Abstract

The monoterpene alcohol α-terpineol is extensively used as the foaming agent in mineral processing and can be released to environment along with the wastewater. This study evaluated the feasibility of eliminating α-terpineol in water by ultraviolet irradiation (UV) in combination with hydrogen peroxide (H2O2). Within an H2O2 dose of 10 mg/L and an UV fluence of 64.8 J/cm2, more than 95% of the α-terpineol can be removed. The reactions fitted well to pseudo-first-order kinetics, and the apparent rate constant was 0.0678 min−1. The effects of matrix species including various anions and humic acid (HA), were evaluated. The degradation rate decreased significantly with the addition of bicarbonate and HA. Further verification was carried out with three types of real water samples. In the ground water and the surface water, the degradation rate decreased likely due to the presence of natural organic matter. Finally, possible degradation pathways were proposed based on the identification of transformation products, and the occurrence of two main transformation products were monitored. This study demonstrated that the UV/H2O2 is an effective technology for the degradation of α-terpineol in water.

INTRODUCTION

The monoterpene alcohol α-terpineol (see Figure 1) is widely used as a foaming agent in mineral processing due to its high surface activity and bubble size distribution (Brossard et al. 2005; Gupta et al. 2007; Vasseghian et al. 2014). It is also used as a fragrance in personal care products (Simonich et al. 2002). When released into the environment, it can lower water quality by depressing gas transfer at gas-liquid interfaces and increasing the solubility of other contaminants (Rosso et al. 2006). Moreover, it is considered as a precursor of disinfection by-products when it is discharged to drinking water sources. In August 2014, an illegal discharge of wastewater from the mineral processing plant to a drinking water source in Wushan, China, threatened the safety of drinking water for more than 50,000 people. In this accident, α-terpineol was found to be one of the main pollutants.

Figure 1

Molecular structure of α-terpineol.

Figure 1

Molecular structure of α-terpineol.

Previous research into the decomposition of α-terpineol has mainly focused on the gas phase or solid-gas interface as it is considered to be an indoor air pollutant (Wells 2005; Ham & Wells 2011; Shu & Morrison 2011). However, because α-terpineol can increase the stability of other pollutants in water as it is a surfactant, and because it can cause taste and odor complaints from customers if spilled into the water supply, studies on its degradation in water are also important. It is known that α-terpineol is stable in aquatic environments because of the ring structure and branched chains (Hou 2013), but it can be decomposed by soil bacteria (Tadasa 1977). Moreover, Hinrichs' group investigated multiphase ozonolysis of aqueous α-terpineol and found that the reaction products and kinetics differ from the gas-phase reaction (Leviss et al. 2016). In this paper, another advanced oxidation process (AOP) technique, UV/H2O2, is introduced for the treatment of α-terpineol in water.

UV/H2O2 is widely used in both wastewater and drinking water treatments due to its high treatment efficiency, non-selectivity of the pollutants and resulting low concentration of disinfection byproducts (Canonica et al. 2008; Souissi et al. 2013; Tan et al. 2013; Aydin 2015; Carlson et al. 2015; Godayol et al. 2015). In the photodegradation process, H2O2 absorbs UV light and then generates highly reactive hydroxyl radicals (•OH), so the aqueous pollutants are decomposed by both UV direct photolysis and •OH-assisted indirect photolysis (Bolton & Cater 1994; Rosenfeldt et al. 2006).

The aims of this study were to investigate the removal efficiency of α-terpineol by UV/H2O2 treatment and to elucidate its transformation during the process. The effects of H2O2 concentration, pH, and matrix species on the degradation of α-terpineol were evaluated, and this treatment technology was also further examined and verified in different types of real water samples. Subsequently, transformation products in this degradation process were identified using gas chromatography-mass spectrometry (GC-MS), based on which two degradation pathways were proposed. Moreover, the changing trends of the two main transformation products were monitored.

