An ultrasound/O₃ and UV/O₃ process for atrazine manufacturing wastewater treatment: a multiple scale experimental study

Atrazine (2-chloro-4-(isopropylamino)-6-(ethylamino)-s-triazine), a herbicide commonly used worldwide, has been identified as an endocrine disruptor by the U.S. Environmental Protection Agency (EPA) (Goldman 1994; Dionne et al. 2021). It has been listed by as the EPA's Hazardous Waste Hazard Classification System as a persistent and toxic compound (Sonnenschein & Soto 1998; Yang et al. 2020). The production of atrazine requires significant amounts of water and trichlorethylene as solvents, leaving a large quantity of wastewater containing organic pollutants such as atrazine, cyanuric chloride, ethylamine and isopropamide (Köck-Schulmeyer et al. 2013). This effluent also contains high concentrations of chemical oxygen demand (COD) and ammonia nitrogen (NH₃-N), up to hundreds of thousands mg L⁻¹ and over 50 mg L⁻¹, respectively (Janssens et al. 1997; Saylor & Kupferle 2019). In addition, high levels of inorganic salt (NaCl), sodium hydroxide (NaOH) and other inorganic substances are commonly observed, which can hinder biochemical treatment processes (Aslan & Şekerdağ 2016; Chen et al. 2019a). Conventionally physical/chemical treatment methods followed by a biochemical process are usually used in dealing with atrazine manufacturing wastewater. However, the performance of the physical/chemical treatment often fails to meet the designed efficacy due to the complicated and inconstant composition of the effluent. Further, the biodegradability of atrazine manufacturing wastewater is low owing to its high toxicity and salinity. All of the above lead to substantial challenges in meeting the discharge regulations (Mannina et al. 2016; Muñoz Sierra et al. 2018; Zhao et al. 2020). Therefore, the research and development of more effective treatment methods for atrazine manufacturing wastewater are urgently desired.

Advanced oxidation processes (AOPs) have gained growing interest due to their strength in efficiently breaking down persistent organic pollutants in water, presenting a great potential for treating atrazine manufacturing wastewater. AOPs generate highly reactive species, particularly hydroxyl radicals (^•OH), which can effectively remove atrazine and reduce the overall toxicity by generating less toxic compounds (Choi et al. 2013; Yang et al. 2014). Various sole and combined AOP technologies have been used for treating atrazine contaminated water, including ozonation (Yang et al. 2016; Wardenier et al. 2019), ultrasound (Collings & Gwan 2010; Liu et al. 2017; Pinto et al. 2019), peroxide (Li et al. 2020a; Song et al. 2021), Photo-Fenton and Photo-Fenton like (Khandarkhaeva et al. 2017), catalytic oxidation (Zhu et al. 2017), and photocatalysis (Yang et al. 2020; Zheng et al. 2021). Among these technologies, UV/O₃ based treatments are more energy-effective, non-selective of wastewater quality and eco-friendly (Jing et al. 2017; Liu et al. 2019; Wardenier et al. 2019). In addition, the integration of ultrasound (US) not only promotes the transformation of O₃ into ^•OH but also generates microbubbles, which provide a large interfacial area for the mass transfer of ozone into liquid phase, thus increase the utility rate of O₃ (Xiong et al. 2019).

