Introduction
N-Nitrosamines (NAs) are a newly recognized category of pollutants in drinking water. NAs include N-nitrosopiperidine (NPIP), N-nitrosomorpholine (NMOR), N-nitrosomethylethylamine (NMEA), N-nitrosodibutylamine (NDBA), N-nitrosodipropylamine (NDPA), N-nitrosodiethylamine (NDEA), N-nitrosodimethylamine (NDMA), and N-nitrosopyrrolidine (NPYR). Six of these compounds (NMEA, NDBA, NDPA, NDEA, NDMA, and NPYR) represent an excess lifetime cancer risk of 10-5 at concentrations of 2-200 ng/L according to the Integrated Risk Information System (IRIS) of the U.S. Environmental Protection Agency (EPA) [1–7]. These six compounds have been categorized as Group B2, indicating their potential to induce cancer in humans. The International Agency for Research on Cancer (IARC) has classified two compounds (NDEA and NDMA) as Group 2A (probably carcinogenic to humans), while the other compounds (NPIP, NMOR, NMEA, NDBA, NDPA, and NPYR) are categorized as Group 2B (potentially carcinogenic to humans) [8]. NDMA is the most frequently detected compound in drinking water and a potent carcinogen. In 2006, the Office of Environmental Health Hazard Assessment (OEHHA) set a public health goal of 3 ng/L for NDMA in drinking water [9]. Furthermore, the World Health Organization (WHO) set a guideline value (GV) of 100 ng/L, associated with an upper-bound excess lifetime cancer risk of 10-5 [10]. The Ministry of Environment (ME) of the Republic of Korea has established a monitoring goal for NDMA and NMEA in drinking water to ensure that their concentrations do not exceed 100 ng/L, as stipulated by the GV [10–11].
In drinking water treatment plants (DWTPs), NDMA occurrence was observed after oxidation using chloramines and ozone. The concentrations ranged from non-detected to 630 ng/L in the EPA’s Unregulated Contaminant Monitoring Regulation Monitoring Program [13] and in studies conducted in China, Japan, and Taiwan [14–17]. NDMA, NDEA, and NDBA are the most frequently detected NAs in raw water in China, with concentration ranges of 6.4–13.9 ng/L, 1.9–16.3 ng/L, and 1.0–19.9 ng/L, respectively. Their concentrations in the finished water range between 4.6 and 20.5 ng/L, 1.9 and 16.3 ng/L, and 0.4 and 3.4 ng/L, respectively. Furthermore, NDMA was reported to increase from undetectable levels to 8.9 ng/L at DWTPs [18].
Secondary aliphatic amines (SAAs) are considered the organic precursors of NAs. Dimethylamine (DMA) and diethylamine (DEA) were detected at concentrations ranging from 200 to 3900 ng/L and 300 to 2400 ng/L in raw water, and 400 to 4000 ng/L and 100 to 1800 ng/L in treated water, respectively [18]. The potential correlation between SAAs and NA formation has been investigated. The absence of DMA in chlorinated water resulted in the non-detection of NDMA. By contrast, NDMA exhibited the highest formation rate (229 µ M/h) when monochloramine and DMA were present in the water [19]. Thus, a relationship between SAAs and NAs is evident based on their detection rate in untreated water, as supported by a coefficient of determination (r2) of 0.87 [18]. In a study conducted in Spain, raw water exhibited an average NDMA concentration of 1.3 ng/L. Subsequently, the concentrations increased to 1.5 ng/L after peroxidation/coagulation, 3.9 ng/L after filtration and chlorination, 4.0 ng/L in the reservoir, and peaked at 4.5 ng/L in the distribution network [20].
Several studies have investigated the formation and precursors of NAs during water treatment. Initially, it was suggested that nitrosation occurs through the reaction between SAAs and nitric oxide; however, this mechanism is notably slow. Mitch and Sedlak [19] proposed an alternative mechanism in which monochloramine reacts slowly with DMA, generating unsymmetrical dimethylhydrazine (UDMH), which can be readily oxidized, resulting in a low NDMA yield (<1%) [21,22]. Choi and Valentine [23] reported NDMA formation through the reaction between DMA and nitrite, noting a yield six times lower than that of the reaction between DMA and monochloramine. The mechanism was modified to emphasize the importance of dichloramine as the primary reactant for generating NDMA. Chlorinated UDMH, a compound produced through the reaction of DMA with dichloramine, is thought to react with O2, dichloramine, or peroxynitrous acid [24]. These reactions yield by-products from the degradation of dichloramine, resulting in the formation of NDMA.
