effluent are discharged in the SPS, which may increase the concentration of sulfonamides in summer. These findings conform to the previous, which indicated that sulfonamide levels in surface water were on average 2.2 times greater during the warmer summer months than in the colder winter periods. The primary reason for this was the low adsorption of sulfonamides into soil, which facilitates their transportation through surface runoff [41,42,43].

High concentrations of antibiotics in the SPS may be attributed to high consumption and their resistance to degradation. Among SAs, SMZ was the highest detected, with a mean concentration of 19.37 ng/L in summer, followed by SMX with a mean concentration of 10 ng/L in summer (Figure 4). This may imply high usage of these SAs in livestock from the SPS neighborhood. Previous studies reported that the veterinary antibiotics sulfamethoxazole (SMX), sulfachloropyridazine (SCP), and sulfadiazine (SD) were found at the highest frequencies (100%) at rural sites, which were situated near various animal husbandry and aquaculture industry operations [44]. The study also concluded that an increase in population and livestock could lead to higher antibiotic usage and greater water input [43,44]. Another study on the seasonal variation in antibiotics in the surface water of the Pudong New Area of Shanghai revealed that the concentrations of sulfonamides in the winter were lower than in summer. In the winter season, their concentration was observed as follows: sulfadiazine (2.18ng/L), sulfamethiadiazole (4.92ng/L), sulfamethazine (2.57ng/L), sulfamethoxazole (3.44 ng/L); however, in (summer) their concentrations were 3.97, 4.95, 2.63, and 3.90 ng/L, respectively. Therefore, this may be attributed to the higher emission of antibiotics in summer from typical wastewater sources such as pig farms, pharmaceutical manufacturing factories (PMFs), or STPs during the winter season, replacing the dilution effect of high flow [22].

In winter, the concentration of SDZ increases from sampling site A to E (Figure 4). SDZ can both used for both veterinary and human medicine, and the rate of use may increase in winter and lead to its increased concentration in the pond. In winter (dry season), low SPS flow patterns and the water body’s inability to self-purify may keep SDZ residues in the SPS water by augmenting their concentration in the pond’s downstream area [45]. SDZ may be photodegraded to some extent; however, during the winter season, the sunlight intensity decreases, which can increase SDZ persistence and high concentrations down the pond.

3.2.2. Seasonal Variation of Fluoroquinolones Within the SPS

In the group of fluoroquinolones, OFX was detected with the highest mean concentration at all sampling points A, B, C, D, and E of the SPS in summer (21.44, 22, 17.47, 17.28, 5.01 ng/L) vs. winter (9.29, 5.58, 3.08, 3.73, 2.6 ng/L), respectively (Figure 5). Other fluoroquinolones (NFX, EFX, and DFX) were seldom detected in summer or winter; however, similarly, their concentration in summer was higher than in winter. For example, the concentration of NFX in summer at sampling point D was 7.8 ng/L, while in winter, it was 3.245 ng/L (Figure 5). The concentration of EFX in summer at sampling point D was 4.57 ng/L, while in winter, it was not detected. The concentration of DFX in summer at sampling point A was 0.47 ng/L, while it was not detected in winter. These findings are similarly reported by Zhao et al. Reports indicate that the residual concentrations of enrofloxacin and ofloxacin in surface water were higher during the summer compared to winter in water samples collected from the Yellow River Delta, China [46].

Due to the physicochemical properties of ofloxacin, it tends to sorb to the solid matrix (sediments, sludge, and soil [35]). During a period of high SPS water flow (summer), the particulate matter is in movement flow in the water phase section, which does not allow sorbed OFX to settle; thus, a higher concentration of OFX was observed in summer in the water phase than in winter [35]. In winter (dry season), there is low flow in the SPS, and OFX can be sorbed into SPS sediments; thus, its concentration is reduced in the water phase and hence lower concentration was observed in winter. Due to the properties of fluoroquinolones, all other FQs investigated in this study (NFX, EFX, DFX) had similar variation patterns to those of OFX. They were almost not detected in winter and were seldom detected in summer, and their concentration in summer was higher than in winter (Figure 5).

