崔 航,傅曉日,顧小鋼,2,呂樹光*,邱兆富,隋 倩

?

二價(jià)鐵催化過碳酸鈉處理水中乙苯

崔 航1,傅曉日1,顧小鋼1,2,呂樹光1*,邱兆富1,隋 倩1

(1.華東理工大學(xué),國(guó)家環(huán)境保護(hù)化工過程環(huán)境風(fēng)險(xiǎn)評(píng)價(jià)與控制重點(diǎn)實(shí)驗(yàn)室,上海市功能性材料化學(xué)重點(diǎn)實(shí)驗(yàn)室,上海 200237；2.華東理工大學(xué),化學(xué)工程聯(lián)合國(guó)家重點(diǎn)實(shí)驗(yàn)室,上海 200237)

采用Fe(II)催化過碳酸鈉(SPC)體系處理水溶液中的乙苯(EB),考察(SPC)、(Fe(II))、陰離子濃度、天然有機(jī)物(NOM)以及溶液初始pH值對(duì)EB降解效果的影響,并確定EB降解過程中起主導(dǎo)作用的自由基.結(jié)果表明,對(duì)于濃度為1mmol/L的EB溶液,(SPC)、(Fe(II))均為12mmol/L時(shí),20min內(nèi)EB可被完全去除; Cl-、HCO3-、NOM的存在均會(huì)抑制EB降解,SO42-和NO3-對(duì)EB降解無影響;溶液初始pH值(pH3.00~11.00) 越高,EB去除率越低,但當(dāng)pH=9時(shí),降解效果仍很顯著,表明該體系能夠在較寬pH值范圍內(nèi)高效降解水溶液中EB;自由基探針試驗(yàn)證實(shí)體系中存在?OH和O2?-,自由基清掃試驗(yàn)表明?OH對(duì)EB降解起主導(dǎo)作用.

乙苯；過碳酸鈉；羥基自由基；地下水修復(fù)

苯系物(包括苯,甲苯,乙苯,二甲苯,BTEX)作為原油的重要組成部分,在工業(yè)生產(chǎn)中常用作工業(yè)溶劑,有機(jī)合成原料以及設(shè)備清洗劑等[1-2].然而,由于人們的不當(dāng)操作,如未經(jīng)處理廢水的排放,燃油泄漏事故等,導(dǎo)致其成為污染場(chǎng)地土壤及地下水中廣泛存在的有機(jī)污染物之一[3-4].乙苯(EB)作為BTEX的一種,對(duì)生物具有遺傳毒性和致癌性[2,5].

受污染的地下水有多種修復(fù)方法:自然衰減,異位修復(fù)和原位修復(fù)等.其中,原位修復(fù)因具有高效性和易操作性而被廣泛采用.一些化學(xué)氧化試劑被用于地下水的原位修復(fù)過程中,如臭氧[6]、高錳酸鉀[7]、過硫酸鹽[8]以及Fenton試劑[9]等.其中Fenton試劑具有氧化能力強(qiáng),成本較低,對(duì)環(huán)境無害以及處理范圍廣等特點(diǎn)而被廣泛研究和應(yīng)用.在Fe(II)的催化作用下,過氧化氫會(huì)產(chǎn)生×OH(E=2.73V)等活性氧自由基[10].×OH具有很強(qiáng)的、無選擇性的氧化能力,可以快速有效降解芳香烴類化合物[9].通常情況下, Fenton試劑反應(yīng)的最適pH值為2~4,然而地下水pH值呈中性或弱堿性,因此, Fenton試劑在地下水原位修復(fù)的應(yīng)用中受到一定限制.

過碳酸鈉(2Na2CO3×3H2O2, SPC)被廣泛用作漂白劑、清洗劑、去污劑等.SPC溶于水中,即分解為過氧化氫和碳酸鈉.因此,SPC擁有和過氧化氫類似的氧化性質(zhì).同時(shí),SPC價(jià)格低廉,無毒,自身及分解產(chǎn)物對(duì)環(huán)境無污染,作為類Fenton的氧化劑,適用的pH值范圍更廣[11-12].近年來,作為過氧化氫的固體載體,已有研究將SPC作為地下水原位修復(fù)的氧化劑[13].

