馮 姝,王鴻斌,2,劉義青*,付永勝

HSO3?強化Fe3+/S2O82?降解水中雙氯芬酸

馮 姝1,王鴻斌1,2,劉義青1*,付永勝1

(1.西南交通大學(xué)地球科學(xué)與環(huán)境工程學(xué)院,四川 成都 611756；2.成都大學(xué)建筑與土木工程學(xué)院,四川 成都 610106)

采用HSO3?強化Fe3+/S2O82?降解水中雙氯芬酸(DCF),考察了溶液初始pH值,Fe3+、HSO3?和S2O82?用量,溶解氧對HSO3?/Fe3+/S2O82?體系降解DCF的影響;通過自由基淬滅實驗,識別了體系中主要的活性物種;最后,探討了DCF在該體系中的降解產(chǎn)物和轉(zhuǎn)化路徑.結(jié)果表明:HSO3?可以明顯促進(jìn)Fe3+/S2O82?對DCF的降解,初始pH 4.0時,DCF降解效果最佳.DCF的降解速率隨Fe3+或S2O82?濃度的增大而增大;適量增加HSO3?濃度可提高DCF的降解,而過量的HSO3?對DCF降解有一定抑制作用.在通入氮氣條件下,DCF去除率僅下降10.4%,無明顯的抑制作用.自由基抑制實驗表明,該體系含有SO4??、HO?和SO5??3種活性自由基,其對DCF降解的貢獻(xiàn)率分別為83.0%、12.8%和4.3%.在HSO3?/Fe3+/S2O82?降解DCF的反應(yīng)中共檢測出4種產(chǎn)物,據(jù)此提出DCF可能的轉(zhuǎn)化路徑為:羥基化、脫羧基、脫水和甲酰化反應(yīng).

雙氯芬酸；亞硫酸氫鹽；過硫酸鹽；鐵離子；硫酸根自由基

雙氯芬酸(DCF)作為新興污染藥物,在地下水和地表水中被頻繁檢測出[1-2].研究表明,痕量的DCF會對水生生物產(chǎn)生不良影響,對生態(tài)系統(tǒng)和人類健康構(gòu)成潛在威脅[3],因此,DCF的水處理技術(shù)受到越來越多的關(guān)注.近年來,基于過硫酸鹽(PS)的高級氧化技術(shù)已發(fā)展成為解決水中難降解有機污染物的有效方法.PS在室溫下具有低反應(yīng)性,通常可通過加熱[4-6]、過渡金屬[7-10]、紫外光(UV)[11-13]、堿[14-15]等方式活化,生成高反應(yīng)性物種,如硫酸根自由基(SO4??)和羥基自由基(HO?),進(jìn)一步降解水中有機污染物.通過研究9種不同過渡金屬離子對PS的激活能力,結(jié)果發(fā)現(xiàn)Ag+和Fe2+活化PS降解有機污染物效果最好[7].但是,在Fe2+/PS體系中,氧化生成的Fe3+還原為Fe2+的速率較慢,故在實際應(yīng)用過程中該體系的持續(xù)反應(yīng)能力不足.近年來,鹽酸羥胺[16]、腐殖酸[17]、抗壞血酸[18]等還原劑被引入到Fe2+/PS體系中,加速了Fe3+/ Fe2+的氧化還原,進(jìn)而提高水中有機污染物的降解效率.Fe3+活化PS的效率非常低,故擬引入亞硫酸鹽還原劑.已有研究表明,Fe3+可以和HSO3?結(jié)合產(chǎn)生硫酸根自由基和Fe2+[19],產(chǎn)生的Fe2+又可以活化PS產(chǎn)生硫酸根自由基.目前,亞硫酸鹽聯(lián)合Fe3+/PS降解水中有機污染物的研究報道較少.因此,本文系統(tǒng)地研究了HSO3?強化Fe3+/PS對水中DCF的降解,考察HSO3?、PS和Fe3+用量、溶液pH值、溶解氧等對HSO3?/Fe3+/PS降解DCF的影響;識別該體系中主要的活性自由基物種;基于UPLC-Q-TOF-MS對DCF降解產(chǎn)物的鑒定,提出了DCF在該體系中可能的轉(zhuǎn)化途徑.

1 材料與方法

1.1 試劑

雙氯芬酸(DCF,純度398%)購自阿拉丁公司;九水硝酸鐵(Fe(NO3)3·9H2O)、亞硫酸氫鈉(NaHSO3)、過硫酸鈉(Na2S2O8)、硫代硫酸鈉(Na2S2O3)、濃硫酸(H2SO4),氫氧化鈉(NaOH)均為分析純級,冰乙酸和叔丁醇為色譜純,購于成都市科隆化學(xué)品有限公司;甲醇為色譜級,購自Fisher Scientific公司;反應(yīng)液和試劑均用去離子水(18MΩ·cm)進(jìn)行配置.

