曾佳敏,鐘仕花,錢 偉,袁圣武,朱小山,3*

水環(huán)境中有機(jī)磷酸酯的污染現(xiàn)狀及其生物毒性

曾佳敏1,鐘仕花2,錢 偉1,袁圣武1,朱小山1,3*

(1.清華大學(xué)深圳國(guó)際研究生院海洋工程研究院,廣東 深圳 518055；2. 深圳市農(nóng)產(chǎn)品質(zhì)量安全檢驗(yàn)檢測(cè)中心,廣東 深圳 518055；3.南方海洋科學(xué)與工程廣東省實(shí)驗(yàn)室(珠海),廣東 珠海 519000)

為有效評(píng)估有機(jī)磷酸酯(OPEs)潛在的生態(tài)健康風(fēng)險(xiǎn),綜述了OPEs在全球水和沉積物中的污染現(xiàn)狀,并重點(diǎn)關(guān)注OPEs對(duì)水生生物的毒性效應(yīng),根據(jù)浮游生物、游泳生物和底棲生物等不同生物類群的特點(diǎn)分析其潛在的毒性作用機(jī)制,進(jìn)而展望本領(lǐng)域未來的研究方向和科學(xué)問題,以期有效評(píng)估OPEs的生態(tài)效應(yīng)和健康風(fēng)險(xiǎn),推動(dòng)我國(guó)OPEs食品安全監(jiān)控和生態(tài)毒理學(xué)研究,為規(guī)范其綠色應(yīng)用提供參考.

水環(huán)境；有機(jī)磷酸酯(POEs)；污染；生物富集；毒性效應(yīng)

有機(jī)磷酸酯(OPEs)是一類磷酸酯合成衍生物,因其良好的阻燃性能廣泛應(yīng)用于建材、紡織、化工以及電子等行業(yè)[1-2].近年來,隨著多溴聯(lián)苯醚(PBDEs)等溴化阻燃劑在全世界范圍內(nèi)被禁止使用, OPEs作為阻燃劑的使用量快速增長(zhǎng),從2011年的500kt增加到2018年的2800kt[1-4].OPEs根據(jù)取代基結(jié)構(gòu)不同可以分為烷烴類OPEs、含氯類OPEs和芳烴類OPEs三大類,表1列舉了常見OPEs及其理化性質(zhì).不同結(jié)構(gòu)的OPEs其使用范圍有所不同,其中含氯類OPEs和芳烴類OPEs主要作為塑料制品、紡織物、電子設(shè)備以及建筑、家裝材料的阻燃劑;烷烴類OPEs則主要用作增塑劑、去泡劑、液壓劑等[5].大部分OPEs主要以物理混合方式而非化學(xué)鍵合成方式添加到聚合物材料中,在產(chǎn)品生命周期中極易通過揮發(fā)、浸出、磨損和溶解等過程從材料中釋放到周圍環(huán)境.目前,在水和沉積物中均已檢測(cè)到大量OPEs的存在[5].通過生物體的呼吸、攝食等生命活動(dòng),也有越來越多的OPEs在生物體內(nèi)累積,并沿食物鏈進(jìn)行傳遞,可能對(duì)整個(gè)生態(tài)系統(tǒng)造成危害.因此,OPEs已被認(rèn)定為一類新型有機(jī)污染物,并受到全球的高度關(guān)注[1].早在2000年,含氯類OPEs被列入歐盟優(yōu)先控制污染物名單[6].2014年歐盟開始限制兒童玩具中磷酸三(2-氯乙基)酯(TCEP)和磷酸三(1,3-二氯異丙基)酯(TDCPP)的使用[7].美國(guó)也針對(duì)消費(fèi)品中的有機(jī)磷阻燃劑紛紛出臺(tái)禁/限用要求,包括所有鹵系有機(jī)磷酸酯阻燃劑,尤其是TCEP、TDCPP等[8].但迄今為止,關(guān)于OPEs在水環(huán)境中的賦存狀態(tài)、污染程度和毒性效應(yīng)的相關(guān)研究仍較少.為更全面有效地評(píng)估OPEs潛在的生態(tài)和健康風(fēng)險(xiǎn),本文在分析全球水環(huán)境中OPEs污染現(xiàn)狀基礎(chǔ)上,重點(diǎn)總結(jié)OPEs對(duì)水生生物的毒性效應(yīng)及其毒性機(jī)制,概括了目前研究中存在的問題,并提出未來的研究方向.

表1 常見OPEs的名稱及理化性質(zhì)

①數(shù)據(jù)來自US EPA(2020),其中l(wèi)ogKOW為正辛醇-水分配系數(shù)(Octanol-Water Partition Coefficient),WS為水中溶解度,Vr為蒸汽壓;②TCP為磷酸三鄰甲苯基酯(Tri?o?cresyl phosphate,78-30-8)、磷酸三間甲苯基酯(Tri?m?cresylphosphate,563-04-2)、磷酸三對(duì)甲苯基酯(Tri?p?cresyl phosphate,78-32-0)混合物,圖中分子結(jié)構(gòu)用磷酸三對(duì)甲苯基酯結(jié)構(gòu)式表示.

1 水環(huán)境中OPEs的污染現(xiàn)狀

水環(huán)境是OPEs的重要?dú)w宿之一.如圖1所示,大量生產(chǎn)的OPEs主要經(jīng)以下幾種途徑進(jìn)入水環(huán)境:①生產(chǎn)、生活中OPEs經(jīng)污水處理廠出水排放[9-10];②固體廢物處理廠等集中處理、填埋區(qū)滲濾液中OPEs的點(diǎn)源排放[11-12];③生產(chǎn)和生活場(chǎng)景中室內(nèi)灰塵、粉塵中OPEs無管理排放,并隨大氣環(huán)流、降雨等過程再輸入陸地及海洋中[13-15].OPEs一旦進(jìn)入水環(huán)境,便受到不同水動(dòng)力作用的影響,發(fā)生復(fù)雜的遷移、轉(zhuǎn)化,其存在形式、分布狀況及毒性效應(yīng)不斷改變.OPEs因其理化性質(zhì)的差異,在水中的穩(wěn)定性各不相同.例如,TMP極性強(qiáng)、水溶解度大且易揮發(fā),而TEHP則難溶于水且不易揮發(fā).在中性條件下,大多數(shù)磷酸三酯不易水解,但是在堿性條件下或有磷酸酯酶存在時(shí),水解過程會(huì)顯著加劇[16].含氯類OPEs如TCEP、TCPP和TDCPP在水環(huán)境中很難發(fā)生轉(zhuǎn)化或者降解[17].水溶性較差的OPEs,可能更容易吸附在沉積物中;即便部分水溶性好的OPEs,也可通過和水體中有機(jī)質(zhì)、膠體和浮游生物等顆粒相結(jié)合,再經(jīng)重力作用沉降到沉積物中.OPEs在水和沉積物中的大量存在,使得水環(huán)境被認(rèn)為是OPEs最終的“匯”.目前,全球各類水體和沉積物中,均已檢測(cè)到大量OPEs的存在(表2和3).

