周宏光,甘艷平,伍德權(quán),楊延梅*,張 揚,王璐瑤
FeMnMg-LDH對污染底泥中As(Ⅴ)的穩(wěn)定化
周宏光1,甘艷平1,伍德權(quán)2,楊延梅1*,張 揚3,王璐瑤3
(1.重慶交通大學(xué)國家內(nèi)河航道整治技術(shù)工程研究中心,重慶 400074;2.黔西南州廣播電視大學(xué),貴州 興義 562400;3.陜西省土地工程建設(shè)集團(tuán)有限責(zé)任公司,陜西 西安 710075)
針對疏浚底泥穩(wěn)定化修復(fù)問題,采用改進(jìn)的共沉淀法制備出FeMnMg-LDH材料以實現(xiàn)對底泥中As的鈍化,通過底泥培養(yǎng)試驗,探究了不同As污染水平和腐殖酸(HA)添加情況下FeMnMg-LDH對底泥As的鈍化修復(fù)效果.結(jié)果表明:FeMnMg-LDH材料促使弱酸提取態(tài)砷(MASF-As)向更穩(wěn)定的形態(tài)轉(zhuǎn)變;隨著FeMnMg-LDH添加水平的增加,污染底泥中As的浸出濃度和生物有效態(tài)含量均逐漸降低,高污染水平下(250mg/kg As)底泥中As的浸出濃度降低至261.35 μg/L,其穩(wěn)定化效率最高達(dá)80.29%,有效態(tài)含量百分比由13.24%降至3.21%(3.0% FeMnMg-LDH添加水平).HA的添加對底泥中As有一定活化效應(yīng),增加了底泥中As的浸出毒性和生物有效性;FeMnMg-LDH鈍化修復(fù)時可減弱HA所帶來的活化效應(yīng),在0%~3.0%的修復(fù)劑添加水平上,隨著添加量的增加,由HA所帶來的活化效應(yīng)逐漸減弱.相關(guān)性分析結(jié)果表明,底泥As的浸出毒性和生物有效性與MASF-As的百分比呈正相關(guān),與可還原態(tài)砷(RF-As)百分比呈負(fù)相關(guān),這說明FeMnMg-LDH材料通過改變底泥中As的形態(tài)分布降低As的浸出毒性和生物有效性.綜上所述,FeMnMg-LDH可作為底泥中As的鈍化劑,為底泥重金屬的鈍化修復(fù)提供新材料和新方法.
砷;底泥;FeMnMg-LDH;腐殖酸;穩(wěn)定化
砷(As)是自然界廣泛存在的類金屬有毒元素[1],一定的自然過程或人為活動會造成環(huán)境中As的富集和污染[2].大氣、土壤、水、生物體等環(huán)境要素中均存在As,在自然豐度中排名第20[3].As污染的來源分為自然來源和人為來源[4],自然來源主要是指礦物、土壤及沉積物中的As通過巖石風(fēng)化、火山活動等途徑[5-6],由原來難溶解、遷移的狀態(tài)變?yōu)橐兹芙?、遷移的狀態(tài),從而對環(huán)境造成危害;人為來源主要是指為滿足人類生產(chǎn)和生活需要而生產(chǎn)的含As產(chǎn)品的使用造成環(huán)境中As的累積與污染[7-9].環(huán)境As污染現(xiàn)象普遍存在,當(dāng)以人體可吸收的形態(tài)和足夠的劑量進(jìn)入人體時將導(dǎo)致嚴(yán)重的健康問題.在美國、中國、智利等國家地區(qū)都有As污染相關(guān)報道[10-11].
物質(zhì)的流動是循環(huán)的,地殼中As通過沉積、風(fēng)化、揮發(fā)及火山等作用在一定環(huán)境介質(zhì)中的物質(zhì)流進(jìn)入底泥.底泥是水環(huán)境中重金屬污染物的源和匯[12-13].一方面,重金屬污染物通過吸附、絡(luò)合、沉淀等作用由水體轉(zhuǎn)移至固相而沉積到底泥中,使底泥受到重金屬污染,另一方面,底泥中的重金屬污染物在一定條件下也會再次釋放進(jìn)入水體,對水環(huán)境造成二次污染[14-15].重金屬對底泥的依附性和危害性尤為嚴(yán)重,并通過生物積累[16]、生物放大[17]等過程影響陸地生物和人類健康.
層狀雙金屬氫氧化物(LDHs)是一種典型的陰離子黏土礦物,因其良好的熱穩(wěn)定性、陰離子交換能力[18]和催化能力[19]而被廣泛用作環(huán)境修復(fù)材料.LDHs的化學(xué)組成通式表示為[M2+ 1?xM3+ x(OH?)2](A)/n?mH2O.其中,M2+和M3+分別表示二價和三價金屬離子;A代表價陰離子;是三價陽離子與陽離子總量的物質(zhì)的量比,一般為0.25~0.33.因金屬陽離子和陰離子的種類、物質(zhì)的量等不同[20-22],LDHs的種類也多種多樣.LDHs的特殊結(jié)構(gòu)決定了其獨特性質(zhì),LDHs呈堿性,在化學(xué)組成方面構(gòu)成元素可選擇種類多,結(jié)構(gòu)方面易于插層[23]、修飾、層間陰離子可交換[24]等,LDHs的獨特性質(zhì)為其實際應(yīng)用提供了理論依據(jù).
目前,國內(nèi)外主要關(guān)注LDHs對水體中重金屬離子的去除[25-27],對底泥中As的鈍化修復(fù)研究相對較少.因此,研究LDHs作為鈍化劑對底泥中As的穩(wěn)定修復(fù)作用具有十分重要的實踐意義,不僅可以為底泥重金屬的化學(xué)改良提供新材料和方法,還可以充分發(fā)掘LDHs在環(huán)境修復(fù)領(lǐng)域的潛力,提高生態(tài)環(huán)境修復(fù)效果.本文以重慶某碼頭附近底泥為研究對象,考慮有機(jī)質(zhì)含量、污染水平、修復(fù)劑含量的影響,探究了外源As進(jìn)入底泥后的形態(tài)分配特征、腐殖酸(HA)對底泥As形態(tài)轉(zhuǎn)化的影響以及以FeMnMg-LDH作為鈍化劑對底泥As賦存形態(tài)轉(zhuǎn)化、浸出毒性、生物有效態(tài)含量的影響.
