Degradation of three nonsteroidal anti-inflammatory drugs by UV/persulfate: Degradation mechanisms, efficiency in effluents disposal
Yingying Fu, Xingsheng Gao, Jinju Geng⁎, Shaoli Li, Gang Wu, Hongqiang RenState Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Jiangsu, PR China
H I G H L I G H T S
• SO4·− contributes more than ·OH in
NSAIDs degradation by UV/PS.
• Hydroxylation and decarboxylationare main pathways in NSAIDs de-gradation.
• The degradation efficiency of NSAIDswas much lower in real wastewaterthan in pure water.
• The degradation of NSAIDs was in-hibited in the presence of humic acidand HCO3
This study investigated the degradation of three nonsteroidal anti-inflammatory drugs (NSAIDs) (diclofenac(DCF), ibuprofen (IBP) and naproxen (NPX)) by UV/persulfate (UV/PS) in pure water and the influence ofwastewater on UV/PS performance. Three NSAIDs have no obvious removal in the PS system, while they allshowed a decrease in the UV system, and DCF showed the most significant decrease among the three NSAIDs inthe UV system. The degradation of NSAIDs in UV/PS system followed the pseudo first-order degradation kinetics,and the kinetic constants orders are: kDCF > kNPX > kIBP. SO4
·− contributes more than ·OH in NSAIDs de-gradation according to tertiary butanol and methanol inhibition tests. According to the degradation productsdetected by LC-MS, hydroxylation and decarboxylation are main pathways in NSAIDs degradation, and de-chlorination is also a main path in DCF degradation. When UV/PS was used to dispose effluents spiked withNSAIDs solution (DCF, NPX, IBP, 0.1 μM each), 10 times of PS dosage and 6 times treatment time were needed toget the same removal in pure water. The water quality parameters (i.e., natural organic matter (NOM), HCO3
−,Cl− and NO3
−) affects the degradation of NSAIDs by UV/PS. The degradation efficiency of NSAIDs was inhibitedin the presence of humic acid and HCO3
−. Meanwhile, Cl− and NO3− barely influenced the degradation effi-
ciency.
1. Introduction
Acidic pharmaceuticals, mainly non-steroidal anti-inflammatorydrugs (NSAIDs) have been detected at the concentration ranges from10 μg·L−1 to 70 μg·L−1 in wastewater treatment plant (WWTP) effluent
and from 0 to 30 μg·L−1 in receiving surface water in Europe [1], Asia[2], North America [3]. Diclofenac (DCF), ibuprofen (IBP), naproxen(NPX) are the three most frequently reported NSAIDs in WWTP influent[4], and they are ubiquitously detected in main river basins in China,including the Pearl River [5], Yellow River, Hai River, and Liao Rive
https://doi.org/10.1016/j.cej.2018.08.013Received 11 June 2018; Received in revised form 1 August 2018; Accepted 3 August 2018
[6]. Residues of NSAIDs in fish and invertebrates captured downstreamfrom WWTPs, indicates that NSAIDs tend to accumulate in the biota[7]. Brozinski [8] studied the NSAIDs in the bile of two wild fish, bream(Abramis brama) and roach (Rutilus rutilus), caught downstream of aWWTP, and found that only DCF, IBP and NPX were detected out of 17different pharmaceuticals. Due to the low concentration of NSAIDs inaquatic water, some effects caused by NSAIDs may only be recognizedin a later phase of life, or are manifested only in higher members of afood-web. Chronic histopathological effects in rainbow trout after28 days of exposure at 5 μg·L−1 of DCF revealed renal lesions (degen-eration of tubular epithelia, interstitial nephritis) and alterations of thegills occurred in rainbow trout, and subtle subcellular effects wereobserved even at 1 μg·L−1 of DCF. Cleuvers [9] also reflected that thetoxicity of mixed NSAIDs is considerable, even at concentrations atwhich the single substances showed no or only very slight effects. Therisk quotient (RQ) was often used to evaluated the environmental riskposed by NSAIDs on aquatic organisms, and the NSAIDs RQs in theYellow River and Liao River are higher than 0.1, which means NSAIDsmay pose medium risk to aquatic organisms, thus it is urgent to controlthe risk of NSAIDs in aquatic environment.
WWTPs represent an obligatory and final treatment step prior to theeffluent discharge, however, conventional WWTPs were designedwithout consideration of micropollutants removal. To reduce the risk ofNSAIDs in aquatic environment, reliable tertiary treatment technologiesare needed to effectively remove most pharmaceuticals [10]. UV/Na2S2O8 (UV/PS), one of typical UV-based AOPs, attracted more in-terests because of high oxidation ability of SO4
·− , with a redox potentialof 2.6 V in neutral pH [11]. UV light [12], ultrasonic [13], heat [14]and transition metals [15] can catalyze persulfate (PS) to produce SO4
·− ,among which UV light has the highest efficiency. PS irradiated by UV254 nm can generate SO4
·− as Eq. (1) [16].
ε φS O hv 2SO 21.1 M cm , 0.7 mol Einstein2 82
4· 1 1 1+ → = =
− − − − − (1)
ε is the molar extinction coefficients at 254 nm, and φ is the quantumefficiency of PS.
Degradation of DCF, IBP and NPX by heated PS, transition metalcatalyzed PS have been studied previously [17], and the studies wereconducted in pure water and focused on the efficiency and degradationmechanisms, whereas the application in wastewater was not in-vestigated thoroughly. In heated PS process, wastewater should beheated over 50 °C consuming lots of energy, and in transition metalcatalyzed PS process, ferrous ions or copper ions act as catalyst, whichmay cause heavy metal pollution in water. UV/PS not only consumesless energy, but also do not cause heavy metal pollution, and the de-gradation mechanisms of NSAIDs by UV/PS are not clear yet. Besides,in aquatic environment, different NSAIDs present simultaneously andthe concentration (mg·L−1) of anions or dissolved organic matters inwastewater is much higher than NSAIDs (ng·L−1–μg·L−1), thus the ef-ficiency of UV/PS will be affected. It is more meaningful to investigatethe efficiency of UV/PS in wastewater.