In addition to residual chlorine, various factors responsible for NDMA formation have been documented, including natural organic matter (NOM), inorganic nitrogen, dissolved organic nitrogen (DON), bromide ion, and dissolved oxygen. Hydrophilic bases show the highest tendency for NDMA formation. Furthermore, hydrophobic acids containing high concentrations of dissolved organic carbon (DOC) are significant contributors (up to 72%) to NDMA formation. Inorganic nitrogen species, such as nitrite, nitrate, and other inorganic salts, are known for their crucial role in nitrosation. An optimal pH of 3.4 is required for NDMA formation through nitrosation. The formation of NDMA and the DOC/DON ratio are correlated, indicating that the NDMA concentration decreases as the DOC/DON ratio increases [25]. DON also demonstrates seasonal fluctuations, increasing as the water temperature rises. In a nationwide survey conducted in Japan, Asami et al. [14] reported that the NDMA concentration in untreated water was 4.6 ng/L, which increased to 280 ng/L following ozonation treatment. High concentrations of nitrogen species, such as ammonia, nitrite, and nitrate, were observed in finished water from a DWTP in the Yodo River Basin. NA formation not only depends on the chlorine dose but also on other precursors; therefore, various precursors should be considered when developing strategies to control NA formation.
In Korea, tap water is disinfected using free chlorine, sodium hypochlorite, or chlorine dioxide, and undergoes a series of treatment steps including pre-chlorination, coagulation, flocculation, sedimentation, sand filtration, and postchlorination [26]. Chlorination is crucial for eliminating pathogenic bacteria; therefore, free and combined chlorine levels in tap water should be maintained above 0.1 and 0.4 mg/L, respectively. Water is pre-chlorinated to reduce the organic matter content and eliminate microorganisms, whereas post-chlorination is performed to eliminate any remaining pathogens. Previous studies have reported that NAs can be generated due to the influence of chloramines, which are produced by the reaction of chlorine and ammonia in chlorinated water [23,27–29]. Upon completion of the water treatment process, water is transported from the reservoir to the faucet through the distribution system. This process helps generate disinfection byproducts (DPBs) through reactions with residual disinfection agents during the contact time. Maqbool et al. [30] validated an increase of 0–20% in the NDMA formation potential (NDMA-FP) within the water distribution system. Thus, examining the occurrence of NAs in untreated and treated water is crucial, considering their progressive formation in chlorinated water.
This study aimed to investigate the spatial distribution of NAs in drinking water at the Soyang drinking water treatment plant (DWTP) and in tap water distributed to Chuncheon, Gangwon-do, Republic of Korea. The goal was to investigate variations in NA concentrations in chlorinated water over a specific time frame and analyze potential correlations with physicochemical parameters.
Materials and Methods
Reagents and materials
All standard solutions of NAs, SAAs, N-ethylbutylamine (EBA), and N-nitrosomethylbutylamine (NMBA) were obtained from Sigma-Aldrich (St. Louis, MO, USA). Stock solutions were individually prepared in methanol at a concentration of 1000 mg/L. EBA and NMBA were used as surrogate standards (SS) for quantitative analysis. Carboxen 572, sodium thiosulfate, and benzenesulfonyl chloride were also purchased from Sigma-Aldrich. All the organic solvents used in this study, including acetone, methanol, and dichloromethane, were of HPLC grade and were purchased from Honeywell Burdick & Jackson Inc. (Muskegon, MI, USA). Sodium hydroxide, sodium bicarbonate, sodium carbonate, sodium sulfate, hydrochloric acid (35%), and acetic acid were obtained from Daejeong (Siheung, Korea). Dansyl chloride (5-(dimethylamino)naphthalene-1-sulfonyl chloride) was obtained from Calbiochem (San Diego, CA, USA). Hydrobromic acid (48%) was obtained from Wako (Osaka, Japan).