The SPS water pH and high temperatures observed in summer may impact the occurrence and degradation of FQs. A previous study reported that many fluoroquinolones exhibited limited water solubility within the pH range of 6 to 8 and demonstrated reduced susceptibility to microbial degradation when subjected to elevated temperatures [47]. Therefore, the pH of the SPS in the range of 6.8–7.8 and the higher temperature observed in summer (31 °C) than in winter (10 °C) may have also impacted the seasonal detection of FQs, as well as their fate in the SPS water.

In addition, the mean concentrations of OFX are higher than those of other investigated FQs both in summer and winter, which may be attributed to the usage and consumption of OFX in this region. Moreover, the degradability of OFX is different from that of other FQs. Reports similarly indicated that the concentrations of OFX in river water, untreated wastewater, and WWTP discharges were notably elevated in the summer compared to the winter months (p < 0.05). This trend may be attributed to increased OFX usage during the summer season [48]. However, seasonal differences in concentrations of NFX were insignificant (p > 0.05) in both seasons [48]; therefore, FQ concentration is higher in summer, probably because is frequently used for the treatment of intestine infections in summer [44,49].

Furthermore, in our study, sulfonamides and fluoroquinolones have both been found at the highest levels in both summer and winter. This shows the extensive use of these antibiotics in the region. Another study found that SMZ and NOF were the most abundant among the SAs and QNs. NOF has been proven to be widely used in medical treatments [50] and was commonly detected in WWTPs. In effluent from a WWTP, OFX and NFX were 37.6 and 74.5 in summer, while in winter, they were 35.1 and 11.6, respectively [44].

3.2.3. Seasonal Variation of Macrolides Within the SPS

Conversely, all investigated macrolide antibiotics were not detected in summer; however, they were detected in winter. ROX and AZT were not detected at any sampling point of the SPS in summer, but in winter, the mean concentration of ROX was 13.87, 10.53, 10.3, 8.4, and 0.15, ERY 1.89, 2.28, 0.8, and 1.37, and AZT 4.81, 3.57, 1.44, 1.7, and 0.86, respectively, for sampling points A, B, C, D, and E of the SPS (Figure S2).

AZT, ROX, and ERY were not only detected in summer but also in winter (Figure S2). Our findings are similar to previous reports [22,40,44]. For example, Jiang et al. reported that roxithromycin was observed at concentrations ranging from 0.13 to 1.86 ng/L in June (summer) and 1.89 to 9.93 ng/L in December (winter). Pan et al. observed a lower concentration of macrolides in summer than in winter in the surface water of Shanghai Pudong New Area. Consequently, the elevated flow during the summer led to a notable reduction in the concentration of antibiotics in the surface water, aligning with the findings of [20,40]. In addition, the dilution observed in summer is enhanced by the water consumption of urban residents and heavy precipitation that is usually much higher in summer than in winter. This results in a significant dilution in the concentrations of antibiotics in WWTP influent and effluent to be discharged to the SPS [20,44,45,51,52,53].

One possible explanation for the reduced levels of antibiotics during the wet season could be the elevated temperatures and increased sunlight associated with summer (June), which may facilitate the bio- and photodegradation of antibiotics in aquatic environments [54]. Lower concentrations of macrolides were observed in June (Summer) than in March (winter).

It is well known that macrolides are used to treat respiratory human diseases during winter, which results in high concentrations, which was observed. He et al. and Zhang et al. also observed a higher concentration of macrolides in winter than in summer. During the cold weather in the winter season, human beings and livestock use many more macrolide antibiotics to cure respiratory tract infections (e.g., influenza) [44,55].

Moreover, during the winter season, there is less precipitation, a smaller SPS flow, and water body self-purification is unlikely to occur, which results in high concentrations of antibiotics in winter [45].

