Degradation of Tryptophan by UV Irradiation: Inﬂuencing Parameters and Mechanisms

: The chlorination of dissolved amino acids can generate disinfection by-products (DBPs). To prevent the formation of DBPs, we examined the UV-induced degradation of tryptophan (Trp). In order to further understand the impact of UV disinfection on Trp, the effects of initial concentrations of Trp, pH, temperature, concentrations of NO 3 − , HCO 3 − and Cl − on Trp removal were investigated, and a degradation mechanism was also proposed. The results demonstrated that degradation ﬁtted a pseudo ﬁrst-order reaction kinetic model. The degradation of Trp was mainly caused by direct UV degradation. The apparent rate constant k obs decreased with the increase in initial Trp concentration and increased with increases in pH and temperature. The thermal degradation activation energy was 19.65 kJ/mol. Anions in water also had a signiﬁcant inﬂuence on the degradation of Trp. HCO 3 − and NO 3 − contributed to the k obs of Trp, but Cl − inhibited the degradation rate. By electron paramagnetic resonance (EPR) spectroscopy, · OH was proven to be formed during the degradation of Trp by UV. Based on the intermediate products of C 11 H 15 NO 3 , C 10 H 15 N and C 9 H 13 N detected by LC-MS-MS, the degradation pathway of Trp was speculated.


Introduction
Dissolved organic nitrogen in surface water bodies primarily consists of amino acids (AAs), which are present in concentrations ranging from 20 to 10,000 µg/L [1]. Cai, et al. [2] detected that the total AA content in East Taihu Lake was 32,158 ng/L during autumn and approximately 3000 ng/L for the rest of the year. Dotson and Westerhoff [3] reported that the total concentration of AAs, free and combined, was in the range of 50-1000 µg/L in rivers, streams and lakes. The AA concentration in natural water bodies is drastically affected by the presence of algae; AA concentrations increased with the occurrence of algal blooms [1,4].
The total AA concentration includes free AAs (0.20-0.25) and combined AAs in the form of proteins, peptides and humic-bound AAs. Although the overall free AA concentration is relatively low, a small fraction remains even after the growth of cyanobacteria or blue-green algae [5]. In addition, these low-molecular-weight AAs can only be removed by the chlorination process in traditional water treatment plants [3]. However, chlorination or chloroamination converts the free AAs to a variety of disinfection by-products (DBPs), including trihalomethanes, haloacetaldehydes, haloacetonitriles, haloacetamides and halonitromethanes [6][7][8]. For example, aspartic acid and Trp can be converted to dichloroacetonitrile [9], while L-tyrosine and aspartic acid form dichloroacetamide [8].
Several methods, such as improved coagulation, nanofiltration, reverse osmosis, ultraviolet (UV) irradiation and advanced oxidation processes, have been proposed to control the number of DBP precursors [10][11][12]. Of these, low-pressure UV irradiation is a promising alternative for disinfection in drinking water treatment plants because it is a costeffective process with an easily implementable system [13]. Recently, UV disinfection has received considerable attention in the field of water treatment [14]. Thus, UV irradiation All chemicals were of analytical grade, except as noted. Trp (≥98%) and methanol were obtained from Aladdin (Shanghai, China). Atrazine and t-butanol were purchased from Sigma Aldrich. Sodium dihydrogen phosphate and disodium phosphate were purchased from Sinopharm Chemical Reagent Co. Ltd. (Shanghai, China), and were used as received. All solutions were prepared using ultrapure water (NW Ultrapure water system, HealForce, Shanghai, China), which was obtained from a Milli-Q system with a resistivity >18 MV·cm.

UV Reactor
UV irradiation experiments were conducted in a UV reactor equipped with a 6 W LP Hg UV lamp (254 nm, 4P-SE, Philips), as shown in Figure S1. A small stirring rotor was placed at the bottom of the reactor to ensure that all molecules in the solution were homogenously exposed to UV light. Atrazine and hydrogen peroxide were used as actinometers to determine the UV fluence rate (I 0 ; 4 ×10 −7 Einstein s −1 L −1 ) and effective path length (b; 4.1 cm) [19,20].

