Introduction

Trimethoprim (TMP), a broad-spectrum antibacterial, has been widely used for its high effective and unique antibacterial properties. TMP can enter the environmental water through the pharmaceutical wastewater generated during industrial production processes and the domestic sewage formed during daily use. TMP concentrations up to 605 μg/L has been detected in the Pearl River basin of China (Bu et al. 2013). Residues and accumulation of TMP in the surface water and groundwater can have a major impact on the ecological environment. The priority control of 39 commonly used medicines and personal care products was studied by considering three factors: consumption, removal efficiency, and potential ecological risk, and it was found that TMP priority control ranks among the top in China (Sui et al. 2012). However, the study of Adams et al. (2002) showed that conventional water treatment processes (aluminum salt or iron salt coagulation process) cannot effectively remove TMP in surface water and deionized water. Zhao et al. (2019) used the activated sludge (AS) process to treat high-concentration antibiotic wastewater and found that the adsorption of trimethoprim was negligible, and no biodegradation occurred during the AS process.

Advanced oxidation technology (AOPs) with •OH and \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) as the main active substance is a promising treatment technology that improves the biodegradability of wastewater through oxidation or directly degrades organic pollutants through mineralization (Hou et al. 2018; Huang et al. 2020; Ji et al. 2018). •OH is a strong oxidant, which has a high redox potential (E0 = 1.89~2.72 V vs. NHE). •OH is effective in the degradation of organic pollutants. However, •OH has a poor selectivity, short half-life, and is easily affected by water constituents (natural organic matter, inorganic salt ions, etc.) (Wan et al. 2019). In contrast, \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) is characterized by a high redox potential (E0 = 2.5~3.1 V vs. NHE), strong selectivity, long half-life (30~40 μs), high quantum yield, and is not easily affected by pH and organic matter (Avetta et al. 2015; Lutze et al. 2015).\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) has a good removal effect on emerging organic compounds (EOCs) (such as pharmaceuticals, disinfectants, and personal care products) in wastewater (Tang et al. 2019).

Persulfate has the characteristics of good stability and strong water solubility. Usually, under the conditions of transition metal ions (e.g., Co2+ and Fe2+), light, heat, ultrasound, PS, or PMS can be activated to produce \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \)(Ji et al. 2016; Naim and Ghauch 2016; Wei et al. 2017). UV-254nm irradiation, which has been widely used for sterilization and disinfection, is also a promising method that activates PS/PMS to generate sulfate radical efficiently (Lin et al. 2020; Zhang et al. 2021). Lin et al. (2020) studied the degradation of organosilicon wastewater by UV-based advanced oxidation processes: UV/H2O2, UV/peroxydisulfate (PDS), and UV/peroxymonosulfate (PMS), and the results showed that UV/PDS achieved a higher COD removal (90.6%) than UV/PMS (80.8%) after a 240 min reaction time. Ran and Li (2020) studied refractory organic compounds from dinitrodiazophenol (DDNP) containing industrial wastewater degraded through two ultraviolet (UV)-based advanced oxidation processes: UV/hydrogen peroxide (UV/H2O2) and UV/potassium persulfate (UV/PS) processes, and the results showed that the UV/PS process exhibited a higher organic compound removal efficiency and better applicability. Bougdour et al. (2020) studied the removal efficiency of colors from wastewater containing mixed primary direct dyes and real textile industry wastewater using PDS (peroxydisulfate)/Fe(II)/UV process, and it was found that PDS/Fe(II)/UV is the best treatment method for real textile wastewater. Zhang et al. (2015) studied the degradation of drugs and metabolites in synthetic human urine by UV, UV/H2O2, and UV/PMS (pH = 6 and pH = 9), respectively. The results showed that UV/H2O2 and UV/PMS had a better degradation effect on TMP. However, the degradation of TMP by UV/PS system under neutral condition has not been reported yet. The effects of complex environmental matrices and transition metal ions present in industrial wastewater on the degradation of TMP by UV/PS are unclear.

This study aims to establish an efficient UV/PS technology to degrade EOCs in wastewater. TMP is used as the target compound. Under the steady-state assumption, the pseudo-first-order reaction kinetic model is used to study the degradation mechanism of TMP in a UV/PS system. Competitive kinetics technique and kinetics model were used to determine \( {k}_{\mathrm{TMP},\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \). The effects of persulfate concentration and matrix components (natural organic matters, inorganic anions) were evaluated by the kinetics model and experiment. Then, the effects of transition metal ions (Cu2+, Zn2+, Co2+) in industrial wastewater on the photodegradation kinetics of TMP were further evaluated.