Although many studies have been reported, only a few investigated atrazine in authentic wastewater with a complex composition of organic and inorganic compounds (Liu et al. 2020b). For example, many researches reported that high levels of chloride ion (Cl⁻) in water matrices affect AOPs by auxo-action (Xu et al. 2013; Monteagudo et al. 2016; Mukimin et al. 2017; Chen et al. 2019b), inhibition (Zhou et al. 2017; Zhang et al. 2018; Li et al. 2020b), and dual functions (Yuan et al. 2012b; Luk 2016; Huang et al. 2017; Chen et al. 2019b). Cl⁻ consume oxidants and radicals in AOPs, which can reduce the efficacy of AOPs by as much as 50% (Kiwi et al. 2000; Chan & Chu 2009). It can further alter the degradation of analytes, resulting in diverse pathways that are hard to monitor (Wiszniowski et al. 2003; Li et al. 2006; Yuan et al. 2011). High COD in wastewater could also diminish the degradation efficiency of analytes by AOPs, owing to the competitive consumption of oxidants by the wastewater substrates (Liu et al. 2020a). The accurate analysis of COD was influenced by the matrices (Wayne 1997). Therefore, in-depth experimental studies on AOPs in treating authentic wastewater such as the effluent from atrazine manufacturing are of great important to reveal these effects of water matrices on the degradation of analytes and mineralization, and further guide industrial applications.

To help address the above questions and further improve the performance of AOPs in treating atrazine manufacturing wastewater, this study developed UV and US enhanced ozonation in multiple-scale systems, and investigated the influence of factors (UV and US), and the matrix effects on the degradation efficiency of atrazine and COD. Factorial analysis of UV, US and ozone was demonstrated at pilot scale. The findings should be able to unveil the possible mechanisms of atrazine and COD degradation in a complex matrix of chloride, and nitrogenous compounds, and guide the design and configurations of energy effective AOPs treatment systems for atrazine manufacturing wastewater at large scale.

Atrazine-d5 standard and sodium thiosulfate was purchased from Anpel Laboratory Technologies (Shanghai) Inc. Trichloromethane (Thermo Fisher Scientific China) was used for the extraction of atrazine from aqueous samples. Ultrapure water was produced on-site from a Direct-Q 3 UV unit (Millipore, France). Potassium dichromate, sulfuric acid, silver sulfate and mercuric sulfate were purchased from Beijing Chemical Works, China. Atrazine manufacturing wastewater was collected onsite from an agrochemical production plant in China. Detailed location information is not available due to client confidentiality and non-disclosure. The physical-chemical characteristics of the atrazine manufacturing wastewater are summarized in Table 1. The COD was reduced to 6,000–8,000 mg L⁻¹ after sand filtration.

As shown in Figure 1, a bench-scale photoreactor has an inner 4 L quartz jar and an outer stainless-steel jacket (Jing et al. 2014). The outer jacket has an aluminum lid for heat and light insulation. The inner diameter, height and wall thickness of the quartz jar are 20, 25 and 0.4 cm, respectively. Eight 3.5 W low-pressure UV lamps, emitting exclusively at 254 nm, are evenly mounted inside the quartz jar near the wall. A 300 W ozone generator with dedicated ozone flow rate monitor are used to produce ozone onsite from ambient air. The ultrasonic system (JY92-IIN, XinZhi Co. Ltd, China) equipped with a probe of which the diameter is 6 mm. The inner quartz jar has a PTFE lid equipped with a stirring rod on which two PTFE six-bladed paddle impellers are mounted to stir the wastewater sample. As to bench-scale experiments, LP-UV lamps were allowed to warm up and reach the stable emission stage for 20 minutes, 2 L water samples were injected into the quartz jar and stirred for uniform mixing. For the UV/O₃ test, an ozone generator at a fixed rate of 15 g/h (7.5 g/L·h) was run with the irradiation device simultaneously, and UV power was provided at three different levels of 3.5, 7 and 14 W/L by using 2, 4 and 8 lamps, with light intensities of 2.88, 5.65, 10.93 mw/cm², respectively. For the UV/O₃/US test, ultrasonic power (20 kHz) was provided at three different levels of 50, 100 and 200 W/L by using an ultrasonic energy converter, and the ultrasonic probe was submerged in the liquid at a level between 15 and 20 mm. During the 180-min reaction period, samples were taken by a peristaltic pump at intervals. In order to eliminate the influence in which water samples remain in tubes, the water samples were taken after pumping for 3 s, transferred into a 20 mL amber vial and immediately quenched by Na₂S₂O₃ to remove residual oxidants.

