Abstract

The advanced oxidation technologies based on OH can effectively degrade the pharmaceutical and personal care products under operating conditions of normal temperature and pressure. In this study, direct photolysis of ibuprofen (IBU) is slow due to the relatively low molar extinction coefficient and quantum yield. Compared to direct photolysis, the degradation kinetics of IBU was significantly enhanced in the UV/H2O2 system, mainly by OH radical mediated oxidation. In the UV/H2O2 system, the degradation rate of ionic IBU was slightly faster than that of the molecular form. Kinetic analysis showed that the second-order reaction rate constant of ionic IBU (5.51 × 109 M−1 s−1) was higher than that of the molecular form (3.43 × 109 M−1 s−1). The pseudo first-order rate constant for IBU degradation (kobs) increased with increasing H2O2 dosage. kobs can be significantly decreased in the presence of natural organic matter (NOM), which is due to (i) NOM radical scavenging effects (dominant role) and (ii) UV absorption. The degradation of IBU was inhibited by HCO3, which was attributed to its scavenging effect. Interestingly, when NO3 was present in aqueous solution, a slight increase in the degradation rate was observed, which was due to NO3 absorbing photons to generate OH at a low quantum yield. No obvious effects were observed when SO42 and Cl were present.

INTRODUCTION

Pharmaceutical and personal care products (PPCPs) are commonly used chemicals, which can be easily found in various water matrices, such as wastewater effluent, seawater and surface water (Kim et al. 2007; Yu et al. 2009; Yuan et al. 2009). PPCPs are usually present in the environment at the order of ng L−1 and μg L−1 (Kim et al. 2007; Yu et al. 2009). The impact on both the human body and ecological environments is unclear under the low residual PPCP concentrations in surface water. Many toxicological studies have shown that fish and other aquatic organisms show reproductive disorders and behavioral changes when exposed to PPCPs (Christensen et al. 2006; Flippin et al. 2007). Ibuprofen (IBU), a common anti-inflammatory analgesic, is chosen as the target recalcitrant PPCP due to its frequent occurrence in surface water and potential long-term adverse effects. The detection of IBU in the surface water has been reported in the concentration range of 0.087 to 5 μg L−1 (Dębska et al. 2004; Kim et al. 2007). Flippin et al. (2007) found that there was a delay in the spawning time and breeding period of medaka (Oryzias latipes) that were exposed to aqueous solutions containing 0–100 μg L−1 IBU for 6 weeks.

However, current wastewater treatment plants are not specially designed to remove PPCPs (Kim et al. 2007; Vieno et al. 2007). Kim et al. (2007) investigated the removal of IBU and other drugs at a pilot scale. The results showed that the traditional water treatment technologies, such as coagulation, precipitation and filtration, cannot effectively remove these compounds. Vieno et al. (2007) reported that rapid sand filtration following coagulation and sedimentation only eliminated an additional 10% of the studied PPCPs, including ibuprofen. More than 95% of ibuprofen was found to be removed in the aeration tank, with aerobic biodegradation being the dominant mechanism (Smook et al. 2008). But the biological removal of PPCPs by secondary treatment process is unreliable, ranging from no removal to complete removal (Yu et al. 2009). Therefore, it is necessary to develop more efficient and stable water treatment methods to remove such compounds.

The OH (E0 = 1.89–2.72 V vs. NHE) is a strong oxidizing agent, which has a high oxidation-reduction potential (Vogna et al. 2004). OH based advanced oxidation technologies (AOTs), including UV/H2O2, Fenton, ozonation, and sonication, can oxidize target contaminants by attacking electron-rich sites on molecules (Crittenden et al. 2005). They can effectively degrade PPCPs under operating conditions of normal temperature and pressure (Vogna et al. 2004). Laboratory studies have demonstrated that OH based AOTs have a high ability to degrade PPCPs in deionized water, surface water, saltwater and wastewater (Grebel et al. 2010; Guo et al. 2013; Kwon et al. 2015; Yang et al. 2016). However, there are still some problems that need to be studied before large-scale application, such as the effect of target compound species and matrix components (i.e. natural organic matter (NOM), inorganic anions).