To prepare the analytical samples, a denitrosation solution was prepared by diluting 1 mL of 48% hydrobromic acid with acetic acid to a final volume of 10 mL. A dansylation solution was prepared by diluting 25 mg of dansyl chloride in 50 mL of acetone. A buffer solution with a pH of 10.5 was prepared by dissolving 0.6 g of sodium hydroxide and 2.0 g of sodium bicarbonate in 50 mL of ultrapure water. All solutions were stored at 4°C and used within 2 weeks.
Collection of water samples
Water samples were collected from July to December 2014 to examine the spatial distribution of NAs, SAAs, and factors influencing the formation of NAs such as chlorine, pH, temperature, DOC, total nitrogen (T-N), nitrate, and NH3-N. Sample information is presented in Table S1. The samples were collected from the Soyang DWTP near the Soyang River and five community centers (S1–S5) in Chuncheon, Gangwon-do, Republic of Korea (Figure 1). At the Soyang DWTP, water samples were obtained at various stages of the treatment process: 1) raw water, 2) after sedimentation (pre-chlorination), 3) after sand filtration, and 4) finished water (post-chlorination). Tap water samples were collected from community centers located approximately 0.9–6.2 km from the DWTP. The treated water was delivered to the tap within 3 h. Water samples were collected 12 times from the DWTP and community centers, resulting in 108 samples. The Soyang intake station, located adjacent to the Soyang River, provides water to the DWTP. Finished water samples were collected from the Soyang DWTP to analyze the temporal distribution of NAs and their precursors. Fifty-two samples were collected and divided into four sets, each containing 13 samples collected at specific time intervals (0, 1, 2, 3, 4, 5, 6, 7, 8, 12, 16, 20, and 24 h). Sodium thiosulfate (0.2 g) was added to each sample to neutralize the residual chlorine.
Water samples were collected in 1 L brown glass bottles without headspace, and approximately 0.2 g of sodium thiosulfate was added to each sample. Water samples for analyzing nitrate, DOC, and UV254, T-N, and ammoniacal nitrogen (NH3-N) were collected in 40 mL dark brown glass vials, ensuring no headspace was present. All samples were transported to the laboratory in a cooler and stored in a refrigerator at 4 °C until analysis. All samples were analyzed within two weeks of collection.
Quantitative analysis
N-Nitrosamines
The pretreatment preceding the analysis of NAs was a modified version of the U.S. EPA Method 521 [31,32]. Each sample (500 mL) was enriched with 1 µ L of NMBA (20 mg/L, SS). The sample was passed through a cartridge containing 2.0 g of Carboxen 572 at a flow rate of 10 mL/min. Subsequently, a cartridge was connected to an impinger with 16 g of silica gel and dried using a vacuum pump (-34 kPa) for 60 min. The analytes (NAs) were eluted from the cartridge using 15 mL of dichloromethane at a flow rate of 4 mL/min. The eluate volume was reduced to 1 mL using a gentle stream of nitrogen.
For instrumental analysis, NAs were converted into dansyl amine derivatives through denitrosation and dansylation reactions (Figure 2). After adding 150 µ L of the denitrosation solution, the concentrate was vigorously mixed for 10 s. The resulting mixture was incubated at 40 °C for 30 min and dried under a gentle stream of nitrogen. Subsequently, 150 µ L of a buffer solution with a pH of 10.5 and 150 µ L of the dansylation solution were combined with the sample and agitated vigorously for 10 s. The mixture was then incubated at 40 °C for an additional 30 min, and 50 µ L of ultrapure water was added. The final sample was transferred to a 1 mL vial and injected into a high-performance liquid chromatograph-fluorescence detector (HPLC-FLD) for analysis. The instrumental conditions and validation results are presented in Tables S2 and S3.
Secondary aliphatic amines
SAAs were derivatized into benzenesulfonamides and analyzed using gas chromatography-mass spectrometry (GC-MS), following the methodology of Park et al. [33]. A 200 mL water sample was spiked with 1 μL of EBA (1015 mg/L, SS). During derivatization, the sample was mixed with 8 mL of 10 M sodium hydroxide and 2 mL of benzenesulfonyl chloride. The mixture was agitated at room temperature for 30 min. Afterward, 5 mL of a 10 M sodium hydroxide solution was added, and the mixture was shaken at 80 °C. After cooling, 18.5% hydrochloric acid was added to adjust the pH to 5.5. The sample was extracted with 25 mL of dichloromethane by agitating for 30 min using a strong shaker (TAITEC SR-2 DS, Koshigaya, Japan), and this process was repeated twice. After separating the organic layer (approximately 50 mL), 50 mL of 0.05 M sodium carbonate was added to neutralize any remaining hydrochloric acid. The sample was dried with sodium sulfate until no water droplets were visible, then concentrated to 0.1 mL using a rotary evaporator and nitrogen gas. The final sample was transferred to a 200 µ L insert in a 2 mL vial and then analyzed using a GC-MS instrument. The instrumental conditions and validation results are detailed in Tables S2 and S3.