A pilot-scale US/UV/O₃ system was developed as elaborated previously (Jing et al. 2017) (Figure 2). Briefly, a continuous flow-through system contained four reaction columns (cylindrical polycarbonate; 3 feet in height and 10 inches’ in internal diameter) with functions of coarse sand filtration, fine sand filtration, ultrasonic ozone treatment and photolytic ozonation, respectively. The low-pressure mercury lamp (UVC-7, LiZhen Co., Ltd, China) was jacked with a quartz filter and the ultraviolet intensity was about 14.5, 15 and 22 mw/cm² at UV power of 6.86, 10.8 and 14.7 W/L, respectively. The irradiation and ultrasonic units (US power of 0–1500 W) cooperated with aerating apparatus which can disperse ozone. Peristaltic pumps (YZ1515x, ChuangRui Co., Ltd, China) are used to propel liquids through this system in order to maintain a rather stable flow rate (1.7 L/h). In the pilot-scale tests, water samples passed through filtration and then treated by US/O₃ and UV/O₃ consequently. The separation of the UV/O₃ and US/O₃ units was because the synergetic effect of UV and US was marginal (Xu et al. 2014). The test procedure was similar to the bench-scale testes, with adjustable factors of UV power (6.86–14.7 W/L), O₃ dose (0.98–2.94 g/L·h) and US intensity (0–294 W/L). The detailed experimental procedure was documented in a published paper (Jing et al. 2017).

All laboratory determinations were implemented, following the Standard Methods for Examination of Water and Wastewater. The determination of COD was referred to the standard COD method regulated by the State Environmental Protection Administration, China after dilution due to the high chloride concentration of wastewater samples. The detection limit of the method is 9 mg/L.

Atrazine was extracted by trichloromethane and measured by gas chromatograph-mass spectrometry (GC-MS). An Agilent Technology 7890A GC equipped with an Agilent 5975C MS detector was employed. The gas chromatograph was equipped with an Agilent 7683 autosampler and splitless injector with electronic pressure control. A HP-5MS column (30 m0.25 mm0.25 μm) capillary column was used, with helium as carrier gas at a constant flow of 1 mL min⁻¹. The temperature of the injection port was 250 °C and a 1 μL volume was injected in splitless mode. The GC temperature was ramped from 70 °C, held for 2 min, then to 230 °C at 20 °C·min⁻¹ and held for 20 min. The mass spectrometer was operated in EI mode with an ionising energy of 70 eV, ion source temperature at 230 °C, and MS Quad temperature at 150 °C. The mass signal was captured in the SIM mode, scanning from m/z 50 to 500 with a solvent delay at 6 min.

In order to optimize the UV power (W/L), O₃ flow rate (g/L·h) and US power (W/L), and investigate the potential interactions of each factor, central composite design (CCD) and response surface methodology (RSM) were employed for experimental design by Design-Expert^® 8.0. According to the parameter levels of the single factor experiment and the treatment capacity of the pilot system, UV power (A), O₃ flow rate (B) and US power (C) were amplified to three levels (Table 2), and used −1, 0, 1 on behalf of factors of low, medium, and high values, respectively.

Response values are COD degradation rate and atrazine removal rate; 20 experiments with three replicates were required for improving response values. The experimental data was further analyzed by assuming a second-order polynomial with linear, quadratic and interaction effects.

In the UV/O₃ system, when the UV power was increasing from 3.5 to 7 W/L, UV power improved the removal rate of COD, in which the COD removal rates were 57 and 47% at 180 min, respectively (Figure 3(a)). When the UV power exceeded 7 W/L, on the contrary, the COD removal rate did not reduce but increased since the beginning of the reaction. The optimum value of UV power was determined as 7 W/L, in which a removal rate of 94.9% of atrazine was achieved. With the increase of US intensity, the removal rate of COD in the atrazine wastewater was 34.8%, 43.5% and 23.1% at 180 min, respectively (Figure 3(b)). An increase of COD value was observed at the highest US intensity. The best energy-effective value of US power in bench-scale treatment was 100 W/L and 96.5% of atrazine was removed in 10 min under the optimal condition.