Experimental Procedure
All UV-induced Trp degradation experiments were performed at a constant pH, which was maintained using 10 mM phosphate buffers.
To investigate the effects of initial concentrations, temperatures and ionic concentrations on Trp degradation, a 700 mL testing solution was prepared, which consisted of Trp (0.6-10 mg/L), nitrate (0-500 mg/L) and bicarbonate ions (0-500 mg/L). The effects of pH and temperature were studied in the ranges of 6-7.5 and 7-27 • C, respectively. The mixture was stirred using a magnetic stirrer. A constant temperature was maintained by circulating cold water. Samples (1 mL) were collected at different retention times and t-butanol was added to quench the reaction. These samples were analyzed by liquid chromatography (LC). Samples (500 mL) collected after 0, 10, 20, 40, 60, 90 and 120 min were analyzed by liquid chromatography tandem mass spectrometry (LC-MS-MS).

Analytical Methods
Trp concentration was measured by high-performance liquid chromatography (HPLC; Agilent, Santa, SA). The HPLC system was equipped with an extended C18 column (4.6 × 250 mm, 5 mm, Waters Co., Milford, USA). Phosphoric acid in methanol (85% v/v) was used as the mobile phase with a flow rate of 1.0 mL/min. The excitation wavelength and emission wavelength of Trp were 260 nm and 340 nm, respectively. The HPLC graphs of Trp can be seen in Figure S2.
·OH was tested with a Bruker A300 EPR Spectrometer manufactured by Bruker, Germany. DMPO was used as an electron spin capture agent. In this experiment, because the phosphoric acid buffer solution could reduce the signal strength of DMPO-·OH, no phosphoric acid buffer solution was added to the reaction solution. The process was as follows: 10 mg/L Trp solution was exposed to UV, then 1 mL of Trp solution was quickly mixed with 1 mL of 16000 mg/L DMPO and transferred into a 200 µL capillary tube for The intermediate products were characterized by LC-MS-MS. The compounds were separated on a ZORBAX-SB C18 (250 mm × 4.6 mm, 5.0 µm) column. The mobile phase was acetonitrile-isopropanol (1:1) elution, the flow rate was 1.0 mL/min, the detection wavelength was 210 nm, the column temperature was 30 • C and the injection quantity was 10 µL. Electrospray ion source positive ion mode was adopted and the scanning range was m/z 100-1500.

Kinetic Model of UV Degradation of Trp
The pseudo first-order reaction kinetic equation is shown in Equation (1): where t is the reaction time, y is the ratio of concentration of Trp at t to the initial concentration and k obs is the reaction rate constant. The degradation curve of Trp is fitted by Equation (1) and the fitting curve is shown in Figure 1. The linear regression coefficient R 2 of the curve was greater than 0.98 for Trp, which indicated that its reaction by UV fitted a new pseudo first-order reaction kinetic model. wavelength was 210 nm, the column temperature was 30 °C and the injection was 10 µL. Electrospray ion source positive ion mode was adopted and the scannin was m/z 100-1500.

Kinetic Model of UV Degradation of Trp
The pseudo first-order reaction kinetic equation is shown in Equation (1): where t is the reaction time, y is the ratio of concentration of Trp at t to the initial tration and kobs is the reaction rate constant. The degradation curve of Trp is fitted by Equation (1) and the fitting curve i in Figure 1. The linear regression coefficient R 2 of the curve was greater than 0.98 which indicated that its reaction by UV fitted a new pseudo first-order reaction model.
UV degradation principles can be divided into direct degradation and indir radation. Direct degradation refers to the direct oxidation, bond breaking, isome and rearrangement of organic matter by UV [21]. Indirect degradation refers to duction of radicals such as ·OH. In order to verify the existence of ·OH, a radical q TBA (tert-butyl alcohol, 0.5 mM) was added during the UV treatment to compare ferences between direct and total degradation with and without TBA. Figure 1 sh the UV degradation could generate ·OH. From two kobs of curves, we inferred tha dation of Trp occurred mainly by direct UV degradation. The first pathway to produce ·OH by UV irradiation is to excite and react T water molecules to generate ·OH [22]. The second pathway is the intermediate rea Trp through hydration electrons, superoxide anion radicals and hydrogen p which finally produces hydroxyl radicals [23,24]. The reaction equation is shown tions (2) and (3). Pathway 1: UV degradation principles can be divided into direct degradation and indirect degradation. Direct degradation refers to the direct oxidation, bond breaking, isomerization and rearrangement of organic matter by UV [21]. Indirect degradation refers to the production of radicals such as ·OH. In order to verify the existence of ·OH, a radical quencher TBA (tert-butyl alcohol, 0.5 mM) was added during the UV treatment to compare the differences between direct and total degradation with and without TBA. Figure 1 shows that the UV degradation could generate ·OH. From two k obs of curves, we inferred that degradation of Trp occurred mainly by direct UV degradation.
The first pathway to produce ·OH by UV irradiation is to excite and react Trp with water molecules to generate ·OH [22]. The second pathway is the intermediate reaction of Trp through hydration electrons, superoxide anion radicals and hydrogen peroxide, which finally produces hydroxyl radicals [23,24]. The reaction equation is shown in Equations (2) and (3).