In this study, the main objectives are to: (1) investigate the de-gradation characteristics of the three NSAIDs in UV/PS process andidentify main degradation products; (2) evaluate the applicability ofUV/PS in degrading NSAIDs in effluent; (3) investigate the influences ofwater quality parameters (i.e., natural organic matter (NOM), HCO3
−,Cl− and NO3
−) on the performance of UV/PS. This research will con-tribute to a better evaluation of UV/PS in degradation NSAIDs, espe-cially the applicability in wastewater.
2. Materials and methods
2.1. Chemicals and reagents
DCF, IBP, NPX and humic acid (HA) were purchased from SigmaAldrich (St. Louis, MO, USA). HPLC grade methanol was supplied by
Merck (Darmstadt, Germany). Sodium persulfate, sodium sulfite, so-dium bicarbonate and sodium chloride were analytical grade and ob-tained from Nanjing Chemical Reagent Factory (Naniing, China). Milli-Q water, with a resistivity of at least 18.2 MΩ/cm, was produced from aMillipore purification system (Billerica, CA, USA).
2.2. UV/PS experiment procedures
UV/PS degradation experiments were carried out in a photoreactionreactor (XPA-7, Nanjing Xujiang Factory, China). The photochemicalreactor was shown in Fig. S1. The photoreaction reactor consisted oftwo parts: quartz reactors and a low-pressure mercury lamp. The low-pressure mercury lamp (22W, 254 nm) supplied by an electronic ballastwas placed in the center of the reactor with a quartz cover. The quartzreactors were vertically placed at a fixed distance of 3.5 cm from thelamp. The UV lamp have an emission irradiance of 1250 μW/cm2 at254 nm measured by a UV radiation meter (Model FZ-A, PhotoelectricInstrument Factory of Beijing Normal University, China). Cappedquartz tubes each containing 50mL of reaction solution were placed ina rotating unit. A stir bar was placed inside each quartz tube to ensurethat the solution was well mixed. In a reaction experiment, 50mL ofreaction solution (NSAIDs solution and PS solution) was added intoeach capped quartz tube, and then the low-pressure mercury lamp wasturned on. In order to make the pH value closer to the wastewater, thereaction in pure water was adjusted to pH 7 by phosphate buffer so-lution (2mM), and the pH value in real wastewater was unadjusted.During the photoreaction, a steady flow of cooling water was used tomaintain a constant temperature of about 25 °C.
In radical inhibition experiments, scavengers (tertiary butanol(TBA) or methanol (MeOH), 2mM) was added in NSAIDs solution. Toidentify the degradation products of each NSAID, UV/PS experimentswere conducted using single constituent NSAID solutions.
To investigate the degradation of NSAIDs in wastewater, effluents offive WWTPs in Nanjing and Wuxi, China, were collected and performedUV/PS experiments. Table S1shows the detailed characteristics of theWWTPs studied, and Table 1 shows the water quality parameters ofeffluents. Effluents added with NSAIDs (DCF, IBP and NPX, 0.1 μMeach) were treated by UV/PS, and samples were treated by a solid phaseextraction (SPE) workstation using CNW Poly-Sery HLB cartridges(CNW, ANPEL, Shanghai, China). NSAIDs in raw effluents (Table 1) isless than 1/10 of NSAIDs addition (0.1 μM), so the NSAIDs in raw ef-fluents were neglected. At first, groups of experiments were conductedto ascertain suitable PS dosage in effluent disposal. Then, removal ofNSAIDs in five different effluents was investigated under optimized PSdosage. HA or inorganic salt stock solutions was added into NSAIDssolution separately when conducting factor-influence experiments.
2.3. Acute toxicity measurements
The acute toxicity assays for NSAIDs and degradation products werecarried out according to the standard toxicity measurement protocol[18]. Photobacterium phosphoreum T3 spp (P. phosphoreum) obtained
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1033
FYY
高亮
from the Nanjing Institute of Soil Science, Chinese Academy of Sci-ences) was used as the test bacteria, and the acute toxicity was de-termined using the Microtox Toxicity Analyzer (DXY-3, Nanjing In-stitute of Soil Science, Chinese Academy of Science). Each sample wastested in quintuplicate. According to the samples luminescence unit (Li)and the controls luminescence unit (L0), the relative luminescence unit(Ii) was calculated as Eq. (2) [19]:
I L L(%) ( / ) 100%i i 0= × (2)
2.4. Analytical methods
NSAIDs was quantified by UPLC-MS (Waters Xevo TQ-S UPLC-MS,USA). The analytical instrument was controlled by the MASS Lynxsoftware (Waters Technologies). Separation was performed using anACQUITY UPLC BEH-C18 column (2.1× 100mm, 1.7 μm, Waters) at30 °C with 0.2 mL/min isocratic elution. The isocratic elution wasperformed by 10% ammonium hydroxide solution (NH4OH, 0.02% v/v)and 90% methanol holding for 3min and the injection volume was20 μL. The triple quadrupole mass spectrometer operated in electro-spray negative (ESI-) mode. Data acquisition was performed by multiplereaction monitoring (MRM), recording the transitions between theprecursor ion and the product ions for each target analyte. All MRMparameters were obtained by successive injections in the full chroma-tographic-spectrometric system, meanwhile compared with publishedresearch to ensure the reliability [20].
The UV/PS degradation products (DPs) of NSAIDs were identifiedusing U-3000 HPLC coupled with LTQ Orbitrap (Thermo, USA). Themobile phase consisted of 10% ammonium acetate solution (0.005mM)and 90% methanol at a flow rate of 0.2mL/min, and the injection vo-lume was 20 μL.