1.2 氧化實驗

所有實驗均在250mL 的玻璃反應(yīng)器中進(jìn)行,溫度控制在25℃.首先將預(yù)定量的試劑(DCF、Na2S2O8、NaHSO3)添加到反應(yīng)器中,利用0.2mol/L NaOH或H2SO4盡快調(diào)節(jié)初始pH值,再加入Fe3+啟動反應(yīng).在反應(yīng)過程中,通過恒溫水浴磁力攪拌器(轉(zhuǎn)子轉(zhuǎn)速為600r/min)確保溶液完全混合.溶解氧影響實驗中,在開始反應(yīng)前,預(yù)先向反應(yīng)液中鼓入氮氣或空氣吹脫溶液15min.反應(yīng)啟動后在特定的時間間隔(0,20,40,60, 90, 120和180s)提取樣品(1mL),并立即用0.2mol/L Na2S2O3(0.05mL)淬滅.所有實驗至少重復(fù)2次,取平均值作圖.

1.3 分析方法

DCF濃度利用高效液相色譜儀(Waters 2695)進(jìn)行測定,具體參數(shù):固定相為C18柱(4.6mm× 150mm,5μm),流動相為1‰乙酸水溶液與甲醇(= 25/75)混合液,流速為1mL/min,檢測波長為276nm,柱溫30℃,進(jìn)樣體積為20μL.DCF降解產(chǎn)物采用UPLC-QTOF-MS(Waters)檢測.具體參數(shù):色譜柱為BEH C18(2.1mm×50mm,1.7μm);流動相為0.1%甲酸水溶液和乙腈混合液,采用梯度洗脫的方式:0～2min,乙腈由10%提高到30%;2～10min,乙腈提高至100%;10～13min,乙腈降至10%,流速為0.5mL/min;進(jìn)樣體積為1 μL;采用電噴霧電離,在正離子模式下采集數(shù)據(jù),掃描范圍為50～500Da.TOC濃度使用總有機碳分析儀(Elementar, vario TOC)進(jìn)行測定.溶液pH值使用PHS-3C型pH計進(jìn)行測定.DCF在HSO3?/Fe3+/PS體系中的降解符合準(zhǔn)一級反應(yīng)動力學(xué)模型,其表觀降解速率常數(shù)(obs)可由式(1)計算:

ln(C/0) = ?obs(1)

式中:obs為表觀降解速率常數(shù),s?1;為反應(yīng)時間, s;0和C分別為初始時間和指定反應(yīng)時間下DCF的物質(zhì)的量濃度,μmol/L.

2 結(jié)果和討論

2.1 HSO3?強化Fe3+/PS降解DCF

圖1 不同體系對DCF降解的影響

[DCF]0= 10μmol/L, [PS]0= 1mmol/L, [HSO3?]0= 200μmol/L, [Fe3+]0= 10μmol/L, pH0= 4.0,= 25oC

如圖1所示,在pH = 4.0條件下,單獨Fe3+、HSO3?/PS和Fe3+/PS體系均不能有效降解水中的DCF,表明Fe3+不能直接氧化DCF,且Fe3+和低濃度HSO3?對PS活化能力均較差,不能產(chǎn)生活性自由基.然而在Fe3+/PS體系中引入HSO3?,極大地促進(jìn)了DCF的降解,反應(yīng)3min 后其去除率達(dá)到91.2%,其原因主要為:1)HSO3?是一種還原劑,其加入可將Fe3+還原為Fe2+,作為電子供體的Fe2+可活化PS產(chǎn)生SO4??,使得DCF快速降解(見式(2)～(4))[20];2)Fe3+可與HSO3?形成FeSO3+配合物,該配體可通過一系列反應(yīng)使體系中產(chǎn)生SO3??、SO5??和SO4??(見式(2)～ (6)[20]),從而使得DCF降解.由圖1可見,在Fe3+/ HSO3?體系中,反應(yīng)3min后DCF的去除率可達(dá)76.5%;3)在HSO3?/Fe3+/PS體系中,Fe3+和Fe2+可以通過一系列氧化還原反應(yīng)(見式(2)～(7)[20])循環(huán)生成,從而促進(jìn)了SO4??的生成.