Experimental method

Materials

TMP (99.0%), 4-chlorobenzoic acid (pCBA, 99.0%), sodium dihydrogen phosphate (99.0%), disodium phosphate (99.0%), sodium persulfate (99.0%), fulvic acid (technical), and t-butanol (99.7%) were obtained from Sigma Aldrich. Copper sulfate (guaranteed reagent, GR), cobalt sulfate (GR), zinc sulfate (GR), sodium chloride (GR), sodium sulfate (analytical grade), sulfuric acid (GR), and hydrogen peroxide (30% by weight) were obtained from Sinopharm Chemical Reagent. Deionized (DI) water was obtained from a Molresearc 1010A molecular water system.

Irradiation experiments

In the kinetics tests, the initial concentrations of the target compound TMP and probe compound pCBA are both set at 10 uM. Phosphate buffer system (NaH2PO4/Na2HPO4, 10mM) was used to stabilize solution pH at 7.55. The value of pH did not change during the experiments. The detailed experimental design, operation process, data collection, and result analysis had been described in our previous articles (Su et al. 2018; Yang et al. 2017).

Analytical methods

According to the measurement method we previously reported, the average light intensity per volume (I0) and effective optical path length (b) were determined to be 7.496 × 10−6 Einstein L −1 s −1 and 0.935 cm, respectively (Beltran et al. 1995; Parker 1953; Yang et al. 2017).

A USB 2000+, Ocean Optics fiber optic spectrometer was used to measure the emission spectrum and light intensity of low-pressure mercury lamp. Shimadzu UV-1800 spectrometer was used to determine the absorption spectra of TMP and pCBA (Fig. 1). A Mettler Toledo S220 pH meter was used to measure the solution pH.

Fig. 1
figure 1

Decadic molar absorption coefficient (ε) of TMP and pCBA (pH = 7.55) with reference to the UV lamp emission spectra from 200 to 400 nm

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The concentration of TMP and pCBA was performed using an ultra-performance liquid chromatography (Waters ACQUITY H-Class) with a BEH C18 column (1.7 μm, 2.1 mm × 50 mm, Waters). The detailed method is shown in Table 1.

Table 1 The detailed method of measure the concentration of TMP and pCBA
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Results and discussion

Degradation of TMP in UV, UV/H2O2, and UV/PS systems

Linear regression analysis of the TMP degradation process showed that it followed the pseudo-first-order reaction kinetics. As shown in Fig. 2, under the reaction conditions of UV intensity 7.496×10−6 Einstein L − 1 s − 1, pH = 7.55, and [TMP]0 = 10 μM, the initial direct photolytic degradation rate of TMP in only the UV system was 0.038 μM min  − 1. Without UV irradiation, the dark reaction experiments showed that TMP had no degradation. The same results were observed in the presence of H2O2 or PS. Therefore, the degradation of TMP under UV radiation was mainly attributed to the direct photolysis of TMP. As shown in Fig. 1, at the wavelength ranging from 250 to 260 nm, TMP has a relative trough in the absorption of light. Therefore, the direct photolysis of TMP under 254 nm ultraviolet mercury lamp irradiation is slower.

Fig. 2
figure 2

Time-dependent degradation kinetics of TMP in the UV, UV/H2O2, and UV/PS systems ([TMP]0 = 10 μM, [H2O2] = [PS] = 100 μM, pH = 7.55, and I0 = 7.496×10−6 Einstein L−1 s−1). The degradation was fit to a pseudo first-order kinetic model (lines)

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The molar absorption coefficient (ε) and quantum yield (φ) are two important factors that affect the kinetics of target compound direct photolysis (Pereira et al. 2007). The molar absorption coefficient is a measure of the absorb light ability of a target compound at a specific wavelength (λ). ε can be calculated as Eq. 1.

$$ A={\varepsilon}_{\mathrm{TMP}}\times \left[\mathrm{TMP}\right]\times l $$
(1)

A is the absorbance of 10 μM TMP solutions. The TMP solution pH is adjusted to 7.55 by phosphate buffer solution. l is the path length of quartz cuvette. In this study, a 1 cm path length quartz cuvette was used to measure the TMP solution absorbance. Figure 1 illustrates the decadic molar extinction coefficient for TMP with reference to Hg lamp emission spectra. At the wavelength of 254 nm, the ε values of TMP were 3078.56 M−1 cm−1, which is consistent with the reported value of 2942 M−1 cm−1 (Baeza and Knappe 2011). εTMP was lower than the reported value of carbamazepine (6070 M−1 cm−1) and sulfamethoxazole (16580 M−1 cm−1), while it was higher than that of ibuprofen (256 M−1 cm−1) and bisphenol (750 M−1 cm−1) (Baeza and Knappe 2011; Pereira et al. 2007; Yuan et al. 2009).