In recent years, studies have reported the degradation of IBU in UV and UV/H2O2 systems (Szabó 2010; Chianese et al. 2016; Peng et al. 2017). The IBU degradation rate is related to initial IBU concentration, initial H2O2 concentration, light intensity, effective optical path, solution pH, IBU properties (i.e. IBU species, absorption characteristics, secondary reaction rate constant with OH), and matrix components (i.e. NOM, inorganic anions) in the UV/H2O2 system (Crittenden et al. 2005). Due to the different structures and properties of IBU species (molecular and ionic forms), Chianese et al. (2016) found that ionic IBU was degraded more rapidly than the molecular form under light conditions. However, the absorption characteristics of IBU species were not determined. Kwon et al. (2015) found that the degradation of IBU was accelerated in UV/H2O2 system in neutral conditions, but the comparison of IBU degradation between the ionic form and the molecular form was not studied. Moreover, previous studies have also identified that the H2O2 dosage and matrix components (i.e. NOM, inorganic anions) have impacts on PPCPs degradation (Guo et al. 2013; Xiao et al. 2015). However, studies of the correlation impacts between IBU and H2O2 dosage and matrix components are studied.

Hence, the degradation kinetics of ionic and molecular IBU were thoroughly investigated under two different treatment methods (i.e. UV and UV/H2O2). The molar absorptivity and quantum yield of IBU species at the wavelength of 254 nm were measured. The second-order reaction rate constants of IBU species with OH were determined by using competitive kinetics. The mechanism of IBU degradation in the UV/H2O2 system was studied by establishing a pseudo first-order reaction kinetics model based on a steady-state approach. Then the effects of H2O2 dosage and matrix components (such as NOM, inorganic anions) on the photodegradation kinetics of IBU were further evaluated.

EXPERIMENTAL METHOD

Materials

Ibuprofen (IBU, 99%), 4-chlorobenzoic acid (pCBA, 99%), disodium phosphate (99%), sodium dihydrogen phosphate (99%), and humic acid (technical) were purchased from Sigma Aldrich. Sodium chloride (guaranteed reagent), sodium sulfate (guaranteed reagent), sodium nitrate (guaranteed reagent), hydrogen peroxide (30% by weight), sulfuric acid (guaranteed reagent), potassium permanganate (analytical grade), sodium sulfate (analytical grade) and sodium oxalate (analytical grade) were purchased from Sinopharm Chemical Reagent, China.

Irradiation experiments

In aqueous solution, anionic (A) and molecular (HA) species of IBU coexist, mainly determined by the solution pH at a constant temperature. The Henderson-Hasselbalch equation can be used to calculate the ratio of anions and molecules in dilute solutions (Hasselbalch 1916).  
formula
(1)
The dissociation constant (pKa) of IBU is 4.9 at the solution temperature of 20°C. As shown in Equation (1), when pH is lower than pKa, IBU is present in the molecular form. And the anionic form is more likely to be present when pH is higher than pKa. In particular, at pH = 3.00, the molecular form of IBU accounts for 98.76%, while at pH = 7.55, the anionic form of IBU accounts for 99.78%.

IBU and pCBA stock solutions were prepared in deionized water and stored at 4 °C in the dark. For the kinetic studies, the initial concentrations of IBU and pCBA in the working solutions were 10 μM (2.06 mg L−1 and 1.57 mg L−1, respectively). The solution pH was adjusted to pH = 7.55 or pH = 3.00 with 10 mM phosphate buffer solution. The selection of pH = 7.55 is based on two factors: (1) pH = 7.55 is environmentally relevant and (2) IBU is a weak acid with pKa = 4.9, thus the majority (i.e. 99.8%) of IBU is in its ion form at pH = 7.55.

The photochemical reactor is a cylindrical glass tube with an effective volume of 450 mL, with two casings. The schematic diagram of the photochemical reactor was shown in our previous study (Yang et al. 2017). The outer casing is connected to the water bath circulator (SC150–A25B, Thermo Fisher Scientific) and maintains a constant system temperature (20 ± 1 °C). A low-pressure UV lamp (GPH212T5 L/4, 10 W, Heraeus) is placed in the quartz cold trap in the center of the inner casing, and a small stirring bar was placed at the bottom of the reactor to ensure homogeneous UV exposure. The water bath circulator and the ultraviolet lamp were pre-opened to stabilize the system temperature and light intensity. One millilitre of solution was taken from the reactor at scheduled times using a 2.5 mL glass syringe (Gastight 1001, Hamilton) and chemical analysis in the solution was immediately carried out by ultra-performance liquid chromatography (Waters ACQUITY H–Class). All kinetic experiments were replicated independently at least three times.