In order to avoid the growth of COD caused by US, the atrazine manufacturing wastewater was treated in the UV/O₃ and US/O₃ processes consequently. The observed and predicted results of atrazine degradation rate (Y₁, %) and COD removal rate (Y₂, %) are listed in Table S1.

According to statistical model fit summary by Analysis of Variance (ANOVA), a quadratic model (second order polynomial) was selected as the best fitted model. The quadratic effects of the factors on atrazine (Figure 4) and COD (Table 3) were obtained. Figure 4 illustrates that the degradation of atrazine is affected by linearity and quadratic terms of the O₃ flow rate. The possibility of three interaction effects was less than 0.01, which are between UV and O₃ (AB), UV and US (AC), and O₃ flow rate and US (BC), demonstrating there exist interactions among them. Table 3 shows the ANOVA results for the response surface quadratic model on COD removal. The COD removal was sensitive to the linearity term of O₃ flow rate, which agreed with the results of atrazine degradation. However, the possibility of three interaction effects were higher than 0.05, demonstrating no interactions.

The degradation of atrazine in AOPs is mainly through dechlorination and dealkylation, in which the chloride atom on the triazine ring and the alkyl groups on the amines can be attacked and substituted by ^•OH, forming hydroxyatrazine (OIET) and hydroxydeethyl atrazine (OIAT) and deethyl atrazine (CIAT), deisopropyl atrazine (CIET), and deethyl deisopropyl atrazine (CAAT) (Choi et al. 2013; Xu et al. 2014; Fan et al. 2017). The end product of atrazine in the UV and US process was ammeline (Xu et al. 2014). It was suggested that in alkaline solution, indirect oxidation (induced by ^•OH) dominated (Cuerda-Correa et al. 2020). Therefore, the concentration of ^•OH is of paramount important for atrazine degradation. Equation (8) shows that the most significant factors affecting the degradation of atrazine are O₃ (B), and the interactive effects of UV and O₃ (AB), O₃ and US (BC), and UV and US (AC). The interactive effect of UV and O₃ was positive. In general, the results indicated that more O₃ resulted in a higher concentration of ^•OH, while higher intensity of UV and US can also produce more ^•OH, as well as the decomposition of O₃ to ^•OH. UV and US (AC) showed a strong negative interaction in terms of atrazine degradation. The phenomenon could result from the interference of TSS, and the generation and decomposition of H₂O₂. US can break down the insoluble particles in atrazine wastewater, which would increase the turbidity, increasing the masking effect on UV (Choi et al. 2020). In the meantime, US generated a higher level of H₂O₂ (Ziembowicz et al. 2017). It thus increased the degradation rate of atrazine. H₂O₂ in alkaline solution can be quickly decomposed in O₂ and water (Nicoll & Smith 1955). UV accelerates the process. It seems that UV with high intensity overaccelerated the decomposition of H₂O₂ generated by US, thus slightly reducting the removal rate of atrazine.

Bench-scale tests showed an increased atrazine removal to 96.5% by an optimal intensity of US. This aligned with Xu's results, US promotes the generation of ^•OH by bubble cavitation and attacks the C-Cl position and/or the alkyl side chains of atrazine (Xu et al. 2014). Figure 5(b) illustrates the relationship between US power and O₃ flow rate at pilot scale. The US intensity at low and high level had a higher degradation of atrazine. However, the quadratic model (Figure 4) showed that the effect of US was not significant (P-value >0.05). This was probably because of the dual effects of US. On one hand, US promoted the decomposition of O₃ and the formation of ^•OH at low intensity in the vapor phase of cavitated bubbles (Jing et al. 2017). On the other hand, the cavitated bubble created a barrier for a biphasic reaction. The interference of the excess ozone could also react with ^•OH in bubbles, resulting in a recombination of free radicals (Gogate 2008). As the US intensity increased, more barriers were created thus reducing the degradation rate of atrazine. The increase of atrazine degradation at the highest US intensity is probably because of the direct generation of ^•OH by bubble cavitation. Therefore, it is important to optimize the US and O₃ ratio for the most energy-efficient solutions.