A multi N/C ®2100 TOC analyzer (Analytik Jena, German) was usedto evaluate the mineralization of NSAIDs during the UV/PS treatment.The concentrations of anions were analyzed using an ion chromato-graphy system (Dionex ICS-5000, Thermo USA). The pH values weremeasured with a pH meter (HQ40d, HACH USA). COD (ChemicalOxygen Demand) was determined according to Standard Methods forthe Examination of Water and Wastewater [21].
2.5. Quality assurance and quality control (QA/QC)
All samples were taken at each designated sampling time in tripli-cate and stored at 4 °C until analysis, and all the analytical results arereported as mean ± RSD. The precision of the analytical method wasvalidated by observing the short-term and long-term relative standarddeviation (RSD) under identical conditions. The RSD for intra- andinter-day precision were 5.6–10.9% and 6.8–12.5%. No quantifiableanalytes were detected in blank samples. The recoveries of threeNSAIDs were 88.7 ± 3.2% for water samples.
3. Results and discussion
3.1. Degradation of NSAIDs by UV or PS
Fig. 1 shows the degradation of NSAIDs by UV or PS alone. In1800 s, no obvious removal of the three NSAIDs were observed by PS.Researchers also found there is no degradation of florfenicol, oxcarba-zepine, geosmin and 2-methylisoborneol by PS [22]. In UV photolysis,92% of DCF was removed in 600 s, and removal of IBP and NPX were28% and 52%, respectively. It can be concluded that, direct photolysiscontributes to NSAIDs degradation, and the photolysis characteristics isDCF > NPX > IBP.
The direct photolysis of a compound M can be described by thereaction (3):
hvM products+ → (3)
And the removal rate of compound M by direct UV photolysis can becalculated with the Eq. (4):
k k εd[M]dT
[M] Φ [M]M M− = = ′(4)
where T is the irradiation time (s), k is the pseudo first order kineticconstant (s−1), k’ is the correction factor, εM is the molar absorptioncoefficient of M at 254 nm (mol−1·cm−1), ΦM is the quantum yield of Mat 254 nm (mol·Einstein−1).
εM and ΦM are two main parameters that influence the direct pho-tolysis of a compound. ΦM is defined as the fraction of photons thatdecompose the compound over the total number of photons absorbedby the compound. Table 2 shows the εM and ΦM of the three NSAIDs.DCF has the highest molar absorption coefficient and quantum yield,and the molar absorption coefficient of NPX is higher than IBP, whereasthe quantum yield of NPX is lower than IBP. The higher the product ofεM and ΦM , the more inclined to be photodegrade, as Eq. (5):
b ε ΦM M= (5)
The b (Einstein−1·cm−1) of DCF, IBP and NPX are 1842, 49 and 58,respectively. From Fig. 1, the half-life period (s−1) of DCF, IBP, NPX are105, 1687 and 3858, respectively, NSAID with longer half-life periodhas lower b, which is correspondence with the results calculated by Eq.(5).
3.2. Degradation of NSAIDs by UV/PS
PS cannot degrade the three NSAIDs, and NPX, IBP were difficult tobe degraded by UV irradiation (254 nm) alone, thus degradation ofNSAIDs by UV/PS was further investigated. The effect of PS dosages(2–10 μM) on NSAIDs degradation in UV/PS system was investigated.
Fig. S2 shows the degradation profile of NSAIDs in UV/PS, when PSdosage is 10 μM, the removal of NSAIDs reached 90% in 300 s, muchhigher than UV or PS treated alone (Fig. 1). Table 3 shows the pseudo
Fig. 1. Degradation of NSAIDs by UV or PS. ([DCF]0= [IBP]0= [NPX]0=0.1 μM, [PS]0= 10 μM, UV 1250 μW/cm2).
Table 2εM and ΦM of NSAIDs [23].
εM (mol−1·cm−1) ΦM(mol·Einstein−1)
DCF 6307 0.292NPX 4492 0.0130IBP 256 0.192
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1034
first-order kinetic constants of NSAIDs under different PS concentra-tions. The kinetic constants of each NSAID linearly increased with thePS dosage. At each PS dosage, the kinetic constants orders are:kDCF > kNPX > kIBP, implying the degradation of DCF is the quickestand IBP is the slowest. SO4
·− , %OH radicals were generated when per-sulfate irradiated by UV, and the radicals played an important role inNSAIDs degradation. The formation rate of SO4
·− increased with theincrease of PS dosage [24], then the ideal steady-state concentrations ofSO4
·− was expected to increase thus contribute more to NSAIDs de-gradation. Luo et al. [25] investigated the oxidation of the odorouscompound 2,4,6-trichloroanisole (TCA) by UV/PS and found the ob-served pseudo first-order kinetic constant (kobs) increased from1.32×10−3 s−1 to 6.55× 10−3 s−1 when PS dosage increased from100 μM to 500 μM.