Fe3++ HSO3?→ FeSO3++ H+log= 2.45 (2)

FeSO3+→ Fe2 ++ SO3??= 0.19s-1(3)

Fe2++ S2O82?→ Fe3++ SO4??+ SO42?=

2.7′101L/(mol·s) (4)

SO3??+ O2→ SO5??= (1.1-2.5)′109L/(mol·s) (5)

SO5??+ HSO3?→ SO42 ?+ SO4??+ H+= 1.2′

104L/(mol·s) (6)

Fe2++ SO4??→Fe3++ SO42?= 4.6′109L/(mol·s) (7)

2.2 HSO3?/Fe3+/PS體系中活性自由基的識別

由反應(yīng)式(8)可知,SO4??可與水反應(yīng)產(chǎn)生羥基自由基[21],故在Fe3+/HSO3?/PS體系中,可能同時存在SO4??、SO3??、SO5??和HO?4種活性自由基[22-23].其中SO3??可快速與氧氣反應(yīng)轉(zhuǎn)化為SO5??,其對有機物的降解作用可忽略.為鑒定體系中降解DCF的主要活性自由基,并研究甲醇(MeOH)和叔丁醇(TBA)對DCF的抑制作用,引入值(為反應(yīng)物的物質(zhì)的量濃度,為反應(yīng)物與自由基之間的二級反應(yīng)速率常數(shù))比較反應(yīng)物之間競爭自由基的能力,見表1.TBA與HO?(= 7.6 × 108L/(mol·s)[24])的反應(yīng)速率常數(shù)遠(yuǎn)高于SO4??(= 4.0 × 105L/(mol·s)[25]),可作為HO?的有效抑制劑;而MeOH可同時作為SO4??(=2.5×107L/(mol·s)[25])和HO?(=9.7×108L/ (mol·s)[24])的抑制劑;另外,MeOH和TBA與SO5??(<103L/(mol·s)[25])的反應(yīng)均較慢,故它們不會抑制SO5??.因此,在一定程度上,TBA和MeOH的抑制實驗可驗證體系中SO4??、SO5??和HO?的存在.

SO4??+ H2O →HO?+ SO42?+ H+=1.1′101L/(mol·s) (8)

表1 反應(yīng)物和自由基的Ck值

由圖2可看出,與未添加抑制劑相比,加入10mmol/L TBA后DCF降解率降低 14.9%,只能抑制HO?,但不能淬滅SO4??;該抑制部分可能與HO?的作用有關(guān),因為10mmol/LTBA和HO?的值遠(yuǎn)大于DCF和HO?的值,而和SO4??的值則小于DCF.隨著TBA濃度增加,其與SO4??的值接近DCF與SO4??的值,故它會與DCF競爭體系中的SO4??,進(jìn)而進(jìn)一步抑制DCF的降解.10mmol/L MeOH與SO4??的值大于100mmol/L TBA與SO4??的值,故它對DCF去除的抑制作用要強于100mmol/L TBA,如圖2所示.隨著 MeOH濃度增加至 1mol/L,其與SO4??的值越來越高,其對DCF降解的抑制作用達(dá)到最大.此時,1mol/L MeOH和SO4??或HO?的值均遠(yuǎn)大于DCF和SO4??或HO?的值,但仍有4.1%的DCF被降解.這一結(jié)果表明體系中可能存在其他反應(yīng)物種(即SO5??),因為 1mol/L MeOH能夠完全淬滅SO4??和HO?.

在加入10mmol/LTBA和1mol/LMeOH的條件下,DCF的表觀反應(yīng)速率常數(shù)分別從 4.7×10?3s?1下降到4.1×10?3和2×10?4s?1.據(jù)此計算得知, SO4??、HO?和SO5??對DCF降解的貢獻(xiàn)率分別為83.0%、12.8%和4.3%[28],表明在HSO3?/Fe3+/PS體系降解DCF過程中同時存在SO4??、SO5??和HO?3種活性自由基,其中SO4??為主要活性自由基.

圖2 不同自由基淬滅劑對DCF降解的影響

[DCF]0= 10μmol/L, [PS]0= 1mmol/L, [HSO3?]0= 200μmol/L, [Fe3+]0= 10μmol/L, pH0= 4.0,= 25oC