圖1 水環(huán)境中OPEs的來源及遷移途徑示意

1.1 水相中OPEs的污染現(xiàn)狀

表2和圖2總結(jié)了全球范圍內(nèi)部分水體包括海洋、河流、湖泊和城鎮(zhèn)污水中OPEs的濃度水平.從水中總濃度來看,OPEs在水相中的全球濃度分布極不均勻,地表水濃度最高已超過1000ng/L,例如澳洲城市地表水[18-19];但也有不少地區(qū)濃度低至1ng/L,甚至未檢出,如美國(guó)密歇根湖水或北極海水[20-21].人類的生產(chǎn)和生活活動(dòng)或是造成OPEs分布不均的關(guān)鍵因素.高度城市化地區(qū)附近的水體中OPEs含量較高,水環(huán)境中可檢測(cè)到數(shù)百ng/L或更高濃度的OPEs,例如韓國(guó)洛東江[22](483.45ng/L)、日本東京灣水體[23](284ng/L)、法國(guó)羅納河[24](128.90ng/L)、悉尼地表水[19](1060ng/L)等.而人類活動(dòng)較低的地區(qū),例如美國(guó)密歇根湖及其支流中檢測(cè)到OPEs的水平相對(duì)較低,范圍從10～50ng/L[20,25].另外,海水中OPEs的濃度比淡水中要低.即便在OPEs濃度較高的香港海域[23],其濃度值(181.93ng/L)仍遠(yuǎn)小于陸地水域中的OPEs濃度.是否由于OPEs在海洋中被大量海水稀釋,還是其他原因,仍有待查證.

表2 全球水體環(huán)境中OPEs的濃度數(shù)據(jù)(ng/L)

續(xù)表2

圖2 水環(huán)境中OPEs的濃度分布

此外,不同地區(qū)和不同類型水環(huán)境中優(yōu)勢(shì)OPEs種類明顯不同.在韓國(guó)河流、美國(guó)五大湖及其支流、日本東京灣水體中,TCEP和TCPP等含氯類OPEs的含量最為豐富[22-23,25-26].這與含氯類OPEs具有更強(qiáng)的持久性,在水中難以降解有關(guān)[6].而與國(guó)外相比,國(guó)內(nèi)檢測(cè)到的OPEs并非都以含氯類為主.TEP等烷烴類OPEs在中國(guó)黃河口、渤海灣近海水樣中分布也十分廣泛[23,27];中國(guó)南方珠江三角洲地區(qū)烷烴類OPEs占48%,含氯類OPEs占21%[23].這反映了不同地區(qū)使用的OPEs數(shù)量、類型和工業(yè)開發(fā)活動(dòng)的差異[1].

1.2 沉積相中OPEs的污染現(xiàn)狀

OPEs通過自身沉降或與水中其他物質(zhì)結(jié)合后共沉降進(jìn)入沉積物中.但目前,相對(duì)水體而言,有關(guān)沉積物中OPEs的調(diào)查還較少.表2和圖3總結(jié)了全球范圍內(nèi),OPEs在部分河流、湖泊及海洋沉積物中的濃度分布.OPEs在所有調(diào)查區(qū)域均有檢出,甚至在海底[28]和南極[29]等人跡罕至之處的沉積物中也檢測(cè)到OPEs的存在,進(jìn)一步證實(shí)OPEs的全球污染.沉積物中OPEs總濃度范圍從幾十到數(shù)百ng/g dw(干重),較其在水中的變化小,可能與OPEs在沉積物中的遷移性能較弱有關(guān),這也說明沉積物可能是OPEs的重要賦存庫(kù),可在長(zhǎng)時(shí)間尺度上不斷積累.有趣的是, OPEs在沉積物中的總濃度與其在水相中總濃度規(guī)律類似,在不同國(guó)家之間以及同一國(guó)家不同地域之間的污染程度存在明顯差異(圖3).這一現(xiàn)象很大程度上歸因于人類活動(dòng)的地域差異,以及不同地區(qū)間OPEs使用量、應(yīng)用類型和經(jīng)濟(jì)發(fā)展水平等方面的差異.例如,在海底巖芯(41.6ng/g dw)和南極沉積物(3.66ng/g dw)中檢測(cè)到較低的OPEs濃度[28-29];而在發(fā)達(dá)的歐洲Evrotas、Adige和Sava三大流域[30]、西班牙流域[31]、荷蘭西謝爾特河口[32]、美國(guó)舊金山灣區(qū)[33]的沉積物樣品中均檢測(cè)到較高濃度的OPEs.

中國(guó)沉積物中OPEs的污染現(xiàn)狀也不可小覷.截至目前,已調(diào)查區(qū)域盡管僅有渤海沿海海域[28,34]、太湖[35]、珠江三角洲地區(qū)[36-37]和廣西沿海灣區(qū)[38]等地.但所有調(diào)查區(qū)域均有檢出,其中渤海萊州灣沉積物中OPEs濃度高達(dá)300ng/g dw[34];珠江三角洲地區(qū)沉積物中也檢測(cè)到高OPEs含量,濃度范圍為8.30～ 470ng/g dw[37];太湖沉積物中OPEs總濃度相對(duì)較低,范圍是3.38～14.26ng/g dw[35].從已有數(shù)據(jù)來看,我國(guó)和世界其他地區(qū)沉積物OPEs的污染程度類似,但是在主要OPEs污染物種類方面稍有不同:歐美沉積物中含氯類OPEs如TCPP、TCEP等占總OPEs濃度的比例較中國(guó)的高;中國(guó)沉積物中主要以TnBP、TBOEP等烷烴類OPEs為主,也有一定比例的TCEP、TCPP等含氯類OPEs存在(圖3).