四種苯系物雖然性質(zhì)相似,但在實(shí)際降解過程中仍存在個(gè)體差異[14].本研究通過改變Fe(II)和SPC的濃度比例來考察Fe(II)和SPC的劑量對(duì)EB降解的影響,分析地下水中所含Cl-、HCO3-、SO42-、NO3-和天然有機(jī)物(NOM)以及溶液初始pH值對(duì)EB降解的影響,并用自由基清掃試驗(yàn)驗(yàn)證起主導(dǎo)作用的活性氧自由基(ROS),以期為-Fe(II)體系修復(fù)BTEX污染地下水提供技術(shù)支撐.

1 材料與方法

1.1 試劑及儀器

試劑:乙苯、過碳酸鈉、硫酸亞鐵、四氯化碳、異丙醇、甲醇、硝基苯、三氯甲烷、硝酸鈉、氯化鈉、硫酸鈉、碳酸氫鈉、碳酸鈉、腐殖酸,以上試劑均為分析純.試驗(yàn)中用水均為自制超純水.

儀器:氣相色譜儀(6890N,美國(guó)安捷倫科技有限公司)、磁力攪拌器(85-2,上海閔行虹浦儀器廠)、低溫恒溫槽(SDC-6,寧波新芝生物科技股份有限公司)、pH計(jì)(DELTA320,瑞士Mettler- Toledo集團(tuán))、超純水儀(Classic DI,英國(guó)ELGA公司).

1.2 試驗(yàn)方法

試驗(yàn)均在內(nèi)徑為6.0cm、高9.0cm,有效容積為250mL帶夾層的玻璃反應(yīng)器中進(jìn)行.采用低溫恒溫槽控制反應(yīng)溫度為20℃,用磁力攪拌器控制溶液攪拌轉(zhuǎn)速為600r/min.依次向反應(yīng)器中加入250mL濃度為1mmol/L的EB、所需量的硫酸亞鐵和過碳酸鈉并開始計(jì)時(shí).既定時(shí)間點(diǎn)取2.5mL樣品加入盛有1mL純甲醇溶液的頂空瓶中(甲醇用于終止反應(yīng)進(jìn)行),并迅速密封以便進(jìn)行樣品分析.除分析溶液初始pH值影響的試驗(yàn),溶液pH值均不做調(diào)節(jié).每組試驗(yàn)設(shè)2個(gè)平行樣,結(jié)果取平均值.

1.3 分析方法

溶液中的(EB)采用頂空進(jìn)樣的分析.氣相色譜條件:自動(dòng)進(jìn)樣器(COMBI-PLA, CTC, Switzerland)、HP-5色譜柱(長(zhǎng)30m、內(nèi)徑320μm、膜厚0.25μm)、FID檢測(cè)器.柱溫:60℃(保持6min);進(jìn)樣口與檢測(cè)器溫度分別為150,250℃;載氣(N2)流速:3.0mL/min;進(jìn)樣量:500μL.

頂空條件:加熱溫度50℃,加熱時(shí)間6min,注射溫度60℃,震蕩速度500r/min.

2 結(jié)果與討論

2.1 Fe(II)催化SPC降解EB的有效性及SPC和Fe(II)濃度對(duì)EB降解效果影響

在SPC-Fe(II)體系中降解初始濃度為的EB,初始pH=6.3,溫度為20℃,結(jié)果如圖1.由圖1可知,當(dāng)體系中沒有SPC和Fe(II)時(shí),EB的去除率非常小(<3%).表明在試驗(yàn)的整個(gè)階段,EB的揮發(fā)損失很小,可以忽略.當(dāng)(SPC)/(Fe(II))/(EB)=2:2:1時(shí),EB的去除率可達(dá)到67%.并且EB的去除率隨著(SPC)/(Fe(II))/(EB)的增加而增大.當(dāng)(SPC)/(Fe(II))/(EB)=12:12:1時(shí),EB未檢出.說明在SPC-Fe(II)體系中可以通過增加濃度比例來完全去除EB.與此同時(shí),在反應(yīng)的前2min,EB的降解速度非?？?2min之后,EB基本無降解.這是由于在初始階段, SPC和Fe(II)快速反應(yīng)產(chǎn)生大量的羥基自由基(×OH) [式(1)和(2)],將EB快速降解.然而,Fe(II)很快被SPC氧化成為Fe(III),在pH>3(本試驗(yàn)中1min后,pH4.8~5.6)的條件下,Fe(III)即開始以氫氧化鐵的形式沉淀,并難以轉(zhuǎn)化為Fe(II),從而抑制了氧化反應(yīng)的進(jìn)行[15].