MATERIALS AND METHODS

Materials and chemicals

The α-terpineol (90% technical grade) was purchased from Sigma-Aldrich (St. Louis, MO, USA). H2O2 (30% w/w), sodium chloride (99.5%), sodium sulfate (99.0%), sodium nitrate (99.0%), sodium bicarbonate (99.8%) and humic acid (90%) was purchased from Sinopharm Chemical Reagent Co., Ltd (Beijing, China). BSTFA (N,O-bis-trimethylsilyl-trifluoroacetamide) containing 1% TMCS (trimethylchlorosilane) was purchased from Sigma-Aldrich (St. Louis, MO, USA). PFBHA (O-(2,3,4,5,6-pentafluorobenzyl) hydroxylamine hydrochloride) and 4-acetylbutyric acid were purchased from TCI (Tokyo, Japan). Ethyl acetate and chloroform were purchased from Sigma-Aldrich (St. Louis, MO, USA) at GC-MS grade purity. Two standards of the transformation products, p-menthane-1,8-diol and 4-methylcyclohex-3-en-1-one were purchased from TCI (Tokyo, Japan) and Toronto Research Chemicals (Toronto, Canada), separately.

Photochemical reacting apparatus (BL-GHX-V, Shanghai) was used as the UV reactor. Ultrapure (UP) water (18.2 MΩ· cm, dissolved oxygen concentration (DOC) <0.1 mg/L) was obtained from a Persee (Beijing) pure water system. Prior to use, all glassware was cleaned with a laboratory detergent, rinsed with methanol and UP water, and baked in muffle furnace at 450 °C for 3 hours.

Photodegradation experiments

Photodegradation was conducted in a cylindrical glass reactor at room temperature. An 8 W low-pressure mercury lamp (Phillips) with the emission peak at 254 nm was placed coaxially with the reactor in a quartz sleeve, and the distance from the lamp to the reactor wall was 7.5 cm. The light intensity was determined as 18 mW/cm2 at the lamp sleeve using a radiometer (Lutron YK-35UV). The molar absorption coefficients of α-terpineol and H2O2 at 254 nm were 43.8 L· mol−1· cm−1 and 18.2 L· mol−1·cm−1 respectively, measured by a UV/Vis spectrophotometer (TU-1901, Persee, China). The reactor was agitated by a magnetic stirrer operated at 300 rpm. The lamp was turned on 5 min before the experiment to ensure that the lamp was stable and had reached the full spectral output.

The initial concentration of α-terpineol was 20 mg/L and the initial concentrations of H2O2 were set at 5, 10, and 20 mg/L. Initial solution pH was adjusted with HCl or NaOH for the range of 2–12. Samples were withdrawn at preselected intervals, and each 1 mL of solution was extracted with 0.5 mL ethyl acetate after sampling. The extraction mixture was vortexed for about 30 seconds to ensure adequate contact between the water and the organic layer. The ethyl acetate layer was then analyzed with GC-MS. All experiments were conducted in triplicate.

Derivatization

Oximation and silylation were applied before GC-MS analysis to identify transformation products generated in the degradation procedure with carbonyl and hydroxyl/carboxyl groups, respectively. Three modes of derivatization strategies were used: oximation only, silylation only, and combined oximation and silylation. For oximation, 100 μL of PFBHA (0.05 mg/L) was added to the aqueous solution. Then the vials were left for 60 min to allow for the completion of the reaction. Silylation was performed after extraction with ethyl acetate. The organic layer was evaporated to dryness under a gentle stream of nitrogen. The residue was then treated with 100 μL of BSTFA + 1% TMCS and allowed to react at 65 °C for 90 min.