3.2.4. Seasonal Variation in Diaminopyrimidine Within the SPS

Trimethoprim is a diaminopyrimidine antibiotic that is prescribed in combination with sulfamethoxazole [56]. Our study observed that trimethoprim was only detected in winter, with a concentration of 0.93–1.27 ng/L. Its detection patterns were similar to those observed for macrolides. These detection values may be due to two factors. First, its detection in winter may be caused by the increased antibiotic consumption in winter to treat bacterial infections of the respiratory tract. Another reason is that in summer, trimethoprim may not be used in huge quantities, and thus it is not detected anymore.

Furthermore, in summer, TMP may bear a great extent of degradation and may not be detected. For example, much water consumption and heavy precipitation during the summer may gradually reduce TMP concentrations by dilution. The higher sunlight intensity in summer and temperature may enhance the extent of photodegradation and biodegradation of TMP, which may result in lower detection. Our findings also agree with those previously reported by [57]. Prior investigations suggested that the half-lives of two common antibiotics, sulfamethoxazole and trimethoprim, were reduced by a factor of two for sulfamethoxazole and a factor of 10 for trimethoprim under summer sunlight conditions compared to sunlight non-wastewater matrices. Effluent organic matter (EfOM) and nitrate were identified as the principal photosensitizers in the wastewater matrix, while the natural organic matter (NOM) present in the surrounding surface water was determined to be an ineffectual photosensitizer [57,58].

3.2.5. Seasonal Variation in Tetracyclines Within the SPS

For the tetracyclines group, OXY was not detected at any sampling point in summer; however, its concentration in winter was n.d.–6.29 ng/L. In contrast, CLT was not detected at any sampling point in winter; however, its concentration in summer was n.d.–6.82 ng/L. Tetracyclines are mainly used to treat animal diseases. Oxytetracycline and chlortetracycline were seldom detected in the SPS, with a detection frequency of 10% and 6.6%, respectively. This perhaps reflects tetracyclines not being used in this region, as no livestock farms surround the artificial stabilization pond system (SPS). In addition, tetracyclines are known to be likely adsorbed into the solid matrix (sludge, solid, or sediments) and depleted by abiotic processes (photodegradation and hydrolysis) and biotic process biodegradation. The low detection of tetracyclines in the water column of the SPS may be primarily caused by self-attenuation processes such as adsorption and photodegradation, as the literature indicates that adsorption and photodegradation are the primary elimination processes [2,37,59]. Chabilan et al. reported a contribution ranging from 44.4% to 52.0% for the adsorption of all tetracyclines. Photodegradation played a comparable role for doxycycline and oxytetracycline, accounting for 41.3% to 43.0%, while hydrolysis contributed 12% [60].

Similarly, based on other studies, tetracyclines are infrequently detected in natural water, which is attributed to their significant adsorption into sediments and particles, as well as their degradation processes. In earlier studies, tetracycline was not found in surface water in Germany or groundwater near confined animal feeding operations in the US [40].

3.2.6. Seasonal Variation in Antibiotics Classified as “Others” Within the SPS

Antibiotics classified as “others” included lincomycin, clindamycin, and florfenicol. Lincomycin detection patterns were the same as for macrolides. The mean concentration of lincomycin in winter was 0.16–0.2 ng/L for all sampling points of the SPS (Figure S3). Clindamycin was detected in both summer and winter. The concentration of clindamycin in summer at all points of the SPS ranged from n.d. to 2.08 ng/L, while in winter, it ranged from n.d. to 8.78 ng/L. Clindamycin and lincomycin belong to the same antibiotic class, “lincosamide antibiotics.” Clindamycin is a lincosamide antibiotic employed in the therapy of head and neck infections, including acute sinusitis, otitis media, and pharyngitis, generally attributed to aerobic or facultative anaerobic respiratory flora such as Streptococcus pneumoniae, Staphylococcus aureus, and Haemophilus influenza [61,62]. Lincomycin is employed to address certain bacterial infections. It is efficacious in pulmonary disorders (e.g., pneumonia), pharyngeal conditions, etc. In our study, clindamycin concentrations were greater in winter (8.78 ng/L) compared to summer (4.04 ng/L). Lincomycin was detected only in winter, with the highest mean concentration 0.2 ng/L (Figure S3). Therefore, this higher detection of these lincosamide antibiotics may be directly linked to their consumption during winter to treat infection, as more people become sick during this period due to cold weather.