2Na2CO3×3H2O2→ 2Na2CO3+ 3H2O2(1)

H2O2+ Fe2+→×OH + OH-+ Fe3+(2)

為了評(píng)估(SPC)對(duì)EB(初始濃度為1mmol/L)降解效果的影響,將(Fe(II))設(shè)定為4,分別加入不同劑量的SPC(0,1,2,5, 10,20mmol/L),如圖2所示,當(dāng)只有Fe(II)存在時(shí),EB無降解.說明單獨(dú)Fe(II)不能降解EB,隨著(SPC)從1mmol/L增加到2mmol/L,EB的去除率由73.3%增加為94.4%,說明SPC濃度的增加有助于EB的快速降解.然而隨著(SPC)的進(jìn)一步升高,EB的去除率反而明顯下降.如當(dāng)(SPC)= 20時(shí),EB去除率只有64.7%.研究表明,(SPC)過高時(shí),在水中分解產(chǎn)生的過量H2O2可以消耗×OH[式(3)],從而降低了×OH對(duì)污染物的氧化降解能力[16].

×OH + H2O2→ HO2?+ H2O

H2O2,×OH= 2.7×107L/(mol×s) (3)

(SPC)增加時(shí),(CO32-)也會(huì)隨之增加,當(dāng)(SPC)=3mmol/L時(shí),反應(yīng)終點(diǎn)pH=3.7,當(dāng)(SPC)= 10mmol/L時(shí),反應(yīng)終點(diǎn)pH=9.6,EB降解受到抑制.通過考察(CO32-)對(duì)Fe(II)催化H2O2降解乙苯的影響(表1,以H2O2代替SPC,并加入不同濃度的碳酸鈉),發(fā)現(xiàn)隨著(CO32-)的增大,溶液pH值升高,EB降解率降低,說明(SPC)增加時(shí)所釋放出的大量CO32-也是抑制EB降解的一個(gè)重要因素.

表1 不同c(CO32-)條件下,Fe(II)催化H2O2降解EB效果Table 1 EB degradation by Fe(II) activated H2O2process under various CO32-concentration

注:反應(yīng)中(H2O2)=(Fe(II))=3mmol/L,反應(yīng)時(shí)間為60min.

考察(Fe(II))對(duì)EB(初始濃度為)降解效果的影響時(shí),固定(SPC)=4mmol/L并加入不同劑量的Fe(II)(0,2,3,5,10,20).如圖3所示,當(dāng)不加入Fe(II)時(shí),EB無降解.說明SPC只有在Fe(II)的催化作用下才能氧化降解有機(jī)污染物.隨著(Fe(II))升高,EB去除率顯著提升.當(dāng)(Fe(II))=5mmol/L時(shí),EB去除率為99.5%.隨著(Fe(II))進(jìn)一步升高,EB去除率出現(xiàn)微弱降低的趨勢(shì).這是由于過量的Fe(II)也會(huì)對(duì)?OH產(chǎn)生清掃作用(式4)[17].