GC-MS analysis

GC-MS analysis was performed on an Agilent 7890A-5975C GC-MS (USA). A HP-5MS fused-silica capillary column (5% diphenyl/95% dimethyl-polysiloxane, 30 m × 0.25 mm × 0.25 μm, Agilent, USA) was used for separation. The GC temperature program was: initial temperature of 50 °C held for 2 min, followed by an increase of 20 °C/min to 105 °C, then increased by 5 °C/min to 115 °C, and finally increased by 20 °C/min to 240 °C and held for 0.5 min. The carrier gas was helium, with a constant flow of 1.0 mL/min. Splitless mode was used and 1 μL was injected. Electron impact with the electron energy of 70 eV was utilized. The m/z range was 50–500. The ion source temperature of the mass spectrometer was 230 °C. Quadrupole temperature and transfer line temperature were 150 °C and 280 °C, respectively.

RESULTS AND DISCUSSION

Effect of H2O2 concentration

Figure 2 shows the removal of α-terpineol versus time in the different experimental conditions tested. In the dark and with 10 mg/L of H2O2, no significant loss of α-terpineol was observed after 60 min, indicating the stability and persistence of the surfactant. The removal of the contaminant was significantly affected by UV irradiation and UV/H2O2. When irradiated with UV only, the percentage removal was increased (approximately 50% in 60 min), and further acceleration was achieved by the addition of H2O2. The degradation of α-terpineol reached 74.2% within 60 min when 5 mg/L H2O2 was added. As the concentration of H2O2 increased to 10 mg/L and 20 mg/L, the degradation of α-terpineol reached 95.2% and 98.1% within 60 min, respectively. The results verified the importance of H2O2 in UV/H2O2 treatment for the generation of •OH radicals to enhance the degradation of α-terpineol. In this system, H2O2 in solution acted as the electron acceptor, absorbed photons under UV irradiation, and produced •OH radicals as a result of O-O bond ruptures. Free •OH radicals could subsequently react with α-terpineol. With the increasing H2O2 concentration, the amount of •OH radicals increased correspondingly (Keen et al. 2014), so the removal efficiencies of α-terpineol were increased.

Figure 2

Removal of α-terpineol in darkness, H2O2 alone, UV irradiation alone and combined UV/ H2O2, respectively (Conditions: [α-terpineol]0 = 20 mg/L, UV light intensity = 18 mW/cm2, pH = 7).

Figure 2

Removal of α-terpineol in darkness, H2O2 alone, UV irradiation alone and combined UV/ H2O2, respectively (Conditions: [α-terpineol]0 = 20 mg/L, UV light intensity = 18 mW/cm2, pH = 7).

The degradation reactions fitted well to the pseudo-first-order kinetics (see Table 1 for the coefficient of determination R2 > 0.99) under UV irradiation without or with H2O2. The apparent rate constant Kobs increased from 0.0231 min−1 to 0.0678 min−1, corresponding to the increase in H2O2 concentration from 5 mg/L to 20 mg/L. This was in agreement with the removal efficiencies observed. In the following experiments, the concentration of H2O2 was set to 10 mg/L due to the satisfactory degradation efficiency.

Table 1

Kinetic parameters of degradation of α-terpineol under UV irradiation and UV/H2O2 (Conditions: [α-terpineol]0 = 20 mg/L, UV light intensity = 18 mW/cm2, pH = 7)

ConditionKinetic equationReaction rate constant Kobs/(min−1)R2
UV ln(C0/C) = 0.012t − 0.0458 0.012 0.990 
UV + 5 mg/L H2O2 ln(C0/C) = 0.023t − 0.0552 0.023 0.995 
UV + 10 mg/L H2O2 ln(C0/C) = 0.050t − 0.0705 0.050 0.997 
UV + 20 mg/L H2O2 ln(C0/C) = 0.068t + 0.0469 0.068 0.993 
ConditionKinetic equationReaction rate constant Kobs/(min−1)R2
UV ln(C0/C) = 0.012t − 0.0458 0.012 0.990 
UV + 5 mg/L H2O2 ln(C0/C) = 0.023t − 0.0552 0.023 0.995 
UV + 10 mg/L H2O2 ln(C0/C) = 0.050t − 0.0705 0.050 0.997 
UV + 20 mg/L H2O2 ln(C0/C) = 0.068t + 0.0469 0.068 0.993 