Florfenicol detection patterns were the same as for CLT. It was not detected at any sampling point in winter; however, it was detected in summer at all sampling points of the SPS, with a concentration of 3.09–3.59 ng/L (Figure S3). FFL was detected only in summer, with a mean concentration of 3.09–3.59 ng/L. This occurrence might be linked to its consumption during this period. Lei et al. found that FFL levels were higher in summer in WWTP influent and effluent than in winter and concluded that there was increased usage and release of most of the FFL during the wet season (summer) compared to the dry season (winter) along the catchment based on the measurements [48].

In brief, the occurrence and antibiotic persistence in the SPS water in different seasons (summer and winter) is influenced by physicochemical (e.g., inherent antibiotic chemical structure, sorption) and environmental factors (SPS water pH, SPS flow patterns, rainfall, temperature, light intensity, and microbial activity). The frequency of antibiotic use for various purposes across different seasons leads to fluctuations in antibiotic concentrations throughout the year.

4. Environmental Risk Assessment of Antibiotics

This study evaluated the potential ecological risks associated with antibiotics by employing the risk quotient (RQ) approach, in accordance with a European technical guidance document (TGD) on risk assessment. The RQ was established as the ratio between the measured environmental concentration (MEC) and the predicted no-effect concentration (PNEC). The PNEC value was evaluated using the toxicity data derived from the ecological structure–activity relationship (ECOSAR), as presented in Table 2. In order to effectively differentiate the ecological risk levels, the individual RQ values were categorized into four distinct levels: <0.01: no risk; 0.01–0.1: low risk; 0.1–1: medium risk; >1: high risk [15].

The risk quotients (RQs) of antibiotics in the artificial stabilization pond system (SPS) are shown in Figure 6 and Table S3. According to the observed RQs, four antibiotics, SMX, EFX, OXT, and ERY, posed a low risk in winter of S. capricornutum, M. aeruginosa, M. aeruginosa, and D. magna, respectively. OFX only posed a medium risk in winter of M. aeruginosa.

In summer, SMX, SMZ, EFX, and CLT posed low risk to the following species: S. capricornutum (SMX); L. gibba and M. macrocopa (SMZ); M. aeruginosa (EFX); and P. subcapitata (CLT). NFX and CLT posed medium risk of the following species in summer: M. wesenbergii (NFX) and M. macrocopa (CLT). In summer, only OFX posed a high risk to M. aeruginosa. As can be observed, OFX posed a medium risk in winter and a high risk in summer to algae (M. aeruginosa). Previous studies have also reported ecological risk results. OFX was observed with a high-risk level in the mainstream and tributaries of the Yangtze River [63]. Hena et al. reported OFX poses high risks of algae, and this may have an impact on their growth [64].

Therefore, specific measures should be taken to limit exposure to algae within the artificial pond system.

Normally, RQs in summer were remarkably higher than in winter. For example, NFX caused medium risk in summer, while no risk was observed in winter. OFX caused high risk in summer, while medium risk was observed in winter. CLT caused medium risk, while no risk was observed in winter (Figure 6). Since the artificial pond water is discharged to a river, these residual traces of antibiotics may also spread resistance genes in the surface water.

Research has shown that residual trace antibiotics in aquatic environments can create selective pressure on microbial communities, potentially accelerating the dissemination of antibiotic-resistance genes (ARGs) [65]. The most prevalent resistance genes found in the aquatic environment include sul I, sul II, tet (C), and tet (G). For instance, the absolute abundance of the sul and tet class genes varied from 10¹⁰ to 10¹² copies/L in raw water collected from the Yangtze River Delta [66]. A total of 11 ARGs were identified at elevated levels, with sul II recorded at a peak concentration of 4.19 × 10⁸ copies/L in a drinking water source located in Shanghai. This phenomenon may indicate the extensive application of sulfonamides in this area [67]. Additionally, the transfer of ARGs among bacteria can occur via transposons, plasmids, and integrons [68]. The observed abundance of ARGs showed a significant correlation with the levels of mobile genetic elements, suggesting that intI-1 and transposons could play a role in the abundance of ARGs in drinking water [69,70,71]. Hence, the prevalence of the antibiotics present in the SPS may pose ecological risks in the water column, as well as the potential for the spread of antibiotic-resistance genes, warranting further investigation in future studies.