在堿性條件下,Mn(OH)2易氧化為MnO(OH)2,易破壞LDHs晶體的生長[28].因此反應(yīng)溶液均采用無氧水進(jìn)行配制,采取氮氣吹脫去離子水制得無氧水.FeMnMg-LDH材料通過改進(jìn)的共沉淀法[29]制備而成,具體操作如下:將FeCl3?6H2O、MnCl2?4H2O和MgCl2?6H2O溶解到無氧水中,按照Fe3+:Mn2+: Mg2+=1:1:2的比例(物質(zhì)的量比)配制成濃度約為2mol/L的金屬離子混合溶液A;另稱取一定質(zhì)量氫氧化鈉固體,配制成2mol/L的堿液B;將金屬離子混合溶液A、堿液B以一定速率(約0.5mL/min)滴加到三頸燒瓶,以速率200~250r/min持續(xù)攪拌反應(yīng)混合溶液,且pH值保持在(9.5±0.1).滴加完成后,反應(yīng)混合物繼續(xù)通入氮氣并攪拌2h,然后密封三頸燒瓶并將其置于60℃的水浴鍋中恒溫水浴2h,使體系繼續(xù)充分反應(yīng),取出讓其自然晶化12h.用無氧水反復(fù)洗滌至上清液呈中性,離心,冷凍干燥,將材料研磨過100目篩并密封保存待用.
通過FeMnMg-LDH材料吸附底泥溶液中的As(Ⅴ),對FeMnMg-LDH吸附As(V)前后的結(jié)構(gòu)性能進(jìn)行表征.FeMnMg-LDH樣品的X射線衍射(XRD)分析采用Bruker D8Advance X射線衍射儀進(jìn)行測定,其陽極材料為Cu靶,掃描范圍10°~70°,掃描步長0.04°;掃描電子顯微鏡(SEM)測試采用捷克TESCAN MIRA LMS;傅里葉紅外光譜(FT-IR)分析采用美國Nicolet 670儀器進(jìn)行測定.能量色散X射線光譜(EDX)分析采用日本日立的EDAX儀器.
供試底泥采自重慶市巴南區(qū)魚洞某碼頭附近,GPS點位儀確定點位(106°31’36”E,29°23’11”N).底泥均采自0~10cm的表層,去除雜質(zhì)后自然風(fēng)干,磨碎過100目篩,保存?zhèn)溆?底泥的基本理化性質(zhì)見表1.
供試腐殖酸(HA)為商用腐殖酸.購買于上海易恩化學(xué)技術(shù)有限公司,其為黑色粉末狀,HA檢測報告顯示其含量為90.4%,pH值為10,氯化鉀、水不溶物、水分含量分別為12.5%、0.05%、10.6%.
表1 供試底泥基本理化性質(zhì)
本研究采用室內(nèi)模擬試驗,先將外源HA添加至風(fēng)干過篩底泥樣品中,實驗室避光培養(yǎng)2個月,HA添加水平分別按照土重的0%、2%、5%(以C計);HA陳化穩(wěn)定后,再向底泥中添加不同污染水平的As,污染水平參考中華人民共和國國家標(biāo)準(zhǔn)《土壤環(huán)境質(zhì)量農(nóng)用地土壤污染風(fēng)險管控標(biāo)準(zhǔn)(試行)》(GB 15618-2018)[30]關(guān)于農(nóng)用地As污染風(fēng)險篩選值與污染風(fēng)險管制值(見表2),選取具代表性的兩個As污染水平(60, 250mg/kg)進(jìn)行底泥培養(yǎng).
表2 農(nóng)用地As土壤風(fēng)險篩選值與管制值(mg/kg)
稱取850g風(fēng)干過篩底泥樣品(添加HA處理后的樣品),按照設(shè)置的As污染水平將一定量的As(V)標(biāo)準(zhǔn)溶液加入到底泥樣品中(添加濃度按土重計)充分拌勻并保持32%左右的含水率,實驗室避光培養(yǎng)并按期取樣(試驗設(shè)計見表3).自培養(yǎng)之日起,分別于培養(yǎng)后第1,7,15,40,60,90d取樣進(jìn)行樣品的形態(tài)分析.為保證取樣的均勻,取樣時多點混合取樣,且需測定含水率.
表3 試驗設(shè)計
上述制備的污染底泥樣品在培養(yǎng)結(jié)束后,自然風(fēng)干、磨碎過100目篩后封裝保存?zhèn)溆?采用制備的FeMnMg-LDH材料作為污染底泥的穩(wěn)定化修復(fù)劑,探討其對污染底泥中As賦存形態(tài)、浸出毒性和生物有效性的影響.試驗分為6組污染底泥,分別稱量100.0g污染底泥于系列干凈塑料盒中,按照質(zhì)量比0%、0.5%、1.5%和3.0%分別加入FeMnMg-LDH材料,充分?jǐn)嚢?使二者混合均勻.加入去離子水調(diào)節(jié)相對含水率為32%左右,室溫下陳化40d,期間通過稱重法判斷添加的水量.陳化結(jié)束后取鮮樣進(jìn)行As形態(tài)順序提取、浸出毒性測定和生物有效態(tài)提取.
表4 As元素形態(tài)順序提取試劑及提取條件
采用中華人民共和國國家標(biāo)準(zhǔn)《土壤和沉積物13個微量元素形態(tài)順序提取程序》(GB/T25282- 2010)[31]方法進(jìn)行順序提取的形態(tài)分析,As主要形態(tài)的提取試劑及提取條件見表4,提取形態(tài)依次包括弱酸提取態(tài)(MASF-As)、可還原態(tài)(RF-As)、可氧化態(tài)(OF-As)和殘渣態(tài)(ResF-As);提取液中的As根據(jù)國家環(huán)境保護(hù)標(biāo)準(zhǔn)《土壤和沉積物汞、砷、硒、鉍、銻的測定微波消解/原子熒光法》(HJ680- 2013)[32]以及文獻(xiàn)[33]測定;底泥中As浸出毒性浸出方法采用中華人民共和國環(huán)境保護(hù)行業(yè)標(biāo)準(zhǔn)HJ/T300-2007醋酸緩沖溶液法[34],以醋酸緩沖溶液為浸提劑,模擬有害組分在滲濾液的影響下浸出的過程;選用0.5mol/L NaHCO3溶液為提取劑,測定底泥有效態(tài)As.所有測定過程均進(jìn)行空白實驗,每個處理設(shè)置3個平行,試驗中采用順序提取程序及原子熒光法測定底泥浸提液中(包括總量和形態(tài))As的加標(biāo)回收率為85.21%~120.69%.