Effect of pH

In the photodegradation process, pH is a significant parameter because it can alter the charge state of the pollutants and influence the formation of active hydroxyl radicals (Chong et al. 2010; Lin et al. 2014; Kuo et al. 2015). In this study, as α-terpineol is a very weak acid (pKa = 15.09), it could remain in its undissociated form in the pH range investigated (2, 4, 7, 10, 12). Thus the change of the reaction rate at different pH values can be attributed to the generation of hydroxyl radicals. As Table 2 shows, the degradation reactions fitted well to pseudo-first-order kinetics under acidic, circumneutral and alkaline conditions. The degradation rate decreased slowly when the initial pH increased from 2 to 10, which can be attributed to a higher redox potential of •OH/H2O in acidic conditions (Xiao et al. 2014). Moreover, the degradation rate decreased quite clearly when the pH increased from 10 to 12, probably due to (1) self-decay of H2O2 being enhanced under high pH conditions; (2) the fraction of hydroperoxide anion (HO2) becoming dominant with the increasing pH and acting as a strong •OH scavenger (Tan et al. 2013). So if the pH of the wastewater was higher than 10, acidification would be necessary for a satisfactory degradation efficiency. However, this condition is not common.

Table 2

Kinetic parameters of degradation of α-terpineol of different pH under UV/H2O2 (Conditions: [α-terpineol]0 = 20 mg/L, [H2O2] = 10 mg/L, UV light intensity = 18 mW/cm2)

pHKinetic equationReaction rate constant Kobs/(min−1)R2
ln(C0/C) = 0.0579t − 0.0683 0.058 0.994 
ln(C0/C) = 0.0534t − 0.0571 0.053 0.992 
ln(C0/C) = 0.0497t − 0.0705 0.050 0.997 
10 ln(C0/C) = 0.0458t − 0.0172 0.046 0.997 
12 ln(C0/C) = 0.0322t − 0.0659 0.032 0.995 
pHKinetic equationReaction rate constant Kobs/(min−1)R2
ln(C0/C) = 0.0579t − 0.0683 0.058 0.994 
ln(C0/C) = 0.0534t − 0.0571 0.053 0.992 
ln(C0/C) = 0.0497t − 0.0705 0.050 0.997 
10 ln(C0/C) = 0.0458t − 0.0172 0.046 0.997 
12 ln(C0/C) = 0.0322t − 0.0659 0.032 0.995 

Effect of matrix species

Further investigation was conducted to evaluate the effects of matrix species including chloride, nitrate, sulfate, bicarbonate and humic acid (HA) as they are commonly found in natural water systems as well as wastewater.

As Figure 3 shows, when sulfate or chloride was spiked into the reaction solution, the degradation kinetics remained the same as that obtained in the UP water solution. In contrast, the degradation rate decreased with the addition of 100 mg/L . It can be rationalized that the UV absorption by nitrate was slightly higher than that of other anions at 254 nm (Xiao et al. 2014). Nitrite was also present in the solution (100 mg/L), and it is a powerful scavenger of hydroxyl radicals (Ward et al. 1996; Mack & Bolton 1999; Sharpless & Linden 2001). Although bicarbonate did not absorb UV light, a significant inhibitory effect was observed. It is known that bicarbonate is a radical scavenger in the AOP process, and can quench •OH at a rate of 8.6 × 106 M−1s−1 (Buxton et al. 1988; Liu et al. 2012), so this result also demonstrated the important role of •OH in UV/H2O2 treatment degradation of α-terpineol.

Figure 3

Effect of matrix species including Cl, , , , and HA on the photodegradation kinetics in UP water (Conditions: [α-terpineol]0 = 20 mg/L, [H2O2] = 10 mg/L, UV light intensity = 18 mW/cm2, pH = 7).