2.3 HSO3?/Fe3+/PS體系降解DCF的影響因素

2.3.1 初始pH值對DCF降解的影響 溶液pH值對HSO3?和Fe3+的形態(tài)分布尤為重要,是影響HSO3?/ Fe3+/PS體系的關(guān)鍵性因素,例如:HSO3?在pH>7.0時轉(zhuǎn)化為SO32?的形式存在,而pH 4.0～6.0條件下主要以HSO3?的形式存在,pH<2.0時轉(zhuǎn)為SO2?H2O的形式存在.如圖3(a)所示,在pH值大于5.0條件下,DCF幾乎不被降解,此條件下Fe3+主要以Fe(OH)2+和Fe(OH)3(aq)形式存在,這2種形態(tài)的Fe3+既不能活化PS也不能活化HSO3?[29],以致體系內(nèi)無活性自由基產(chǎn)生.pH值為 4.0時,體系中存在大量的HSO3?,其與Fe3+通過式(2)形成配體物FeSO3+,該配體物經(jīng)分解后產(chǎn)生Fe2+和SO3??(式(3));進(jìn)一步生成SO5??和SO4??,使DCF快速被降解.然而,DCF的降解隨著pH值的再度降低而受到抑制,pH值為2.0時,反應(yīng)3min,僅24.5%的DCF被降解.這主要是由于在該pH值下,HSO3?轉(zhuǎn)化為SO2?H2O,降低了FeSO3+的形成和Fe2+的再生[30-31],進(jìn)而抑制了SO4??和SO5??的生成.

[DCF]0= 10μmol/L, [PS]0= 1mmol/L, [HSO3?]0= 200μmol/L, [Fe3+]0= 10μmol/L, pH0= 4.0,= 25oC

2.3.2 Fe3+濃度對DCF降解的影響 如圖3(b)所示,當(dāng)Fe3+濃度由5μmol/L增加至20μmol/L時,反應(yīng)3min,DCF的去除率由68.2%提高至100%,繼續(xù)增加Fe3+濃度至160μmol/L,DCF在90s內(nèi)完全降解.當(dāng)Fe3+用量小于20μmol/L時,反應(yīng)速率常數(shù)obs隨著Fe3+濃度提高呈線性增大,線性回歸斜率為0.0012 (2=0.99).而當(dāng)Fe3+濃度超過20μmol/L時,雖然obs仍在增加,但它與Fe3+濃度的線性增長關(guān)系消失.此現(xiàn)象可解釋為,1)Fe3+用量小于20μmol/L時,其濃度的增加加速了對PS和HSO3?的活化作用,產(chǎn)生更多的SO4??,使DCF降解更快;2)Fe3+濃度在20～ 160μmol/L時,obs增加速率變慢歸因于體系中產(chǎn)生過量的Fe2+對SO4??的淬滅作用(式(7))以及SO4??的再結(jié)合作用(式(9)).王鴻斌等[32-33]在研究Fe2+/PS和Fe2+/HSO3?體系降解DCF的過程中也都觀察到了類似的現(xiàn)象.

SO4??+ SO4??→S2O82?= (4～8.1)′108L/(mol·s) (9)

2.3.3 PS濃度對DCF降解的影響 如圖3(c)顯示,當(dāng)體系中的PS濃度由100μmol/L增加到1.5mmol/L時,DCF的去除率由82.8%增大至94.4%;同時其k和PS濃度有著較好的線性關(guān)系,線性回歸斜率為0.0093(2= 0.99).這主要是因為PS在Fe2+的活化作用下分解產(chǎn)生強氧化性的活性物種SO4??(式(4)),促進(jìn)了DCF的降解,所以PS濃度會直接影響該體系對DCF的去除.

2.3.4 HSO3?濃度對DCF降解的影響 如圖3(d),HSO3?用量從0.05mmol/L增加到0.4mmol/L時,DCF降解效率從62.0%提高到96.4%,繼續(xù)投加HSO3?用量至0.8mmol/L,DCF的去除率下降至94%.當(dāng)HSO3?投加量小于0.4mmol/L時,obs隨著HSO3?濃度的提高逐漸增加,而當(dāng)HSO3?濃度超過0.4mmol/L時,obs明顯減小.此現(xiàn)象可解釋為:1) HSO3?用量小于0.4mmol/L時,HSO3?濃度增加會導(dǎo)致FeSO3+濃度增加,進(jìn)而產(chǎn)生較多的SO4??和Fe2+,產(chǎn)生的Fe2+又可以活化S2O82?產(chǎn)生更多的SO4??,因而導(dǎo)致DCF降解速率增快;2)HSO3?濃度大于0.4mmol/L時,obs下降歸因于體系中過量的HSO3?對SO4??的淬滅作用以及SO4??的再結(jié)合作用[34-35].