The quantum yield described the ratio of the total numbers of molecules of the compound destroyed to the total numbers of photons absorbed by the system. The quantum yield of TMP can be calculated as follows (Pereira et al. 2007):

$$ {\varphi}_{\mathrm{TMP}}=\frac{r_{\mathrm{UV}}}{I_0\times \left(1-{10}^{-{\varepsilon}_{\mathrm{TMP}}\ b\ \left[\mathrm{TMP}\right]}\right)} $$
(2)

where φTMP is the quantum yield of TMP at 254 nm (mol Einstein−1), ruv (M s−1) is the direct photolytic degradation rate at an initial concentration of 10 μM. I0 is the incident UV intensity, εTMP is the molar extinction coefficient of TMP at wavelength 254 nm, and b is the reactor light path. φTMP was calculated to be 1.29×10−3 mol Einstein−1, which was close to the reported value of 1.18×10−3 M−1 cm−1 (Baeza and Knappe 2011). Different Pharmaceutical and Personal Care Products (PPCPs) have different molecular structures, which can result in different quantum yield values (Yuan et al. 2009). The value of φTMP was relatively low compared to the values of other PPCPs (ibuprofen, 0.192 M−1 cm−1 and sulfamethoxazole, 0.0215 M−1 cm−1) (Yang et al. 2016; Yuan et al. 2009).

Due to the relatively low molar absorptivity and extremely low quantum yield, the direct photolysis degradation rate of TMP was slow. As shown in Fig. 2, the degradation kinetics of TMP was significantly enhanced by adding 100 μM H2O2/PS compared to direct photolysis. The initial degradation rate in the UV/H2O2 and UV/PS systems was 1.657 μM min−1 and 2.581 μM min−1, respectively. These results indicated that the degradation of TMP in the UV/H2O2 or UV/PS systems included direct photolysis and radical degradation, but radical degradation played a major role (more than 95%). The enhanced degradation of TMP with the addition of H2O2 or PS was due to the fact that TMP degradation mainly contributed to •OH/\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) radical-mediated oxidation. The results were consistent with the results reported by Kwon et al. (2015), and they also demonstrated that •OH/\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) played a major role in the degradation of ibuprofen by UV/H2O2 or UV/PS. Therefore, the degradation rate constants of TMP largely depended on the formation of •OH/\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) in the UV system.

The competitive kinetics method was used to determine the second-order rate constants of TMP with \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) (\( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \)), and pCBA (\( {k}_{{\mathrm{SO}}_4^{\bullet -},p\mathrm{CBA}} \) = 3.60×108 M−1 s−1) was chosen as a reference compound in this study (Kwon et al. 2015). The 1 mM t-butanol was added to suppress any contribution of •OH in oxidizing TMP. The competitive kinetics method had been described in detail in our previous study (Yang et al. 2017). Figure 3 shows that an average reaction rate constant ratio between TMP and pCBA with \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) was 10.58. The \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \) values were determined to be 3.81×109 M−1 s−1. It was lower than the reported value of 7.71×109 M−1 s−1 (Zhang et al. 2015). pH values determine the morphology of TMP in the system. The possible reason was that our pH conditions (pH = 7.55) are different from those of Zhang (pH = 6).

Fig. 3
figure 3

Competitive oxidation of TMP and pCBA in UV/H2O2 and UV/PS system

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Pseudo first-order reaction kinetics

The steady-state approximation for the kinetic description of radicals was also used to estimate the \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{IBU}} \) (Luo et al. 2016a; Yang et al. 2017). The model was developed based on the hypothesis that the degradation of the target compound depended primarily on radicals (i.e., \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) and •OH) generated from the irradiation of PS. \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) was produced from activation of \( {\mathrm{S}}_2{\mathrm{O}}_8^{2-} \) by UV and •OH was simultaneously formed from the reaction of \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) with H2O or OH−. The reactions in the UV/PS system and these second-order rate constants are presented in Table 2.