The average light intensity per volume (I0) in the UV reactor was estimated to be 6.16 × 10−6 Einstein L−1 s−1 (which corresponds to a power output of 3.83 mW cm−2) with potassium ferrioxalate chemical actinometry (Parker 1953) and the effective optical path length (b) was 1.32 cm measured by H2O2 chemical actinometry (Beltrán et al. 1995) in this study.

Analytical methods

The emission wavelength and intensity of the UV lamp were determined using fiber optic spectrometers (USB 2000+, Ocean Optics) and the absorption spectra of IBU and pCBA in solution were measured by a UV–1800 spectrometer (Shimadzu, Japan). The pH of the solutions was measured by a S220 pH meter (Mettler Toledo). The concentration of H2O2 was measured by the KMnO4 titration method (Klassen et al. 1994).

Analysis of IBU and pCBA was performed using ultra-performance liquid chromatography (Waters ACQUITY H–Class) with a BEH C18 column (1.7 μm, 2.1 mm × 50 mm, Waters). The column temperature was set at 35 °C. An isocratic mobile phase of 50% acetonitrile and 50% phosphate buffer (20 mM, pH = 3.00) was used for the quantification of IBU and 30% acetonitrile and 70% phosphate buffer (20 mM, pH = 3.00) was used for the quantification of pCBA. The sample injection volume was 5 μL and a flow rate was 0.3 mL min−1. The UV detection wavelengths of IBU and pCBA were set at 220 nm and 238 nm, respectively.

RESULTS AND DISCUSSION

Degradation of IBU in UV and UV/H2O2 systems

The degradation of IBU in the UV/H2O2 system at pH = 7.55 fitted the first-order kinetic models with a higher R2 (0.994) than zero-order (R2 = 0.826) and second-order kinetics (R2 = 0.708), indicating that IBU degradation follows first-order kinetics. The same result was obtained under direct UVs and at pH = 3.00 in the UV/H2O2 system. Under UV intensity of 3.83 mW cm−2 and initial IBU concentration of 10 μM (2.06 mg L−1), the initial direct photolysis degradation rates of IBU were 4.25 × 10−3 μM min−1 (pH = 3.00) and 28.80 × 10−3 μM min−1 (pH = 7.55), respectively. As shown in Figure 1, IBU absorbed light with a relatively low molar absorption coefficient at the wavelengths of 200–230 nm, thus it is expected that direct photolysis of IBU by UV light at 254 nm is low.

Figure 1

Decadic molar absorption coefficient (ɛ) of IBU with reference to the UV lamp emission spectra from 200 to 400 nm.

Figure 1

Decadic molar absorption coefficient (ɛ) of IBU with reference to the UV lamp emission spectra from 200 to 400 nm.

Molar extinction coefficient and the quantum yield have a significant effect on the direct photolysis of a compound (Pereira et al. 2007; Szabó 2010). The capacity of a target compound to absorb incident light determines the kinetics of its photolysis. The decadic molar extinction coefficient (ɛ) represents the ability of the compound to absorb light at a certain wavelength (λ). ɛ can be calculated by measuring the absorbance (A) of 10 μM (2.06 mg L−1) IBU solutions at pH = 3.00 and pH = 7.55 with a 1 cm path length (l) quartz cuvette in this study:  
formula
(2)

Figure 1 illustrates the decadic molar extinction coefficient for IBU with reference to Hg lamp emission spectra. At the wavelength of 254 nm, the ɛ values of IBU were 248.41 M−1 cm−1 (pH = 3.00) and 283.64 M−1 cm−1 (pH = 7.55), which were relatively low compared to the values of other PPCPs (Yang et al. 2016). The value of at 254 nm determined by Kwon et al. (2015) was 256 M−1 cm−1 (pH = 7.00), which is between the two values we measured. The molar extinction coefficient of IBU was positively correlated with the pH value, indicating that the was dependent on the species of IBU.