In addition, higher energy US process could result in the release of higher amounts of UV-absorbing compounds, which would reduce the energy-effectiveness of UV significantly. It can be observed in the two configurations of the system with the best atrazine degradation rates. Firstly, an atrazine degradation rate of 96.85% was achieved in the condition of 10.8 W/L UV, 10 g/h O₃ and 294 W/L US. Meanwhile, without US, the atrazine degradation rate can reach 93.93% in the condition of 14.7 W/L UV and 15 g/h O₃.

UV irradiation can effectively promote the decomposition of O₃ and the generation of ^•OH (Beltrán et al. 2000; Liu et al. 2020a). In bench-scale tests, the degradation of atrazine achieved 94.94% within 10 min, indicating the high efficacy with the optimal UV intensity. However, UV intensity alone was not a significant factor for atrazine degradation at pilot scale. Under the condition of low amount of O₃, UV intensity limited enhanced the degradation rate of atrazine, while under the condition of high amount of O₃, UV intensity accelerated the degradation rate of atrazine significantly. It suggested that at low concentration, O₃ was quickly consumed in a high salinity, alkaline and COD environment. The fast decomposition of O₃ in high salinity could limited the effect of UV. When the level of O₃ increased to medium, the acceleration of O₃ decomposition of UV as indicated by the degradation rate is also enhanced. However, higher O₃ dose resulted in a decline of atrazine removal, probably because the excess O₃ might quench ^•OH.

High salinity wastewater impacts on the performance of AOPs, as was found in many studies, the summary is shown in Table S2. Organic compounds, including a variety of dye wastewaters and pharmaceutical wastewater, were affected by Cl⁻. Some studies suggested that the inhibitory effect of the chloride ion was due to the performance of ^•OH scavengers (4.3 × 10⁹M⁻¹ s⁻¹). It competed with organics for the photo-oxidizing species (Guillard et al. 2003; Konstantinou & Albanis 2004; Ghodbane & Hamdaoui 2010). In addition, the influence of the chloride ion depended on the pH of the solution. For ^•OH based-AOP, chloride was found to accelerate the degradation of dyes under lower pH conditions, but significant inhibitory effect was found in alkaline condition (Shi-ying et al. 2005; Yuan et al.2012a, 2012b). For ^•SO₄⁻ based-AOP, normal pH level was found to result in a higher degradation rate of atrazine, because of the inactivation of PMS at acidic pH and precipitation at basic pH (Chan & Chu 2009). Also, the concentration of Cl⁻ was observed to have a dual effect on AOPs. Studies indicated that lower concentration of Cl⁻ promoted the discoloration of dyes, which might be attributed to a surface chain-transfer mechanism involving chlorine radicals. But higher Cl⁻ level reduced the further degradation, which may be caused by the deactivation of the photocatalyst and reduction in photon receiving efficiency (Yuan et al. 2012b).

The presence of chloride slightly affected the degradation of atrazine. Chloride quickly reacted with O₃ and •OH, forming less oxidative free chlorine radicals (e.g., •Cl, and •Cl₂⁻). But the reaction rate constants of •Cl and •Cl₂⁻ with atrazine were 6.87 × 10⁹ M⁻¹s⁻¹ and 5 × 10⁴ M⁻¹s⁻¹, respectively (Luo et al. 2015; Kong et al. 2020). Considering the abundance of free radicals, the overall degradation of atrazine was changed slightly. Nevertheless, the presence of Cl could change the degradation pathways of atrazine. The presence of free chlorine radicals inhibited the dechlorination process, leading to dealkylation as the main process. Atrazine was then gradually decomposed to the intermediates (e.g., CIAT, CIET, CAAT, ammeline) with less toxicity to Daphnia magna (Choi et al. 2013) and Vibrio fischeri (Li & Zhou 2019), which can be further treated by activated sludge (Li et al. 2018).