圖1為FeMnMg-LDH吸附底泥溶液中As(Ⅴ)前后的XRD圖譜.由圖可知在2為11.38°、22.6°、34.2°、59.66°和60.8°處均有特征峰,這些峰分別對應(yīng)(003)、(006)、(009)、(110)和(113)晶面,與LDHs特征衍射峰相符,表明制備的材料為層狀金屬氫氧化物.并且,2為11.38°、22.6°、34.2°處表現(xiàn)出良好的倍數(shù)關(guān)系,(110)和(113)處的峰形尖銳且沒有雜峰,說明制備的FeMnMg-LDH材料結(jié)晶度高、晶體結(jié)構(gòu)好.FeMnMg-LDH吸附As(V)后各處特征峰發(fā)生不同程度的偏移,但仍存在對應(yīng)的關(guān)系,說明FeMnMg-LDH吸附As后仍保留典型層狀結(jié)構(gòu).
圖1 FeMnMg-LDH吸附As(V)前后XRD圖譜
FeMnMg-LDH吸附As(V)前后SEM圖像如圖2所示.明顯可以看出FeMnMg-LDH材料呈現(xiàn)出層狀結(jié)構(gòu)(圖2(a),(b)),表明本文制備的材料具備LDHs典型片層結(jié)構(gòu).As-FeMnMg-LDH材料的片狀結(jié)構(gòu)上被顆粒狀物體團(tuán)聚(圖2(c),(d)),這些顆粒狀物體很有可能是FeMnMg-LDH材料中的金屬離子與水中砷酸根或亞砷酸根離子所形成的配位絡(luò)合物.
FeMnMg-LDH吸附As(V)前后的FTIR圖譜如圖3所示.FeMnMg-LDH在3400和1630cm-1兩處附近出現(xiàn)了較寬的吸收峰,分別為O—H和H2O的收縮振動峰[35].As-FeMnMg-LDH圖譜中O—H處的峰強(qiáng)度上升,可能由于砷酸根離子與層板上-OH反應(yīng)生成了水[36];在1363cm-1處出現(xiàn)較強(qiáng)的吸收峰是O—C—O的伸縮振動峰[37],表明吸附后可能有金屬碳酸鹽的形成.在592.43cm-1的吸收峰為材料的金屬骨架M—O的收縮振動峰[38],吸附后的峰為595.1cm-1,發(fā)生不同程度的偏移,這可能是由于As(V)與M—O發(fā)生了配位絡(luò)合反應(yīng).
圖2 FeMnMg-LDH吸附As(V)前后SEM圖像
FeMnMg-LDH吸附As(V)前后的EDX圖如圖4所示.由圖可知,吸附后的圖譜中代表Cl-的譜峰迅速下降,同時出現(xiàn)As元素的譜峰,說明在砷酸根或亞砷酸根離子和層間的Cl-間發(fā)生了離子交換,且該機(jī)制可能占主導(dǎo)作用.
2.2.1 采樣點底泥As的形態(tài) 采樣原始底泥中As的各形態(tài)含量如圖5所示,可以看出底泥中As以ResF-As為主,占總As含量的80.37%;其次是RF-As,占總As含量的14.39%;MASF-As、OF-As的含量相對較少,分別占總As含量的1.13%、4.11%.采樣點底泥中As主要以難遷移轉(zhuǎn)化形態(tài)存在,而易遷移轉(zhuǎn)化形態(tài)占比則很低.
圖5 取樣點底泥中As各形態(tài)百分比
風(fēng)險評價準(zhǔn)則(RAC)是基于沉積物中重金屬的不同形態(tài)對其不同的結(jié)合力而提出的[39-40],其將重金屬中可交換態(tài)和碳酸鹽結(jié)合態(tài)(本研究的弱酸提取態(tài))所占百分?jǐn)?shù)分為5個等級(表5),原始底泥中MASF-As占比為1.13%,依據(jù)RAC進(jìn)行風(fēng)險評價,底泥As的可利用性低,環(huán)境風(fēng)險為低風(fēng)險.
表5 風(fēng)險評價準(zhǔn)則
2.2.2 外源As在底泥中的形態(tài)分異 外源As進(jìn)入底泥后其順序提取的各形態(tài)分布如圖6所示,在兩種As污染水平下(60和250mg/kg),底泥中MASF- As、RF-As、OF-As和ResF-As變化趨勢基本保持一致.As-60污染底泥中,MASF-As和RF-As占比呈現(xiàn)不斷下降的趨勢,MASF-As前15d內(nèi)變化較大,15d后趨于穩(wěn)定,而RF-As在60d后才趨于穩(wěn)定. OF-As和ResF-As占比呈現(xiàn)不斷上升的趨勢,在60d后趨于穩(wěn)定.外源As進(jìn)入底泥90d時其形態(tài)以RF- As和ResF-As為主,OF-As和MASF-As占比相對較低,As-60和As-250污染底泥中MASF-As占比分別為0.83%和2.19%,根據(jù)風(fēng)險評價準(zhǔn)則,As-60污染底泥無潛在環(huán)境風(fēng)險,As-250污染底泥存在低等程度的環(huán)境風(fēng)險.在培養(yǎng)過程中,各形態(tài)之間總體呈現(xiàn)由易遷移形態(tài)向難遷移形態(tài)分配,底泥As活性逐漸降低;污染濃度更高的As-250中MASF-As和RF-As明顯高于As-60,高濃度外源As進(jìn)入底泥后,易遷移轉(zhuǎn)化態(tài)占比更高,說明高濃度As不易固定,這可能是由底泥本身的污染承載能力決定,隨著污染濃度增大,因污染承載能力有限,污染風(fēng)險就更高.
與原始底泥中As的賦存形態(tài)比較,培養(yǎng)后的外源As污染底泥中呈現(xiàn)易遷移形態(tài)占比增加,而難遷移形態(tài)占比減少,說明外源As進(jìn)入底泥后在底泥中的轉(zhuǎn)化與As添加量有密切關(guān)系,低濃度As在底泥中易于固定[41].當(dāng)?shù)啄喟l(fā)生外源As污染事件時,其賦存形態(tài)在一定環(huán)境條件將隨時間的推移不斷轉(zhuǎn)化分配,且在發(fā)生污染的一定時期內(nèi),易遷移轉(zhuǎn)化形態(tài)的存在使得底泥可能存在更高的活性,增加底泥利用的潛在風(fēng)險.
2.2.3 腐殖酸(HA)添加對外源As在底泥中形態(tài)分異的影響 HA是天然有機(jī)質(zhì)的一種重要組分[42],在控制重金屬的環(huán)境地球化學(xué)行為中起關(guān)鍵作用[43],可通過吸附[44]、螯合[45]、配位絡(luò)合[46]等作用影響土壤及底泥環(huán)境中重金屬的有效態(tài).不同HA添加水平(0%、2%、5%)處理時,外源As進(jìn)入底泥后其賦存形態(tài)變化如圖7所示.