Other measurements
Each sample was analyzed in the field for residual chlorine, pH, and temperature using a portable MARTINI Mi411 device (Milwaukee Instruments, Inc., Rocky Mount, USA). T-N, nitrate, and NH3-N were determined using the official methods (ES 04363.1a, ES 04355.1c, and ES 04350.1b, respectively) outlined by the ME [34]. The water samples were filtered through a 0.45 µ m nylon filter and analyzed using a Sievers 5310C TOC analyzer (General Electric Co., Boulder, CO, USA) to determine the DOC, and a UV-VIS spectrophotometer (UV-9100, Human Co., Seoul, Korea) was used to determine the UV254.
Statistical analysis
Statistical analysis was conducted using IBM SPSS Statistics 29 software with a confidence level of 95%. The NDMA concentrations in water were assessed using Pearson’s correlation and regression analysis to explore their relationship with precursor compounds and other potential influencing factors. Student’s t-test and one-way analysis of variance (ANOVA) were conducted to assess variations in the concentration distribution. ANOVA was followed by post-hoc analysis using Tukey’s HSD test.
Results and Discussion
NA concentrations in the DWTP and tap water
Water samples from the Soyang DWTP contained five nitrosamines (NMOR, NDMA, NMEA, NDEA, and NDBA) (Tables 1 and 2). NDMA and NDEA were identified in 91.7% (11/12) and 66.7% (8/12) of the raw water samples, respectively. NDBA was found in all samples collected from the DWTP after sedimentation, filtration, and the final stages of water treatment. Moreover, NDMA and NDEA were detected at rates exceeding 66.7% at all sampling points. NDBA and NDEA were found in untreated water at average concentrations of 11.6 ng/L and 9.78 ng/L, respectively. The levels of both compounds slightly decreased during the treatment, with average concentrations of 9.30 ng/L and 6.60 ng/L in the finished water samples, respectively. The NDMA concentration in the raw water was 15.2 ng/L, which increased to 21.2 ng/L in the finished water. NMOR was found below the limit of quantification (LOQ) in three samples: one in raw water, one after filtration, and one in finished water. NMEA was found at a concentration of 8.77 ng/L in a raw water sample but was not detected in the other samples. However, the increase in NDMA levels is not clearly explainable due to the insignificant concentration differences observed between treatment processes (ANOVA, p > 0.05). This finding is consistent with a previous study that reported a slight increase in NDMA within the treatment system, although it lacked statistical significance. In that study, the influence of ammonia during chlorination was examined, with average NH3-N and Cl2:NH3- N ratios of 0.6 mg/L and 5:1, respectively, in the influent sample. This Cl2:NH3-N ratio was lower than the breakpoint (10:1) necessary for the conversion of chloramines to free chlorine and trichloramine [20]. Identifying clear evidence of NDMA formation during treatment remains challenging due to its complex formation mechanisms and the presence of various precursors.