Fe2++×OH → Fe3++ HO-(4)

2.2 無機(jī)陰離子和天然有機(jī)物對(duì)SPC-Fe(II)體系降解EB的影響

無機(jī)陰離子(Cl-、HCO3-、SO42-和NO3-)對(duì)SPC-Fe(II)體系降解EB的影響 考察地下水中常見陰離子(Cl-、HCO-、SO42-和NO-)對(duì)SPC-Fe(II)體系降解EB的影響.分別配制初始濃度為1mmol/L的EB和不同濃度(1,10, 100mmol/L)各陰離子的混合溶液,-Fe(II)體系下反應(yīng),其中(SPC)和(Fe(II))均為3mmol/L.結(jié)果如圖4.由圖4可知,SO42-和NO3-的存在對(duì)乙苯的降解無影響.當(dāng)(Cl-)為1mmol/L和10mmol/L時(shí),EB的降解情況基本無變化.而當(dāng)(Cl-)為100mmol/L時(shí),EB的降解效率有所降低.這是因?yàn)榇罅緾l-的存在,可消耗×OH.反應(yīng)如式5~8[18].另有研究認(rèn)為在反應(yīng)過程中生成了含氯的Fe(III)絡(luò)合物,從而影響了Fe(III)轉(zhuǎn)化為Fe(II),抑制了氧化反應(yīng)的進(jìn)行[19-20].并且在酸性條件下,Cl-對(duì)氧化反應(yīng)的抑制作用更加顯著[21].

Cl-+?OH → ClOH×-= 4.3×109L/(mol×s) (5)

ClOH×-+ H+→ HClOH×

= 3.0×1010L/(mol×s) (6)

ClOH×→ Cl×+ H2O= 5.0×104L/(mol×s) (7)

HClOH×+ Cl-→ Cl2-+ H2O

= 8.0×109L/(mol×s) (8)

與Cl-相比,HCO3-對(duì)EB降解效果的抑制作用更明顯.如圖4,沒有HCO3-時(shí),乙苯的去除率約為82.3%,而當(dāng)HCO3-濃度為1mmol/L時(shí),乙苯去除率降低為68.5%.HCO3-濃度為10,100mmol/L時(shí),乙苯的降解率只有11.6%和8.6%.這是由于HCO3-存在時(shí),溶液的初始pH值會(huì)發(fā)生很大的變化.HCO3-濃度越高,pH值越高,高pH值會(huì)使Fe(II)在初始階段便以氫氧化亞鐵的形式沉淀下來,導(dǎo)致催化劑含量降低,從而抑制Fenton反應(yīng)的進(jìn)行.此外,HCO3-與×OH的反應(yīng)速率常數(shù)為8.5×106L/(mol×s)[22],因此對(duì)×OH有較強(qiáng)的清掃效果(式9,10)[23-24].

HCO3-+×OH → HCO3×+ HO-

= 8.5×106L/(mol×s)(9)

CO32-+×OH → CO3-×+ HO-

= 4.2×108L/(mol×s)(10)

2.2.2 天然有機(jī)物(NOM)對(duì)SPC-Fe(II)體系降解EB的影響 除了一些陰離子外,地下水中還存在有大量天然有機(jī)物(NOM).腐殖酸(HA)是一種常見的自然有機(jī)物,本試驗(yàn)以HA為代表考察NOM對(duì)反應(yīng)的影響.配制不同濃度(1,10, 100mg/L)的HA和1mmol/L EB的混合溶液.在-Fe(II)體系下反應(yīng),其中(SPC)和(Fe(II))均為3mmol/L.從圖5中可以看出,低濃度(1,10mg/L)的HA對(duì)EB的降解幾乎無影響.(HA)為100mg/L時(shí),EB的去除率有所降低.這可能是由于HA可以與×OH反應(yīng),大量HA的存在,與EB的氧化產(chǎn)生競(jìng)爭(zhēng),因此降低了EB的去除率.值得一提的是,Wang等[25]發(fā)現(xiàn),在Fenton反應(yīng)中,HA不僅可以爭(zhēng)奪×OH,而且可以阻止Fe(III)再生為Fe(II),從而減少×OH的生成,因此降低目標(biāo)污染物的去除率.