陳化周期內(nèi)各形態(tài)的轉(zhuǎn)化結(jié)果表明,外源As進(jìn)入底泥后由易遷移形態(tài)向其他形態(tài)轉(zhuǎn)化,隨著HA添加量的增加,底泥中MASF-As和ResF-As的占比逐漸增加,RF-As的占比逐漸降低.在90d時,與對照組(HA-0%)相比,HA-2%和HA-5%處理的60mg/kg As污染底泥中,MASF-As占比分別增加0.65%、2.26%,相應(yīng)增加了底泥的環(huán)境風(fēng)險程度,ResF-As占比分別增加7.75%、6.49%,RF-As占比分別降低6.74%、9.21%,250和60mg/kg污染底泥中As各形態(tài)變化趨勢基本一致.HA的添加對OF-As的作用相對不明顯,僅影響了其占比.這可能與培養(yǎng)過程中環(huán)境條件有關(guān),陳化穩(wěn)定后的As在一定環(huán)境條件下其MASF-As、RF-As和OF-As三種形態(tài)會相互轉(zhuǎn)化.
HA處理的外源As污染底泥中MASF-As占比增加,可能是由于HA的低分子量、高酸性[47]等特征,使As結(jié)合形成的絡(luò)合物溶解性和移動性更強(qiáng)[48],從而底泥中MASF-As保持更高比例;RF-As占比降低可能是因為砷酸根或亞砷酸根帶負(fù)電荷,難以直接被HA所吸附,而HA對底泥中Fe3+、Al3+等離子具有很強(qiáng)的絡(luò)合能力[49],使其減少與As的絡(luò)合從而降低了RF-As分配比例.
2.2.4 FeMnMg-LDH對底泥中As賦存形態(tài)轉(zhuǎn)化的影響添加FeMnMg-LDH材料重新培養(yǎng)40d后,底泥中As賦存形態(tài)隨FeMnMg-LDH添加水平的變化趨勢如圖8所示.
由圖8可知,隨FeMnMg-LDH添加水平的增加,污染底泥中MASF-As和OF-As占比逐漸降低,RF-As和ResF-As占比逐漸增加.僅添加FeMnMg-LDH材料(即HA-0%)時,60mg/kg As污染底泥中MASF-As占比由2.25%降低至0.11%,250mg/kg As污染底泥中MASF-As占比由6.88%降低至0.21%,占比均小于1%,根據(jù)風(fēng)險評價準(zhǔn)則,其均不存在潛在的環(huán)境風(fēng)險.污染底泥中As的各賦存形態(tài)相互轉(zhuǎn)化,FeMnMg-LDH材料的加入使得MASF-As向RF-As和ResF-As轉(zhuǎn)化,從易遷移形態(tài)向難遷移形態(tài)轉(zhuǎn)化,活性形態(tài)占比降低,毒性風(fēng)險減小.
圖7 HA添加對外源As在底泥中形態(tài)分異的影響
圖8 FeMnMg-LDH材料穩(wěn)定修復(fù)下兩種As污染濃度的底泥中As的賦存形態(tài)變化
由圖8可知,HA的添加影響FeMnMg-LDH材料對底泥中As賦存形態(tài)的調(diào)控作用,具體表現(xiàn)為:增加MASF-As和ResF-As占比,減少RF-As占比,對OF-As占比有影響但趨勢不明顯.以HA-0為對照相比,在0~3.0% 4種FeMnMg-LDH材料添加水平上,60和250mg/kg的污染底泥中MASF-As分別由0.11%~2.25%和0.21%~6.88%增至0.24%~9.76%和0.33%~18.27%,As潛在釋放風(fēng)險增加.
由As的賦存形態(tài)轉(zhuǎn)化結(jié)果可以看出,FeMnMg- LDH材料的加入使MASF-As占比降低,向其他更穩(wěn)定形態(tài)轉(zhuǎn)化和分配,且增大FeMnMg-LDH材料添加水平,HA對MASF-As的影響逐漸減弱,潛在環(huán)境風(fēng)險相應(yīng)降低.
根據(jù)HJ/T300-2007醋酸緩沖溶液法以酸性浸提劑浸提(浸提劑2#: 0.3mol/L 醋酸溶液,溶液pH值為(2.64±0.05)),測定As浸出毒性,計算FeMnMg- LDH材料的穩(wěn)定化效率
從圖9可見,添加FeMnMg-LDH材料重新培養(yǎng)40d后,FeMnMg-LDH材料對底泥As的穩(wěn)定化效果顯著,在0~3.0%的FeMnMg-LDH材料添加水平下,隨著材料添加水平的增加,污染底泥中As浸出濃度逐漸降低,60和250mg/kg污染底泥中As的浸出濃度分別由164.85和1291.95μg/L降低至47.39和261.35μg/L,穩(wěn)定化效率最高分別可達(dá)71.25%和80.29%,且60mg/kg As污染底泥中浸出濃度達(dá)到我國Ⅲ類地表水環(huán)境質(zhì)量標(biāo)準(zhǔn)(50μg/L).
圖9 FeMnMg-LDH材料穩(wěn)定修復(fù)下底泥中As的浸出毒性變化
由圖10可見,HA的添加影響FeMnMg- LDH材料對污染底泥中As的浸出毒性調(diào)控,在相同F(xiàn)eMnMg-LDH材料添加水平,不同HA添加水平下,材料對兩種污染濃度底泥浸出毒性的穩(wěn)定化程度不同.在較低FeMnMg-LDH材料添加水平(0%、0.5%和1.5%)時,60和250mg/kg底泥中As浸出濃度均隨HA的增加而呈現(xiàn)增長趨勢;而在FeMnMg- LDH材料添加水平為3%時,60和250mg/kg底泥中As浸出濃度均隨HA的增加而減少,由47.39和261.35μg/L分別降至35.67和163.22μg/L,HA對FeMnMg-LDH穩(wěn)定化作用的負(fù)效應(yīng)變?yōu)檎?yīng).
底泥有效態(tài)As的測定目前尚無公認(rèn)的統(tǒng)一方法,化學(xué)提取和生物指示的研究表明以NaHCO3作為提取劑提取的石灰性紫色土的有效態(tài)As與植物含As量具有良好的相關(guān)性,本文所用底泥為弱堿性,因此選用0.5mol/L NaHCO3溶液為提取劑,測定As生物有效態(tài)含量.