The detection rates of NAs in tap water were similar to those in the DWTP samples (Table 1). NDMA was present in all samples, whereas NDEA and NDBA had high detection rates ranging from 70.0% (42/60) to 98.3% (59/60). The concentrations of NDMA, NDEA, and NDBA ranged from 4.57 to 62.6 ng/L, 1.22 to 23.9 ng/L, and 2.74 to 38.0 ng/L, respectively (Table 2). The average NDMA concentrations at S1–S5 ranged from 19.3 to 27.2 ng/L, exceeding the NDEA and NDBA levels by 2.4 to 7.5 times (6.20–8.97 ng/L and 3.51–11.4 ng/L, respectively). Hashemi et al. (2022) conducted a comprehensive study to monitor 33 DWTPs in Korea from 2013 to 2020. The concentrations of NDMA, NMEA, NDEA, NMOR, and NDBA ranged from 0.3 to 11.4 ng/L, 0.1 to 3.6 ng/L, 0.1 to 8.0 ng/L, 0.1 to 9.46 ng/L, and 0.15 to 15.3 ng/L, respectively. NDMA had the highest detection rate at 44.4%. The monitoring process in DWTPs analyzes treated water but overlooks the potential formation of contaminants during water distribution. However, NDMA was detected in all tap water samples, with levels up to 5.5 times higher than those found in a previous study. This could potentially be due to the contact time during the water supply process. In addition, Hashemi et al. [35] confirmed that the concentration of NAs in finished water has decreased since the monitoring regulation began in 2018. However, they noted that it was difficult to clearly explain the reason for this decrease and emphasized the need for continued monitoring of tap water. Previous studies also highlighted a limitation in that monitoring information for raw water and influencing parameters was not provided [35,36]. Contact time is crucial for the formation of NDMA in the presence of DMA, its primary precursor [37]. NAs, including NDMA, can be found in raw water from residential, industrial, and agricultural areas [38,39]. Although this study does not accurately reflect the current status of NDMA distribution in drinking water, as it was conducted in 2014, there remains a significant deficiency in domestic research focused on ensuring the safety of drinking water. In various countries, NDMA has been identified in drinking water at concentrations ranging from 1.12 to 45.36 ng/L in China [30,41], 8 to 67 ng/L in Brazil [48], and 1.3 to 9.1 ng/L in Spain [20]. In contrast, nationwide monitoring in South Korea has predominantly reported NDMA levels as 'not detected' in drinking water [49]. To address this research gap, it is essential to conduct further investigations and implement regular cross-validation across different laboratories to produce more reliable results and to confirm both the contamination of source water and the occurrence of NDMA during treatment processes.
The physicochemical parameters of all samples were tested, and the results are presented in Table 3. The pH, TN, and nitrate levels were consistent across all sampling points. After sedimentation, DOC, UV254 absorption, and NH3-N exhibited minimal variation. The residual chlorine levels gradually increased during the water treatment process, resulting in free chlorine levels in tap water ranging from 0.20 to 0.71 mg/L. West et al. [40] reported NDMA formation at concentrations of 10.05 ng/L and 7.49 ng/L with free chlorine levels of 0.14 mg/L and 0.62 mg/L, respectively. In the future, more diverse water quality parameters need to be considered. Previous studies have shown that COD and Fe had a significant correlation with NAs in both raw and treated water, and these were proposed as indicator materials [41].
Spatial distribution and relationship between NDMA and DMA
The distances from the DWTP to the sampling points are labeled as S1 to S5. The concentrations observed from July to December were insufficient for a comprehensive spatial analysis of chlorination by-products (Table 2). Hence, seasonal fluctuations in the water conditions were considered for the spatial distribution (Figure 3). The seasons can be divided as follows: summer (July and August), fall (September and October), and winter (November and December). Summer is characterized by the highest temperatures and precipitation, with the highest temperature recorded at 35.1 °C, and more than 40% of the annual precipitation occurring during this season. Winter, on the other hand, is the coldest season with the lowest temperature recorded at -20.1 °C in 2014 [42]. The average NDMA concentrations in raw water were 13.6 ng/L and 9.18 ng/L in the summer and fall, respectively, and 20.3 ng/L in the winter. The concentrations in tap water increased to 25.7 ng/L in the summer and fall, and 35.9 ng/L in the winter. Most samples exhibited a gradual increase in NDMA concentrations during the fall, while no fluctuation was observed during the winter. The concentrations of DMA in raw water were 126,107 ng/L and 12753 ng/L in the summer and fall, respectively. This concentration was halved during treatment. The detected concentrations are listed in Table S4. The concentration of DMA in raw water was 33,144 ng/L in the winter, with no decrease observed in the DWTP. Therefore, the distribution of NDMA in the fall may be influenced by the high levels and variability of DMA, which can affect the spatial distribution of NDMA in the DWTP and tap water samples.