Table 2. Toxicity data for algae, invertebrates, and fish.

Antibiotic Compound	Taxonomic Group	Species	Toxicity Data (mg/L)	PNEC (ng/L)	References
Sulfamethoxazole	Algae	C. meneghiniana	0.00125	1250	[72]
	Algae	S. capricornutum	0.0000146	146
	Invertebrate	D. magna	0.0252	25,200
	Fish	D. rerio	0.008	8000
Sulfamethazine	Algae	L. gibba	0.001277	1277	[73]
	Invertebrate	M. macrocopa	0.001277	1277
	Fish	Danio rerio	0.0326	32,600	[74]
Sulfapyridine	Algae	Freshwater algae	0.02079	20,790
	Invertebrate	D. magna	0.0007	700
	Fish	Zebrafish embryos	0.0454	45,400
Sulfachloropyridazine	Algae	Freshwater algae	0.02808	28,080
	Invertebrate	D. magna	0.0008	800
	Fish	Zebrafish embryos	0.0789	78,900
Sulfacetamide	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.	N.O.
Sulfadiazine	Algae	S. capricornutum	2.2	2200	[75]
	Invertebrate	N.O.	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.	N.O.
Trimethoprim	Algae	R. saline	0.016	16,000	[72]
	Algae	M. aeruginosa	0.112	112,000
	Algae	S. capricornutum	0.0255	25,500
	Invertebrate	D. magna	0.06	6000
	Fish	B. rerio	0.1	100,000
Ofloxacin	Algae	M. aeruginosa	0.021	21	[76]
	Invertebrate	C. dubia	3.13	3130
	Fish	D. rerio	>1000	1,000,000
Norfloxacin	Algae	M. wesenbergii	0.038	38	[77]
	Invertebrate	D. magna	0.88	880	[78]
	Fish	N.O.	20,081.355	20,081,355	ECOSAR
Enrofloxacin	Algae	M. aeruginosa	0.000049	49	[72]
	Invertebrate	D. magna	0.01	10,000
	Fish	O. mykiss	0.01	10,000
Difloxacin	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.
Clindamycin	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.
Lincomycin	Algae	P. subcapitata	0.07	70	ECOSAR
	Invertebrate	Thamnocephalus platyurus	33	33,000	ECOSAR
	Fish	Danio rerio	1000	10,000,000	ECOSAR
Roxithromycin	Algae	Freshwater algae	0.00466	4660	[74]
	Invertebrate	D. magna	0.006	6000
	Fish	D. magna	0.023	23,000
Erythromycin	Algae	Freshwater algae	0.0023	2300
	Invertebrate	D. magna	0.0002	200	[79]
	Fish	D. rerio	0.0615	61,500	[76,80]
Azithromycin	Freshwater algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.
Oxytetracycline	Algae	M. aeruginosa	0.23	230	ECOSAR
	Invertebrate	N.O.	3.08	308,000
	Fish	Oryzias latipes	50	500,000
Chlortetracycline	Algae	P. subcapitata	0.00017	170	[81]
	Invertebrate	M. macrocopa	0.00005	50	[73]
	Fish	M. saxatilis	0.15	150,000
Florfenicol	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.

PNECs were calculated from the toxicity data using an assessment factor (AF) of 1000 for acute toxicity. ECOSAR: ecological structure–activity relationship. N.O.: not observed.

Table 2. Toxicity data for algae, invertebrates, and fish.