3.3. Roles of sulfate radicals and hydroxyl radicals
SO4·− and ·OH are main radicals reacting with NSAIDs in UV/PS
system. TBA and MeOH were used as scavengers to distinguish thecontributions of ·OH and SO4
·− . TBA [26] is effective quencher for ·OH,and MeOH is effective quencher for both ·OH and SO4
·− [27]. The re-lative contributions of ·OH, SO4
·− and others to the degradation ofNSAIDs in the UV/PS system were modeled with Eq. (6). In order toquench radicals sufficiently, the concentrations of TBA and MeOH were2mM, 400 times higher than PS concentration. Fig. 2 shows the con-tributions of different radicals in NSAIDs degradation, 52–63% de-gradation of the three NSAIDs was accomplished by SO4
·− , and 20–38%degradation of the three NSAIDs was accomplished by ·OH. Photo-degradation takes 28% in DCF degradation, which is much higher thanNPX(5%) and IBP(8%), that is because DCF is more inclined be pho-todegraded.
k t k k k tln [T][T]
· ( )·obs T OH SO others0
, · 4·= − = − + +−
(6)
where [T]0 is the initial concentration of target compound and [T] isthe concentration after treatment; kobs,T was the observed pseudo-first-order degradation rate constant (s−1) for the target compound; and thek·OH, kSO4
·− and kothers were pseudo-first-order degradation rate constants(s−1) with ·OH, SO4
·− and others.NPX [28] is more susceptible to oxidation via electron transfer, due
to its structure shows a more electron-rich environment (e.g. presenceof a naphthalene moiety). However, IBP is abundance with hydrogenatoms (three –CH3 in IBP) that can undergo quickly ·OH abstraction.Thus, when ·OH was inhibited by TBA, kobs of IBP decreased by 38%,higher than NPX (32%). When ·OH and SO4 were both inhibited byMeOH, kobs of IBP, NPX decreased by 92%, 95%, indicating SO4 play amore important role in NPX degradation than in IBP degradation.
Besides, to investigate the difference of k·OH, and kSO4·− , the con-
centrations of ·OH and SO4·− were quantified. It can be calculated with
Eqs. (7) and (8) based on the kinetic of IBP whose second-order rate
Table 3Pseudo first-order kinetic constants of NSAIDs under different concentrations ofPS.
Concentrations of PS/μM kinetic constants/s−1
DCF NPX IBP
10 1.25× 10−2 9.83×10−3 7.01× 10−3
5 1.14× 10−2 6.08×10−3 4.32× 10−3
2 1.02× 10−2 3.41×10−3 2.82× 10−3
linearity of kinetic constants/R2 0.97 0.99 0.99
Fig. 2. Contributions of different radicals in NSAIDs degradation.
Fig. 3. Removal and mineralization of NSAIDs by UV/PS. ([DCF]0= [IBP]0=[NPX]0= 10.0 μM, [PS]0= 0.5mM, UV 1250 μW/cm2).
Fig. 4. Toxic effect of NSAIDs treated by UV/PS on P. phosphoremu.
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1035
constants with ·OH and SO4·− is known.
kk
[·OH] ·OH
·OH,IBP=
(7)
kk
[SO ]4· SO
SO ,IBP
4
4·
=−
− (8)
where k·OH,IBP and kSO ,IBP4·− represent the second-order rate constants
(M−1s−1) of ·OH and SO4·− with IBP, respectively. And the value of
k·OH,IBP and kSO ,IBP4·− were 5.23×109M−1s−1 and 1.32× 109M−1s−1
[29]. [·OH] and [SO ]4·− were the steady-state concentration of ·OH and
SO4·− . By calculation, the concentrations of ·OH and SO4
·− were2.66×10−14 M and 9.33× 10−13 M, respectively. The concentrationof SO4
·− was much higher than that of ·OH. So the higher contribution ofSO4 than ·OH may caused by the higher concentration of SO4 and thehigher second-order rate constants with SO4
·− .
3.4. Mineralization and toxicity of NSAIDs in UV/PS
High removal of NSAIDs was achieved when treated by UV/PS,however, the decline of the concentration of NSAIDs cannot stand forthe diminish of ecotoxicity, because new degradation products willgenerate during the degradation process. The change of TOC canmanifest the mineralization of pollutants. Fig. 3 shows the removal andmineralization of NSAIDs by UV/PS. The removal of NSAIDs were allover 99% when treated by UV/PS for 1800 s, but the TOC removal wasonly 27%. The high removal of NSAIDs and the low removal of TOC
proved that some degradation products generated and the degradationproducts were hard to be removed by UV/PS. Similar results that de-gradation products cannot be removed by UV/PS, can also be found insulfamethazine [30], florfenicol [31], clofibric acid [32] degradationprocess.
Fig. 4 shows the toxic effect of NSAIDs (10 μM and 0.1 μM) treatedby UV/PS on P. phosphoremu. The removal of TOC had no distinctchange after 1800 s, it means that increasing time has little effect on thefurther mineralization, moreover, considering the time cost, 1800 s oftreatment time is enough. At beginning, the inhibition ratios of un-treated 10 μM NSAIDs and 0.1 μM NSAIDs on P. phosphoremu were 37%and −11%, respectively. During the treatment, the inhibition increasedwith time. At 1800 s, the inhibition of 10 μM NSAIDs group increased to100%, and the inhibition of 0.1 μM NSAIDs group increased to 12%,much lower than the 10 μM NSAIDs group. According to the results ofFigs. 3 and 4, it can be concluded that the degradation products aremore toxic than NSAIDs. Although the TOC changed little after 1800 s,the toxicity of NSAIDs treated by UV/PS for longer time needs furtherassessment due to the change of the intermediates.
3.5. Degradation pathways of NSAIDs
3.5.1. Degradation pathways of DCFMass spectrum and proposed structures of three main products of
DCF detected by LTQ-Orbitrap (ESI-) are shown in Fig. S3. The isotopicdensity of 35Cl and 37Cl satisfies the ratio 3:1, which facilitates theidentification of the degradation products of DCF (containing two
Fig. 5. Proposed degradation pathways of DCF.
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1036
chlorine atoms) [33]. DCF showed three isotopic peaks (m/z 294, 296,298), with the relative intensity ratio of 9:6:1. DCF-DP1 showed twoisotopic peaks (m/z 258, 260), and the m/z of the isotopic peaks were38 or 36 Da (HCl) less than DCF. The same product with DCF-DP1 wasidentified during the transformation of DCF by UV photolysis [34].DCF-DP2 was 44 Da (–COO) less than DCF, and it showed three peaks(m/z 250, 252, 254), implicating there were two Cl in DCF-DP2.Compared with DCF-DP2, DCF-DP3 (m/z 214, 216) presented 38 or36 Da (HCl) less.