The quantum yield describes 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 IBU can be calculated as follows (Pereira et al. 2007):  
formula
(3)
where is the quantum yield of IBU at 254 nm (mol Einstein−1), ruv (M s−1) is the direct photolytic degradation rate at an initial concentration of 10 μM (2.06 mg L−1). I0 is the incident UV intensity, is the molar extinction coefficient of IBU at wavelength 254 nm, and b is the reactor light path. The was calculated to be 0.015 mol Einstein−1 (pH = 3.00) and 0.091 mol Einstein−1 (pH = 7.55), respectively. Different PPCPs have different molecular structures, which can result in different quantum yield values. (Yuan et al. 2009). The is higher than the reported value of carbamazepine (0.0006 mol Einstein−1) and naproxen (0.0093 mol Einstein−1), while it is lower than that of phenytoin (0.279 mol Einstein−1) and clofibric acid (0.5390 mol Einstein−1) (Pereira et al. 2007; Yuan et al. 2009).

Due to both and at pH = 3.00 being lower than at pH = 7.55, the direct photolysis degradation rates of molecular IBU is slower than ionic IBU. As shown in Figure 2, the degradation kinetics of IBU was significantly enhanced by adding 625 μM (21.25 mg L−1) H2O2 compared to direct photolysis. In the UV/H2O2 system, the initial degradation rate of ionic IBU at pH = 3.00 (10.62 μM min−1) was also slightly faster than that of the molecular form at pH = 7.55 (8.61 μM min−1). The dark reaction experiments showed that IBU did not degrade without UV irradiation, and the same results were obtained by adding 625 μM (21.25 mg L−1) H2O2 to the dark reaction experiments. These results indicate that degradation of IBU in UV/H2O2 systems includes direct photolysis and radical degradation, but radical degradation plays a major role (more than 95%). The enhanced degradation of IBU with the addition of H2O2 is due mainly to OH radical mediated oxidation. The results are consistent with the results reported by Xiao et al. (2015), and they also demonstrated that OH plays a major role in the degradation of iodinated trihalomethanes by UV/H2O2. Therefore, the degradation rate constants of IBU largely depend on the formation of OH in the UV system.

Figure 2

Time-dependent degradation kinetics of IBU in the UV and UV/H2O2 systems ([IBU] = 10 μM, [H2O2] = 625 μM, and I0 = 3.83 mW cm−2, b = 1.32 cm). The degradation was fitted to a first-order kinetic model (shown by the lines).

Figure 2

Time-dependent degradation kinetics of IBU in the UV and UV/H2O2 systems ([IBU] = 10 μM, [H2O2] = 625 μM, and I0 = 3.83 mW cm−2, b = 1.32 cm). The degradation was fitted to a first-order kinetic model (shown by the lines).

Competitive kinetics

The second-order rate constant of IBU with OH can be determined by competitive kinetics (Packer et al. 2003; Kwon et al. 2015). pCBA was chosen as a reference compound in this study for two significant reasons. First, k values for pCBA reacting with OH are known (5.0 × 109 M−1 s−1 at neutral and 4.5 × 109 M−1 s−1 at acidic) (Buxton et al. 1988). Second, the reaction of pCBA with OH is less affected by other substances (Katsoyiannis et al. 2011). The k values of IBU reacting with OH was calculated as follows (Yuan et al. 2009):  
formula
(4)
where ktot, IBU and ktot, pCBA are the first-order rate constants for the total degradation of IBU and pCBA, kUV, IBU and kUV, pCBA are the first-order rate constant for the direct UV photolysis of IBU and pCBA, respectively. The value of ktotkUV will then be the degradation rate of IBU by the radical reaction. A straight line with the slop and the intercept of zero was plotted with ktot, IBUkUV, IBU against ktot, pCBAkUV, pCBA. As shown in Figure 3, an average reaction rate constant ratio of IBU and pCBA with OH was 0.686 (unitless) and 1.047 (unitless) at pH = 3.00 and pH = 7.55, respectively. The values of molecular and ionic IBU with OH were determined to be 3.43(±0.09) × 109 M−1 s−1 (pH = 3.00) and 5.93(±0.15) × 109 M−1 s−1 (pH = 7.55), respectively. Packer et al. (2003) have determined the value of to be 6.50 × 109 M−1 s−1 (pH = 3.00) in the Fenton system, which is different from our measurement (3.43 × 109 M−1 s−1). Some of the difference stem from two factors: (i) Packer et al. (2003) determined the value of in the Fenton system, whereas we measured it in the UV/H2O2 system, (ii) acetophenone was used as a reference substance in the study of Packer et al. (2003), whereas pCBA was used in this study. However, under the same conditions of the UV/H2O2 system and reference substance (pCBA), the (5.93 × 109 M−1 s−1 at pH = 7.55) is consistent with the result (5.57 × 109 M−1 s−1 at pH = 7.00) reported by Kwon et al. (2015).
Figure 3

Competitive kinetics of IBU and pCBA in the UV/H2O2 system ([IBU] = 10 μM, [pCBA] = 10 μM, [H2O2] = 100 μM, and I0 = 3.83 mW cm−2, b = 1.32 cm, pH = 7.55).