Table 2 Summary of the reactions in the UV/PS system (pH = 7.55 and 10 mM phosphate buffer)
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Under steady-state conditions, \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \), \( {\left[\mathrm{S}{\mathrm{O}}_4^{\bullet -}\right]}_{\mathrm{SS}} \) and [HO•]SS can be expressed (for detailed derivation process, see previous paper, Yang et al. 2017):

$$ {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}}=\frac{r_{0,\mathrm{S}{\mathrm{O}}_4^{\bullet -}}\ {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}}\ \alpha -\beta \left({k}_{\mathrm{app}}-{k}_{\mathrm{UV}}\right)}{\left({k}_{\mathrm{app}}-{k}_{\mathrm{UV}}\right)\left[\mathrm{TMP}\right]-{r}_{0,\mathrm{S}{\mathrm{O}}_4^{\bullet -}}} $$
(3)
$$ {\left[\mathrm{S}{\mathrm{O}}_4^{\bullet -}\right]}_{\mathrm{SS}}=\frac{\left({k}_{\mathrm{app}}-{k}_{\mathrm{UV}}\right)\left[\mathrm{TMP}\right]-{r}_{0,\mathrm{S}{\mathrm{O}}_4^{\bullet -}}}{\alpha\ {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}}\ \left[\mathrm{TMP}\right]-\beta } $$
(4)
$$ {\left[\mathrm{H}{\mathrm{O}}^{\bullet}\right]}_{\mathrm{SS}}=\frac{\alpha\ \left(\left({k}_{\mathrm{app}}-{k}_{\mathrm{UV}}\right)\ \left[\mathrm{TMP}\right]-{r}_{0,\mathrm{S}{\mathrm{O}}_4^{\bullet -}}\right)}{\alpha {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}}\ \left[\mathrm{TMP}\right]-\beta } $$
(5)
$$ \alpha =\frac{k_2\left[{\mathrm{O}\mathrm{H}}^{-}\right]+{k}_3\left[{\mathrm{H}}_2\mathrm{O}\right]}{k_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}}\left[\mathrm{TMP}\right]+{k}_7\left[{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\right]+{k}_{11}\left[{\mathrm{H}}_2{\mathrm{PO}}_4^{-}\right]+{k}_{12}\left[{\mathrm{H}\mathrm{PO}}_4^{2-}\right]} $$
(6)
$$ \beta ={k}_2\left[{\mathrm{O}\mathrm{H}}^{-}\right]+{k}_3\left[{\mathrm{H}}_2\mathrm{O}\right]+{k}_4\left[{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\right]+{k}_9\left[{\mathrm{H}}_2{\mathrm{PO}}_4^{-}\right]+{k}_{10}\left[{\mathrm{H}\mathrm{PO}}_4^{2-}\right] $$
(7)

The average of \( {r}_{0\_\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \) and kapp was 4.623×10−8 M s−1 and 4.301×10−3 s−1, respectively. The value of α was 8.78×10−3 (unitless), and β was 9.00×103 s−1 as calculated. \( {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}} \) was 6.02±0.13×109 M−1s−1 from competition kinetics method in this study. \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \) was determined to be 3.88±0.07×109 M−1 s−1, which was consistent with the value determined by competitive kinetics. The average of \( {\left[\mathrm{S}{\mathrm{O}}_4^{\bullet -}\right]}_{\mathrm{SS}} \) and [HO•]SS was 9.46×10−15 M and 1.08×10−12 M, respectively.

With the values of \( {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}} \), \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \), [HO•]SS, \( {\left[\mathrm{S}{\mathrm{O}}_4^{\bullet -}\right]}_{\mathrm{SS}} \), the contribution of direct UV photolysis, •OH, and \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) in degrading TMP was calculated. In the UV/PS system, the values of ruv, \( {r}_{\mathrm{H}{\mathrm{O}}^{\bullet }} \), and \( {r}_{\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \) were determined to be 6.22×10−10 M s−1, 5.69×10−10 M s−1, and 4.18×10−8 M s−1, respectively. Thus, approximately 1.45% of TMP removal was attributed to direct photolysis and •OH only contributed to about 1.32% of the total degradation, indicating that\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) played a dominant role during the degradation process.