Figure 3

Effect of matrix species including Cl, , , , and HA on the photodegradation kinetics in UP water (Conditions: [α-terpineol]0 = 20 mg/L, [H2O2] = 10 mg/L, UV light intensity = 18 mW/cm2, pH = 7).

HA is often selected as a matrix species to investigate the effect of natural organic matter (NOM) on the degradation of contaminants in water (Prados-Joya et al. 2011). As shown in Figure 3, HA had an obvious negative effect on the degradation rate of α-terpineol. On the one hand, HA consists of a great number of aromatic rings and oxygen related functional groups leading to a competitive absorbance of photons. On the other hand, HA can act as a strong scavenger of •OH (Crittenden et al. 1999). These two effects resulted in a significant inhibition of the degradation of α-terpineol in the presence of HA.

Photodegradation of α-terpineol in different types of water

To assess the degradation effectiveness of α-terpineol by UV/H2O2, experiments were carried out in real water samples, including groundwater (collected from Yuelu Mountain, Changsha), surface water (collected from Xiangjiang River, Changsha), and the effluent of a mineral processing plant (collected from Chenzhou City, Changsha). Here the nitrate concentration was detected by an ion chromatograph with electrical conductivity detector, the total alkalinity was analyzed by titration with hydrochloric acid, and the total organic carbon (TOC) was detected by a TOC analyzer (Shimadzu, TOC-VCSH). It should be noted that the TOC values of groundwater and surface water were measured before 20 mg/L α-terpineol was spiked, while the concentration of α-terpineol in the effluent was 23 mg/L, contributing 73% of the TOC. No more α-terpineol was added to the effluent.

As shown in Table 3, the pH values of the four water samples varied in a small range and the concentrations of nitrate and bicarbonate were all less than 20 mg/L, so according to the preceding discussion, the reduction of degradation rate in groundwater and surface water was mainly attributed to the NOM as they had higher TOC values than that of UP water. Interestingly, the TOC of the effluent was six times higher than that of the surface water, but the degradation rates in these two water samples were quite close. A possible explanation was that the TOC in the effluent was mainly contributed by the foaming agent (α-terpineol itself) and collecting agents (used to change the hydrophobicity of the minerals). In China, xanthate is the most widely used collecting agent, which decomposes very quickly under UV irradiation (Fu et al. 2016), so the high TOC value in the effluent did not mean high content of NOM.

Table 3

Effect of different types of water on degradation kinetics of α-terpineol (Conditions: [H2O2] = 10 mg/L, UV light intensity = 18 mW/cm2, the spiked concentration of α-terpineol was 20 mg/L except for the effluent)

UP waterGroundwaterSurface waterEffluent
pH 6.2 7.5 6.6 8.7 
Nitrate (mg·L−1<0.05 2.20 3.52 7.41 
Total alkalinity (as CaCO3, mg·L−1<5 20 
TOC (mg·L−10.084 1.87 4.02 24.6 
Kobs (min−10.050 0.046 0.035 0.036 
UP waterGroundwaterSurface waterEffluent
pH 6.2 7.5 6.6 8.7 
Nitrate (mg·L−1<0.05 2.20 3.52 7.41 
Total alkalinity (as CaCO3, mg·L−1<5 20 
TOC (mg·L−10.084 1.87 4.02 24.6 
Kobs (min−10.050 0.046 0.035 0.036 

Identification of transformation products

In order to elucidate the possible degradation pathways of α-terpineol in UV/H2O2 treatment, the transformation products were identified by direct GC-MS and GC-MS after derivatization. The mass spectra were checked in the database from NIST MS Search 2.0 and compared with the GC-MS results of the degradation products of α-terpineol in the related references (Wells 2005; Ham & Wells 2008; Ham & Wells 2011). Ten transformation products (marked P1 to P10 in Figure 4 and Table 4) were found after UV/H2O2 degradation, among which P1 and P2 were much more abundant than the others identified. The mass spectra of their corresponding derivatives are presented in the Supplementary Material (see Figures S1–S10, available with the online version of this paper). It is important to note that the hydroxylated product (P1), which was confirmed by silylation experiments, has never been reported in the study of the degradation of gaseous α-terpineol (Wells 2005; Ham & Wells 2008; Ham & Wells 2011), indicating an additional step in the degradation pathway of aqueous α-terpineol with an AOP system.