DCF may undergo two possible reaction processes (Fig. 5). Electrontransfer is a main pathway in reaction between pollutants and SO4
·−
[35]. In reaction Path 1, SO4·− abstracted one electron from DCF, and
DCF converted into cationic state (DCF+) which is unstable. Then DCF-DP1 generated by dechlorination. In Path 2, the – COOH group of DCFwas decarboxylated to form intermediate DCF-DP2, and DCF-DP2 wasdechlorinated to form DCF-DP3. Yu [36] studied degradation me-chanism by adding isopropanol in DCF solutions to inhibit ·OH and H+,and found that dechlorination was induced by solvated electron.
3.5.2. Degradation pathways of NPXMass spectrum and proposed structures of three main products of
NPX detected by LTQ-Orbitrap (ESI+) are shown in Fig. S4, and theproposed degradation pathway is shown in Fig. 6. NPX molecule isionized upon electron abstraction by SO4
·− to form radical cation
[NPX]+. Electron transfer mostly occur on naphthalene ring since itsionization energy (8.144 eV) is less than that of the nonbonding elec-tron pairs on oxygen atoms (≥10.0 eV). Then, NPX-DP3 (m/z 185)formed through decarboxylation. NPX-DP3 underwent oxidative reac-tion in the presence of dissolved oxygen and generated NPX-DP4 (m/z200). In another pathway, NPX was first hydroxylated to NPX-DP2, andNPX-DP2 transferred to NPX-DP3 by decarboxylation. NPX-DP3 has thehighest relative abundance of all NPX-DPs (Fig. S4), implicating dec-arboxylation is the main pathway in NPX degradation. NPX-DP3, NPX-DP4 are also main DPs in UV or VUV [37], UV/TiO2 [38], heated PSprocesses.
3.5.3. Degradation pathways of IBPMass spectrum and proposed structures of five main products of IBP
detected by LTQ-Orbitrap (ESI-) are shown in Fig. S5. And the proposeddegradation pathway is shown in Fig. 7. IBP-DP3 (m/z 221) is hydro-xylated product of IBP, and further hydroxylation on the ring of IBP-DP3 led to the formation of the multi-hydroxylated compound IBP-DP2(m/z 257), which is 36 Da (2 H2O) higher than IBP-DP3·H2O decom-posed to H+ and ·OH under SO4
·− attack, and then addition reaction tookplace on IBP-DP3. A bond of carbon formed when IBP-DP2 (m/z 257)was oxidized, generating IBP-DP1. Decarboxylation on the side chainsof IBP produced IBP-DP5 (m/z 159), and the decarboxylation was ac-complished by electron transfer, the same with DCF and NPX. Further
Fig. 6. Proposed degradation pathways of NPX.
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1037
oxidation of IBP-DP5 generated IBP-DP4 (m/z 175).IBP-DP1, IBP-DP2, IBP-DP3 are three DPs with high relative abun-
dance, which suggests that hydroxylation is a main path in IBP de-gradation. IBP-DP3 is also a kind of main DPs in UV/Cl2 [39], UV/TiO2,and sonophotocatalysis [40] processes.
The DPs described above can be detected when NSAIDs (10 μM)treated by UV/PS (PS 1mM) for less than 180 s, whereas the removal ofTOC was only 10% at 300 s, that is because main degradation productswere further degraded under SO4 or ·OH attack to produce a range ofsmaller molecule products. Corresponding to the results of toxicity as-says, the inhibition increased with time during the whole treatment(1800 s). Hama Aziz [41] compared the efficiency of ozonation, pho-tocatalysis and photocatalysis ozonation on DCF and IBP degradation,and concluded that photocatalysis ozonation had the highest TOC re-moval. However, there are still 5 mg·L−1 of acetate and oxalate re-maind, even after a complete decomposition of the two pollutants.Arany’s [37] study also suggested the formation of aliphatic acids whenNPX (0.1mM) irradiated by UV (26.1mW cm−2) for 1 h, and theidentified aliphatic acids were acetic, propionic, oxalic, malic andsuccinic acids. In summary, the increase of aliphatic acids during thedisposal increased the toxicity.
3.6. Degradation of NSAIDs in WWTP effluent
Fig. 8 shows the removal of NSAIDs under 10 μM and 100 μM PSdosage. When treated by UV/PS ([PS]0= 10 μM) for 1800 s, removal ofDCF, NPX and IBP in effluent are 64%, 46% and 12%, respectively.Under the same PS dosage, in pure water, the removal of NSAIDsreached 90% in 300 s, much higher than those in effluent. When PSdosage raised to 100 μM in effluent, higher removal rate (90%, 85%,41% for DCF, IBP, NPX, respectively) achieved. A moderate removal at100 μM PS dosage is suitable for the analysis of effluent quality on UV/PS efficiency, thus the following experiments were performed at100 μM PS dosage.
The removal of NSAIDs in different effluents are shown in Fig. 9.Three NSAIDs shown lowest removal (80%, 77%, 37% for DCF, NPXand IBP, respectively) in WWTP2 effluent and highest removal (97%,91%, 81% for DCF, NPX and IBP, respectively) in WWTP4 effluent.From Table 1, effluent of WWTP2 are high in HA, COD and HCO3
−
concentration compared with other effluents. In each effluent, the re-moval of NSAIDs are DCF > NPX > IBP. The result was in goodagreement with the degradation of NSAIDs in pure water above. It maybe caused by the higher second-order rate constants of DCF with ·OH
Fig. 7. Proposed degradation pathways of IBP.