The pseudo first-order reaction kinetics model can be used to study and simulate the effects of other factors on TMP degradation. In this study, we used the pseudo first-order reaction kinetics model to predict and explain the effects of PS concentration and matrix components on TMP degradation by UV/PS process. Then the contributions of •OH and\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) to TMP degradation (i.e., kcal, UV, kcal, • OH and \( {k}_{\mathrm{cal},\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \)) under various experimental conditions were calculated by Eqs. 8, 9, and 10, respectively.

$$ {k}_{\mathrm{cal},\mathrm{UV}}={\varphi}_{\mathrm{TMP}}\times {E}_0\times \frac{\ b\ {\varepsilon}_{\mathrm{TMP}}}{A}\times \left(1-{\mathrm{e}}^{-2.303A}\right) $$
(8)
$$ {k}_{\mathrm{cal},\mathrm{H}{\mathrm{O}}^{\bullet }}={k}_{\mathrm{TMP},\mathrm{H}{\mathrm{O}}^{\bullet }}\times {\left[\mathrm{H}{\mathrm{O}}^{\bullet}\right]}_{\mathrm{SS}} $$
(9)
$$ {k}_{\mathrm{cal},\mathrm{S}{\mathrm{O}}_4^{\bullet -}}={k}_{\mathrm{TMP},\mathrm{S}{\mathrm{O}}_4^{\bullet -}}\times {\left[\mathrm{S}{\mathrm{O}}_4^{\bullet -}\right]}_{\mathrm{SS}} $$
(10)

The total contribution of •OH and\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) to TMP degradation (kcal) can be expressed by Eq. 11.

$$ {k}_{\mathrm{cal}}={k}_{\mathrm{cal},\mathrm{UV}}+{k}_{\mathrm{cal},\mathrm{H}{\mathrm{O}}^{\bullet }}+\kern0.5em {k}_{\mathrm{cal},\mathrm{S}{\mathrm{O}}_4^{\bullet -}} $$
(11)

Effect of the initial oxidant concentration

The degradation rates of TMP were affected by the PS concentration in the UV/PS system. As shown in Fig. 4, the observed pseudo first-order rate constants (in the unit of s−1) for TMP degradation (kobs) increased from 4.95×10−3 s−1 to 25.12×10−3 s−1 when PS concentration increases from 100 μM to 500 μM. This result could be accurately predicted by the kinetic model (the calculated result, kcal). The contributions of direct UV, •OH, and\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) to TMP degradation were also calculated. As the dose of PS increased, the contribution rate of direct photolysis was significantly reduced due to the competition of PS for UV, and there was no significant change in the contribution rate of •OH. However, the results showed that \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) was the main reactive species in the UV/PS system at pH 7.55, with a contribution to TMP degradation (\( {k}_{\mathrm{cal},\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \)) always greater than 97%. This result was also consistent with the previous study of Luo et al. (2016a) who found that the contribution of \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) was increased from 82.6 to 92.5% with the increase of PS concentration from 100 to 500 μM in the degradation of 2,4,6-trichloroanisole. However, Xie et al. (2015) found that the contribution of •OH was about 3.5 times and 2.0 times higher than \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) for 2-methylisoborneol and geosmin degradation, respectively. This discrepancy might be ascribed to the secondary reaction rate constant of radicals with the target compound and the concentration of free radicals in the system. The k•OH/\( {k}_{\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \) of 2,4,6-trichloroanisole and TMP were 1.37 and 1.55, which is smaller than the value of 2-methylisoborneol and geosmin (10.24 and 7.50, respectively). On the other hand, \( {\left[\mathrm{S}{\mathrm{O}}_4^{\bullet -}\right]}_{\mathrm{SS}} \) was expected to be about two orders of magnitude higher than [HO•]SS based on Eqs. 4 and 5, which was consistent with Luo’s study (Luo et al. 2016a). For instance, the calculated steady-state concentrations of \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) and •OH were about 1.08×10−12 M and 9.46×10−15 M at [PS] = 100 μM, respectively.

Fig. 4
figure 4

Impacts of PS concentration on the pseudo first-order constants of TMP (k). ([TMP]0 = 10 μM, pH = 7.55, and I0 = 7.496×10−6 Einstein L−1 s−1)