2.3.5 溶解氧對DCF降解的影響 溶解氧可直接與亞硫酸根自由基反應(yīng),從而影響硫氧根自由基之間的鏈?zhǔn)椒磻?yīng),見式(5)和式(6).因此,考察溶解氧對HSO3?/Fe3+/PS體系降解DCF的影響.如圖3(e)所示,與自然攪拌條件相比,曝入空氣會促進(jìn)DCF的降解,使其在2min內(nèi)即可完全降解.這是因為曝入空氣時,反應(yīng)過程中溶液中的溶解氧濃度要高于自然攪拌條件,反應(yīng)結(jié)束后其殘留的溶解氧濃度分別為8.9,7.9mg/L,故曝氣加快了SO3??向SO4??的轉(zhuǎn)化歷程,增加了體系中SO4??的濃度.一般來說,在缺氧條件下,通過氮氣吹掃,溶液中的溶解氧少于2mg/L,會直接抑制SO3??的氧化,進(jìn)一步影響硫氧根自由基之間鏈?zhǔn)椒磻?yīng),減少SO4??的生成[20].然而在HSO3?/ Fe3+/PS體系中通入氮氣對DCF的降解并沒有較強的抑制作用,其降解率僅下降了10.4%,這可能是由于:1)缺氧條件下,HSO3?主要以還原作用為主,使體系中Fe3+不斷轉(zhuǎn)化為Fe2+,此外,Fe3+與HSO3?形成的配體物亦可分解為Fe2+和SO3??,繼而Fe2+活化PS不斷產(chǎn)生SO4??和Fe3+,故DCF仍有較好的去除率;2)生成的SO3??與苯胺類有機物的二級反應(yīng)速率常數(shù)約為106L/(mol·s)[25],DCF含有苯胺官能團,故其可能與SO3??發(fā)生反應(yīng).

2.4 DCF降解產(chǎn)物及轉(zhuǎn)化途徑的鑒定

表2 HSO3?/Fe3+/PS體系中DCF的降解產(chǎn)物

在HSO3?/Fe3+/PS降解DCF的過程中共檢測到4種反應(yīng)產(chǎn)物,其分子質(zhì)量、質(zhì)荷比()、分子式及可能的結(jié)構(gòu)見表2.基于這些產(chǎn)物提出HSO3?/ Fe3+/PS降解DCF可能的反應(yīng)機理,主要包括4種不同的降解途徑,分別為(1)羥基化反應(yīng),(2)脫羧反應(yīng),(3)脫水反應(yīng)和(4)甲?；磻?yīng),具體如圖4所示.DCF通過途徑(1)、(2)、(3)可以分別生成產(chǎn)物312、252和278.生成的產(chǎn)物252通過途徑(4)可以進(jìn)一步生成產(chǎn)物266.

羥基化反應(yīng)(途徑(1))可能是通過DCF結(jié)構(gòu)中含乙酸基苯環(huán)上短暫添加和消除硫酸根離子而發(fā)生,從而生成自由基陽離子,該自由基陽離子可通過親核反應(yīng)形成羥基化產(chǎn)物312[36].脫羧作用(途徑(2))是指DCF的側(cè)鏈上損失-COOH基團,從而導(dǎo)致252的產(chǎn)生.在脫羧反應(yīng)之后發(fā)生甲?；磻?yīng)(途徑(4)),產(chǎn)物252中的-CH3基團被氧化為-CHO,從而形成266[37].脫水反應(yīng)(途徑(3))是由DCF上的N原子和-COOH基團偏離而導(dǎo)致羧酸與仲胺分子內(nèi)酰胺化脫水而引發(fā)[38],生成產(chǎn)物278.

圖4 HSO3?/Fe3+/PS體系中DCF可能的降解途徑

2.5 HSO3?/Fe3+/PS體系對實際水體中DCF的降解

實際水體包括沱江河河水和西南交通大學(xué)犀浦校區(qū)犀湖湖水,水質(zhì)參數(shù)如表3所示.由圖5(a)可知,在未調(diào)節(jié)pH值的實際水體中,DCF幾乎不被降解,這可能主要是由于pH值的影響.如前文所述,在pH值>5時,DCF幾乎不被降解,而河水和湖水的pH值分別為7.6和7.7,在此pH值條件下,Fe3+主要以Fe(OH)3形式存在,此形態(tài)的Fe3+既不能活化PS也不能活化HSO3?,以致體系內(nèi)無活性自由基產(chǎn)生,故DCF不被降解.將實際水體pH值調(diào)整為4.0后, HSO3?/Fe3+/PS體系能夠降解實際水體中的DCF,但與純水中的相比,DCF在實際水體中的去除呈現(xiàn)抑制作用,且河水的抑制作用明顯強于湖水.由表3可知,河水的堿度、UV254和CODcr均高于湖水,表明河水中的HCO3?和溶解性有機物濃度均高于湖水,而這兩種物質(zhì)能夠與體系中的活性自由基反應(yīng),從而抑制DCF的降解,故DCF在河水中的去除要弱于湖水,更弱于純水.此外,本文也研究了DCF在純水和實際水體中的礦化,結(jié)果如圖5(b)所示.在反應(yīng)60min后,純水中的總有機碳(TOC)衰減了11.6%,表明DCF在HSO3?/Fe3+/PS體系中不能被完全礦化,DCF在自由基的攻擊下首先被分解為其他中間產(chǎn)物,這些中間產(chǎn)物的降解礦化可能需要更長的反應(yīng)時間或者更高的氧化劑用量[36].而河水和湖水中TOC分別降解了17.5%和12.5%,高于純水,這可能是因為實際水體中本身含有溶解性有機物,這些有機物在自由基的攻擊下被氧化降解,從而導(dǎo)致實際水體的礦化率更高.