Table 4

Molecular structure and confirmation methods of α-terpineol photodegradation transformation products

Product nameMolecular structureCharacteristic ions (m/z)Identification method
p-Methane-1,8-diol (P1)  73, 81, 131, 211 Silylation 
4-Methyl-3-cyclohexen-1-one (P2)  79, 107, 181, 288, 305 Oximation 
1,3-Dihydroxyacetone (P3)  117, 161, 181, 195 Oximation 
4-Methyl-2-oxopentanoic acid (P4)  73, 181, 200, 355, 382 Combined oximation and silylation 
3-Methyl-2-oxobutanoic acid (P5)  73, 181, 366, 383 Combined oximation and silylation 
2,3-Butanedione (P6)  181, 279, 476 Oximation 
Isobutyraldehyde (P7)  56, 117, 181, 195, 250 Oximation 
Butyraldehyde (P8)  161, 181, 195, 239 Oximation 
Propionaldehyde (P9)  161, 181, 195, 236 Oximation 
Acetone (P10)  72, 117, 181, 253 Oximation 
Product nameMolecular structureCharacteristic ions (m/z)Identification method
p-Methane-1,8-diol (P1)  73, 81, 131, 211 Silylation 
4-Methyl-3-cyclohexen-1-one (P2)  79, 107, 181, 288, 305 Oximation 
1,3-Dihydroxyacetone (P3)  117, 161, 181, 195 Oximation 
4-Methyl-2-oxopentanoic acid (P4)  73, 181, 200, 355, 382 Combined oximation and silylation 
3-Methyl-2-oxobutanoic acid (P5)  73, 181, 366, 383 Combined oximation and silylation 
2,3-Butanedione (P6)  181, 279, 476 Oximation 
Isobutyraldehyde (P7)  56, 117, 181, 195, 250 Oximation 
Butyraldehyde (P8)  161, 181, 195, 239 Oximation 
Propionaldehyde (P9)  161, 181, 195, 236 Oximation 
Acetone (P10)  72, 117, 181, 253 Oximation 
Figure 4

Proposed degradation pathways of α-terpineol via identified transformation products with (a) GC-MS directly and (b) GC-MS after derivatization.

Figure 4

Proposed degradation pathways of α-terpineol via identified transformation products with (a) GC-MS directly and (b) GC-MS after derivatization.

Proposed photodegradation pathways

Based on these identified products, two degradation pathways were proposed and illustrated in Figure 4. P1 and P2 were the two representative transformation products of the two degradation pathways, respectively. P1 suggested that the absorption of UV by the C-C bond initially led to the generation of the biradical, which rearranged to M1 (see Figure 4) due to the higher stability of the tertiary radical. One hydroxyl group was then added to the side chain, which had a lower steric hindrance, to form P1. P2, which was also found in the degradation of α-terpineol in the gas phase and at the gas-solid interface, suggested a different pathway, which started with the hydroxyl radical attacking the active hydrogen atom in the hydroxyl group of α-terpineol. After the production of P2, it underwent the cleavage of the C-C bond, hydroxylation of M2 and oxidation of M3 (see Figure 4). After the formation of P1 and P2, other small molecular weight products were generated via ring cleavage induced by hydroxyl radical attack or UV adsorption, combined with a series of rearrangements. The C-C bond cleavage by UV led to continual loss of carbon, and the hydroxylation-by-hydroxyl radical that followed generated smaller organic fragments containing aldehyde and carboxylic acid groups, such as the observed products (P3-P9) presented in the proposed pathways.