Figure 3

Competitive kinetics of IBU and pCBA in the UV/H2O2 system ([IBU] = 10 μM, [pCBA] = 10 μM, [H2O2] = 100 μM, and I0 = 3.83 mW cm−2, b = 1.32 cm, pH = 7.55).

Pseudo first-order reaction kinetics

The degradation of target compounds in UV-based processes can be predicted and explained by the method of the steady-state approximation for the kinetic description of radicals (Yuan et al. 2009). The method is established based on the assumption that the radicals (i.e. OH) produced by UV photolysis of H2O2 play a major role on the degradation of target compounds (Yuan et al. 2009). The reactions in the UV/H2O2 system and their rate constants are presented in Table 1.

Table 1

Summary of the reactions in the UV/H2O2 system (10 mM phosphate buffer)

Reaction k (M−1 s−1Reference or note 
 , s−1 Crittenden et al. (1999)  
 k1 = 2.7 × 107 Buxton et al. (1988)  
 k2 = 7.5 × 109 Crittenden et al. (1999)  
 k3 = 2.51 × 10−12 Crittenden et al. (1999)  
 pKa1 = 2.17 unitless Stumm & Morgan (1996)  
 pKa2 = 7.21 unitless Stumm & Morgan (1996)  
 pKa3 = 12.35 unitless Stumm & Morgan (1996)  
 kH1 = 2.0 × 104 Buxton et al. (1988)  
 kH2 = 1.5 × 105 Buxton et al. (1988)  
10  kH3 < 1.5 × 107 Buxton et al. (1988)  
11  kH4 = 2.7 × 106 Buxton et al. (1988)  
 In the presence of NOM   
12   = 1.40 × 104 L mgC−1 s−1 Lutze et al. (2015)  
 Degradation of IBU   
13  rUV, M s−1 Measured in this study 
14   Measured in this study 
Reaction k (M−1 s−1Reference or note 
 , s−1 Crittenden et al. (1999)  
 k1 = 2.7 × 107 Buxton et al. (1988)  
 k2 = 7.5 × 109 Crittenden et al. (1999)  
 k3 = 2.51 × 10−12 Crittenden et al. (1999)  
 pKa1 = 2.17 unitless Stumm & Morgan (1996)  
 pKa2 = 7.21 unitless Stumm & Morgan (1996)  
 pKa3 = 12.35 unitless Stumm & Morgan (1996)  
 kH1 = 2.0 × 104 Buxton et al. (1988)  
 kH2 = 1.5 × 105 Buxton et al. (1988)  
10  kH3 < 1.5 × 107 Buxton et al. (1988)  
11  kH4 = 2.7 × 106 Buxton et al. (1988)  
 In the presence of NOM   
12   = 1.40 × 104 L mgC−1 s−1 Lutze et al. (2015)  
 Degradation of IBU   
13  rUV, M s−1 Measured in this study 
14   Measured in this study 
Under steady-state conditions, the reaction rate (rtot, M s−1) of IBU degradation in the UV/H2O2 process can be expressed with the following equation (Crittenden et al. 1999):  
formula
(5)
where rUV is the initial IBU direct photolytic degradation rate in the UV/H2O2 process and is the IBU reaction rate with OH.  
formula
(6)
 
formula
(7)
 
formula
(8)
where I0 is the incident UV intensity (3.83 mW cm−2), and b is the reactor light path (1.32 cm). A is the absorbance of the reaction solution. is the molar extinction coefficient of H2O2 (19.6 M−1 cm−1), and (in Equation (11) below) is the quantum yield of H2O2 (0.5 mol Einstein−1) at wavelength 254 nm (Crittenden et al. 1999).
The degradation of IBU in UV/H2O2 systems follows pseudo first-order kinetics (the unit for pseudo first-order is s−1), and can be described by an equation as follows:  
formula
(9)
Under steady-state conditions, the formation rate of OH is equal to the rate of consumption. Thus, the steady-state concentrations for OH (i.e. ) can be calculated as follows:  
formula
(10)
In the UV/H2O2 system, can be calculated as follows (Crittenden et al. 1999):  
formula
(11)
 
formula
(12)

The averages of and were 2.24 × 10−7 M s−1 and 1.42 × 10−2 s−1 at pH = 3.00, then 2.24 × 10−7 M s−1 and 1.77 × 10−2 s−1 at pH = 7.55, respectively.