圖11 FeMnMg-LDH材料穩(wěn)定修復(fù)下底泥中As的生物有效態(tài)含量變化
由圖11可知,FeMnMg-LDH材料對不同濃度As污染底泥中As的生物有效態(tài)含量影響顯著,隨著材料添加水平的增加,污染底泥中As元素的生物有效態(tài)含量逐漸降低.與不添加FeMnMg-LDH材料相比,3.0% FeMnMg-LDH添加水平的底泥提取液中60和250mg/kg As污染底泥的生物有效態(tài)含量分別由5.91和33.09mg/kg降低至1.67和8.02mg/kg,有效態(tài)含量百分比分別由9.85%和13.24%降至2.79%和3.21%.
由圖12可知在相同F(xiàn)eMnMg-LDH材料添加水平下,60和250mg/kg As污染濃度底泥中有效態(tài)As含量均隨HA的增加而呈現(xiàn)增長趨勢.與對照HA-0%相比,在0~3.0% FeMnMg-LDH材料添加水平上,60和250mg/kg兩種污染濃度的底泥中As的生物有效態(tài)含量百分比分別增加了5.44%、2.88%、3.57%、2.37%和4.82%、4.19%、2.93%、0.33%,可以看出HA的添加對底泥As的調(diào)控呈現(xiàn)負(fù)效應(yīng),但隨著FeMnMg-LDH材料添加量的增加,這種負(fù)效應(yīng)逐漸減弱.
本文用origin在0~3.0%的FeMnMg-LDH材料添加水平上分別對浸出毒性、生物有效態(tài)和順序提取形態(tài)做相關(guān)性分析,明確賦存形態(tài)與浸出毒性及生物有效性的關(guān)聯(lián)性.
2.5.1 底泥中As的浸出濃度變化和生物有效態(tài)含量變化的相關(guān)性分析 以弱酸提取的重金屬形態(tài)有時可看作重金屬的生物有效態(tài),從二者之間的定義來看存在一定的關(guān)聯(lián)性,但具體在量方面的關(guān)系還需進(jìn)一步研究.本試驗制備的污染底泥顯堿性, NaHCO3為酸式鹽,溶于水顯弱堿性,而浸提劑2#顯酸性,兩種提取劑浸提的污染底泥中As的結(jié)果如圖13所示.可以看出兩種浸提劑提取As的百分含量變化趨勢基本一致,NaHCO3提取的含量百分比始終高于浸提劑2#,這與As污染底泥pH值本身呈堿性(測定樣品pH值均為7.89~10.48)有關(guān),高pH值下有機(jī)酸作為浸提劑時,重金屬的浸出濃度偏低.有機(jī)酸促進(jìn)As釋放的主要機(jī)制是其陰離子與重金屬螯合而促進(jìn)重金屬釋放,高pH值條件下,因螯合而釋放的較少;而NaHCO3與重金屬形成的配位體相對穩(wěn)定,難以被底泥吸附,浸出就相對容易.
對0~3.0% FeMnMg-LDH材料添加水平內(nèi)底泥中As的浸出濃度變化和生物有效態(tài)含量變化作相關(guān)性分析,由表6可知生物有效態(tài)與浸出濃度呈顯著正相關(guān)(<0.05),可以看出NaHCO3提取的生物有效態(tài)含量百分比與浸提劑2#提取的浸出濃度質(zhì)量占比具有良好相關(guān)性.
表6 6種污染底泥中As的浸出濃度與生物有效態(tài)含量百分比的Pearson相關(guān)性
注:**代表<0.01,*代表<0.05.
2.5.2 底泥中As的賦存形態(tài)變化與浸出毒性和生物有效態(tài)含量變化的相關(guān)性分析 對0~3.0% FeMnMg-LDH材料添加水平內(nèi)底泥中As的賦存形態(tài)變化和浸出毒性變化作相關(guān)性分析,由表7可知浸出毒性與MASF-As呈顯著正相關(guān)(<0.01),與RF-As呈顯著負(fù)相關(guān)(<0.05),FeMnMg-LDH材料降低浸出毒性在于其減少了MASF-As占比,促使MASF-As向更穩(wěn)定的形態(tài)轉(zhuǎn)化.
表7 6種污染底泥中As的賦存形態(tài)與浸出毒性質(zhì)量百分?jǐn)?shù)的Pearson相關(guān)性
注:**代表<0.01,*代表<0.05.
對0~3.0%FeMnMg-LDH材料添加水平內(nèi)底泥中As的賦存形態(tài)變化和生物有效態(tài)含量變化作相關(guān)性分析,由表8可知生物有效態(tài)與MASF-As呈顯著正相關(guān)(<0.05),與RF-As呈顯著負(fù)相關(guān)(<0.05=,FeMnMg-LDH材料降低生物有效性在于其減少了MASF-As占比,促使MASF-As向更穩(wěn)定的形態(tài)轉(zhuǎn)變.
表8 6種污染底泥中As的賦存形態(tài)與生物有效態(tài)質(zhì)量百分?jǐn)?shù)的Pearson相關(guān)性
注:**代表<0.01,*代表<0.05.
MASF-As對環(huán)境條件的變化敏感性強(qiáng),尤其是pH值,pH值變化時MASF-As易重新釋放,RF- As主要被氧化鐵、氧化錳等吸持,在As賦存形態(tài)變化研究中發(fā)現(xiàn),添加FeMnMg-LDH材料會使MASF-As向穩(wěn)定性更高的RF-As轉(zhuǎn)化,說明底泥中活性較高的As與FeMnMg-LDH材料結(jié)合,以更穩(wěn)定的鐵、錳氧化物復(fù)合形式存在,因而浸出濃度降低,起到顯著的穩(wěn)定化作用[50].這可能是因為FeMnMg-LDH與As發(fā)生了離子交換、配位絡(luò)合等作用,錳氧化物具有將As(III)氧化為As(V)的作用,對As有較高親和力和選擇性,可提高As(V)和未被氧化的As(III)的吸附能力[51].SEM中顆粒物以及FTIR中As-O鍵的出現(xiàn)表明砷酸鹽和FeMnMg- LDH之間的特定相互作用是通過配體交換以及外/內(nèi)球表面復(fù)合物的形成.此外As與Fe表面的-OH形成羥基化金屬離子([MOH]+)[52],Mg2+的存在也會與As形成Mg3(AsO4)2和MgHAsO3沉淀[53], FeMnMg-LDH表面-OH參與配體交換和絡(luò)合反應(yīng),是決定材料對重金屬離子吸附和離子交換能力的主要因素之一[54],材料對As的穩(wěn)定化主要通過表面-OH基團(tuán)的吸附作用[55],As被高效吸附到層間,底泥As浸出毒性及生物有效性降低,起到鈍化穩(wěn)定修復(fù)作用.