Antibiotic Compound	Taxonomic Group	Species	Toxicity Data (mg/L)	PNEC (ng/L)	References
Sulfamethoxazole	Algae	C. meneghiniana	0.00125	1250	[72]
	Algae	S. capricornutum	0.0000146	146
	Invertebrate	D. magna	0.0252	25,200
	Fish	D. rerio	0.008	8000
Sulfamethazine	Algae	L. gibba	0.001277	1277	[73]
	Invertebrate	M. macrocopa	0.001277	1277
	Fish	Danio rerio	0.0326	32,600	[74]
Sulfapyridine	Algae	Freshwater algae	0.02079	20,790
	Invertebrate	D. magna	0.0007	700
	Fish	Zebrafish embryos	0.0454	45,400
Sulfachloropyridazine	Algae	Freshwater algae	0.02808	28,080
	Invertebrate	D. magna	0.0008	800
	Fish	Zebrafish embryos	0.0789	78,900
Sulfacetamide	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.	N.O.
Sulfadiazine	Algae	S. capricornutum	2.2	2200	[75]
	Invertebrate	N.O.	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.	N.O.
Trimethoprim	Algae	R. saline	0.016	16,000	[72]
	Algae	M. aeruginosa	0.112	112,000
	Algae	S. capricornutum	0.0255	25,500
	Invertebrate	D. magna	0.06	6000
	Fish	B. rerio	0.1	100,000
Ofloxacin	Algae	M. aeruginosa	0.021	21	[76]
	Invertebrate	C. dubia	3.13	3130
	Fish	D. rerio	>1000	1,000,000
Norfloxacin	Algae	M. wesenbergii	0.038	38	[77]
	Invertebrate	D. magna	0.88	880	[78]
	Fish	N.O.	20,081.355	20,081,355	ECOSAR
Enrofloxacin	Algae	M. aeruginosa	0.000049	49	[72]
	Invertebrate	D. magna	0.01	10,000
	Fish	O. mykiss	0.01	10,000
Difloxacin	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.
Clindamycin	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.
Lincomycin	Algae	P. subcapitata	0.07	70	ECOSAR
	Invertebrate	Thamnocephalus platyurus	33	33,000	ECOSAR
	Fish	Danio rerio	1000	10,000,000	ECOSAR
Roxithromycin	Algae	Freshwater algae	0.00466	4660	[74]
	Invertebrate	D. magna	0.006	6000
	Fish	D. magna	0.023	23,000
Erythromycin	Algae	Freshwater algae	0.0023	2300
	Invertebrate	D. magna	0.0002	200	[79]
	Fish	D. rerio	0.0615	61,500	[76,80]
Azithromycin	Freshwater algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.
Oxytetracycline	Algae	M. aeruginosa	0.23	230	ECOSAR
	Invertebrate	N.O.	3.08	308,000
	Fish	Oryzias latipes	50	500,000
Chlortetracycline	Algae	P. subcapitata	0.00017	170	[81]
	Invertebrate	M. macrocopa	0.00005	50	[73]
	Fish	M. saxatilis	0.15	150,000
Florfenicol	Algae	N.O.	N.O.	N.O.	N.O.
	Invertebrate	N.O.	N.O.	N.O.
	Fish	N.O.	N.O.	N.O.

PNECs were calculated from the toxicity data using an assessment factor (AF) of 1000 for acute toxicity. ECOSAR: ecological structure–activity relationship. N.O.: not observed.

5. Discussion

5.1. Antibiotic Occurrence and Seasonal Variation Within the SPS

Sulfonamides showed higher concentrations in summer than winter, likely due to their resistance to degradation and low sorption, which may have resulted in their transportation through various environmental media. Fluoroquinolones were also more prevalent in summer, particularly OFX, indicating extensive regional usage and persistence. In addition, runoff due to higher precipitation in summer and changes in SPS flow patterns due to higher precipitation and high water usage may contribute to the high detection observed for both sulfonamides and fluoroquinolones in summer. Macrolides were not detected in summer, possibly due to dilution from summer water flow (high water usage and high precipitation) and increased degradation under summer conditions (high temperature and light intensity). Macrolides were detected in winter due to low flow patterns, light intensity, and temperature, which may decrease the rate of photodegradability and microbial activity. In addition, high antibiotic consumption in winter to treat influenza may be another reason for their high detection. Trimethoprim showed a similar pattern to macrolides, likely reflecting increased winter usage for respiratory infections. Oxytetracycline and chlortetracycline were seldom detected, suggesting limited use in the region, as there are no nearby livestock farms surrounding the SPS. Lincomycin and clindamycin were predominantly detected in winter, possibly due to higher consumption during colder months to treat infections such as influenza. Florfenicol was absent in winter, but detected in summer, suggesting greater usage and release during the wet season.