From the above formula, we calculated the secondary reaction rate constant of IBU with OH and the steady-state concentration of OH with Equations (13) and (14):  
formula
(13)
 
formula
(14)

The of molecular and ionic IBU with OH were calculated to be 3.47(±0.11) × 109 M−1 s−1 (pH = 3.00) and 5.89(±0.19) × 109 M−1 s−1 (pH = 7.55), which were consistent with the values determined by competitive kinetics. The average of were calculated to be 4.06 × 10−12 M and 2.93 × 10−12 M at pH = 3.00 and pH = 7.55, respectively. Kwon et al. (2015) has determined the to be 0.27 × 10−12 M in the UV/H2O2 system ([H2O2]0 = 0.5 mM, I0 = 0.5 mW cm−2, b = 0.79 cm, [IBU]0 = 10 μM, pH = 7.00), which was lower than the value we have measured (4.06 × 10−12 M at pH = 3.00 and 2.93 × 10−12 M at pH = 7.55). This is mainly due to both I0 and b in the research of Kwon et al. (2015) are lower than that in this study (I0 = 3.83 mW cm−2, b = 1.32 cm).

Here we used the mean value of , , and Equations (5)–(7) to calculate the contribution of direct photolysis and OH in the UV/H2O2 system.  
formula
(15)
In the UV/H2O2 system, the rUV and were determined to be 6.95 × 10−10 M s−1 and 1.41 × 10−7 M s−1 at pH = 3.00, respectively. Therefore,  
formula
(16)
Meanwhile, the rUV and were determined to be 4.72 × 10−9 M s−1 and 1.72 × 10−7 M s−1 at pH = 7.55, respectively. Therefore,  
formula
(17)
The calculated contributions of direct photolysis and OH to IBU degradation are shown in Equations (16) and (17). OH is the main reactive species in the UV/H2O2 system, which is responsible for more than 97% of IBU degradation at both pH = 3 and pH = 7.55 (i.e. the contribution of direct photolysis (rUV) was less than 3%). The results support the previous conclusion that OH play a major role in IBU degradation in the UV/H2O2 system.

Thus, under the same initial conditions, the secondary reaction rate constants of molecular and ionic IBU with OH and the steady-state concentration of OH determine the IBU degradation rate in the UV/H2O2 system. These also explain why the IBU degradation rate at pH = 7.55 is faster than at pH = 3.00 in UV/H2O2 system.

The pseudo first-order reaction kinetics model can also be used to study and simulate the effects of other factors on IBU degradation (Xiao et al. 2015). In this study, we used the pseudo first-order reaction kinetics model to predict and explain the effects of H2O2 dosage and NOM on IBU degradation in the UV/H2O2 system. Then the contributions of direct photolysis and OH to IBU degradation (i.e. and ) under various experimental conditions were calculated according to Equations (18) and (19), respectively.  
formula
(18)
 
formula
(19)
The total contribution of direct photolysis and OH to IBU degradation can be expressed by Equation (20).  
formula
(20)

Effect of the initial H2O2 dosage

The concentration of H2O2 in the UV/H2O2 system affected the degradation rates of IBU. As shown in Figure 4, the experimental results of the pseudo first-order rate constants for IBU degradation increased from 2.25 × 10−3 s−1 to 10.24 × 10−3 s−1 when the H2O2 concentration increased from 45 to 375 μM (1.53 mg L−1–12.75 mg L−1) at pH = 7.55. This result could be well predicted by the kinetic model. In Figure 4, the was consistent with the modeling results . The calculated contributions of direct photolysis and OH to IBU degradation are also shown in Figure 4. The OH was the main reactive species in the UV/H2O2 system, whose contribution to IBU degradation was always greater than 97% with the increase of H2O2 concentration from 45 μM to 375 μM (from 1.53 mg L−1 to 12.75 mg L−1) (i.e. the contribution of direct photolysis was less than 3%). The calculated contributions of direct photolysis and OH were 2.67% and 97.33% at the initial concentration of H2O2 (625 μM) at pH = 7.55, which were consistent with experimental data (Equation (17)). Guo et al. (2013) found that the apparent rate constant (kapp) of ciprofloxacin (CIP) increased from 0.39 × 10−3 s−1 to 3.72 × 10−3 s−1 when H2O2 concentrations were in the range of 0–5 mM. The modeling result also showed that a negative effect would be found as the concentration of H2O2 further increased, but only when the concentration of H2O2 was above 10 mM, when the decreasing trend of occurred (the inset in Figure 4). The possible explanation is that OH could be scavenged by the excess H2O2 with the second-order rate constant of 1.4 × 109 M−1 s−1 (Buxton et al. 1988).