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1038
and SO4·− [29], and that DCF is most inclined to be degraded by UV is
also an important reason. The low efficiency of UV/PS in effluent maybe due to the quench of SO4
·− by constituents in effluent. The keyscavenging species presented in effluents were verified by performing adetailed investigation of the scavenging effects of the analyzed water
using synthetic solutions.
3.7. Effect of HA and anions on the degradation of NSAIDs
HA are typical dissolved organic matter in wastewater, and HCO3- ,
Cl−, NO3- are common anions in wastewater, so it is necessary to verify
the effects of these anions on UV/PS performance. The apparentpseudo-first-order rate constants (kobs) of NSAIDs decreased by in-creasing the concentrations of HA (Fig. 10 (a)). It has been reported thatHA could either promote or inhibit the UV/H2O2 process [42]. In thisstudy, degradation of NSAIDs were inhibited most evidently by HA,compared with other anions. Organic matter absorbed UV photons,thereby decreasing the UV transmittance, and can also scavenge radi-cals species competing with NSAIDs. kobs of NSAIDs also decreased byincreasing the concentrations of HCO3 (Fig. 10(b)). Table 4 exhibits thereactions between anions and radicals. HCO3 and CO3 react with ·OHand SO4 to produce weak oxidant carbonate radical [43]. Thus theoxidation reactions are insignificant and can be neglected for treatingmost organic pollutants, so it is reasonable that degradation of NSAIDswere inhibited significantly by HCO3.
Slight inhibition of NSAIDs degradation was observed in the pre-sence of NO3
- and Cl−. Though kobs of DCF decreased with the in-creasing of NO3
- and Cl− (Fig. 10(c) and (d)), the removal of DCF didnot declined much (less than 3% decrease) (Table S2). The weak in-fluence of NO3
- is due to the low reaction rate between NO3- and SO4.
Cl· generated by SO4 and Cl− can have negative or positive impact oncompounds’ degradation, and it depends on the reactive of compoundwith SO4
·− and Cl.
4. Conclusion
Degradation of NSAIDs were accelerated by the combination of UVand PS compared to UV alone. Hydroxylation and decarboxylation aremain pathways in NSAIDs degradation, and they are accomplished byelectron transfer between NSAIDs and PS. Dechlorination is also a mainpath in DCF degradation. Main products of NSAIDs were further de-graded to aliphatic acids, and the increase of aliphatic acids during thedisposal increased the toxicity. The efficiency of UV/PS in wastewater ismuch lower than that in pure water for the quenching effect of watermatrix. HA and HCO3
− have obvious inhibition on the efficiency ofUV/PS, other anions (NO3
- and Cl−) have slight inhibition effect.Longer reaction period or higher PS dosage are needed to satisfy theapplication of UV/PS in effluent disposal, but whether the completeminzeration of degradation products needs to be further tested.
Fig. 8. Removal of NSAIDs in WWTP1 effluent treated by UV/PS([DCF]0= [IBP]0= [NPX]0=0.1 μM, UV 1250 μW/cm2).
Fig. 9. Removal rate of NSAIDs in effluent of five WWTPs treated by UV/PS for1800 s. ([DCF]0= [IBP]0= [NPX]0= 0.1 μM, [PS]0= 100 μM, UV 1250 μW/cm2).
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
1039
Acknowledgement
This work was supported by the National Science Foundation ofChina (No. 21677071) and the Jiangsu Natural Science Foundation(Nos. BK20161474, SBK2018010218).
Appendix A. Supplementary data
Supplementary data associated with this article can be found, in theonline version, at https://doi.org/10.1016/j.cej.2018.08.013.
References
[1] M. Gros, M. Petrović, A. Ginebreda, D. Barceló, Removal of pharmaceuticals duringwastewater treatment and environmental risk assessment using hazard indexes,Environ. Intern. 36 (2010) 15–26.
[2] W.-J. Sim, J.-W. Lee, E.-S. Lee, S.-K. Shin, S.-R. Hwang, J.-E. Oh, Occurrence anddistribution of pharmaceuticals in wastewater from households, livestock farms,
hospitals and pharmaceutical manufactures, Chemistry 82 (2011) 179–186.[3] Y. Yu, L. Wu, A.C. Chang, Seasonal variation of endocrine disrupting compounds,
pharmaceuticals and personal care products in wastewater treatment plants, Sci.Total Environ. 442 (2013) 310–316.
[4] P. Verlicchi, M. Al Aukidy, E. Zambello, Occurrence of pharmaceutical compoundsin urban wastewater: removal, mass load and environmental risk after a secondarytreatment—A review, Sci. Total Environ. 429 (2012) 123–155.
[5] J.-L. Zhao, G.-G. Ying, L. Wang, J.-F. Yang, X.-B. Yang, L.-H. Yang, X. Li,Determination of phenolic endocrine disrupting chemicals and acidic pharmaceu-ticals in surface water of the Pearl Rivers in South China by gas chromato-graphy–negative chemical ionization–mass spectrometry, Sci. Total Environ. 407(2009) 962–974.
[6] L. Wang, G.-G. Ying, J.-L. Zhao, X.-B. Yang, F. Chen, R. Tao, S. Liu, L.-J. Zhou,Occurrence and risk assessment of acidic pharmaceuticals in the Yellow River, HaiRiver and Liao River of north China, Sci. Total Environ. 408 (2010) 3139–3147.
[7] A. Ruhí, V. Acuña, D. Barceló, B. Huerta, J.R. Mor, S. Rodríguez-Mozaz, S. Sabater,Bioaccumulation and trophic magnification of pharmaceuticals and endocrine dis-ruptors in a Mediterranean river food web, Sci. Total Environ. 540 (2016) 250–259.
[8] J.-M. Brozinski, M. Lahti, A. Meierjohann, A. Oikari, L. Kronberg, The anti-in-flammatory drugs diclofenac, naproxen and ibuprofen are found in the bile of wildfish caught downstream of a wastewater treatment plant, Environ. Sci. Technol. 47(2013) 342–348.