An environmental risk assessment of investigated antibiotics in the SPS was also conducted. Most of the investigated antibiotics posed no risk in either season; however, in summer, norfloxacin (NFX) posed medium risk on the species such as M. wesenbergii whereas OFX posed high risk in summer and medium risk in winter to algae (M. aeruginosa). This shows that special attention could be given to fluoroquinolones in order to reduce their environmental risks in SPS and downstream surface water (e.g., river)

Therefore, understanding the seasonal variability in antibiotic concentrations in the SPS is essential for designing targeted monitoring programs and implementing appropriate mitigation measures. Management strategies should consider the dynamic nature of antibiotic contamination, accounting for seasonal influences on the fate and transport of antibiotics within stabilization pond systems. This may involve adjusting operational practices, optimizing treatment processes, and implementing source control measures to mitigate the environmental impacts of seasonal antibiotic variability.

5.2. Summary of the Stabilization Pond Configuration and Mechanism

The treatment of waste stabilization ponds through algal–bacterial symbiosis is a biological process aimed at attenuating contaminant levels in wastewater. This includes a reduction in organics, nutrients, heavy metals, and various pollutants facilitated by the mechanism of algal–bacterial symbiosis [82,83]. Various types of bacteria found in the pond, including Achromobacter, Proteus, Alcaligenes, Pseudomonas, Thiospirillum, and Rhodothecae, oxidize organic matter, resulting in the release of CO₂, H₂O, ammonia, and nitrates [84]. The growth of algal biomass occurs through the utilization of sunlight and CO₂, while also leveraging the byproducts generated from the decomposition of organic matter in the pond. The presence of these compounds enhances the growth of algae, thereby promoting its photosynthesis process and leading to a subsequent rise in oxygen levels within the pond [14]. Chlorella, Euglena, Scenedermus, and Microcystis represent the predominant species of algae found in waste stabilization ponds. With the increasing growth of algae in the pond, oxygen is generated through photosynthesis, which is facilitated by sunlight. This heightened rate of photosynthesis allows bacteria to decompose more waste, thereby lowering the organic levels in the water. Mechanical aerators are incorporated into the SPS to enhance oxygen diffusion from the atmosphere and may improve the efficiency of antibiotic removal [12]. In a pond, the symbiotic relationship between algae and bacteria is referred to as a redox process, and such a pond is identified as a redox pond.

Therefore, the effectiveness of treatment in the SPS is significantly affected by the characteristics of the incoming treated wastewater effluent, the level of organic loading, the design and physical layout of the pond system, and the hydraulic dynamics involved. The performance of SPS treatment is influenced by various physical, chemical, and biological processes, including sedimentation, adsorption, biodegradation, photodegradation, and hydrolysis, all of which are affected by environmental factors like the presence of algae, light intensity, pH, and temperature [14].

Compared to previous studies, the detected concentration in the SPS was much lower than that of wastewater stabilization ponds in Tanzania, India, Ghana, and in the aquaculture pond (China)

Table 3. Comparison of detected concentrations (ng/L) within the SPS and other ponds worldwide.