表3 實際水體水質(zhì)參數(shù)

[DCF]0= 10μmol/L, [HSO3?]0= 200μmol/L, [Fe3+]0= 10μmol/L, [PS]0= 1mmol/L,= 25oC

3 結(jié)論

3.1 HSO3?可明顯強化Fe3+/S2O82?體系對DCF的降解,且pH值為 4.0時,HSO3?/Fe3+/PS對DCF的去除效果最優(yōu),但礦化效果不佳,反應(yīng)60min后TOC僅降解11.6%.

3.2 DCF的降解速率隨Fe3+或PS濃度的增大而增大,且與PS濃度成線性正相關(guān);適量增加HSO3?的濃度可提高DCF的降解,而過量的HSO3?對DCF降解有一定抑制作用.在通入氮氣條件下,DCF降解率僅下降10.4%,并無明顯的抑制作用.

3.3 自由基抑制實驗表明,HSO3?/Fe3+/S2O82?體系含有SO4??、SO5??和HO?3種活性自由基,其中SO4??為主要活性自由基.該體系降解DCF的反應(yīng)中共檢出4種產(chǎn)物,由此提出DCF可能的轉(zhuǎn)化路徑,分別為羥基化反應(yīng)、脫羧基反應(yīng)、脫水反應(yīng)和甲酰化反應(yīng).

3.4 由于實際水體的pH值在中性或者弱堿性,故HSO3?/Fe3+/S2O82?體系對實際水體中的DCF幾乎無降解.

[1] Halling S B, Nielsen S N, Lanzky P, et al. Occurrence fate and effects of pharmaceutical substances in the environment-A review [J]. Chemosphere, 1998,36(2):357-393.

[2] Ebele A J, Abdallah M A, Harrad S. Pharmaceuticals and personal care products (PPCPs) in the freshwater aquatic environment [J]. Emerging Contaminants, 2017,3:1-16.

[3] Letzel M, Metzner G, Letzel T. Exposure assessment of the pharmaceutical diclofenac based on long-term measurements of the aquatic input [J]. Environment International, 2009,35(2):363-368.

[4] Waldemer R H, Tratnyek P G, Johnson R L, et al. Oxidation of chlorinated ethenes by heat-activated persulfate: kinetics and products [J]. Environmental Science & Technology, 2007,41(3):1010-1015.

[5] Huang K C, Couttenye R A, Hoag G E. Kinetics of heat-assisted persulfate oxidation of methyl tert-butyl ether (MTBE) [J]. Chemosphere, 2002,49(4):413-420.

[6] 鐘 敏,李 孟,盧 芳,等.熱活化過硫酸鹽聯(lián)合混凝處理微乳濁液的機理 [J]. 中國環(huán)境科學(xué), 2021,41(2):704-712.

Zhong M, Li M, Lu F, et al. Mechanism of thermal-activated persulfate combined with coagulation in the treatment of microemulsion [J]. China Environmental Science, 2021,41(2):704- 712.

[7] Anipsitakis G P, Dionysiou D D. Radical generation by the interaction of transition metals with common oxidants [J]. Environmental Science & Technology, 2004,38(13):3705-3712.

[8] Liang C J, Bruell C J, Marley M C, et al. Persulfate oxidation for in situ remediation of TCE. I. Activated by ferrous ion with and without a persulfate–thiosulfate redox couple [J]. Chemosphere, 2004,55(9): 1213-1223.

[9] Zhang N, Kong X, Zhang M, et al. Study on treatment of methyl- orange in water by derivable oxidation of peroxydisulfate [J]. Journal of Advanced Oxidation Technologies, 2007,10(1):193-195.

[10] 劉美琴,宋秀蘭.Fe2+激活過硫酸鹽耦合活性炭深度處理焦化廢水 [J]. 中國環(huán)境科學(xué), 2018,38(4):1377-1384.