圖3 沉積環(huán)境中OPEs的濃度分布

表3 全球沉積物中OPEs的濃度數(shù)據(jù)(ng/g dw)

續(xù)表3

2 OPEs對(duì)水生生物的毒性效應(yīng)及機(jī)制

OPEs已在水環(huán)境中大量存在并隨時(shí)間推移不斷富集,其對(duì)水生生物和人類健康的風(fēng)險(xiǎn)受到全球環(huán)境和毒理學(xué)家的高度重視.目前,已有大量報(bào)道證實(shí)OPEs在水生生物體內(nèi)累積,但是有關(guān)OPEs毒性效應(yīng)及其作用機(jī)制的研究仍較少[34,39].最新研究證實(shí)OPEs暴露能對(duì)浮游生物、魚類和底棲生物等造成傷害(表4),但總體所用受試生物及被研究OPEs種類仍較少,且缺乏環(huán)境濃度下長(zhǎng)期慢性暴露等毒性數(shù)據(jù),相關(guān)毒理學(xué)機(jī)制也有待進(jìn)一步明確.

表4 不同OPEs對(duì)水生生物毒性效應(yīng)

注:① L(E)C50值、NOEC值來源于文獻(xiàn)[17].

2.1 OPEs對(duì)浮游生物的毒性效應(yīng)

目前,OPEs對(duì)浮游生物的毒性效應(yīng)研究仍較少,已知僅有TnBP、TDCPP和TPHP 3種OPEs的浮游植物毒性以及TnBP、TBOEP、TCEP、TCPP、TDCPP、TPHP和TCP 7種OPEs的浮游動(dòng)物毒性研究結(jié)果(表4).

如圖4所示,OPEs對(duì)浮游植物的毒性表現(xiàn)為:生長(zhǎng)抑制、滲透壓調(diào)節(jié)干擾、細(xì)胞膜損傷和細(xì)胞形態(tài)的改變,其機(jī)制與OPEs引起的氧化應(yīng)激和脂質(zhì)過氧化以及細(xì)胞的代謝功能受損有關(guān).對(duì)不同的OPEs種類而言,TDCPP(2～10mg/L)顯著抑制斜生柵藻和三角褐指藻的生長(zhǎng),隨暴露劑量增加生物量減少,表現(xiàn)出明顯的劑量-效應(yīng)關(guān)系[42-43].在植物細(xì)胞中,葉綠體被認(rèn)為是活性氧物質(zhì)(ROS)產(chǎn)生的主要場(chǎng)所[44].進(jìn)入藻類細(xì)胞后,TDCPP通過抑制光合作用、葉綠素合成和光系統(tǒng)中的光捕獲,導(dǎo)致藻體內(nèi)ROS增加,脂質(zhì)過氧化水平升高,產(chǎn)生膜破壞使個(gè)體凋亡,最終抑制種群增長(zhǎng).代謝組學(xué)和轉(zhuǎn)錄組學(xué)鑒定出52種差異代謝物,其中與脂質(zhì)代謝相關(guān)的代謝物水平下調(diào)[42];與光合作用天線蛋白、碳代謝、細(xì)胞周期和過氧化物酶體有關(guān)的基因表達(dá)也下調(diào),為上述現(xiàn)象提供了關(guān)鍵證據(jù)[43].TPHP的暴露也會(huì)抑制斜生柵藻和小球藻的生長(zhǎng),但小球藻隨著暴露時(shí)間增加,細(xì)胞生長(zhǎng)呈現(xiàn)恢復(fù)趨勢(shì).代謝組學(xué)分析表明,TPHP誘導(dǎo)小球藻脂質(zhì)代謝物(MGDG和DGDG)累積,通過呼吸增強(qiáng)補(bǔ)償應(yīng)激反應(yīng)中的能量損失,并改善膜結(jié)構(gòu)以抵消滲透壓增強(qiáng)壓力,提升耐受性;而暴露于TPHP后斜生柵藻出現(xiàn)滲透壓增強(qiáng)(纈氨酸,脯氨酸和棉子糖增加)和膜脂解代謝反應(yīng)增強(qiáng)[45]現(xiàn)象,使耐受性下降.Liu等同樣發(fā)現(xiàn),TnBP暴露下三角褐指藻細(xì)胞的變形和細(xì)胞膜受損的現(xiàn)象,并證實(shí)TnBP暴露誘導(dǎo)ROS增加,然后通過脂質(zhì)過氧化引發(fā)膜損傷,改變膜的完整性和通透性,最終觸發(fā)細(xì)胞凋亡[46].實(shí)際上,脂質(zhì)成分是細(xì)胞膜的重要結(jié)構(gòu)成分,也是細(xì)胞滲透調(diào)節(jié)的關(guān)鍵因子,此類OPEs通過氧化傷害對(duì)細(xì)胞脂質(zhì)代謝產(chǎn)生影響可能是其重要的致毒機(jī)制.

圖4 OPEs對(duì)浮游植物的毒性途徑示意(修改自文獻(xiàn)[43])