Occurrences of transformation products

P1 and P2 were two main transformation products, the concentrations of which were monitored in the UV/H2O2 process. As Figure 5 shows, the reactions generating P1 and P2 occurred in the initial stage of the photodegradation. The amounts of P1 reached a maximum at 50 min and progressively decreased in the following 70 min, whereas the amounts of P2 reached a maximum at 15 min and disappeared at 120 min. Considering that α-terpineol had degraded almost completely at 60 min when the H2O2 concentration was 10 mg/L and the initial concentration of α-terpineol was 20 mg/L (see Figure 2), P1 and P2 remained much longer than α-terpineol. It should be noted that although α-terpineol and P1 have low toxicity, P2 is suspected to be acutely toxic via the oral route and hazardous to the aquatic environment according to European Chemicals Agency (ECHA). So if a drinking water source was polluted by α-terpineol as in the accident in Wushan described above, particular attention should be paid to the occurrence of P2 in the relevant drinking water plants which apply UV/H2O2 as a disinfection technique.

Figure 5

The peak area vs. time profile of P1 and P2 on degradation of α-terpineol under UV/H2O2 (Conditions: [α-terpineol]0 = 20 mg/L, [H2O2] = 10 mg/L, UV light intensity = 18 mW/cm2, pH = 7).

Figure 5

The peak area vs. time profile of P1 and P2 on degradation of α-terpineol under UV/H2O2 (Conditions: [α-terpineol]0 = 20 mg/L, [H2O2] = 10 mg/L, UV light intensity = 18 mW/cm2, pH = 7).

CONCLUSIONS

This study demonstrated that UV/H2O2 is a promising technology for the degradation of α-terpineol in aqueous solution. With an H2O2 dose of 10 mg/L and a UV fluence of 64.8 J/cm2 (light intensity: 18 mW/cm2, irradiation time: 60 min), more than 95% of the α-terpineol can be removed. A high pH value was found to be negative to the degradation. The investigation into the matrix effects of Cl, , , and HA suggested that the degradation could be significantly inhibited by HCO3 and HA. Thus in the treatment of wastewater from mineral dressing plants, the water should be acidified and aerated to remove bicarbonate before photodegradation if bicarbonate or carbonate were applied in the flotation process.

Two possible degradation pathways were proposed based on the identification of the transformation products: the first one started from hydroxylation of the C-C bond and the second one began with cleavage of the side chain. Moreover, two main transformation products were monitored. They remained much longer under the UV/H2O2 treatment than α-terpineol, and one of them is suspected to be toxic. To break down this potentially toxic byproduct, nearly twice the UV fluence was required than to remove α-terpineol.

ACKNOWLEDGEMENTS

This work was financially supported by the Special Fund for Agro-scientific Research in the Public Interest of China (No. 201503108) and the Key Laboratory of Hunan Province for Water Environment and Agriculture Product Safety.

CONFLICT OF INTEREST STATEMENT

The authors declare no potential conflicts of interest.