2.3 溶液初始pH值對(duì)EB降解效果的影響

采用0.1mol/L NaOH或0.1mol/L H2SO4將1mmol/L乙苯溶液的pH值分別調(diào)節(jié)至3.0,7.0, 9.0,11.0,空白條件的pH值為6.3,(SPC)和(Fe(II))均為3mmol/L.EB在不同pH值條件下的降解效果如圖6所示.從圖6可以看出,隨著pH值的升高,EB的去除率不斷下降.在pH=3時(shí),乙苯的去除率最高可達(dá)96.8%.Fenton反應(yīng)的最佳pH值為2~4,在此pH值下可產(chǎn)生最大量的?OH[26].因此在pH=3的條件下,達(dá)到很好的降解效果.當(dāng)初始pH值升高為9時(shí),和不調(diào)節(jié)pH值時(shí)EB的降解效果相比有所下降,但不明顯.因此也說明了SPC-Fe(II)體系降解EB可以在更大的pH值范圍內(nèi)進(jìn)行.

2.4 EB在SPC-Fe(II)體系中的降解機(jī)制

自由基探測(cè)試驗(yàn) SPC在Fe(II)催化作用下產(chǎn)生活性氧自由基是一個(gè)十分復(fù)雜的過程.在傳統(tǒng)的Fenton反應(yīng)中,×OH被認(rèn)為在氧化有機(jī)物的過程中起主要作用.但同時(shí)O2?-,HO2?等也被證明對(duì)有機(jī)物的降解有一定貢獻(xiàn)[27-28].利用自由基探測(cè)試驗(yàn)證明SPC-Fe(II)體系中存在×OH和O2?-.硝基苯(NB)作為氧化劑探針,和×OH有非常高的反應(yīng)性(×OH=3.9×109L/(mol×s)).四氯化碳(CT)很容易被還原劑降解,如O2?-(e=1.6×1010L/(mol×s)),而不與×OH反應(yīng)(×OH<2.0×106L/(mol×s))[19].在本試驗(yàn)中,配制濃度分別為2,0.05mmol/L的NB和CT溶液,(SPC)和(Fe(II))均為12mmol/L.降解結(jié)果如圖7所示,NB和CT均有明顯降解,說明在SPC-Fe(II)體系中存在×OH和O2?-.

2.4.2 自由基清掃確定起主導(dǎo)作用的活性氧自由基 為進(jìn)一步驗(yàn)證EB降解過程中起主導(dǎo)作用的自由基,考察了不同自由基清掃劑對(duì)1mmol/L EB降解效果的影響.異丙醇(IPA)和×OH具有很高的反應(yīng)速率常數(shù)(×OH=3.0×109L/(mol×s)),而和O2?-的反應(yīng)速率常數(shù)很低(e=1.0×106L/(mol×s)).相反,三氯甲烷(CF)和O2?-具有很高的反應(yīng)速率常數(shù)(e=3.0×1010L/(mol×s)),而與×OH基本不反應(yīng)(×OH= 7.0×106L/(mol×s))[29].因此,選擇IPA(65mmol/L)和CF(65mmol/L)分別作為×OH和O2?-的清掃劑.各組試驗(yàn)(SPC)/(Fe(II))/(EB)均為3:3:1.結(jié)果如圖8所示.與對(duì)照組相比,分別加入IPA和CF后,EB去除率由84.6%分別降低為74.5%和15.4%.表明:在SPC-Fe(II)體系中,與O2?-相比,×OH在EB降解過程中起主導(dǎo)作用.

3 結(jié)論

3.1 SPC-Fe(II)體系能夠有效降解EB,對(duì)于初始濃度為1mmol/L的EB溶液,(SPC)、(Fe(II))均為12mmol/L時(shí),20min內(nèi)EB可被完全去除.

3.2 SPC和Fe(II)過量或少量均會(huì)抑制EB降解.因此,在修復(fù)被EB污染的地下水過程中,需要根據(jù)水質(zhì)情況調(diào)配適當(dāng)?shù)谋壤赃_(dá)到最佳的修復(fù)效果.

3.3 水中Cl-、HCO3-、NOM的存在均會(huì)抑制EB降解.溶液初始pH值越高,EB在SPC-Fe(II)體系中的去除率越小.溶液初始pH=9時(shí),EB仍有較高的去除率.

3.4 SPC-Fe(II)體系中,與O2?-相比,×OH在EB降解過程中起主導(dǎo)作用.

-10- 04.

[2] ICPS. Environmental Health Criteria 186: Ethylbenzene. World Health Organization/International Programme on Chemical Safety,http://www.inchem.org/documents/ehc/ehc/ehc186.htm, 2015-10-04.