HA占據(jù)FeMnMg-LDH材料與As的結(jié)合位點,與As形成競爭吸附[56].從而導(dǎo)致材料吸附As的能力降低,As浸出毒性濃度和生物有效態(tài)含量也因HA的增加而增加,但隨著FeMnMg-LDH材料添加水平的增加,在量上減弱HA添加對修復(fù)所帶來的負(fù)面效應(yīng).
3.1 原始底泥中As以ResF-As為主;外源As進(jìn)入底泥后主要以RF-As和ResF-As形式存在,OF-As和MASF-As占比相對較低;HA的添加使外源As污染底泥中的MASF-As增加,增加了底泥中As的環(huán)境風(fēng)險.
3.2 FeMnMg-LDH材料降低浸出毒性和生物有效性在于其減少了MASF-As占比,促使MASF-As向穩(wěn)定性更高的形態(tài)轉(zhuǎn)化;隨著FeMnMg-LDH材料添加水平的增加,污染底泥中As的浸出濃度和生物有效態(tài)含量均逐漸降低,穩(wěn)定處理后的60和250mg/kg污染底泥As穩(wěn)定化效率分別最高可達(dá)71.25%和80.29%,降低了潛在環(huán)境風(fēng)險.
3.3 HA的添加對底泥中As有一定活化效應(yīng),HA增加了底泥中As的浸出毒性和生物有效性; FeMnMg-LDH鈍化修復(fù)時可減少HA所帶來的活化效應(yīng),在0%~3.0%的修復(fù)劑添加水平上,由HA所帶來的活化效應(yīng)隨著FeMnMg-LDH添加水平的增加逐漸減弱.
3.4 底泥As的浸出毒性和生物有效性與MASF- As的百分比呈正相關(guān),與RF-As百分比呈負(fù)相關(guān),FeMnMg-LDH材料通過促使底泥中As由弱酸提取態(tài)轉(zhuǎn)化為其他形態(tài)的方式降低As的浸出毒性和生物有效性,從而修復(fù)底泥As污染.
[1] Mohan D, Pittman C U, J R. Arsenic removal from water/wastewater using adsorbents--A critical review [J]. Journal of Hazardous Materials, 2007,142(1/2):1-53.
[2] Grover K, Komarneni S, Katsuki H. Uptake of arsenite by synthetic layered double hydroxides [J]. Water Research, 2009,43(15):3884- 3890.
[3] Asere T G, Stevens C V, Du Laing G. Use of (modified) natural adsorbents for arsenic remediation: A review [J]. Science of the Total Environment, 2019,676:706-720
[4] Li L, Pohl C, Ren J L, et al. Revisiting the biogeochemistry of arsenic in the Baltic Sea: Impact of anthropogenic activity [J]. Science of the Total Environment, 2018,613-614:557-568.
[5] Bundschuh J, Schneider J, Alam M A, et al. Seven potential sources of arsenic pollution in Latin America and their environmental and health impacts [J]. Science of the Total Environment, 2021,780(5):146274.
[6] Yazdani M, Tuutuj?rvi T, Bhatnagar A, et al. Adsorptive removal of arsenic(V) from aqueous phase by feldspars: kinetics, mechanism, and thermodynamic aspects of adsorption [J].Journal of Molecular Liquids, 2016,214:149-156.
[7] Wang S, Mulligan C N. Occurrence of arsenic contamination in Canada: sources, behavior and distribution [J]. Science of the Total Environment, 2006,366(2/3):701-721.
[8] Hong J, Zhu Z, Lu H, et al. Synthesis and arsenic adsorption performances of ferric-based layered double hydroxide with α-alanine intercalation [J]. Chemical Engineering Journal, 2014,252:267-274.
[9] Punshon T, Jackson B P, Meharg A A, et al. Understanding arsenic dynamics in agronomic systems to predict and prevent uptake by crop plants [J]. Science of the Total Environment, 2017,581-582:209-220.
[10] Bessaies H, Iftekhar S, Doshi B, et al. Synthesis of novel adsorbent by intercalation of biopolymer in LDH for the removal of arsenic from synthetic and natural water [J]. Journal of Environmental Sciences (China), 2020,91:246-261.
[11] Chen X, Zeng X C, Wang J, et al. Microbial communities involved in arsenic mobilization and release from the deep sediments into groundwater in Jianghan plain, Central China [J]. Science of the Total Environment, 2017,579:989-999.
[12] Niu Y, Chen F, Li Y, et al. Trends and sources of heavy metal pollution in global river and lake sediments from 1970 to 2018 [J]. Reviews of Environmental Contamination and Toxicology, 2021, 257:1-35.
[13] Zhang C, Yu Z G, Zeng G M, et al. Effects of sediment geochemical properties on heavy metal bioavailability [J]. Environment International, 2014,73:270-281.
[14] Xu Y, Wu Y, Han J, et al. The current status of heavy metal in lake sediments from China: Pollution and ecological risk assessment [J]. Ecology and Evolution, 2017,7(14):5454-5466.
[15] Zhao X, Gao B, Xu D, et al. Heavy metal pollution in sediments of the largest reservoir (Three Gorges Reservoir) in China: A review [J]. Environmental Science and Pollution Research, 2017,24(26):20844- 20858.
[16] Wang S L, Xu X R, Sun Y X, et al. Heavy metal pollution in coastal areas of South China: A review [J]. Marine Pollution Bulletin, 2013, 76(1/2):7-15.
[17] Affandi F A, Ishak M Y. Impacts of suspended sediment and metal pollution from mining activities on riverine fish population-a review [J]. Environmental Science and Pollution Research, 2019,26(17): 16939-16951.
[18] Guo Y, Zhu Z, Qiu Y, et al. Synthesis of mesoporous Cu/Mg/Fe layered double hydroxide and its adsorption performance for arsenate in aqueous solutions [J]. Journal of Environmental Sciences, 2013,25 (5):944-953.
[19] Nguyen T H, Tran H N, Nguyen T V, et al. Single-step removal of arsenite ions from water through oxidation-coupled adsorption using Mn/Mg/Fe layered double hydroxide as catalyst and adsorbent [J]. Chemosphere, 2022:295.
[20] Wang J, Kang D, Yu X, et al. Synthesis and characterization of Mg–Fe–La trimetal composite as an adsorbent for fluoride removal [J]. Chemical Engineering Journal, 2015,264:506-513.
[21] Rouahna N, Barkat D, Ouakouak A, et al. Synthesis and characterization of Mg-Al layered double hydroxide intercalated with D2EHPA: Application for copper ions removal from aqueous solution [J]. Journal of Environmental Chemical Engineering, 2018,6(1):1226- 1232.