[9] M. Cleuvers, Mixture toxicity of the anti-inflammatory drugs diclofenac, ibuprofen,naproxen, and acetylsalicylic acid, Ecotoxicol. Environ. Saf. 59 (2004) 309–315.
[10] R.I.L. Eggen, J. Hollender, A. Joss, M. Scharer, C. Stamm, Reducing the discharge ofmicropollutants in the aquatic environment: the benefits of upgrading wastewatertreatment plants, Environ. Sci. Technol. 48 (2014) 7683–7689.
[11] R. Li, J. Kong, H. Liu, P. Chen, Y. Su, G. Liu, W. Lv, Removal of indomethacin usingUV-vis/peroxydisulfate: kinetics, toxicity, and transformation pathways, Chem.Eng. J. 331 (2018) 809–817.
[12] W. Chu, D. Li, Y. Deng, N. Gao, Y. Zhang, Y. Zhu, Effects of UV/PS and UV/H2O2pre-oxidations on the formation of trihalomethanes and haloacetonitriles duringchlorination and chloramination of free amino acids and short oligopeptides, Chem.Eng. J. 301 (2016) 65–72.
[13] H. Hori, Y. Nagano, M. Murayama, K. Koike, S. Kutsuna, Efficient decomposition ofperfluoroether carboxylic acids in water with a combination of persulfate oxidantand ultrasonic irradiation, J. Fluorine Chem. 141 (2012) 5–10.
[14] C. Tan, N. Gao, Y. Deng, N. An, J. Deng, Heat-activated persulfate oxidation of
Fig. 10. Degradation of NSAIDs in the presences of HA or anions. (a)HA, (b) HCO3−, (c) Cl−, (d) NO3−. ([DCF]0= [IBP]0= [NPX]0= 0.1 μM, UV 1250 μW/cm2,
[PS]0= 5 μM).
Table 4Reactions between anions and radicals.
Reaction k/(M−1·s−1) Refs.
(1) SO HCO CO HSO4·
3 3·
4+ → +− − − − 9.1× 106 [44]
(2) SO CO SO CO4·
32
42
3·
+ → +− − − − 6.1× 106 [45]
(3) SO NO NO SO4·
3 3·
42
+ → +− − − 5.0× 104 [46]
(4) SO Cl Cl SO4· ·
42
+ → +− − − 3.1× 108 [47]
(5) HO· HCO CO OH3 3·
+ → +− − − 8.5× 106 [44]
(6) HO· CO CO OH32
3·
+ → +− − − 3.9× 108 [44]
(7) HO· Cl ClOH ·+ →− − 4.3× 109 [44]
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
diuron in water, Chem. Eng. J. 203 (2012) 294–300.[15] C.J. Liang, C.J. Bruell, M.C. Marley, K.L. Sperry, Persulfate oxidation for in situ
remediation of TCE. II. Activated by chelated ferrous ion, Chemistry 55 (2004)1225–1233.
[16] X. Lu, Y. Shao, N. Gao, J. Chen, H. Deng, W. Chu, N. An, F. Peng, Investigation ofclofibric acid removal by UV/persulfate and UV/chlorine processes: kinetics andformation of disinfection byproducts during subsequent chlor(am) ination, Chem.Eng. J. 331 (2018) 364–371.
[17] A. Ghauch, A.M. Tuqan, N. Kibbi, Ibuprofen removal by heated persulfate in aqu-eous solution: a kinetics study, Chem. Eng. J. 197 (2012) 483–492.
[18] I. 11384, Water quality-Determination of the inhibitory effect of water samples onthe light emission of Vibrio fischeri (luminescent bacteria test), InternationalOrganization for Standardization: Geneva; 1994.
[19] C. Wu, Y. Zhou, Q. Sun, L. Fu, H. Xi, Y. Yu, R. Yu, Appling hydrolysis acidification-anoxic–oxic process in the treatment of petrochemical wastewater: from bench scalereactor to full scale wastewater treatment plant, J. Hazard. Mater. 309 (2016)185–191.
[20] B.J. Vanderford, R.A. Pearson, D.J. Rexing, S.A. Snyder, Analysis of endocrinedisruptors, pharmaceuticals, and personal care products in water using liquidchromatography/tandem mass spectrometry, Anal. Chem. 75 (2003) 6265–6274.
[21] W.E. Federation, A.P.H. Association, Standard methods for theexamination of waterand wastewater, American Public Health Association: Washington, DC, USA; 2005.
[22] L. Bu, S. Zhou, Z. Shi, L. Deng, G. Li, Q. Yi, N. Gao, Degradation of oxcarbazepine byUV-activated persulfate oxidation: kinetics, mechanisms, and pathways, Environ.Sci. Pollut. R. 23 (2015) 2848–2855.
[23] R. Marotta, D. Spasiano, I. Di Somma, R. Andreozzi, Photodegradation of naproxenand its photoproducts in aqueous solution at 254 nm: a kinetic investigation, WaterRes. 47 (2013) 373–383.
[24] P. Xie, J. Ma, W. Liu, J. Zou, S. Yue, X. Li, M.R. Wiesner, J. Fang, Removal of 2-MIBand geosmin using UV/persulfate: contributions of hydroxyl and sulfate radicals,Water Res. 69 (2015) 223–233.
[25] C.W. Luo, J. Jiang, J. Ma, S.Y. Pang, Y.Z. Liu, Y. Song, C.T. Guan, J. Li, Y.X. Jin,D.J. Wu, Oxidation of the odorous compound 2,4,6-trichloroanisole by UV activatedpersulfate: kinetics, products, and pathways, Water Res. 96 (2016) 12–21.