Effect of Initial Trp Concentration
The degradation of Trp was examined at different initial concentrations (i.e., 0.6, 2, 4, 6, 8 mg/L). As displayed in Figure 2, increasing the initial Trp concentration decreased the Trp removal rate, and these findings are consistent with those of previous studies [25,26].
The degradation of Trp was examined at different initial concentrations (i.e., 6, 8 mg/L). As displayed in Figure 2, increasing the initial Trp concentration decre Trp removal rate, and these findings are consistent with those of previous studie On the one hand, the light quantum generated by unit time UV irradiation ef remains unchanged. The higher the Trp concentration, the less light quantum cap a single Trp molecule, which leads to the chemical bond of Trp being less like broken and slows down the direct degradation rate of Trp. On the other hand, th ·OH generation remains constant if the UV intensity is maintained, while the rat consumption increases with an increase in the initial Trp concentration [27,28]. M an increase in the initial Trp concentration increases the amount of organic matter volume, which decreases the exposure of a molecule to UV irradiation. Thus, the tration of ·OH also decreases, causing the reaction rate to decrease.

Figure 2. Effect of initial Trp concentration on Trp degradation
In order to analyze the relationship between initial Trp concentration and experimental data were fit to Equation (4), as shown in Figure 3. It was shown th the increase in initial Trp concentration, the kobs became smaller and the rate of slowed down. On the one hand, the light quantum generated by unit time UV irradiation effectively remains unchanged. The higher the Trp concentration, the less light quantum captured by a single Trp molecule, which leads to the chemical bond of Trp being less likely to be broken and slows down the direct degradation rate of Trp. On the other hand, the rate of ·OH generation remains constant if the UV intensity is maintained, while the rate of ·OH consumption increases with an increase in the initial Trp concentration [27,28]. Moreover, an increase in the initial Trp concentration increases the amount of organic matter per unit volume, which decreases the exposure of a molecule to UV irradiation. Thus, the concentration of ·OH also decreases, causing the reaction rate to decrease.
In order to analyze the relationship between initial Trp concentration and k obs , the experimental data were fit to Equation (4), as shown in Figure 3. It was shown that, with the increase in initial Trp concentration, the k obs became smaller and the rate of descent slowed down. k obs = −0.01ln(C 0 ) + 0.03 (4) Water 2021, 13, x FOR PEER REVIEW

Effect of pH
It is known that pH can affect the mechanism and pathways of degradation clarify how pH affected the degradation of Trp during UV irradiation, the samp different pH (i.e., 6.0, 6.5, 7, 7.5, 8) were prepared; the results are shown in Figur degradation rate increased with an increase in pH, and was the highest at pH 8, i that increasing the solution pH promoted Trp removal.
According to a widely acknowledged previous study, increasing the solutio duces the ratio of ·OH, as shown in Eq (5). When the pH was increased from 4.7 to ·OH concentration decreased by 80% [29]. However, this experiment found that, increase in pH, the degradation rate of Trp increased, which indicates that the ma anism of Trp degradation is not indirect degradation through the generation of · direct UV degradation. Bergman [30] found that the hydrogen bonded to the firs in a nitrogen-containing heterocyclic ring is highly acidic, implying that the sy effect of alkaline conditions and ultraviolet light could promote the ring-opening gen-containing heterocycles, thus accelerating direct degradation.

Effect of pH
It is known that pH can affect the mechanism and pathways of degradation [21].To clarify how pH affected the degradation of Trp during UV irradiation, the samples with different pH (i.e., 6.0, 6.5, 7, 7.5, 8) were prepared; the results are shown in Figure 4. The degradation rate increased with an increase in pH, and was the highest at pH 8, implying that increasing the solution pH promoted Trp removal.