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Effect of NOM

NOM is a mixture of macromolecular organic compounds prevalent in the natural environment, widely distributed in soil, lake, river, and ocean. The impact of NOM on the degradation of organic contaminants has received more and more attention. Fulvic acid (FA) is a mixture of non-homogeneous compounds that can be dissolved in a base but cannot be dissolved in acid (Fu et al. 2006). Since FA was the main component of NOM, the effect of NOM in the UV/PS processes was examined by adding different concentrations of FA (0~2.88 mgC L−1) in this study. Figure 5 shows that kobs decreased from 11.36×10−3 s−1 to 7.07×10−3 s−1 with FA concentrations increasing from 0 to 2.88 mgC L−1. The experimental data was basically consistent with the modeling results (kcal). Two factors could be responsible for the inhibitory effect of FA on TMP degradation. First, FA would exert an inner filter effect for the photolysis of persulfate, and A can be modified to A = b(εTMP[TMP] + \( {\varepsilon}_{{\mathrm{S}}_2{\mathrm{O}}_8^{2-}} \)[\( {\mathrm{S}}_2{\mathrm{O}}_8^{2-} \)] + εFA[FA]) (εFA = 0.10 L mgC−1 cm−1 measured in this work). Second, FA acted as a radical scavenger as described by Eqs. 20 and 21 in Table 2. The kinetic model was used to estimate the relative contributions of the inner filter effect and radical scavenger in decreasing kobs values. If the inner filter effect of FA was not accounted for k, then εFA was equal to zero. As we can see from Fig. 5, kcal (green line) slightly changed as compared to the experimental data (black dots) by assuming εFA = 0. If FA did not have a radical scavenging effect (i.e., k17 and k18 values in Table 2 were equal to zero), kcal (blue line) greatly deviated from the experimental data (black dots) in Fig. 5. These results indicated that the radical scavenger effect of FA played a more significant role than the inner filter effect in decreasing TMP degradation rates. At the same time, the model calculation results showed that as the concentration of fulvic acid increased, the contribution rate of direct photolysis and \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) to TMP degradation did not change significantly, and the contribution rate of •OH dropped significantly. The possible reason was that kFA, • OH (3.0×108 \( {\mathrm{M}}_{\mathrm{C}}^{-1}{\mathrm{s}}^{-1} \)) was larger than \( {k}_{\mathrm{FA},\mathrm{S}{\mathrm{O}}_4^{\bullet -}} \) (2.35×107 \( {\mathrm{M}}_{\mathrm{C}}^{-1}{\mathrm{s}}^{-1} \)) (Weishaar et al. 2003; Xie et al. 2015).

Fig. 5
figure 5

Impacts of fulvic acid on the pseudo first-order constants of TMP (k) ([TMP]0 = 10 μM, [PS]0 = 200 μM, pH = 7.55, and E0 = 7.496×10−6 Einstein L−1 s−1)

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Effect of inorganic anions

Inorganic anions present in environmental water (e.g., \( \mathrm{S}{\mathrm{O}}_4^{2-},\mathrm{N}{\mathrm{O}}_3^{-} \), Cl−, and \( \mathrm{HC}{\mathrm{O}}_3^{-} \)) also have a greater effect on the degradation of TMP by UV/PS. In this study, the effect of inorganic ions on the degradation kinetics of \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) was studied by adding anion concentrations close to those in the environment in deionized water (\( \mathrm{S}{\mathrm{O}}_4^{2-} \) and \( \mathrm{N}{\mathrm{O}}_3^{-} \) concentration of 0.5 mM, Cl− and \( \mathrm{HC}{\mathrm{O}}_3^{-} \) the concentration is 1 mM). The results (Fig. 6) show that \( \mathrm{S}{\mathrm{O}}_4^{2-} \) has little effect on the degradation of TMP, while \( \mathrm{HC}{\mathrm{O}}_3^{-} \) will promote the degradation of TMP (11.5%). The main reason why \( \mathrm{HC}{\mathrm{O}}_3^{-} \) promotes the TMP degradation in the UV/PS system is that \( \mathrm{HC}{\mathrm{O}}_3^{-} \) can react with \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) to produce \( {\mathrm{CO}}_3^{\bullet -} \) (1.59 V NHE), which has certain oxidation ability. The oxidation potential of \( {\mathrm{CO}}_3^{\bullet -} \) was 1.59 V vs NHE (Giannakis et al. 2021). The secondary free radicals \( {\mathrm{CO}}_3^{\bullet -} \) can continue to react with TMP (Lian et al. 2017; Luo et al. 2016b; Zuo et al. 1999). \( \mathrm{N}{\mathrm{O}}_3^{-} \) and Cl− will inhibit the degradation of TMP, and the inhibition rates are 32.1% and 20.4%, respectively. \( \mathrm{N}{\mathrm{O}}_3^{-} \) can produce •OH, \( {\mathrm{NO}}_3^{\bullet -} \), and \( {\mathrm{NO}}_2^{\bullet -} \) under a series of complex reactions under UV and sunlight (Dong and Rosario-Ortiz 2012; Ji et al. 2012; Keen et al. 2012; Vione et al. 2006; Xiao et al. 2014). The oxidation potentials of \( {\mathrm{NO}}_3^{\bullet -} \) and \( {\mathrm{NO}}_2^{\bullet -} \) were 2.3–2.5 V and 1.03 V, respectively (Giannakis et al. 2021). However, •OH, \( {\mathrm{NO}}_2^{\bullet -}, \) and \( {\mathrm{NO}}_2^{-} \) have lower quantum yields, \( {\varPhi}_{\bullet \mathrm{OH}/{\mathrm{NO}}_2^{\bullet -}} \)=0.24, \( {\varPhi}_{{\mathrm{NO}}_2^{-}} \)=0.015~0.028 mol Einstein−1 (Keen et al. 2012). According to Keen’s calculation, the [HO•]SS produced by UV excitation 0.5 mM \( \mathrm{N}{\mathrm{O}}_3^{-} \) is only 0.74×10−18 M (Keen et al. 2012). At the same time, \( \mathrm{N}{\mathrm{O}}_3^{-} \) can also quench •OH/\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \). In the UV/\( {\mathrm{S}}_2{\mathrm{O}}_8^{2-} \) system, Cl− and \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) can react to produce a series of new secondary free radicals, such as Cl•, ClOH•−, and \( {\mathrm{Cl}}_2^{\bullet -} \) (Luo et al. 2016b). The oxidation potentials of Cl•, ClOH•− and \( {\mathrm{Cl}}_2^{\bullet -} \) were 1.9V, 2.47 V, and 2.0 V, respectively (Giannakis et al. 2021). On the one hand, Cl− can quench \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) (Lian et al. 2017; Yang et al. 2014; Yang et al. 2016). On the other hand, the generated secondary free radicals can continue to oxidatively degrade TMP. Cl• and \( {\mathrm{Cl}}_2^{\bullet -} \) have strong oxidation ability. The oxidation potentials of Cl• and \( {\mathrm{Cl}}_2^{\bullet -} \) were 2.47 V and 2.0 V, respectively (Beitz et al. 1998). Cl• is also a selective oxidant, which can react with electron-rich components through electron transfer, hydrogen extraction, and addition (Beitz et al. 1998). Therefore, the contribution of secondary free radicals (Cl•, ClOH•−, and \( {\mathrm{Cl}}_2^{\bullet -} \)) and the inhibition of Cl− comprehensive determine the effect on TMP degradation.