[26] Y. Wang, H. Liu, Y. Xie, T. Ni, G. Liu, Oxidative removal of diclofenac by chlorinedioxide: reaction kinetics and mechanism, Chem. Eng. J. 279 (2015) 409–415.
[27] G.P. Anipsitakis, D.D. Dionysiou, Radical generation by the interaction of transitionmetals with common oxidants, Environ. Sci. Technol. 38 (2004) 3705–3712.
[28] A. Ghauch, A. Tuqan, N. Kibbi, Naproxen abatement by thermally activated per-sulfate in aqueous systems, Chem. Eng. J. 279 (2015) 861–873.
[29] L. Lian, B. Yao, S. Hou, J. Fang, S. Yan, W. Song, Kinetic study of hydroxyl andsulfate radical-mediated oxidation of pharmaceuticals in wastewater effluents,Environ. Sci. Technol. 51 (2017) 2954–2962.
[30] Y.-Q. Gao, N.-Y. Gao, Y. Deng, Y.-Q. Yang, Y. Ma, Ultraviolet (UV) light-activatedpersulfate oxidation of sulfamethazine in water, Chem. Eng. J. 195 (2012) 248–253.
[31] Y.-Q. Gao, N.-Y. Gao, Y. Deng, D.-Q. Yin, Y.-S. Zhang, Degradation of florfenicol inwater by UV/Na2S2O8 process, Environ. Sci. Pollut. R. 22 (2015) 8693–8701.
[32] I.Y. Kim, M.K. Kim, Y. Yoon, J.K. Im, K.D. Zoh, Kinetics and degradation mechanismof clofibric acid and diclofenac in UV photolysis and UV/H2O2 reaction, Desal.Water Treatment 52 (2014) 6211–6218.
[33] H. Cheng, D. Song, H. Liu, J. Qu, Permanganate oxidation of diclofenac: the pH-dependent reaction kinetics and a ring-opening mechanism, Chemistry 136 (2015)297–304.
[34] O.S. Keen, E.M. Thurman, I. Ferrer, A.D. Dotson, K.G. Linden, Dimer formationduring UV photolysis of diclofenac, Chemistry 93 (2013) 1948–1956.
[35] M.G. Antoniou, A.A. de la Cruz, D.D. Dionysiou, Intermediates and reaction path-ways from the degradation of Microcystin-LR with sulfate radicals, Environ. Sci.Technol. 44 (2010) 7238–7244.
[36] H. Yu, E. Nie, J. Xu, S. Yan, W.J. Cooper, W. Song, Degradation of diclofenac byadvanced oxidation and reduction processes: kinetic studies, degradation pathwaysand toxicity assessments, Water Res. 47 (2013) 1909–1918.
[37] E. Arany, R.K. Szabo, L. Apati, T. Alapi, I. Ilisz, P. Mazellier, A. Dombi, K. Gajda-Schrantz, Degradation of naproxen by UV, VUV photolysis and their combination, J.Hazard. Mater. 262 (2013) 151–157.
[38] N. Jallouli, K. Elghniji, O. Hentati, A.R. Ribeiro, A.M.T. Silva, M. Ksibi, UV and solarphoto-degradation of naproxen: TiO2 catalyst effect, reaction kinetics, productsidentification and toxicity assessment, J. Hazard. Mater. 304 (2016) 329–336.
[39] Y. Xiang, J. Fang, C. Shang, Kinetics and pathways of ibuprofen degradation by theUV/chlorine advanced oxidation process, Water Res. 90 (2016) 301–308.
[40] I. Michael, A. Achilleos, D. Lambropoulou, V.O. Torrens, S. Pérez, M. Petrović,D. Barceló, D. Fatta-Kassinos, Proposed transformation pathway and evolutionprofile of diclofenac and ibuprofen transformation products during (sono)photo-catalysis, Appl. Catal. B 147 (2014) 1015–1027.
[41] K.H. Hama Aziz, H. Miessner, S. Mueller, D. Kalass, D. Moeller, I. Khorshid,M.A.M. Rashid, Degradation of pharmaceutical diclofenac and ibuprofen in aqueoussolution, a direct comparison of ozonation, photocatalysis, and non-thermal plasma,Chem. Eng. J. 313 (2017) 1033–1041.
[42] J.R. Garbin, D.M.B.P. Milori, M.L. Simões, W.T.L. da Silva, L.M. Neto, Influence ofhumic substances on the photolysis of aqueous pesticide residues, Chemistry 66(2007) 1692–1698.
[43] Z. Zuo, Z. Cai, Y. Katsumura, N. Chitose, Y. Muroya, Reinvestigation of the acid–-base equilibrium of the (bi)carbonate radical and pH dependence of its reactivitywith inorganic reactants, Radiat. Phys. Chem. 55 (1999) 15–23.
[44] G.V. Buxton, C.L. Greenstock, W.P. Helman, A.B. Ross, Critical-review of rateconstants for reactions of hydrated electrons, hydrogen-atoms and hydroxyl radicals(·OH/O-·) in aqueous-solution, J. Phys. Chem. Ref. Data 17 (1988) 513–886.
[45] Z.H. Zuo, Z.L. Cai, Y. Katsumura, N. Chitose, Y. Muroya, Reinvestigation of the acid-base equilibrium of the (bi)carbonate radical and pH dependence of its reactivitywith inorganic reactants, Radiat. Phys. Chem. 55 (1999) 15–23.
[46] T. Logager, K. Sehested, J. Holcman, Rate constants of the equilibrium reactionsSO4
-·+HNO3→HNO4-+NO3
-, Radiat. Phys. Chem. 41 (1993) 539–543.[47] O.P. Chawla, R.W. Fessenden, Electron spin resonance and pulse radiolysis studies
of some reactions of peroxysulfate, J. Phys. Chem. 79 (1975) 2693–2700.
Y. Fu et al. Chemical Engineering Journal 356 (2019) 1032–1041