Effect of pH
It is known that pH can affect the mechanism and pathways of degradation clarify how pH affected the degradation of Trp during UV irradiation, the samp different pH (i.e., 6.0, 6.5, 7, 7.5, 8) were prepared; the results are shown in Figu degradation rate increased with an increase in pH, and was the highest at pH 8, i that increasing the solution pH promoted Trp removal.
According to a widely acknowledged previous study, increasing the solutio duces the ratio of ·OH, as shown in Eq (5). When the pH was increased from 4.7 to ·OH concentration decreased by 80% [29]. However, this experiment found that, increase in pH, the degradation rate of Trp increased, which indicates that the ma anism of Trp degradation is not indirect degradation through the generation of · direct UV degradation. Bergman [30] found that the hydrogen bonded to the firs in a nitrogen-containing heterocyclic ring is highly acidic, implying that the sy effect of alkaline conditions and ultraviolet light could promote the ring-opening gen-containing heterocycles, thus accelerating direct degradation.  The effect of temperature on the degradation of Trp is shown in Figure 5 demonstrates that Trp degradation is facilitated at a higher temperature. Owin According to a widely acknowledged previous study, increasing the solution pH reduces the ratio of ·OH, as shown in Equation (5). When the pH was increased from 4.7 to 8.4, the ·OH concentration decreased by 80% [29]. However, this experiment found that, with the increase in pH, the degradation rate of Trp increased, which indicates that the main mechanism of Trp degradation is not indirect degradation through the generation of ·OH, but direct UV degradation. Bergman [30] found that the hydrogen bonded to the first carbon in a nitrogen-containing heterocyclic ring is highly acidic, implying that the Water 2021, 13, 2368 6 of 12 synergistic effect of alkaline conditions and ultraviolet light could promote the ring-opening of nitrogen-containing heterocycles, thus accelerating direct degradation.

Effect of Temperature
The effect of temperature on the degradation of Trp is shown in Figure 5, which demonstrates that Trp degradation is facilitated at a higher temperature. Owing to the low activation energy, the rate constant for the reaction between ·OH and Trp is largely unaffected by reaction temperature [31,32], suggesting that reaction temperature directly contributes to Trp degradation.
Water 2021, 13, x FOR PEER REVIEW low activation energy, the rate constant for the reaction between ·OH and Trp i unaffected by reaction temperature [31,32], suggesting that reaction temperature contributes to Trp degradation. To further investigate the influence of reaction temperature on Trp degrada calculated rate constant kobs, was fitted with the Arrhenius equation [31], as sh Equation (6): where A is the pre-exponential factor, Ea is the apparent activation energy, R is versal gas constant (8.314 J/mol/K) and T is the temperature in Kelvin. As shown in Figure 6, Ea was calculated to be approximately 19.65 kJ/mol. Ge most activation energy of the chemical reaction ranged from 50 to 250 kJ/mol, lower the activation energy, the easier it is for the reaction to take place. Therefo perature has a significant effect on the UV degradation of Trp and it is specula temperature promotes the direct degradation of Trp. To further investigate the influence of reaction temperature on Trp degradation, the calculated rate constant k obs, was fitted with the Arrhenius equation [31], as shown in Equation (6): where A is the pre-exponential factor, Ea is the apparent activation energy, R is the universal gas constant (8.314 J/mol/K) and T is the temperature in Kelvin. As shown in Figure 6, Ea was calculated to be approximately 19.65 kJ/mol. Generally, most activation energy of the chemical reaction ranged from 50 to 250 kJ/mol, and the lower the activation energy, the easier it is for the reaction to take place. Therefore, temperature has a significant effect on the UV degradation of Trp and it is speculated that temperature promotes the direct degradation of Trp.
versal gas constant (8.314 J/mol/K) and T is the temperature in Kelvin.
As shown in Figure 6, Ea was calculated to be approximately 19.65 kJ/mol. G most activation energy of the chemical reaction ranged from 50 to 250 kJ/mol, lower the activation energy, the easier it is for the reaction to take place. Therefo perature has a significant effect on the UV degradation of Trp and it is specula temperature promotes the direct degradation of Trp.