OPEs對(duì)浮游動(dòng)物的毒性效應(yīng)已經(jīng)開始受到關(guān)注.浮游動(dòng)物中大型溞和鹵蟲是兩類公認(rèn)的標(biāo)準(zhǔn)實(shí)驗(yàn)生物.表4總結(jié)了7種OPEs(TnBP、TBOEP、TCEP、TCPP、TDCPP、TPHP和TCP)對(duì)大型溞的急性毒性效應(yīng).可見,不同的OPEs其毒性效應(yīng)相差甚遠(yuǎn),芳烴類TPHP的急性48h-EC50(50% Effiective Concentration,半數(shù)有效濃度(下同))為1～1.35mg/L,其次是含氯類TDCPP為3.8～4.6mg/L,而同為含氯類的TCPP則毒性較小,其48h-EC50為131mg/L. OPEs對(duì)浮游動(dòng)物的毒性大小還與其濃度高低和暴露時(shí)間長(zhǎng)短呈正相關(guān),隨著暴露濃度增高和暴露時(shí)間增加而增強(qiáng).例如,TnBP的6h-EC50(52～93mg/L)高于24h-EC50(5.8～35mg/L);同樣,TBOEP的24h- LC50(50% Lethal Concentration,半數(shù)致死濃度(下同)) (84mg/L)高于48h-LC50(75mg/L).鹵蟲作為受試生物的OPEs急性毒性數(shù)據(jù)較少,TnBP的急性24h-EC50為54.6mg/L,TDCPP的急性48h-EC50為15～17mg/L.對(duì)比急性暴露,低濃度慢性暴露條件下浮游動(dòng)物的響應(yīng)更能反映OPEs的生態(tài)毒理效應(yīng).Giraudo等人率先報(bào)道了環(huán)境相關(guān)濃度OPEs對(duì)大型蚤的長(zhǎng)期毒性效應(yīng)[47].發(fā)現(xiàn)10μg/L TBOEP可以在大型溞中產(chǎn)生氧化應(yīng)激和內(nèi)分泌干擾作用,其效應(yīng)可在世代間遷移:例如親代(F0)中編碼過氧化氫酶()和谷胱甘肽S～轉(zhuǎn)移酶()的基因均被下調(diào),隨后在F1代中過氧化氫酶(CAT)等活性顯著降低;F0代中的卵黃蛋白原兩種亞型(和)、蛻皮激素受體()、激素核受體()和血紅蛋白()基因轉(zhuǎn)錄水平顯著升高,隨后影響了大型溞多世代生長(zhǎng)發(fā)育.對(duì)比急性暴露(48h)和慢性暴露條件(20d)下鹵蟲()的生長(zhǎng)發(fā)育,發(fā)現(xiàn)急性暴露至8mg/L TDCPP并不會(huì)影響鹵蟲體長(zhǎng),但長(zhǎng)期暴露于環(huán)境相關(guān)濃度(200 μg/L TDCPP),會(huì)顯著降低體長(zhǎng),提高死亡率,影響鹵蟲生長(zhǎng)發(fā)育;此外,兩種條件均會(huì)引起代謝變化,急性暴露觸發(fā)甘油和海藻糖等與滲透作用相關(guān)的代謝物水平升高,以調(diào)節(jié)滲透壓變化;慢性暴露則下調(diào)甘油磷脂代謝,這可能導(dǎo)致能量存儲(chǔ)機(jī)制和脂質(zhì)結(jié)構(gòu)的改變[48].可見,低濃度長(zhǎng)期暴露,會(huì)影響浮游動(dòng)物生長(zhǎng)、發(fā)育與生殖,而氧化應(yīng)激、內(nèi)分泌干擾以及參與遺傳信息處理、生物系統(tǒng)、細(xì)胞過程等生理生化代謝途徑的差異可能是其主要的毒性機(jī)制[49].總體而言,在OPEs浮游動(dòng)物毒性效應(yīng)研究中,受試的物種和OPEs種類仍偏少,環(huán)境相關(guān)濃度的暴露和慢性長(zhǎng)期暴露研究不多,亟需進(jìn)一步開展毒性機(jī)制的探討和深入分析.

此外,基于不同OPEs物化性質(zhì)的差異,在復(fù)雜的實(shí)際環(huán)境中其毒性效應(yīng)也會(huì)產(chǎn)生不同影響.例如,TPHP的LogOW(4.59)比TCEP和TBOEP的LogOW值(1.44、3.75)高,TPHP更容易吸附在溶解性有機(jī)物(DOM)上,從而改變OPE的生物利用度和亞致死毒性.Kovacevic等[50]也發(fā)現(xiàn)DOM不會(huì)改變單獨(dú)TCEP和TBOEP暴露的代謝變化,但會(huì)產(chǎn)生與單獨(dú)暴露TPHP不同的代謝反應(yīng),例如,在5mg 有機(jī)碳/L DOM,存在下,暴露于TPHP后大型溞體內(nèi)代謝產(chǎn)物百分比發(fā)生變化,其中葡萄糖含量顯著降低,亮氨酸顯著升高,而僅TPHP暴露則未觀察到.而且,不同種類OPEs間可能存在聯(lián)合毒性作用,有必要對(duì)幾種OPEs進(jìn)行聯(lián)合毒性實(shí)驗(yàn),并關(guān)注浮游生物體內(nèi)OPEs的生物轉(zhuǎn)化機(jī)制和生物轉(zhuǎn)化產(chǎn)物的毒性.例如Choi等[51]基于非目標(biāo)篩選表征大型溞中TPHP的生物轉(zhuǎn)化產(chǎn)物和途徑,發(fā)現(xiàn)TPHP主要的生物轉(zhuǎn)化機(jī)制是半胱氨酸結(jié)合和硫酸化,某些生物轉(zhuǎn)化產(chǎn)物(如磷酸二苯酯、羥基化磷酸三苯酯和巰基磷酸三苯酯)可能會(huì)引起生物毒性.

2.2 OPEs對(duì)魚類的毒性效應(yīng)

表4總結(jié)了目前OPEs的魚類毒性效應(yīng)研究結(jié)果.可見,不同的OPEs對(duì)同種魚類的L(E)C50值差別很大,反映了不同OPEs的毒性存在顯著的差異.例如TCP和TCEP對(duì)虹鱒()以及TCP和TCPP對(duì)藍(lán)鰓魚()的96h-LC50值,差距均高達(dá)1000倍以上.而且同種OPEs在不同物種中也存在差異.TPHP暴露下,美洲原銀漢魚()的96h-LC50(95mg/L)是雜色鱂()(0.31～0.56mg/L)的300倍.總體而言,TnBP、TDCPP、TPHP、TCP的毒性較強(qiáng),對(duì)水生生物具有輕微至高毒性.而暴露于TDCPP中的斑馬魚胚胎、暴露于TCP中的藍(lán)鰓魚以及暴露于TPHP中的鯽魚則十分敏感,其96h-L (E)C50均小于1,表明最容易受到侵害.