Liu M Q, Song X L. Advanced treatment of bio-treated coking wastewater by coupling of ferrous-activated persulfate oxidation and activated carbon adsorption [J]. China Environmental Science, 2018, 38(4):1377-1384.

[11] Mark G, Schuchmann M N, Schuchmann H, et al. The photolysis of potassium peroxodisulphate in aqueous solution in the presence of tert-butanol: a simple actinometer for 254nm radiation [J]. Journal of Photochemistry and Photobiology A: Chemistry, 1990,55(2):157-168.

[12] Gao Y, Gao N, Deng Y, et al. Ultraviolet (UV) light-activated persulfate oxidation of sulfamethazine in water [J]. Chemical Engineering Journal, 2012,195-196:248-253.

[13] 黃麗坤,李 哲,王廣智,等.紫外催化過硫酸鹽深度處理垃圾焚燒廠滲濾液 [J]. 中國環(huán)境科學(xué), 2021,41(1):161-168.

Huang L K, Li Z, Wang G Z, et al. Advanced treatment of landfill leachate by ultraviolet catalytic persulfate [J]. China Environmental Science, 2021,41(1):161-168.

[14] Liang C, Su H W. Identification of sulfate and hydroxyl radicals in thermally activated persulfate [J]. Industrial & Engineering Chemistry Research, 2009,48(11):5558-5562.

[15] Furman O S, Teel A L, Watts R J. Mechanism of base activation of persulfate [J]. Environmental Science & Technology, 2010,44(16): 6423-7428.

[16] 鄒 景.羥胺對Fe2+/過硫酸鹽體系的強化效能與機理研究 [D]. 哈爾濱:哈爾濱工業(yè)大學(xué), 2016.

ZHOU J. Enhanced effectiveness and mechanism of Fe2+/persulfate system with hydroxylamine [D]. Harbin: Harbin Institute of Technology, 2016.

[17] Li D W, Chen D Z, Yao Y Y, et al. Strong enhancement of dye removal through addition of sulfite to persulfate activated by a supported ferric citrate catalyst [J]. Chemical Engineering Journal, 2016,288:806-812.

[18] Curtin M A, Taub I A, Kustin K, et al. Ascorbate-induced oxidation of formate by peroxodisulfate: product yields, kinetics and mechanism [J]. Research on Chemical Intermediates, 2004,30(6):647-661.

[19] Wang S X, Wang G S, Fu Y S, et al. A simple Fe3+/bisulfite system for rapid degradation of sulfamethoxazole [J]. RSC Advances, 2020,10(50):30162-30168.

[20] Liu Z Z, Guo Y Z, Shang R, et al. A triple system of Fe(III)/sulfite/ persulfate: Decolorization and mineralization of reactive Brilliant Red X-3B in aqueous solution at near-neutral pH values [J]. Journal of the Taiwan Institute of Chemical Engineers, 2016,68:162-168.

[21] Herrmann H, A Reese, R Zellner. Time-resolved UV/VIS diode array absorption spectroscopy of SOx??(= 3, 4, 5) radical anions in aqueous solution [J]. Journal of Molecular Structure, 1995,348:183-186.

[22] Xu J, Ding W, Wu F, et al. Rapid catalytic oxidation of arsenite to arsenate in an iron(III)/sulfite system under visible light [J]. Applied Catalysis B: Environmental, 2016,186(5):56-61.

[23] Chen L, Peng X, Liu J, et al. Decolorization of orange II in aqueous solution by an Fe (II)/sulfite system: replacement of persulfate [J]. Industrial & Engineering Chemistry Research, 2012,51(42):13632-13638.

[24] Buxton G V, Greenstock C L, Helman W P, et al. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (?OH/O??) in aqueous solution [J]. Journal of Physical and Chemical Reference Data, 1988,17(2):513-886.

[25] Neta P, Huie R E, Ross A B. Rate constants for reactions of inorganic radicals in aqueous solution [J]. Journal of Physical and Chemical Reference Data, 1988,17(3):1027-1284.

[26] Huber M M, Canonica S, Park G Y, et al. Oxidation of pharmaceuticals during ozonation and advanced oxidation processes [J]. Environmental Science & Technology, 2003,37(5):1016-1024.

[27] Ahmed M M, Barbati S, Doumenq P, et al. Sulfate radical anion oxidation of diclofenac and sulfamethoxazole for water decontamination [J]. Chemical Engineering Journal, 2012,197:440-447.

[28] Liu Y Q, He X X, Fu Y S, et al. Degradation kinetics and mechanism of oxytetracycline by hydroxyl radical-based advanced oxidation processes [J]. Chemical Engineering Journal, 2016,284:1317-1327.