Effect of Water Matrix
HCO3 − is a common anion (50-200 mg/L) in natural water [33] that can affec moval rate of organic pollutants [34]. HCO3 − can react with ·OH to form CO3 2− a (Equation (7)). It should be noted that pH can significantly affect the form of bica ions, and when the pH is too low, hydrolyzed bicarbonate change to carbon dio

Effect of Water Matrix
HCO 3 − is a common anion (50-200 mg/L) in natural water [33] that can affect the removal rate of organic pollutants [34]. HCO 3 − can react with ·OH to form CO 3 2− and H 2 O (Equation (7)). It should be noted that pH can significantly affect the form of bicarbonate ions, and when the pH is too low, hydrolyzed bicarbonate change to carbon dioxide. In order to better study the effect of bicarbonate on the UV degradation of Trp, pH was set at 7; the result is presented in Figure 7.
Water 2021, 13, x FOR PEER REVIEW order to better study the effect of bicarbonate on the UV degradation of Trp, pH at 7; the result is presented in Figure 7.
·OH + HCO3 − →CO3 2− + H2O As shown in Figure 7, HCO3 − can promote the degradation of Trp. Unlike ·OH has a high selectivity and preferentially reacts with compounds with electron-ri ties, such as AAs, via an electron transfer or hydrogen abstraction reaction [35,3 dition to the functional groups of AAs themselves, Trp has electron-rich ind chains, which are more likely to react with HCO3 − to accelerate the degradation o Similar to HCO3 − , NO3 − is also widely present in natural water bodies at co tions ranging from 0.01 mmol/L to 1 mmol/L. The photolysis of NO3 − could gene (Equation (8)), which could degrade the organic matter present in natural wate [37]. Figure 8 demonstrates that NO3 − accelerates the photodegradation of Trp As shown in Figure 7, HCO 3 − can promote the degradation of Trp. Unlike ·OH, HCO 3 − has a high selectivity and preferentially reacts with compounds with electron-rich moieties, such as AAs, via an electron transfer or hydrogen abstraction reaction [35,36]. In addition to the functional groups of AAs themselves, Trp has electron-rich indole side chains, which are more likely to react with HCO 3 − to accelerate the degradation of Trp. Similar to HCO 3 − , NO 3 − is also widely present in natural water bodies at concentrations ranging from 0.01 mmol/L to 1 mmol/L. The photolysis of NO 3 − could generate ·OH (Equation (8)), which could degrade the organic matter present in natural water bodies [37]. Figure 8 demonstrates that NO 3 − accelerates the photodegradation of Trp Because the concentration of Cl − in natural water is typically 1-100 mg/L, t of different concentrations of Cl − on the UV degradation of Trp was also investiga experimental results are shown in Figure 9.
With the increase in Cl − concentration from 1 mmol/L to 10 mmol/L, the ko decreased from 0.021 to 0.015 min −1 , as shown in Figure 9. Chlorine-containing fre species are known as active chlorine species (RCS), including ·Cl, HOCl − and Cl2 − idation potentials of 2.47 V, 1.6-1.8 V and 2.0 V [38,39], respectively. These radic a lower potential than that of ·OH (E = 2.8 V) [40]. The method of generating RC seen in Equation (9). A number of studies have shown that RCS has a weak a oxidize electron-rich organics [38,41,42]; thus, it is speculated that ·OH is conve active chlorine species with a weaker oxidization ability by the increase in Cl − co tion, which slows down the degradation rate of Trp.  Because the concentration of Cl − in natural water is typically 1-100 mg/L, t of different concentrations of Cl − on the UV degradation of Trp was also investiga experimental results are shown in Figure 9.
With the increase in Cl − concentration from 1 mmol/L to 10 mmol/L, the ko decreased from 0.021 to 0.015 min −1 , as shown in Figure 9. Chlorine-containing fre species are known as active chlorine species (RCS), including ·Cl, HOCl − and Cl2 − idation potentials of 2.47 V, 1.6-1.8 V and 2.0 V [38,39], respectively. These radic a lower potential than that of ·OH (E = 2.8 V) [40]. The method of generating RC seen in Equation (9). A number of studies have shown that RCS has a weak a oxidize electron-rich organics [38,41,42]; thus, it is speculated that ·OH is conve active chlorine species with a weaker oxidization ability by the increase in Cl − co tion, which slows down the degradation rate of Trp.  With the increase in Cl − concentration from 1 mmol/L to 10 mmol/L, the k obs of Trp decreased from 0.021 to 0.015 min −1 , as shown in Figure 9. Chlorine-containing free radical species are known as active chlorine species (RCS), including ·Cl, HOCl − and Cl 2 − with oxidation potentials of 2.47 V, 1.6-1.8 V and 2.0 V [38,39], respectively. These radicals have a lower potential than that of ·OH (E = 2.8 V) [40]. The method of generating RCS can be seen in Equation (9). A number of studies have shown that RCS has a weak ability to oxidize electron-rich organics [38,41,42]; thus, it is speculated that ·OH is converted into active chlorine species with a weaker oxidization ability by the increase in Cl − concentration, which slows down the degradation rate of Trp.