魚是人類食品的重要來源,也是OPEs水生生態(tài)毒理學(xué)研究的重點(diǎn).盡管OPEs對(duì)魚類的毒性效應(yīng),已有一定數(shù)量的研究報(bào)道,但有關(guān)OPEs的潛在致毒機(jī)制目前尚未有定論.相關(guān)的研究主要集中在OPEs對(duì)模式生物稀有鮈鯽()和斑馬魚()的神經(jīng)毒性、發(fā)育和生殖毒性等方面. Hong等[52]近期報(bào)道了芳烴類OPEs(TPHP)對(duì)成年雄性鮈鯽的毒性:TPHP顯著提高了魚腦中血腦屏障的通透性,激活了神經(jīng)炎癥反應(yīng),不利于阻止有害物質(zhì)進(jìn)出腦組織;TPHP還明顯抑制細(xì)胞增殖,并顯著降低魚小腦垂體神經(jīng)元(Ce)的樹突狀喬化,導(dǎo)致魚的學(xué)習(xí)和記憶功能受損.其他OPEs,如TDCPP,在200 μg/L處理下,神經(jīng)營(yíng)養(yǎng)因子(例如,,和)的轉(zhuǎn)錄受到明顯抑制,這表明OPEs可能通過靶向神經(jīng)營(yíng)養(yǎng)因子和神經(jīng)膠質(zhì)蛋白引起神經(jīng)毒性作用[53].除了神經(jīng)毒性,OPEs對(duì)水生生物的遺傳毒性也日益受到關(guān)注.以TDCPP為例(圖5),它顯著激活了DNA損傷反應(yīng)(DDR)信號(hào),進(jìn)而改變與DNA損傷相關(guān)的途徑,包括DNA修復(fù)抑制以及促進(jìn)細(xì)胞凋亡和細(xì)胞周期停滯等[54].Hong等[55]則通過高通量測(cè)序(HTS)評(píng)估m(xù)icroRNA和isomiR (microRNA的變體)轉(zhuǎn)錄效果,并以此來評(píng)估三類OPEs(TBOEP、TDCPP、TPHP)對(duì)稀有鮈鯽的肝臟毒性,發(fā)現(xiàn)TBOEP,TDCPP和TPHP暴露條件下均分別檢測(cè)到不同種microRNA的差異表達(dá).

圖5 TDCPP誘導(dǎo)的稀有鮈鯽DNA損傷的系統(tǒng)觀點(diǎn)[54]

OPEs對(duì)水生生物生長(zhǎng)和發(fā)育的毒性研究主要以斑馬魚為受試生物.Zeng等[56]還報(bào)道了環(huán)境相關(guān)濃度烷烴類TBOEP對(duì)斑馬魚生長(zhǎng)的影響,發(fā)現(xiàn)雌性體內(nèi)的TBOEP平均含量均高于雄性,體長(zhǎng)和體重顯著減少,且相關(guān)基因在生長(zhǎng)激素/胰島素樣生長(zhǎng)因子(GH/IGF)軸和下丘腦～垂體～甲狀腺(HPT)軸上的轉(zhuǎn)錄受到影響,雌性血漿中甲狀腺素(T4)含量降低,同時(shí)促甲狀腺激素β亞基()mRNA水平降低.為了更加明確TBOEP暴露后的性別差異和親本轉(zhuǎn)移, Huang等[57]進(jìn)一步研究后認(rèn)為TBOEP暴露所呈現(xiàn)出的明顯性別差異,可能與產(chǎn)卵期間TBOEP的排出有關(guān).斑馬魚子代幼魚中檢測(cè)到TBOEP殘留也證明上述觀點(diǎn).此外,暴露于TBOEP,斑馬魚性腺發(fā)育出現(xiàn)明顯延遲,下丘腦～垂體～性腺(HPG)軸基因轉(zhuǎn)錄發(fā)生改變[57].HPG軸在生殖系統(tǒng)的發(fā)育和調(diào)節(jié)中起著至關(guān)重要的作用.下丘腦分泌促性腺激素釋放激素刺激垂體分泌卵泡刺激素和黃體生成激素,進(jìn)而刺激性腺釋放17-雌二醇(E2)和睪丸激素(T).TBOEP暴露導(dǎo)致血漿中E2和T濃度上升,且較低濃度的TBOEP僅能改變雌性的T/E2比值,表明TBEOP對(duì)女性的生殖影響更大[57].除此之外,斑馬魚暴露于含氯類和芳烴類OPEs也可以顯著增加血漿中T和E2的濃度[58-59].因此,OPEs可能是通過影響生物體內(nèi)激素的平衡,進(jìn)而影響生長(zhǎng),產(chǎn)生發(fā)育毒性;長(zhǎng)期OPEs暴露還將導(dǎo)致該污染物向其后代轉(zhuǎn)移,產(chǎn)生代際毒性.

在OPEs的發(fā)育毒性方面.近來已探明含氯類OPE(TDCPP)導(dǎo)致斑馬魚尾鰭畸形的分子機(jī)制與尾鰭發(fā)育相關(guān)的轉(zhuǎn)錄因子的錯(cuò)誤表達(dá)有關(guān)[60].TDCPP還是斑馬魚胚胎中血管和肌肉發(fā)育的有效破壞者.Dishaw等[61]發(fā)現(xiàn)10μmol/L TDCPP具有致畸性和明顯的毒性,100μmol/L可顯著改變斑馬魚的游泳活動(dòng),顯示TDCPP的毒性可能與神經(jīng)發(fā)育和肌肉發(fā)育的破壞有關(guān).Li等[62]也發(fā)現(xiàn)短鏈含氯類OPEs (TCPP和TCEP)對(duì)斑馬魚胚胎神經(jīng)發(fā)育的不利影響.與前人對(duì)鮈鯽的神經(jīng)毒性研究[53]相似,TCEP和TCPP對(duì)乙酰膽堿的含量和乙酰膽堿酯酶的活性沒有影響,但污染物暴露顯著下調(diào)了與神經(jīng)發(fā)育相關(guān)的選定基因和蛋白質(zhì)的表達(dá),造成顯著的發(fā)育毒性.