Figure 4

Impacts of H2O2 dosages on the pseudo first-order constants of IBU (kobs) ([IBU]0 = 10 μM, pH = 7.55, and I0 = 3.33 mW cm−2, b = 0.93 cm).

Figure 4

Impacts of H2O2 dosages on the pseudo first-order constants of IBU (kobs) ([IBU]0 = 10 μM, pH = 7.55, and I0 = 3.33 mW cm−2, b = 0.93 cm).

Effect of organic matter

Since humic acid is the main component of organic matter, the effected of organic matter on the degradation of IBU in the UV/H2O2 system at pH = 7.55 was studied by adding different concentrations of humic acid (0–2.87 mgC L−1). Figure 5 shows that decreased from 3.55 × 10−3 s−1 to 1.79 × 10−3 s−1 with humic acid concentration increasing from 0 mgC L−1 to 2.87 mgC L−1. Moreover, of the model calculation slowed down slightly faster than that of the experimental data with the increasing humic acid concentration. These indicated that humic acid not only acted as a radical scavenger and the inner filter in the UV/H2O2 system, but also reacted with hydroxyl radicals to form secondary radicals that can continue to react with IBU. Then, the kinetic model was used to estimate the relative contributions of the inner filter effect (Equation (8) can be modified to A = b(ɛIBU[IBU] + [H2O2] + ɛNOM [humic acid]), the ɛNOM of humic acid being 0.10 L mgC−1 cm−1 as measured in this work) and radical scavenger to the decrease of kobs. As shown in Figure 5, kcal,obs changed slightly as compared to the experimental result (magenta line), if the inner filter effect of humic acid was ignored (i.e. assuming that ɛ of humic acid was zero). While the calculated values of kcal,obs greatly deviated from the experimental result (green line), if the radical scavenging effect of humic acid was left out (i.e. assuming that the second-order rate constant of reactions between humic acid and OH ( in Table 1) in Equation (12) were zero). These contrasting results indicate that the radical scavenger effect of humic acids has a greater significant role than its inner filter effect in decreasing IBU degradation rate in the UV/H2O2 process. Even in the presence of humic acid, the calculated contributions of OH was still clearly higher than the direct photolysis (Figure 5). This phenomenon suggests that OH is still the most important candidate regarding IBU degradation.

Figure 5

Impacts of humic acid dosages on the pseudo first-order constants of IBU (kobs) ([IBU]0 = 10 μM, [H2O2]0 = 100 μM, pH = 7.55, and I0 = 3.33 mW cm−2, b = 0.93 cm). Please refer to the online version of this paper to see this figure in color: http://dx.doi.org/10.2166/wst.2018.129.

Figure 5

Impacts of humic acid dosages on the pseudo first-order constants of IBU (kobs) ([IBU]0 = 10 μM, [H2O2]0 = 100 μM, pH = 7.55, and I0 = 3.33 mW cm−2, b = 0.93 cm). Please refer to the online version of this paper to see this figure in color: http://dx.doi.org/10.2166/wst.2018.129.