[22] Shi X, Kang L, Hong J, et al. Strong selectivity and high capacity in the adsorption of As (V) from wastewater by glycine-modified Fe/Cu- layered double hydroxides [J]. Journal of Alloys and Compounds, 2021,865:158956.
[23] Ma L, Islam S M, Liu H, et al. Selective and efficient removal of toxic oxoanions of As(III), As(V), and Cr(VI) by layered double hydroxide intercalated with MoS42–[J]. Chemistry of Materials, 2017,29(7): 3274-3284.
[24] Zubair M, Daud M, Mckay G, et al. Recent progress in layered double hydroxides (LDH)-containing hybrids as adsorbents for water remediation [J]. Applied Clay Science, 2017,143:279-292.
[25] Kameda T, Kondo E, Yoshioka T. Equilibrium and kinetics studies on As(V) and Sb(V) removal by Fe2+-doped Mg-Al layered double hydroxides [J]. Journal of Environmental Management, 2015,151:303- 309.
[26] Rojas R, Perez M R, Erro E M, et al. EDTA modified LDHs as Cu2+scavengers: Removal kinetics and sorbent stability [J]. Journal of Colloid and Interface Science, 2009,331(2):425-431.
[27] Yin C, Li S, Liu L, et al. Structure-tunable trivalent Fe-Al-based bimetallic organic frameworks for arsenic removal from contaminated water [J]. Journal of Molecular Liquids, 2022,346:117101.
[28] Zhou H, Tan Y, Yang Y, et al. Application of FeMgMn layered double hydroxides for phosphate anions adsorptive removal from water [J]. Applied Clay Science, 2021,200:105903.
[29] Zhou H, Jiang Z, Wei S. A new hydrotalcite-like absorbent FeMnMg- LDH and its adsorption capacity for Pb2+ions in water [J]. Applied Clay Science, 2018,153:29-37.
[30] GB15618-2018 土壤環(huán)境質(zhì)量農(nóng)用地土壤污染風(fēng)險管控標(biāo)準(zhǔn)(試行) [S]. GB15618-2018 Soil environmental quality-risk control standard for soil contamination of agricultural land [S].
[31] GB/T25282-2010 土壤和沉積物13個微量元素形態(tài)順序提取程序[S]. GB/T25282-2010 Soil and sediment-Sequential extraction procedure of speciation of 13trace elements [S].
[32] HJ680-2013 土壤和沉積物汞,砷,硒,鉍,銻的測定微波消解/原子熒光法[S]. HJ680-2013 Soil and sediment-determination of mercury, arsenic, selenium, bismuth, antimony-Microwave dissolution/Atomic Fluorescence Spectrometry [S].
[33] 林海蘭,朱日龍,于 磊,等.水浴消解-原子熒光光譜法測定土壤和沉積物中砷、汞、硒、銻和鉍[J]. 光譜學(xué)與光譜分析, 2020,40(5): 1528-1533. Lin H L, Zhu R L, Yu L, et al. Determination of arsenic, mercury, selenium, antimony and bismuth in soil lan sediments by water bath digestion-atomic fluorescence spectrometry [J]. Spectroscopy and Spectral Analysis, 2020,40:1528-1533.
[34] HJ/T300-2007 固體廢物浸出毒性浸出方法醋酸緩沖溶液法 [S]. HJ/T300-2007 Solid waste-extraction procedure for leaching toxicity-Acetic acid buffer solution method [S].
[35] 廖玉梅,余 杰,魏世強(qiáng),等.FeMnNi-LDHs對水中As(Ⅲ)的吸附性能與機(jī)制[J]. 環(huán)境科學(xué), 2021,42(1):293-304. Liao Y M, Yu J, Wei S Q, et al. Adsorption effect and mechanism of aqueous arsenic on FeMnNi-LDHs [J]. Environmental Science, 2021, 42:293-304.
[36] Xu W, Wang J, Wang L, et al. Enhanced arsenic removal from water by hierarchically porous CeO(2)-ZrO(2) nanospheres: role of surface- and structure-dependent properties [J]. Journal of Hazardous Materials, 2013,260:498-507.
[37] Cheng X, Huang X, Wang X, et al. Influence of calcination on the adsorptive removal of phosphate by Zn-Al layered double hydroxides from excess sludge liquor [J]. Journal of Hazardous Materials, 2010, 177(1-3):516-523.
[38] Maziarz P, Matusik J, STR?CZEK T, et al. Highly effective magnet- responsive LDH-Fe oxide composite adsorbents for As(V) removal [J]. Chemical Engineering Journal, 2019,362:207-216.
[39] Jain C K. Metal fractionation study on bed sediments of River Yamuna, India [J]. Water Research, 2004,38(3):569-578.
[40] 喬敏敏,季宏兵,朱先芳,等.密云水庫沉積物中重金屬形態(tài)分析及風(fēng)險評價[J]. 農(nóng)業(yè)環(huán)境科學(xué)學(xué)報, 2013,32(7):1423-1431. Qiao M M, Ji H B, Zhu X F, et al. Fraction distribution and risk assessment of heavy metal in sediments of Miyun Reservoir [J]. Journal of Agro-Environment Science, 2013,32:1423-1431.
[41] 王 俊,王青清,蔣珍茂,等.腐殖酸對外源砷在土壤中形態(tài)轉(zhuǎn)化和有效性的影響[J]. 土壤, 2018,50(3):522-529. Wang J, Wang Q Q, Jiang Z M, et al. Transformation and bioavailability of exogenous as in soil as influenced by humic acids and its active components [J]. Soils, 2018,50(3):522-529.
[42] Tang W W, Zeng G M, Gong J L, et al. Impact of humic/fulvic acid on the removal of heavy metals from aqueous solutions using nanomaterials: a review [J]. Science of The Total Environment, 2014, 468-469:1014-1027.
[43] Li H, Wang J, Zhao B, et al. The role of major functional groups: Multi-evidence from the binding experiments of heavy metals on natural fulvic acids extracted from lake sediments [J]. Ecotoxicology and Environmental Safety, 2018,162:514-520.
[44] Lee S S, Nagy K L, Park C, et al. Heavy metal sorption at the muscovite (001)-fulvic acid interface [J]. Environmental Science & Technology, 2011,45(22):9574-9581.
[45] Bahemmat M, Farahbakhsh M, Kianirad M. Humic substances- enhanced electro remediation of heavy metals contaminated soil [J]. Journal of Hazardous Materials, 2016,312:307-318.
[46] Xie Y, Lu G, Ye H, et al. Fulvic acid induced the liberation of chromium from CrO42?-substituted schwertmannite [J]. Chemical Geology, 2017,475:52-61.