A previous study found higher levels of NDMA in the winter (3.2–20 ng/L) than in the summer (1.5–3.5 ng/L) and fall (0.89–9.2 ng/L) [20]. This phenomenon is thought to be influenced by factors such as precipitation or drought and is consistent with the findings of a previous study [14]. It is similar to the present study, which shows the highest NDMA concentrations in winter. Although the limited sample size constrains the statistical comparison of seasonal distributions, notable differences were observed between the winter and summer/fall seasons. In winter, the average concentrations of NDMA in raw and finished water were recorded at 20.3 ng/L and 32.2 ng/L, respectively. These concentrations exceeded those measured during the summer and fall seasons. Specifically, the average concentrations of NDMA in raw and finished water during the summer were 13.6 ng/L and 16.1 ng/L, while in the fall, they were 9.18 ng/L and 9.54 ng/L, respectively.
In addition, further research on seasonal monitoring is necessary, as seasonal factors may influence the levels of NAs and their precursors, as noted in previous studies. It has been suggested that NAs and their precursors could increase due to the stagnation of river flow in winter, while photolysis and biodegradation may be more pronounced in summer compared to winter [39,41].
NDMA, one of the most potent carcinogens among NAs, is the predominant NA detected in tap water. DMA has been identified as the precursor of NDMA. Thus, Pearson correlation analysis was conducted to investigate the relationship between NDMA and DMA to validate the involvement of DMA in NDMA formation. A negative correlation was observed between NDMA and DMA in the total sample, with a correlation coefficient (r) of -0.577 (p < 0.001). The correlation coefficients were -0.614 (p < 0.001) in fall and -0.424 (p < 0.01) in winter; however, no significant correlation was found in summer. This investigation into spatial distribution focused solely on DMA as a primary precursor. However, several other factors may also influence the formation of NDMA. In the proposed mechanism for NDMA formation in drinking water, dichloramine has been identified as a primary reactant with DMA, and the Cl2:NH3-N ratio is a significant factor associated with the breakpoint of chloramination [20,25]. Additionally, organic matter plays a crucial role in the formation of DBPs, with an increase in NDMA reported as DOC and DON decrease. Furthermore, inorganic nitrogen species can contribute to NDMA formation through nitrosation reactions [25,47].
The most significant correlation was observed in fall, and NDMA concentration was higher in tap water than in finished water (t-test, p < 0.05). The observed increase in NDMA levels may indicate further formation within the supply system due to reactions among its precursors over time. The concentrations of DMA in the DWTP gradually decreased from 126,107 ng/L in raw water to 37,696 ng/L in finished water. A significant variance was observed among the sampling points, as determined by a one-way ANOVA test (p < 0.05). Furthermore, Tukey’s HSD test revealed a significant difference between the raw and treated water samples (p < 0.05). Despite this substantial decrease, the potential for NDMA formation in the supply system could be further mitigated through enhanced DMA removal methods, such as biofiltration and adsorption using beta zeolite [45,46].
The distribution of DMA in tap water was consistent across sampling points S1–S5, as indicated by an ANOVA test (p > 0.05). The average concentrations at these points were 41,124, 42,320, 44,859, 48,510, and 53,172 ng/L, respectively (Table S4). This finding may be attributed to the lesser influence of weather conditions such as heavy rain, drought, and temperature, suggesting potential formation during the water supply process. However, there are limitations in providing clear evidence of NDMA formation over time and its correlation with key factors such as free chlorine, chloramines, pH, DOC, DON, and nitrite. To address these limitations, it is essential to accurately determine the contact time from the DWTP to tap water, increase the sample size, and measure water quality parameters that may influence NDMA formation.
Temporal distribution and relationship between NAs and precursors
The formation of DBPs in chlorinated water samples was examined over a specified time period. Sampling was carried out on four occasions. Table 4 shows the variations within a 24-h period. NDMA, NDEA, and NDBA were detected at average concentrations of 9.44, 3.64, and 2.50 ng/L, respectively, following the water treatment process. The NDMA concentration gradually increased to 30.7 ng/L, whereas NDEA showed no significant variation. The NDBA concentration increased to 24.8 ng/L after 12 h and then decreased to 8.76 ng/L after 24 h. Measurements were conducted simultaneously to assess the variations in residual chlorine, pH, temperature, nitrate, and DOC. The chlorine levels exhibited a gradual decrease, while the concentration of dissolved nitrate showed an increase. By contrast, the pH and DOC levels remained fairly constant. The NDMA concentration doubled after 4 h and tripled after 7 h.