Effect of inorganic anions

In aquatic environments, Cl, , and are common inorganic anions. Therefore, the effects of Cl, , and on IBU degradation in the UV/H2O2 system at pH = 7.55 were studied at the concentration range of 0–5 mM. As shown in Figure 6, the degradation rate of IBU was slightly increased in the presence of . Xiao et al. (2015) found that can promote the degradation of CHCl2I in UV/H2O2 system. The redox potentials of (2.3–2.6 V) is almost similar to that of OH/H2O (2.39 V) (Xiao et al. 2015) in neutral conditions. Thus, reacts slowly with OH and its scavenging effect is limited. Nitrate also absorbs UV photons, and then OH can be generated at a low quantum yield during the photolysis of nitrate as shown in Equations (21)–(25) (Keen et al. 2012).  
formula
(21)
 
formula
(22)
 
formula
(23)
 
formula
(24)
 
formula
(25)
The degradation of IBU was not affected by adding and Cl in the UV/H2O2 system. The results are consistent with the previous studies of Xiao et al. (2015) and Guo et al. (2013). In the presence of and Cl with concentration ranging from 1 to 5 mM, Xiao et al. (2015) found that the degradation rates of CHCl2I remained constant. Guo et al. (2013) reported that had no significant influence on CIP degradation. The redox potentials of (2.43 V) and Cl/Cl (2.41 V) (Huie et al. 1991) are almost similar to that of OH/H2O (2.39 V) (Xiao et al. 2015) in neutral conditions. These indicate that the scavenging effect of can be ignored and the scavenging effect of Cl is limited. However, previous studies reported that the fast reaction of OH with Cl yielded secondary reactive chlorine radical species such as Cl, , and through complex chain reactions and can convert to OH and Cl at pH higher than 7.2 (Yang et al. 2014). This suggests that the effect of those secondary radicals with compounds could not be ignored in the degradation of IBU in the UV/H2O2 process in the presence of Cl. For instance, Cl is a highly selective oxidant (Grebel et al. 2010). and Cl have high oxidation capacities, with oxidation potentials of 2.0 V and 2.47 V, respectively (Beitz et al. 1998). Thus, the scavenging effect of Cl is offset by the release of OH from the transient intermediate under basic conditions and the reaction of secondary radicals with compounds (Deng et al. 2013). These explain the observation that the photodegradation rates of IBU in the UV/H2O2 system at pH = 7.55 were not affected by the presence of Cl.
Figure 6

Impacts of inorganic anions on the pseudo first-order constants of IBU (kobs) ([IBU]0 = 10 μM, [H2O2]0 = 100 μM, pH = 7.55, and I0 = 3.33 mW cm−2, b = 0.93 cm).

Figure 6

Impacts of inorganic anions on the pseudo first-order constants of IBU (kobs) ([IBU]0 = 10 μM, [H2O2]0 = 100 μM, pH = 7.55, and I0 = 3.33 mW cm−2, b = 0.93 cm).

As reported in a previous study, is the predominant carbonate species in neutral conditions (Kwon et al. 2015). As shown in Figure 6, the presence of has an inhibitory effect on IBU degradation. The kobs of IBU decreased significantly when was present. When the initial concentration of was increased to 5 mM, the kobs of IBU was further decreased by 25%. The inhibitory effect of on other PPCPs degradation in the UV/H2O2 system has also been reported in other literature (Grebel et al. 2010; Yang et al. 2014; Kwon et al. 2015). This could be mainly attributed to the scavenging effects of ( = 8.5 × 106 M−1 s−1) (Buxton et al. 1988; Grebel et al. 2010).

In the UV/H2O2 process, although the PPCPs can easily be degraded in most cases, the complete mineralization of PPCPs may be an uneconomical goal for treatment (Yuan et al. 2009). The cost limits the wide application of the process in practical engineering at full scale (Yu et al. 2009). Hence, two aspects should be further addressed in future studies, including the full-scale operation cost and the toxicity of byproducts.

CONCLUSIONS

The degradation rate of IBU was significantly increased after the addition of H2O2 in the UV reactor, indicating the major primary role of OH. Due to the higher molar absorption coefficient and quantum yield, direct UV photolysis of ionic IBU is faster than of the molecular form. In the UV/H2O2 system, the degradation rate of ionic IBU is also slightly faster than that of molecular IBU, and the main determinants are and . The effect of matrix components on IBU degradation in the UV/H2O2 system was predicted and simulated by developing a steady-state approximation and kinetic parameters method. The k values and effect of matrix components are helpful for predicting and explaining IBU degradation mechanisms and selecting better performing treatment processes for the removal of PPCPs in wastewater.

ACKNOWLEDGEMENTS

Funding from the National Nature Science Foundation of China (No. 51674305), Hunan Graduate Student Innovation program (No. CX2017B465), and the Innovation-Driven Project (2017CX010) of Central South University are gratefully acknowledged.

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