近年來,借助組學(xué)技術(shù),OPEs的毒理學(xué)機(jī)制研究取得了新的進(jìn)展.以TPHP為例,高通量蛋白質(zhì)組學(xué)研究表明其急性暴露對(duì)斑馬魚幼魚的發(fā)育毒性與視蛋白基因的表達(dá)改變有關(guān)[63]:TPHP暴露引起五種視蛋白基因(視紫紅質(zhì)zfrho,紫外線視蛋白,藍(lán)色視蛋白,綠色視蛋白,,,,紅色視蛋白)的轉(zhuǎn)錄顯著下調(diào).而代謝組學(xué)分析發(fā)現(xiàn),TPHP還顯著抑制斑馬魚氨基酸代謝,降低纈氨酸、亮氨酸和異亮氨酸水平,抑制氨酰tRNA生物合成過程,同時(shí)引起葡萄糖糖酵解過程和三羧酸循環(huán)發(fā)生障礙[64].以上原因可能是TPHP引起斑馬魚發(fā)育畸形的主要原因.盡管目前僅針對(duì)TPHP開展研究,但上述研究結(jié)果為深入理解OPEs的水生生物毒性機(jī)制奠定了堅(jiān)實(shí)的基礎(chǔ).

2.3 OPEs對(duì)底棲生物的毒性效應(yīng)

與水相相比,沉積物中含有更多的OPEs,底棲動(dòng)物可能更易于受到OPEs的污染威脅.但目前大多數(shù)OPEs的毒性研究都以水層生物為受試生物,有關(guān)OPEs對(duì)底棲生物的毒性仍缺乏系統(tǒng)性認(rèn)知.目前僅有少量研究以亞洲淡水蛤()和海洋貽貝()為受試生物考察OPEs對(duì)底棲生物的潛在負(fù)面影響.

烷烴類OPEs(TBOEP和TnBP)長(zhǎng)期暴露對(duì)亞洲淡水蛤的抗氧化酶活性(CAT、SOD)和丙二醇(MDA)含量產(chǎn)生影響,具體表現(xiàn)為CAT和SOD活性隨暴露劑量(20～2000 μg/L)的增加呈現(xiàn)先降低,后增加的趨勢(shì)[3].可能是SOD和CAT在低濃度時(shí)無法及時(shí)清除過量的ROS,導(dǎo)致機(jī)體抗氧化活性下降,但在高濃度暴露下,由于淡水蛤閉殼效應(yīng)的影響產(chǎn)生虹吸行為導(dǎo)致抗氧化酶活性呈增加趨勢(shì),并使MDA含量降低.對(duì)于含氯類OPEs(TCPP和TDCPP),蛤消化腺通過大量攝取TDCPP產(chǎn)生ROS,誘導(dǎo)SOD活性上升,使MDA含量顯著增加,隨后干擾細(xì)胞正常的氧化還原平衡并降解蛋白質(zhì)、DNA和脂質(zhì),導(dǎo)致細(xì)胞凋亡[65];而貽貝血細(xì)胞內(nèi)SOD活性和MDA含量的升高,同樣證實(shí)TCPP的氧化傷害是導(dǎo)致細(xì)胞凋亡的原因之一[66].芳烴類OPE(TPHP)與含氯類OPEs類似,可導(dǎo)致貽貝消化腺中抗氧化酶(GPX、SOD和CAT)活性與基因轉(zhuǎn)錄水平的上調(diào),表現(xiàn)出明顯的氧化應(yīng)激現(xiàn)象[67-68].上述研究揭示,OPEs對(duì)底棲生物的毒性效應(yīng)可能在于引發(fā)機(jī)體氧化應(yīng)激,使抗氧化功能紊亂,最終產(chǎn)生損害;該過程受OPEs種類、暴露時(shí)間、劑量以及生物生理生化行為(如閉殼效應(yīng)和虹吸行為)等因素的多重影響.值得一提的是,雙殼類生物虹吸行為對(duì)其具有一定的保護(hù)功能,反之,該行為的抑制則可作為機(jī)體功能紊亂的標(biāo)志[3].

底棲生物對(duì)OPEs暴露的響應(yīng),除了氧化應(yīng)激之外,體內(nèi)OPEs的生物轉(zhuǎn)化或解毒過程是決定其毒性效應(yīng)的另一個(gè)關(guān)鍵.一般而言,生物體通過激活I(lǐng)期和II期生物轉(zhuǎn)化和多異生物抗性(MXR)系統(tǒng)來保護(hù)生物免受外源污染物侵害.例如,研究證實(shí)TDCPP可在亞洲淡水蛤消化腺內(nèi)容易發(fā)生生物轉(zhuǎn)化,BDCPP是其主要的I期轉(zhuǎn)化產(chǎn)物;與母體TDCPP相比,BDCPP減低消化腺細(xì)胞凋亡、顯著緩解虹吸抑制,降低對(duì)底棲生物的毒性[65].此外,GST(谷胱甘肽S轉(zhuǎn)移酶)和細(xì)胞色素P450作為重要的生物轉(zhuǎn)化酶,也參與了底棲生物消化腺OPEs的生物轉(zhuǎn)化和解毒過程[3].例如,TDCPP暴露下,亞洲淡水蛤細(xì)胞色素P450家族蛋白基因()、谷胱甘肽系統(tǒng)基因()的表達(dá)水平隨暴露劑量(50～5000μg/L)顯著降低[65];暴露于TBOEP后,谷胱甘肽系統(tǒng)基因(、)表達(dá)水平也顯著降低[3].除上述生物轉(zhuǎn)化酶外,其他生物標(biāo)記物(如神經(jīng)肽[69]、乙酰膽堿酯酶(AChE)[70]等)也受到重視.已證實(shí),在長(zhǎng)期低水平TDCPP(10μg/L)暴露下,貽貝體內(nèi)AChE活性的抑制和蛋白質(zhì)分子水平上受體型酪氨酸蛋白磷酸酶N2型(PTPRN2)的下調(diào)均標(biāo)志著該化合物可引起神經(jīng)毒性[70].

上述研究表明,OPEs的底棲生物毒性不容忽視,盡管已有少量前瞻性研究,但明確OPEs底棲生物毒性機(jī)制、開發(fā)更有效的生物標(biāo)志物等仍有待更多的研究.