[29] Kolthoff I M, Miller I K. The chemistry of persulfate. I. The kinetics and mechanism of the decomposition of the persulfate ion in aqueous medium1 [J]. Journal of the American Chemical Society, 1951,73(7): 3055-3059.

[30] Graedel T E, Weschler C J. Chemistry within aqueous atmospheric aerosols and raindrops [J]. Reviews of Geophysics, 1981,19(4):505- 539.

[31] Grgi?c I, Pozni?c M, Bizjak M. S(IV) autoxidation in atmospheric liquid water: The role of Fe(II) and the effect of oxalate [J]. Journal of Atmospheric Chemistry, 1999,33(1):89-102.

[32] 王鴻斌,王 群,劉義青,等.亞鐵活化過硫酸鹽降解水中雙氯芬酸鈉 [J]. 環(huán)境化學(xué), 2020,39(4):869-875.

Wang H B, Wang Q, Liu Y Q, et al. Degradation of diclofenac by ferrous activated persulfate [J]. Environmental Chemistry, 2020, 39(4):869-875.

[33] Wang H B, Wang S X, Liu Y Q, et al. Degradation of diclofenac by Fe(II)-activated bisulfite: Kinetics, mechanism and transformation products [J]. Chemosphere, 2019,237:124518.

[34] Huie R E, Neta P. Rate constants for some oxidations of S(IV) by radicals in aqueous solutions [J]. Atmospheric Environment, 1987, 21(8):1743-1747.

[35] Huie R E, Clifton C L, Altstein N. A pulse radiolysis and flash photolysis study of the radicals SO2??, SO3??, SO4??and SO5??[J]. International Journal of Radiation Applications and Instrumentation Part C Radiation Physics and Chemistry, 1989,33(4):361-370.

[36] Liu Y Q, He X X, Fu Y S, et al. Kinetics and mechanism investigation on the destruction of oxytetracycline by UV-254nm activation of persulfate [J]. Journal of Hazardous Materials, 2016,305:229-239.

[37] Pérez-Estrada L A, Malato S, Gernjak W, et al. Photo-Fenton degradation of diclofenac: identification of main intermediates and degradation pathway [J]. Environmental Science & Technology, 2005,39(21):8300-8306.

[38] Zhou T, Feng K, Xiang W, et al. Rapid decomposition of diclofenac in a magnetic field enhanced zero-valent iron/EDTA Fenton-like system [J]. Chemosphere, 2018,193:968-977.

Bisulfite enhanced degradation of diclofenac in Fe3+/persulfate system.

FENG Shu1, WANG Hong-Bin1,2, LIU Yi-Qing1*, FU Yong-Sheng1

(1.Faculty of Geosciences and Environmental Engineering, Southwest Jiaotong University, Chengdu 611756, China；2.School of Architecture and Civil Engineering, Chengdu University, Chengdu 610106, China)., 2021,41(6)：2677～2684

The degradation of diclofenac (DCF) by bisulfite enhanced Fe3+/persulfate system was investigated. The influence of pH, Fe3+dosage, HSO3?dosage, persulfate (PS) dosage and dissolved oxygen on DCF degradation in HSO3?/Fe3+/S2O82?system was explored. The main reactive radical species for DCF removal in this system was also identified by scavenging experiments. Finally, The degradation products and transformation mechanism of DCF by HSO3?/Fe3+/S2O82?were evaluated. DCF could be effectively degraded by the introduction of HSO3?in Fe3+/PS process, and the optimal pH was 4.0. The increased initial Fe3+, HSO3?or PS concentration promoted DCF degradation while excessive HSO3?could inhibit its degradation by acting as a SO4??scavenger. The degradation rate of DCF was only reduced by 10.4% with bubbling nitrogen, and there was no obvious inhibitory effect in this system. According to the radical scavenging experiments, the contribution of SO4??, HO?and SO5??to DCF degradation in HSO3?/Fe3+/S2O82?system were calculated to be 83.0%, 12.8% and 4.3%, respectively. Four transformation products were detected using UPLC-Q-TOF-MS. The potential degradation mechanism of DCF was thus proposed showing four reaction pathways including hydroxylation, decarboxylation, dehydration and formylation.

diclofenac；bisulfite；persulfate；ferric ion；sulfate radical

X703

A

1000-6923(2021)06-2677-08

2020-11-04

四川省科技廳重大專項(2018SZDZX0026);中央高校基本科研業(yè)務(wù)費科技創(chuàng)新項目(2682018CX32)

* 責(zé)任作者, 講師, liuyq@swjtu.edu.cn

馮 姝(1996-),女,四川廣元人,西南交通大學(xué)碩士研究生,研究方向為水污染控制.