Fig. 6
figure 6

Inhibition of inorganic anions on TMP degradation in DI water

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The effect of Cl− on TMP degradation was studied using concentrations ranging from 0 to 5 mM. As shown in Fig. 7, for the effect of Cl−, kobs slightly decreased from 11.36×10−3 s−1 in the absence of Cl− to 5.73×10−3 s−1 in the presence of 5 mM Cl−. The negative effect of Cl− in the system might result from the fact that significant amounts of \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) react with Cl− forming chloride-derived radicals. For instance, the influence of Cl− depended on its fast rate constants with both •OH and \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) (Eqs. 22 and 23 in Table 2). Usually, the fast reaction of \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) with Cl− yields secondary reactive chlorine radical species, such as Cl•, ClOH•, and \( {\mathrm{Cl}}_2^{\bullet -} \) through complex chain reactions (i.e., Eqs. 24 to 31 in Table 2).

Fig. 7
figure 7

Impacts of Cl− on the degradation of TMP ([TMP]0 = 10 μM, [PS]0 = 200 μM, pH = 7.55, and E0 = 7.496×10−6 Einstein L−1 s−1)

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Since most of those rate constants of reactive chlorine radicals with TMP are relatively unknown, it is difficult to accurately predict how Cl− affected TMP degradation in the UV/PS process by the steady-state kinetic model. Thus, in this study, only Cl− was considered the scavenger of •OH and \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) for calculating the kcal values. As shown in Fig. 6, kcal was significantly smaller than the experimental data (red line), which contradicts the observation that Cl− had a minor inhibitory effect on TMP degradation. Still, those secondary radicals might play a role in the degradation of TMP in UV/PS process, since Cl• and \( {\mathrm{Cl}}_2^{\bullet -} \) were strong oxidants with oxidation potentials of 2.47 V and 2.0 V, respectively (Beitz et al. 1998). Cl• is a selective oxidant that reacts with electron-rich moieties through one-electron oxidation, H-abstraction, and addition to unsaturated bonds (Grebel et al. 2010). Previous studies had shown conflicting results of Cl− in the UV/\( {\mathrm{S}}_2{\mathrm{O}}_8^{2-} \) process, depending on the target compound. Yuan et al. (2011) reported that a dual effect of chloride (i.e., inhibitory and accelerating effect) on azo dye (Acid Orange 7) degradation in an emerging cobalt/peroxymonosulfate (Co/PMS) advanced oxidation process. Ghauch et al. (2017) showed that the k value of chloramphenicol increased first and then decreased with the concentration increase of chloride anion in UV/PS system. The reaction between •OH/\( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) and Cl− produces secondary chlorine radical species, but the influence of Cl− was compound dependent, either promoting target compound removal (Criquet and Leitner 2009; Fang et al. 2012), or inhibiting the degradation (Liang et al. 2006; Shah et al. 2013).