3 結(jié)語

OPEs已經(jīng)在許多國(guó)家和地區(qū)、甚至是極地水域中被檢出,檢出濃度逐漸升高,其潛在的負(fù)面效應(yīng)不容忽視;檢出的OPEs以烷烴類和含氯類為主,但具體種類及濃度水平在不同的國(guó)家和地區(qū)間有較大差異,這與人類活動(dòng)的地域差異、不同地區(qū)間OPEs使用量、應(yīng)用類型和經(jīng)濟(jì)發(fā)展水平等相關(guān).

進(jìn)一步對(duì)OPEs的水生生物毒性及其作用機(jī)制進(jìn)行分析,發(fā)現(xiàn)OPEs暴露會(huì)對(duì)包括浮游生物、魚類和底棲生物在內(nèi)的整個(gè)水生生態(tài)系統(tǒng)造成傷害,主要表現(xiàn)為生長(zhǎng)抑制、組織病變、發(fā)育和生殖毒性、神經(jīng)毒性和細(xì)胞凋亡等;其毒性作用機(jī)制十分復(fù)雜,既與基因的轉(zhuǎn)錄和表達(dá)、蛋白質(zhì)的表達(dá)有關(guān),也有激素代謝調(diào)節(jié)和神經(jīng)遞質(zhì)因子的變化有關(guān),酶活性的抑制以及氧化應(yīng)激也被證明是可能的致毒機(jī)制.

當(dāng)前,OPEs的環(huán)境與生物毒性數(shù)據(jù)不夠完善, 缺乏OPEs水生生物富集及沿食物鏈傳遞和生物放大的研究,且濃度的大小、暴露時(shí)間、環(huán)境條件以及毒性實(shí)驗(yàn)的設(shè)計(jì)和條件等多種因素均會(huì)影響OPEs的毒性表現(xiàn),難以得出OPEs致毒機(jī)制的一致規(guī)律,更難以對(duì)其環(huán)境效應(yīng)和健康風(fēng)險(xiǎn)實(shí)施有效的評(píng)估.

為深入了解OPEs的生態(tài)毒性及其健康效應(yīng),未來需建立一套相對(duì)完整, 科學(xué)的OPEs毒性測(cè)試標(biāo)準(zhǔn)方法,包括OPEs物理化學(xué)性質(zhì)表征, 模型生物選取, 暴露方式和毒性效應(yīng)指標(biāo)等,保證不同的調(diào)查和實(shí)驗(yàn)結(jié)果間的可比性. 考慮到OPEs存在較高的生物富集及其食物鏈傳遞潛能,在關(guān)注OPEs高劑量急性效應(yīng)(當(dāng)前毒性研究的主要內(nèi)容)的同時(shí),更需要關(guān)注OPEs長(zhǎng)期低劑量或環(huán)境水平的暴露及其生態(tài)毒性效應(yīng),包括在生物體內(nèi)的歸趨、富集和遺傳毒性等, 也包括OPEs沿食物鏈傳遞的影響,及對(duì)種群,群落及生態(tài)系統(tǒng)產(chǎn)生的生態(tài)風(fēng)險(xiǎn),以使研究結(jié)果更加貼近真實(shí)環(huán)境的情況.

另外,在自然環(huán)境中,污染物并非孤立存在,而是與其他污染物混合存在且受各種環(huán)境因子影響,不僅會(huì)改變其遷移、轉(zhuǎn)化和生物毒性,往往還對(duì)生物體產(chǎn)生聯(lián)合毒性效應(yīng).因此需要重視OPEs在環(huán)境條件下的生物轉(zhuǎn)化和水解或光降解過程.經(jīng)過轉(zhuǎn)化或降解,OPEs是否可能生成毒性更強(qiáng)/更弱的代謝產(chǎn)物或副產(chǎn)物(如OP二酯和羥基化代謝物等)亟需深入研究.也需要加強(qiáng)研究OPEs在水環(huán)境中與其他共存污染物可能產(chǎn)生的協(xié)同、拮抗等復(fù)合污染生態(tài)效應(yīng).

[1] Van der Veen I, De Boer J. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis [J]. Chemosphere, 2012,88:1119-1153.

[2] Brandsma S H, Boer J D, Leonards P E G, et al. Organophosphorus flame-retardant and plasticizer analysis, including recommendations from the first worldwide interlaboratory study [J]. Trends in Analytical Chemistry, 2013,43(2):217-228.

[3] Yan S, Wu H, Qin J, et al. Halogen-free organophosphorus flame retardants caused oxidative stress and multixenobiotic resistance in Asian freshwater clams () [J]. Environmental Pollution, 2017,225:559-568.

[4] Israel Chemicals Limited. Tel Aviv, Israel. Worldwide flame retardants market to reach 2.8million tonnes in 2018 [J]. Additives for Polymers, 2015(4):11.

[5] 王曉偉,劉景富,陰永光.有機(jī)磷酸酯阻燃劑污染現(xiàn)狀與研究進(jìn)展[J]. 化學(xué)進(jìn)展, 2010,22(10):1983-1992.

Wang X W, Liu J F, Yin Y G. The pollution status and research progress on organophosphate ester flame retardants [J]. Progress in Chemistry, 2010,22(10):1983-1992.

[6] Reemtsma T, Quintana J B, Rodil R,et al. Organophosphorus flame retardants and plasticizers in water and air I. Occurrence and fate [J]. Trends in Analytical Chemistry, 2008,27(9):727-737.

[7] 菁 菁.歐盟玩具安全法例收緊阻燃劑限制[J]. 中國(guó)質(zhì)量技術(shù)監(jiān)督, 2014:82.

Jing J. EU toy safety legislation tightens flame retardant restrictions [J]. China Quality Supervision, 2014:82.

[8] Bay area compliance labs corp.全球法規(guī)對(duì)阻燃劑禁限用要求匯總[EB/OL]. http://www.baclcorp.com.cn/show.asp?para=cn_2_7_2680/ 2019-12-03.

Bay area compliance labs corp. Summary of global regulations on the prohibition and restriction of flame retardants [EB/OL]. http://www. baclcorp.com.cn/show.asp?para=cn_2_7_2680/2019-12-03.