The atrazine manufacturing wastewater contained a considerable amount of cyanuric chloride, ethylamine and isopropamide that contributed to the COD value. Equation (3) shows that O₃ dose is the only positive factor in the oxidation of these compounds by providing more oxidants such as O₃ and •OH. The increasing trend of COD removal by O₃ indicated that the O₃ dose has not reached the optimal level in the treatment system.

It should be noticed that a high US power resulted in the increase of COD value after treatment. The phenomenon was rarely observed in the clean water studies. For bench-scale tests, the negative effect of COD removal was observed when US power was at 200 W/L. The factorial analysis results of the pilot-scale tests also showed that high US power (294 W/L) had the strongest negative impact to COD removal. The same phenomenon was also found in bench or pilot-scale tests when higher UV power was applied.

Previous research has proved that atrazine-2-hydroxy is the main products of UV photolysis, higher UV power may accelerate the replacement of Cl by OH on atrazine, thus increasing the concentration of free chlorine radicals. In this study, the formation of free chlorine radicals could react with the nitrogenous compounds, forming undesired chloramine, or initiate additional reaction between •Cl and organic matter, such as isopropamide in atrazine wastewater (Equations (11) and (12)), which could contribute to the increase of COD value. In addition, the sono-degradation of ATZ was initiated by side chain oxidation; alkylic oxidation products and the dealkylation products prevailed in the early stage and accumulated to their maximum in the middle stage (Xu et al. 2014), US can also accelerate the de-chlorination step under ozonation (Bianchi et al. 2006). The low frequency US was able to break down the suspended solids in wastewater (Fetyan & Salem Attia 2020). However, compared with the chemical breakdown, the physical breakdown of the particles was the dominant approach at a low frequency (Mason et al. 2011). Therefore, higher power of US may introduce more organic matter in the suspended solids to the aqueous phase without oxidising them, leading to the reduced efficiency of COD removal.

The substrate in the atrazine manufacturing wastewater may interfere with the COD tests in many ways. Although the atrazine manufacturing wastewater has been sand filtrated before oxidation, the soluble residue in the filtrate could still affect the result of COD. The coexist of chloride and nitrogenous compounds (e.g., ammonia and organic amine) in wastewater can produce monochloramine (NH₂Cl) during the acidic K₂Cr₂O₇ oxidation, which can severely interfere with the COD value (Wayne 1997). Although the addition of excess silver nitrate during the COD test can remove free chloride ions in the treated effluent, the presence of chloramine in atrazine manufacturing wastewater would affect the COD value. In addition, the COD values of nitrogen-containing heterocycles can be underestimated due to the incomplete oxidation by K₂Cr₂O₇. The ratio of N₂ and NH₃ as oxidation products of heterocycles was structure dependent (Chudoba & Dalešický 1973).

The transformation of atrazine, ammeline, triazine (an oxidation product from cyanuric chloride) in the UV/O₃/US system includes deformation, polymerization and oxidation, which could result in both increase and decrease of the COD value; it also explained why the linearity term of UV power and US power had insignificant impact on COD degradation at pilot scale (Figure 5). It was more obvious under the condition of high energy input (e.g., 14 W/L UV, 200 W/L US) but limited oxidant level in the bench-scale experiments, in which the deformation of nitrogen-containing heterocycles by UV was the major reaction instead of oxidation.

The EEO of atrazine under different AOPs combinations are shown in Table 4. The processes with US generally consumed more energy than other processes due to the high energy input of US. However, it did not show a corresponding enhancement of atrazine degradation. Most of the US-enhanced processes have EEO values higher than 1,000 kWh/m³. The process with the highest atrazine removal (Run 16, 96.9%) consumed 684.29 kWh/m³, while the process with the second highest atrazine removal (Run 7, 93.9%) only had an EEO value of 181.6 kWh/m³. It is suggested that US was not an energy-effective addition for ATZ degradation.