[47] Sun C Y, Liu J S, Wang Y, et al. Effect of long-term cultivation on soil organic carbon fractions and metal distribution in humic and fulvic acid in black soil, Northeast China [J]. Soil Research, 2012,50(7): 562-569.
[48] 李光林.腐殖酸與幾種重金屬離子的相互作用及影響因素研究[D]. 重慶:西南農(nóng)業(yè)大學(xué), 2002. Li G L. On the reaction of humic acid on some heavy metal ions and the affecting factors [D]. Chongqing:Southwest Agricultural University, 2002.
[49] 王 強(qiáng).腐殖酸與鐵錳鋁及其氧化物的相互作用機(jī)理研究[D]. 重慶:西南農(nóng)業(yè)大學(xué), 2005. Wang Q. Interactions of humic acids with Fe3+、Al3+、Mn2+irons their oxides and the mechanisms [D]. Chongqing: Southwest Agricultural University, 2005.
[50] 費 楊,閻秀蘭,廖曉勇,等.鐵錳雙金屬材料對砷和重金屬復(fù)合污染土壤的穩(wěn)定化研究[J]. 環(huán)境科學(xué)學(xué)報, 2016,36(11):4164-4172. Fei Y, Yan X L, Liao X Y, et al. Stabilization effects and mechanisms of Fe-Mn binary oxide on arsenic and heavy metal co-contaminated soils [J]. Acta Scientiae Circumstantiae, 2016,36(11):4164-4172.
[51] Lou Z, Cao Z, Xu J, et al. Enhanced removal of As(III)/(V) from water by simultaneously supported and stabilized Fe-Mn binary oxide nanohybrids [J]. Chemical Engineering Journal, 2017,322:710-721.
[52] 袁 林.鐵錳復(fù)合氧化物對重金屬鉛鎘吸附解吸特征及其影響因素研究[D]. 重慶:西南大學(xué), 2010. Yuan L. The environmental behavior of lead and cadmium as influenced by Fe-Mn composite oxide from different factor [D]. Ghongqing: Southwest University, 2010.
[53] Zhang D, Yuan Z, Wang S, et al. Incorporation of arsenic into gypsum: Relevant to arsenic removal and immobilization process in hydrometallurgical industry [J]. Journal of Hazardous Materials, 2015,300:272-280.
[54] 張 昱,豆小敏,楊 敏,等.砷在金屬氧化物/水界面上的吸附機(jī)制Ⅰ.金屬表面羥基的表征和作用[J]. 環(huán)境科學(xué)學(xué)報, 2006,10:1586- 1591. Zhang Y, Dou X M, Yang M, et al. Adsorption mechanism of arsenic onmetal oxide adsorbentⅠ.characterization and the role of metal surface hydroxyl groups [J]. Acta Scientiae Circumstantiae, 2006,10: 1586-1591.
[55] 霍麗娟.水鐵礦納米材料對土壤中砷的吸附固定及其穩(wěn)定化反應(yīng)機(jī)制[D]. 北京:中國農(nóng)業(yè)科學(xué)院, 2017. Huo L J. Study on the mechanisms of arsenic sorption and stabilizationin soils using ferrihydrite nanoparticles [D]. Beijing: Chinese Academy of Agricultural Sciences, 2017.
[56] 羅 暢.鐵氧化物吸持腐殖酸對AsO43-、Hg2+次級吸附行為的影響研究[D]. 重慶:西南大學(xué), 2015. Luo C. The secondary adsorption of AsO43-and Hg2+on iron oxides complexed with humic acids [D]. Chongqing: Southwest University, 2015.
Stabilization of As(Ⅴ) in contaminated sediment by FeMnMg-LDH.
ZHOU Hong-guang1, GAN Yan-ping1, WU De-quan2,YANG Yan-mei1*, ZHANG Yang3, WANG Lu-yao3
(1.National Engineering Research Center for Inland Waterway Regulation, Chongqing Jiaotong University, Chongqing 400074, China;2.Qianxinan Radio and Television University, Xingyi 562400, China;3.Shaanxi Province Land Engineering Construction Group, Xi'an 710075, China)., 2023,43(11):6102~6114
To address the problem of stabilized remediation of dredged sediment, FeMnMg-LDH materials were prepared by an improved co-precipitation method to achieve passivation of As in the sediment. The effect of FeMnMg-LDH on the passivation of As in dredged sediment under different As contamination levels and humic acid (HA) addition through sediment incubation tests was investigated. The results showed that the FeMnMg-LDH material induced the transformation of mild acido-soluble fraction arsenic (MASF-As) to a more stable form. With the increase of FeMnMg-LDH addition level, the leaching concentration of As in the sediment at high contamination level (250mg/kg As) was reduced to 261.35 μg/L, the stabilization efficiency was up to 80.29%, and the percentage of effective state content decreased from 13.24% to 3.21% (3.0% FeMnMg-LDH addition level). The addition of HA had a certain activation effect on As in the sediment, which increased the leaching toxicity and biological effectiveness of As in the sediment. The activation effect brought by HA could be weakened by FeMnMg-LDH passivation remediation at the activation effect brought by HA was gradually weakened with the increase of the addition amount at the restorer addition levels of 0%~3.0%. The results of correlation analysis showed that the leaching toxicity and bioeffectiveness of sediment As were positively correlated with the percentage of MASF-As and negatively correlated with the percentage of reducible arsenic (RF-As), which indicated that FeMnMg-LDH materials reduced the leaching toxicity and bioeffectiveness of As by changing the morphological distribution of As in the sediment. In conclusion, FeMnMg-LDH can be used as a passivating agent for As in substrates, providing new materials and methods for the passivation remediation of heavy metals in sediment.
arsenic;sediment;FeMnMg-LDH;humic acid;stabilization
X52
A
1000-6923(2023)11-6102-13
周宏光(1987-),男,四川達(dá)州人,副教授,博士,主要從事環(huán)境功能材料研制及水污染控制.發(fā)表論文9篇.hgzhou@cqjtu.edu.cn.
周宏光,甘艷平,伍德權(quán),等.FeMnMg-LDH對污染底泥中As(Ⅴ)的穩(wěn)定化 [J]. 中國環(huán)境科學(xué), 2023,43(11):6102-6114.
Zhou H G, Gan Y P, Wu D Q, et al. Stabilization of As(Ⅴ) in contaminated sediment by FeMnMg-LDH [J]. China Environmental Science, 2023,43(11):6102-6114.
2023-03-15
中國博士后科學(xué)基金資助項目(2018M643421);陜西省自然科學(xué)基礎(chǔ)研究計劃項目(2021JQ-959);重慶市自然科學(xué)基金資助面上項目(2022NSCQ-MSX2283)
* 責(zé)任作者, 教授, yymeicq@163.com