[3] 楊明星,楊悅鎖,杜新強(qiáng),等.石油污染地下水有機(jī)污染組分特征及其環(huán)境指示效應(yīng) [J]. 中國(guó)環(huán)境科學(xué), 2013,33(6):1025-1032.

[4] 劉玉蘭,程莉蓉,丁愛中,等.NAPL泄漏事故場(chǎng)地地下水污染風(fēng)險(xiǎn)快速評(píng)估與決策 [J]. 中國(guó)環(huán)境科學(xué), 2011,31(7):1219-1224.

[5] 王延讓,楊德一,張 明.乙苯遺傳毒性的研究概述 [J]. 中華勞動(dòng)衛(wèi)生職業(yè)病雜志, 2007,25:702-704.

[6] Sunder M, Hempel D C. Oxidation of tri- and perchloroethene in aqueous solution with ozone and hydrogen peroxide in a tube reactor [J]. Water Research, 1997,31(1):33-40.

[7] Waldemer R H, Tratnyek P G. Kinetics of contaminant degradation by permanganate [J]. Environmental Science & Technology, 2006,40(3):1055-1061.

[8] Xu M, Gu X, Lu S, et al. Role of Reactive Oxygen Species for 1,1,1-Trichloroethane Degradation in a Thermally Activated Persulfate System [J]. Industrial & Engineering Chemistry Research, 2014,53(3):1056-1063.

[9] Lee Y, Lee W. Degradation of trichloroethylene by Fe(II) chelated with cross-linked chitosan in a modified Fenton reaction [J]. Journal of Hazardous Materials, 2010,178(4):187–193.

[10] Pignatello J J, Oliveros E, MacKay A. Advanced Oxidation Processes for Organic Contaminant Destruction Based on the Fenton Reaction and Related Chemistry [J]. Critical Reviews in Environmental Science & Technology, 2006,36(1):1-84.

[11] Zhang Y H, Xue C M, Guo C H. Application Sodium Percarbonate to Oxidative Degradation Trichloroethylene Contamination in Groundwater [J]. Procedia Environmental Sciences, 2011,10:1668–1673.

[12] Miao Z, Gu X, Lu S, et al. Perchloroethylene (PCE) oxidation by percarbonate in Fe2+-catalyzed aqueous solution: PCE performance and its removal mechanism [J]. Chemosphere, 2015, 119:1120–1125.

[13] Calle R G D L, Gimeno O, Rivas J. Percarbonate as a Hydrogen Peroxide Carrier in Soil Remediation Processes [J]. Environmental Engineering Science, 2012,29(10):951-956.

[14] Angela S, Brown D D, Yuriy F. Quantitative determination of toluene, ethylbenzene, and xylene degradation products in contaminated groundwater by solid-phase extraction and in-vial derivatization [J]. International Journal of Environmental Analytical Chemistry, 2005,85(14):1075-1087.

[15] Wenyu H, Marcello B, Feng W, et al. Assessment of the Fe(III)-EDDS complex in Fenton-like processes: from the radical formation to the degradation of bisphenol A [J]. Environmental Science & Technology, 2013,47(4):1952-1959.

[16] Gallard H, Laat J D. Kinetic modelling of Fe(III)/H2O2oxidation reactions in dilute aqueous solution using atrazine as a model organic compound [J]. Water Research, 2000,34(12):3107–3116.

[17] Stuglik Z, Pawe?zagórski Z. Pulse radiolysis of neutral iron (II) solutions: oxidation of ferrous ions by OH radicals [J]. Radiation Physics & Chemistry, 1981,17(4):229-233.

[18] Yu X Y, Barker J R. Hydrogen Peroxide Photolysis In Acidic Aqueous Solutions Containing Chloride Ions. I. Chemical Mechanism [J]. J. phys. chem. a, 2003,107(9):1313-1324.

[19] Siedlecka E M, Wi?ckowska A, Stepnowski P. Influence of inorganic ions on MTBE degradation by Fenton's reagent [J]. Journal of Hazardous Materials, 2007,147(1/2):497-502.