REFERENCES

Bolton
J.
&
Cater
S.
1994
Surface and Aquatic Environmental Photochemistry, Chapter 33, Boca Raton, FL: CRC Press
,
467
490
.
Brossard
L. E.
,
Varela
N.
,
Brossard
L. E.
,
Brossard
C. O.
,
Cortez
L. A. B.
&
Magalhaes
P. G.
2005
The surface-active bio oil solutions in sulfured copper mineral benefit
.
Braz. J. Chem. Eng.
22
(
1
),
1
7
.
Canonica
S.
,
Meunier
L.
&
von Gunten
U.
2008
Phototransformation of selected pharmaceuticals during UV treatment of drinking water
.
Water Res.
42
(
1–2
),
121
128
.
Chong
M. N.
,
Jin
B.
,
Chow
C. W.
&
Saint
C.
2010
Recent developments in photocatalytic water treatment technology: a review
.
Water Res.
44
(
10
),
2997
3027
.
Crittenden
J. C.
,
Hu
S.
,
Hand
D. W.
&
Green
S. A.
1999
A kinetic model for H2O2/UV process in a completely mixed batch reactor
.
Water Res.
33
(
10
),
2315
2328
.
Gupta
A. K.
,
Banerjee
P. K.
,
Mishra
A.
,
Satish
P.
&
Pradip
,
, .
2007
Effect of alcohol and polyglycol ether frothers on foam stability, bubble size and coal flotation
.
Int. J. Miner Process
82
(
3
),
126
137
.
Leviss
D. H.
,
Van Ry
D. A.
&
Hinrichs
R. Z.
2016
Multiphase ozonolysis of aqueous alpha-terpineol
.
Environ. Sci. Technol.
50
(
21
),
11698
11705
.
Liu
X.
,
Garoma
T.
,
Chen
Z.
,
Wang
L.
&
Wu
Y.
2012
SMX degradation by ozonation and UV radiation: a kinetic study
.
Chemosphere
87
(
10
),
1134
1140
.
Mack
J.
&
Bolton
J. R.
1999
Photochemistry of nitrite and nitrate in aqueous solution: a review
.
J. Photoch. Photobio. A
128
,
1
13
.
Prados-Joya
G.
,
Sánchez-Polo
M.
,
Rivera-Utrilla
J.
&
Ferro-Garcia
M.
2011
Photodegradation of the antibiotics nitroimidazoles in aqueous solution by ultraviolet radiation
.
Water Res.
45
(
1
),
393
403
.
Rosso
D.
,
Huo
D. L.
&
Stenstrom
M. K.
2006
Effects of interfacial surfactant contamination on bubble gas transfer
.
Chem. Eng. Sci.
61
(
16
),
5500
5514
.
Simonich
S. L.
,
Federle
T. W.
,
Eckhoff
W. S.
,
Rottiers
A.
,
Webb
S.
,
Sabaliunas
D.
&
De Wolf
W.
2002
Removal of fragrance materials during US and European wastewater treatment
.
Environ. Sci. Technol.
36
(
13
),
2839
2847
.
Souissi
Y.
,
Bouchonnet
S.
,
Bourcier
S.
,
Kusk
K. O.
,
Sablier
M.
&
Andersen
H. R.
2013
Identification and ecotoxicity of degradation products of chloroacetamide herbicides from UV-treatment of water
.
Sci. Total Environ.
458–460
,
527
534
.
Tadasa
K.
1977
Intermediates in the bacterial degradation pathway of α-Terpineol
.
Agr. Bio. Chem.
41
(
10
),
2095
2096
.
Tan
C.
,
Gao
N.
,
Deng
Y.
,
Zhang
Y.
,
Sui
M.
,
Deng
J.
&
Zhou
S.
2013
Degradation of antipyrine by UV, UV/H2O2 and UV/PS
.
J. Hazard Mater
260
,
1008
1016
.
Vasseghian
Y.
,
Heidari
N.
,
Ahmadi
M.
,
Zahedi
G.
&
Mohsenipour
A. A.
2014
Simultaneous ash and sulfur removal from bitumen: experiments and neural network modeling
.
Fuel Process Technol.
125
,
79
85
.
Ward
M. H.
,
Mark
S. D.
,
Cantor
K. P.
,
Weisenburger
D. D.
,
Correa-Villasenor
A.
&
Zahm
S. H.
1996
Drinking water nitrate and the risk of non-Hodgkin's lymphoma
.
Epidemiology
7
(
5
),
465
471
.
Xiao
Y.
,
Fan
R.
,
Zhang
L.
,
Yue
J.
,
Webster
R. D.
&
Lim
T.-T.
2014
Photodegradation of iodinated trihalomethanes in aqueous solution by UV 254 irradiation
.
Water Res.
49
,
275
285
.

Supplementary data