Antibiotic Compounds	SPS	Waste Stabilization Pond	Waste Stabilization Pond	Aquaculture Pond	Waste Stabilization Pond	Aquaculture Pond	Aquaculture Pond
	China	Tanzania	India	China	Ghana	China	Italy
SMX	10.51	336	0–2000	121.34	103–320	0.02–30,792	0.74
SMZ	19.37	–	20–120	–	–	0.02–162.5	0.1
SPY	6.93	–	–	–	–	–	–
SPZ	12.2	–	–	–	–	–	–
STD	7.19	–	–	–	–	–	–
SDZ	0.42	–	–	138.75		0.02–984	145.06
TMP	1.27	8480	3–710	–	31–255	0.02–106.7	34.5
OFX	22	–	–	189.57	–	0.38–2208	–
NFX	7.8	–	30–251,000	185.70	–	0.1–160.95	0.02
EFX	4.57	–	0–488	218.12	–	0.07–4863	–
DFX	0.47	–	–	–	–	–	–
LFXL	0.94	–	–	–	–	–	–
ROX	13.87	–	–	–	–	0.02–40.33	–
ERY	2.28	–	1.41–15.9	–	47–882	–	1.1
AZT	4.81	–	–	–	–	–	–
OXT	6.29	–	110–4200	–	2.4–24	0.31–1015	18.2
CLT	6.82	–	0–600	–	6.0–19	0.1–670.04	–
CLN	8.78	–	–	–	–	–	–
LCN	0.2	–	3.13–248.9	–	–	–	–
FFL	3.5	–	–	–	–	–	–
Reference	This study	[85]	[1]	[86]	[87]	[88]	[89]

–: Not analyzed.

. This is primarily due to these ponds directly receiving wastewater from pollution sources (homes, hospitals, or industries) without any prior treatment, because in many developing countries, waste stabilization ponds are still used as the sole wastewater treatment technology, while the SPS receive treated effluents from the Jinxi wastewater treatment plant. In addition, antibiotic discharge standards in developing countries have not yet been implemented, and thus higher concentrations of antibiotics are still observed in their aquatic environment. Aquaculture ponds in China show higher concentrations of antibiotics than those observed in an aquaculture pond in Italy. This may be primarily due to every country having a different amount and type of antibiotics, as well as different guidelines for antibiotic application in aquaculture.

6. Conclusions

With a focus on long-lasting seasons (summer and winter), the influence of seasonal variation on antibiotic concentrations and their fate and transport in an artificial stabilization pond system (SPS) was investigated for 32 antibiotics using an Agilent 6410B triple-quadrupole LC-MS/MS (Thermo Fisher Scientific, Waltham, MA, USA). Twelve antibiotics were not detected in any sampling sites in summer or winter, thirteen antibiotics were detected one season, and seven antibiotics were detected in both seasons. Sulfonamides and fluoroquinolones were the most detected antibiotic classes in the SPS. It was noted that the prevalence and persistence of antibiotics in SPS water across the seasons (summer and winter) are affected by physicochemical factors (e.g., the antibiotic’s inherent chemical structure, sorption properties) and environmental variables (such as the pH of SPS water, flow patterns, rainfall, temperature, light intensity, and microbial activity). In addition, the seasonal variation in antibiotic usage for diverse purposes results in changes in antibiotic concentrations between seasons.

Therefore, seasonal variability significantly influences the fate and transport of antibiotics in the artificial stabilization pond system (SPS) by affecting their concentrations, degradation kinetics, sorption behavior, and ecological interactions. Recognizing these seasonal dynamics is crucial for effectively managing antibiotic contamination in aquatic environments.

According to the findings from the environmental risk assessment study, ofloxacin (fluoroquinolone class) was observed to pose high risks to algae within the SPS, which subsequently require immediate attention.

This work substantially enhances the understanding of the influence of seasonal variations on the persistence, distribution, and elimination of antibiotics in aquatic ecosystems. The findings have significant implications for advancing water and wastewater treatment technologies in various respects, including optimizing treatment design for antibiotic elimination; refining risk assessment and regulatory frameworks; promoting nature-based solutions (development of sustainable, low-energy treatment methods that improve antibiotic removal); and crucially, mitigating the dissemination of antibiotic resistance.

Future perspectives: Although the other seasons (autumn and spring) are short, their influence on the fate and transport of antibiotics in stabilization pond systems cannot be overlooked. Future research could also focus on these seasons. The pond configuration is also an important feature that may improve the attenuation of antibiotics in stabilization ponds in different seasons, thereby impacting the antibiotics’ fate and transport. Future studies could also investigate the extent to which algae–bacteria symbiosis in SPS contributes to antibiotic removal throughout the seasons.