Among the three NAs, a significant temporal increase was observed only for NDMA (r = -0.306, p < 0.05). NDMA shows a significant negative correlation with temperature and a positive correlation with nitrate (Figure 4). The negative correlation between NDMA and temperature is explained by the 4th sample set, which showed higher NDMA concentrations and lower temperatures compared to the other sample sets. In the 4th sample set, the temperature ranged from 10.0 to 14.9 °C, and the initial NDMA concentration was 33.3 ng/L, which increased to 46.6 ng/L after 24 h. In contrast, the other sample sets had temperatures ranging from 15.6 to 21.0 °C, with initial NDMA concentrations below the method detection limit to 2.95 ng/L. In all sample sets, the concentrations of nitrate gradually increased over time, and it is suspected to play a crucial role in the formation of NDMA in chlorinated water. This hypothesis could be supported by a previous study, which found that NDMA was formed at 8.0 ng/L after 5 h when DMA, free chlorine, and nitrate were added to the water [29].
A multiple regression analysis was conducted to determine the main factor following water treatment. The aforementioned factors such as chlorine, pH, temperature, nitrate, and DOC were considered, leading to the following stepwise regression model: NDMA (ng/L) = -18.203 - 4.099 × temperature (°C) + 9.495 × nitrate (mg/L) + 67.245 × total chlorine (mg/L). The model had an R2 value of 0.477, indicating that the model could explain approximately 47.7% of the variation in the NDMA concentration. Temperature showed a negative correlation with NDMA, whereas nitrate and total chlorine exhibited a positive correlation. Among the factors considered, NDMA exhibited the most notable correlation with total chlorine, with nitrate and temperature following in significance. The negative relationship between NDMA and temperature is consistent with the findings of Park et al. [36]. Moreover, NDMA may be more affected by total chlorine than free chlorine because of the complex interactions involving Cl2, chloramines, and other associated factors.
While this model does not consider significant time dependence, NDMA is expected to increase over time (Table 4). Wang et al. [18] conducted calculations to determine the molar yields (Ym) of NAs. The maximum recorded the concentrations of NDMA, NDEA, NMOR, and NDBA within a 24-h timeframe were 0.24, 1.18, 2.22, and 0.12 nM, respectively. The Ym values ranged from 0.17% to 5.9%. Previous research has suggested that the occurrence of NAs is more prevalent in chloraminated systems than in those using chlorine alone [20]. Establishing a significant association during chlorination poses challenges because of the intricate reaction mechanism that occurs following the interaction between chlorine and ammonia, resulting in the formation of chloramines [44]. In our study, finished water was considered accessible within a 3-h timeframe at the tap water sampling points. Three hours after sample collection, the average NDMA concentration rose to 14.5 ng/L, ranging from 5.00 to 22.7 ng/L. Concentrations exceeding 7 ng/L suggest a potential risk to human health, with a 10 5 risk level according to the U.S. EPA [7]. This cancer risk of 10-5 indicates an essentially negligible level of risk for humans exposed over a lifetime [44]. However, NDMA has been shown to induce carcinogenic effects in the liver, Leydig cells, lungs, and kidneys in animal studies [10]. It may pose a significant concern if its concentrations exceed the threshold corresponding to this risk level. Thus, to mitigate the formation of by-products during water treatment, potential factors such as SAAs, residual chlorine, and nitrogen should be monitored in advance.
Conclusions
This study investigated the presence of N-nitrosamines (NAs) in water collected from a drinking water treatment plant (DWTP) and tap water samples from several locations. The samples were collected during chlorination and after the potential formation of by-products with contact time. Five NAs were detected, with concentrations up to 62.6 ng/L in tap water. The spatial distribution of N-nitrosodimethylamine (NDMA) increased in the fall and exhibited a significant positive correlation with dimethylamine (DMA). After chlorination, NDMA increased to 30.7 ng/L within 24 h and exhibited significant correlations with temperature, nitrate, and total chlorine. This study investigated NDMA by observing its seasonal and spatial distribution and the significant increase after chlorination. In the future, a deeper understanding of the water supply process is necessary to study occurrence characteristics by contact time. For example, factors such as pH, water temperature, water supply time, the presence of organic and inorganic substances like DOC, DON, nitrite, and other parameters, as well as chlorination breakpoints, could be considered. Additionally, more factors influencing NDMA formation must be considered. To strengthen public health, the monitoring system must be improved to account for potential formation after chlorination.
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