On the other hand, the wastewater matrix can strongly interfere with the removal efficiency of target compounds. The competitive consumption of wastewater substrates reduces the overall oxidation rates (Liu et al. 2020a), which on other hand require much higher energy to reach the same removal rates of target compounds in the distilled water system. Miklos et al. (2018) summarized the EEOs of O₃ based AOPs. They pointed out that the median value of EEO in these processes was <1 kWh/m³. However, the EEO value for treating the real wastewater using ozonation (188.01 kWh/m³) can be found at two orders of magnitude higher (Jiménez et al. 2019). EEO omits the effects of the water matrix thus is not suitable to evaluate the energy-effectiveness of different treatment processes for the cleanup of water with different matrices.

For COD, the electron energy consumption per COD removal would be a better indicator. Table 4 shows the energy consumption rates of different runs. It is indicated that US has significantly increased the energy consumption rates. The lowest rate of 0.1 kWh/g can be found in Run 1, which was with an energy input of 6.86 W/L from UV and 2.94 g/h from O₃. However, its EEO value for atrazine was 888.26 kWh/m³, which was much higher that Run 7. In the consideration of both atrazine and COD removal, Run 7 is the most energy-effective treatment option for atrazine manufacturing wastewater.

Conventional physical/chemical/biological treatments of atrazine manufacturing wastewater containing high levels of chloride, COD and alkaline that usually failed to meet the discharge regulations. The efficiency and the energy-effectiveness of the UV/O₃/US processes were investigated in treating the atrazine manufacturing wastewater through the bench- and pilot-scale systems. Bench-scale tests showed that at the levels of 7 W/L of UV and 100 W/L of US, the integration of UV and US in ozonation increased the atrazine degradation and COD removal. A 96.5% removal of atrazine was achieved in the best condition. However, the excessive energy inputs (e.g., UV14 W/L and US 200 W/L) could increase the COD value of the effluent. In the pilot-scale system, the significant factors were O₃, and the interactive effects of UV and O₃, O₃ and US, and UV and US. •OH induced oxidation is the dominated process for atrazine degradation in the alkaline condition. High US intensity might enhance atrazine degradation by directly generating •OH from cavitation. However, high US intensity could also promote the UV absorbance of the background water matrix, thus reducing the energy-effectiveness of UV. In terms of matrix effect, high salinity slightly reduced the removal of atrazine. By calculating the energy consumption rates of different configurations in the pilot-scale system, the most energy-effective option for atrazine and COD removal is 14.7 W/L UV and 7.5 g/L·h O₃, which have the removal rates of 93.9% and 23.7% for atrazine and COD, respectively.

In summary, the UV/O₃/US process could effectively degrade atrazine in the manufacturing wastewater, while US is less energy effective. The COD value of the effluent during and after treatment was affected by O_3. The bidirectional effects of energy inputs (e.g., UV, US) in the presence of insoluble particles, chloride, and nitrogenous compounds were observed in the removal of COD in the wastewater.

Although the system showed a promising performance in the removal of atrazine from the wastewater, the removal of COD can be further improved by optimizing the operational factors, quantifying the interference caused by the wastewater matrices, and reducing the ‘noise’ to the COD analysis. In addition, the primarily treated wastewater is usually collected and further treated in industrial wastewater treatment plants, which is dominated by the biochemical treatment systems (e.g., activated sludge). Therefore, further investigation on the end products and other composition in treated atrazine wastewater, as well as its toxicity is necessary.

Special thanks go to Natural Sciences and Engineering Research Council of Canada (NSERC) and Canada Foundation for Innovation (CFI) for supporting this research.

All relevant data are included in the paper or its Supplementary Information.