Effect of transition metal ions

Transition metal ions (Mn+) were very common in industrial wastewater, and some research results showed that the presence of Mn+ had a very important effect on the degradation of organic matter by UV-activated persulfate system. As shown in Fig. 8, as the concentration of transition metal ions (Cu2+, Zn2+, Co2+) in the UV/PS system increased from 0 to 100 μM, kexp increased from 1.96×10−3 to 2.44×10−3, 3.11×10−3, 2.62×10−3 s−1, respectively. The results showed that the presence of transition metal ions can significantly promote the degradation of TMP. Take Co2+ for example, The main reason was that transition metal ions can activate PS to produce more \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \), such as Eq. (12) (Liu et al. 2012; Nfodzo and Choi 2011).

$$ {\mathrm{M}}^{\mathrm{n}+}+{\mathrm{S}}_2{\mathrm{O}}_8^{2-}\to {\mathrm{M}}^{\left(\mathrm{n}+1\right)+}+\mathrm{S}{\mathrm{O}}_4^{\bullet -}+\mathrm{S}{\mathrm{O}}_4^{2-} $$
(12)
Fig. 8
figure 8

Impacts of transition metal ions on the pseudo first-order constants of TMP (k) ([TMP]0 = 10 μM, [PS]0 = 100 μM, pH = 7.55, and E0 = 4.556×10−6 Einstein L−1 s−1).

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At the same time, the generated M(n + 1)+ is not stable, and M(n + 1)+ oxidative degradation of TMP may be another main reason for the significant increase in kexp (Liang et al. 2013).

However, as the concentration of transition metal ions (Cu2+, Zn2+, Co2+) in the system continued to increase from 100 to 400 μM, kexp did not continue to increase but gradually decreased. The main reason was that excess transition metal ions can be combined with \( \mathrm{S}{\mathrm{O}}_4^{\bullet -} \) to continue the reaction and consume free radicals (Eq. 24) (Furman et al. 2010; Nfodzo and Choi 2011), and excessive transition metal ions can also form hydrated ions with water (such as [Zn(H2O)6 ]2+) as UV contenders affected TMP and \( {\mathrm{S}}_2{\mathrm{O}}_8^{2-} \) absorption of UV.

$$ {\mathrm{M}}^{\mathrm{n}+}+\mathrm{S}{\mathrm{O}}_4^{\bullet -}\to {\mathrm{M}}^{\left(\mathrm{n}+1\right)+}+\mathrm{S}{\mathrm{O}}_4^{2-} $$
(13)

Transition metal ions showed a good synergistic effect in the UV/PS system. Only a very small amount of transition metal ions can significantly accelerate the degradation of organic pollutants, which provided a meaningful exploration for the collaborative and comprehensive treatment of industrial wastewater.

Conclusions

The lower molar absorption coefficient and quantum yield limited the direct photolysis of TMP. After adding H2O2 or PS to the reaction system, the degradation effect of TMP was significantly enhanced, among which free radicals played a major role. \( {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}} \) and \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \) showed good agreement with previously reported values, which was measured by competitive kinetics method and steady-state assumption model. Steady-state approximation and kinetic model were also developed in order to predict and simulate the destruction of TMP by a variety of water matrices in UV-activated persulfate system. kobs of TMP degradation decreased in the presence of FA and Cl−, but kobs was increased with the concentration of PS. Transition metal ions had a good synergistic effect in UV/PS degradation of TMP. However, excessive transition metal ions could reduce the TMP degradation due to the trapping effect of free radicals and the competition of formed hydrated ions against UV. The reported \( {k}_{\mathrm{H}{\mathrm{O}}^{\bullet },\mathrm{TMP}} \) and \( {k}_{{\mathrm{SO}}_4^{\bullet -},\mathrm{TMP}} \) values and effect of matrix components are more beneficial to predict and explain TMP degradation mechanism and select more efficient radical-based advanced oxidation processes in engineered water.