Generation of Radical Species
The presence of ·OH in the reaction mixture was further investigated by the addition of a spin trapping agent, 5,5-dimethyl-1-pyrroline-1-oxide (DMPO), followed by electron paramagnetic resonance (EPR) spectroscopy [43]. During this experiment, the phosphate buffer solution was not added in the reaction mixture because it could decrease the amount of DMPO and the signal intensity of the DMPO-OH adduct [44]. The results from the EPR analysis( Figure 10) further verified the formation of ·OH.
Water 2021, 13, x FOR PEER REVIEW

Generation of Radical Species
The presence of ·OH in the reaction mixture was further investigated b of a spin trapping agent, 5,5-dimethyl-1-pyrroline-1-oxide (DMPO), follow paramagnetic resonance (EPR) spectroscopy [43]. During this experiment, buffer solution was not added in the reaction mixture because it could amount of DMPO and the signal intensity of the DMPO-OH adduct [44]. T the EPR analysis( Figure 10) further verified the formation of ·OH.

Degradation Products
Trp degradation was examined using LC-MS-MS. The secondary mas of intermediates is presented in Figures S3-S5, and the atom potential map o in Figure S6. The N (0.38) on the indole group has a significant difference in between adjacent Cs (−0.07 and −0.04); thus, it is easier for the heterocyclic s to initiate a ring-opening reaction. Moreover, carboxyl can be cleaved ea the large potential difference in its C-C bond. Based on these intermediates tential map, the proposed degradation pathway of Trp by UV treatment is e shown in Figure 11.

Degradation Products
Trp degradation was examined using LC-MS-MS. The secondary mass spectrometry of intermediates is presented in Figures S3-S5, and the atom potential map of Trp is shown in Figure S6. The N (0.38) on the indole group has a significant difference in atom potential between adjacent Cs (−0.07 and −0.04); thus, it is easier for the heterocyclic structure of Trp to initiate a ring-opening reaction. Moreover, carboxyl can be cleaved easily because of the large potential difference in its C-C bond. Based on these intermediates and atom potential map, the proposed degradation pathway of Trp by UV treatment is established and shown in Figure 11. of intermediates is presented in Figures S3-S5, and the atom potential map of Trp is shown in Figure S6. The N (0.38) on the indole group has a significant difference in atom potential between adjacent Cs (−0.07 and −0.04); thus, it is easier for the heterocyclic structure of Trp to initiate a ring-opening reaction. Moreover, carboxyl can be cleaved easily because of the large potential difference in its C-C bond. Based on these intermediates and atom potential map, the proposed degradation pathway of Trp by UV treatment is established and shown in Figure 11.  Figure 11. Proposed degradation pathway of Trp by UV treatment.
During the mechanism, Trp was initially converted to C11H16N2O2 via the cleavage of C-N bonds. The aromatic ring in C11H16N2O2 was attacked by ·OH to afford C11H15NO3, after which C11H15NO3 was further oxidized to form C10H15N through de-hydroxy and decarboxylic reactions. Finally, C9H13N was formed in the cleavage loss of methyl. During the mechanism, Trp was initially converted to C 11 H 16 N 2 O 2 via the cleavage of C-N bonds. The aromatic ring in C 11 H 16 N 2 O 2 was attacked by ·OH to afford C 11 H 15 NO 3 , after which C 11 H 15 NO 3 was further oxidized to form C 10 H 15 N through de-hydroxy and de-carboxylic reactions. Finally, C 9 H 13 N was formed in the cleavage loss of methyl.

Conclusions
The degradation curve of Trp fit a pseudo first-order reaction kinetic model, primarily by direct degradation. Degradation was promoted by an increase in pH, temperature, and HCO 3 − and NO 3 − concentrations. In contrast, increasing the initial Trp and Cl − concentrations decreased the degradation rate. The Ea value was estimated to be 19.65 kJ/mol. The EPR results further verified the presence of ·OH in the reaction mixture. The degradation pathway was proposed on the basis of three intermediate products.
Author Contributions: Conceptualization, W.F. and Y.Y.; data curation, W.F. and K.Z.; methodology, Y.Y.; supervision, J.J. All authors have read and agreed to the published version of the manuscript.