[20] Laat J D, Le G T, Legube B. A comparative study of the effects of chloride, sulfate and nitrate ions on the rates of decomposition of H2O2and organic compounds by Fe(II)/H2O2and Fe(III)/H2O2[J]. Chemosphere, 2004,55(5):715–723.

[21] Laat J D, Le T G. Effects of chloride ions on the iron (III)-catalyzed decomposition of hydrogen peroxide and on the efficiency of the Fenton-like oxidation process [J]. Applied Catalysis B Environmental, 2006,66(1/2):137–146.

[22] 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/O2×-) in aqueous solution [J]. Journal of physical and chemical reference data, 1988,17(2): 513-886.

[23] Liao C H, Kang S F, Wu F A. Hydroxyl radical scavenging role of chloride and bicarbonate ions in the H2O2/UV process [J]. Chemosphere, 2001,44(5):1193–1200.

[24] Richardson D E, Yao H, Frank K M, et al. Equilibria, Kinetics, and Mechanism in the Bicarbonate Activation of Hydrogen Peroxide: Oxidation of Sulfides by Peroxymonocarbonate [J]. Journal of the American Chemical Society, 2000,122(8):1729- 1739.

[25] Dou A X, Wang X Q, Dou M W. Kinetic effect of humic acid on alachlor degradation by anodic Fenton treatment [J]. Journal of Environmental Quality, 2004,33(6):2343-2352.

[26] Neyens E, Baeyens J. A review of classic Fenton's peroxidation as an advanced oxidation technique [J]. Journal of Hazardous Materials, 2003,98(1-3):33–50.

[27] 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.

[28] Smith B A, Teel A L, Watts R J. Identification of the reactive oxygen species responsible for carbon tetrachloride degradation in modified Fenton's systems [J]. Environmental Science & Technology, 2004,38(20):5465-5469.

[29] Teel A L, Watts R J. Degradation of carbon tetrachloride by modified Fenton's reagent [J]. Journal of Hazardous Materials, 2002,94(2):179–189.

* 責(zé)任作者, 教授, lvshuguang@ecust.edu.cn

CUI Hang1, FU Xiao-ri1, GU Xiao-gang1,2, LU Shu-guang1*, QIU Zhao-fu1, SUI Qian1

(1.State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, Shanghai Key Laboratory of Functional Materials Chemistry, East China University of Science and Technology, Shanghai 200237, China；2.State Key Laboratory of Chemical Engineering, East China University of Science and Technology, Shanghai 200237, China)., 2016,36(5)：1449~1455

Ethylbenzene (EB) degradation performance in Fe (II) activated(SPC) system was investigated in this study. The effects of various factors, such as the initial SPC and Fe (II) concentrations, anions (Cl-, HCO3-, SO42-, and NO3-) concentration, natural organic matters (NOM), and initial solution pH were evaluated. The results showed that EB (1mmol/L) could be degraded completely in 20min with both SPC and Fe (II) dosages of 12mmol/L. Both Cl-and HCO3-anions and NOM had significant inhibitive effect on EB degradation, while the influence of SO-and NO3-was negligible at the tested ionic strength ranges. The EB removal was still significant at the initial solution pH of 9 even though the degradation of EB decreased with the increasing of initial solution pH (from 3.0 to 11.0), suggesting that Fe(II) activated SPC process was an effective technique for EB degradation at a wider pH range. In addition, the results of free radical probe tests and free radical scavenger tests indicated that×OH was the predominant species responsible for EB degradation even though both×OH and O2?-were generated in the SPC-Fe(II) system.

ethylbenzene；sodium percarbonate；hydroxyl radical；groundwater remediation

X523

A

1000-6923(2016)05-1449-07

崔 航(1991-),男,河南焦作人,華東理工大學(xué)碩士研究生,主要從事污染場(chǎng)地修復(fù)治理研究.

2015-10-12

國(guó)家自然科學(xué)基金(41373094,51208199);中國(guó)博士后科學(xué)基金(2015M570341);中央高?；究蒲袠I(yè)務(wù)費(fèi)專項(xiàng)資金(22A201514057)