[9] Wang Y, Kannan P, Halden R U, et al. A nationwide survey of 31organophosphate esters in sewage sludge from the United States [J]. Science of the Total Environment, 2019,655:446-453.

[10] Lee S, Cho H J, Choi W, et al. Organophosphate flame retardants (OPFRs) in water and sediment: Occurrence, distribution, and hotspots of contamination of Lake Shihwa, Korea [J]. Marine Pollution Bulletin, 2018,130:105-112.

[11] Schwarzbauer J, Heim S, Brinker S,et al. Occurrence and alteration of organic contaminants in seepage and leakage water from a waste deposit landfill [J]. Water Research, 2002,36:2275–2287.

[12] Yadav I C, Devi N L, Li J, et al. Organophosphate ester flame retardants in Nepalese soil: Spatial distribution, source apportionment and air-soil exchange assessment [J]. Chemosphere, 2018,190:114- 123.

[13] 劉 琴,印紅玲,李 蝶,等.室內(nèi)灰塵中有機(jī)磷酸酯的分布及其健康風(fēng)險(xiǎn)[J]. 中國(guó)環(huán)境科學(xué), 2017,37(8):2831-2839.

Liu Q, Yin H L, Li D, et al. Distribution characteristic of OPEs in indoor dust and its health risk [J]. China Environmental Science, 2017, 37(8):2831-2839.

[14] Huang Y, Tan H, Li L, et al. A broad range of organophosphate tri- and di-esters in house dust from Adelaide, South Australia: Concentrations, compositions, and human exposure risks [J]. Environment International, 2020,142:105872.

[15] Tan H, Peng C, Guo Y, et al. Organophosphate flame retardants in house dust from South China and related human exposure risks [J]. Bulletin of Environmental Contamination and Toxicology, 2017,99: 344-349.

[16] Arabameri M, Mohammadi M M, Monjazeb M L, et al. Pesticide residues in pistachio nut: a human risk assessment study [J]. International Journal of Environmental Analytical Chemistry, 2020: 1-14.

[17] Verbruggen E M J, Rila J P, Traas T P, et al. Environmental risk limits for several phosphate esters, with possible application as flame retardant [R]. National Institute for Public Health and the Environment. 2005.

[18] Martinez-Carballo E, Gonzalez-Barreiro C, Sitka A, et al. Determination of selected organophosphate esters in the aquatic environment of Austria [J]. Science of the Total Environment, 2007, 388:290-299.

[19] Teo T L L, Mcdonald J A, Coleman H M, et al. Analysis of organophosphate flame retardants and plasticisers in water by isotope dilution gas chromatography–electron ionisation tandem mass spectrometry [J]. Talanta, 2015,143:114-120.

[20] Guo J, Venier M, Salamova A, et al. Bioaccumulation of dechloranes, organophosphate esters, and other flame retardants in Great Lakes fish [J]. Science of the Total Environment, 2017,583:1-9.

[21] Gao X, Huang P, Huang Q, et al. Organophosphorus flame retardants and persistent, bioaccumulative, and toxic contaminants in Arctic seawaters: On-board passive sampling coupled with target and non-target analysis [J]. Environmental Pollution, 2019,253:1-10.

[22] Choo G, Cho H S, Park K, et al. Tissue-specific distribution and bioaccumulation potential of organophosphate flame retardants in crucian carp [J]. Environmental Pollution, 2018,239:161-168.

[23] Lai N L S, Kwok K Y, Wang X H, et al. Assessment of organophosphorus flame retardants and plasticizers in aquatic environments of China (Pearl River Delta, South China Sea, Yellow River Estuary) and Japan (Tokyo Bay) [J]. Journal of Hazardous Materials, 2019,371:288-294.

[24] Schmidt N, Castro-Jimenez J, Fauvelle V, et al. Occurrence of organic plastic additives in surface waters of the Rhone River (France) [J]. Environmental Pollution, 2020,257:113637.

[25] Guo J, Romanak K, Westenbroek S, et al. Current-use flame retardants in the water of Lake Michigan tributaries [J]. Environmental Science & Technology, 2017,51:9960-9969.

[26] Lee S, Jeong W, Kannan K, et al. Occurrence and exposure assessment of organophosphate flame retardants (OPFRs) through the consumption of drinking water in Korea [J]. Water Research, 2016, 103:182-188.

[27] Wang R, Tang J, Xie Z, et al. Occurrence and spatial distribution of organophosphate ester flame retardants and plasticizers in 40 rivers draining into the Bohai Sea, north China [J]. Environmental Pollution, 2015,198:172-178.

[28] Liao C, Kim U J, Kannan K. Occurrence and distribution of organophosphate esters in sediment from northern Chinese coastal waters [J]. Science of the Total Environment, 2020,704:135328.

[29] Fu J, Fu K, Gao K, et al. Occurrence and trophic magnification of organophosphate esters in an Antarctic ecosystem: Insights into the shift from legacy to emerging pollutants [J]. Journal of Hazardous Materials, 2020,396:122742.

[30] Giulivo M, Capri E, Kalogianni E, et al. Occurrence of halogenated and organophosphate flame retardants in sediment and fish samples from three European river basins [J]. Science of the Total Environment, 2017,586:782-791.

[31] Cristale J, Garcia Vazquez A, Barata C, et al. Priority and emerging flame retardants in rivers: occurrence in water and sediment, Daphnia magna toxicity and risk assessment [J]. Environment International, 2013,59:232-243.

[32] Brandsma S H, Leonards P E G, Leslie H A, et al. Tracing organophosphorus and brominated flame retardants and plasticizers in an estuarine food web [J]. Science of the Total Environment, 2015, 505:22-31.

[33] Sutton R, Chen D, Sun J, et al. Characterization of brominated, chlorinated, and phosphate flame retardants in San Francisco Bay, an urban estuary [J]. Science of the Total Environment, 2019,652:212- 223.

[34] Bekele T G, Zhao H, Wang Q, et al. Bioaccumulation and trophic transfer of emerging organophosphate flame retardants in the marine food webs of Laizhou Bay, North China [J]. Environmental Science & Technology, 2019,53:13417-13426.

[35] Cao S, Zeng X, Song H, et al. Levels and distributions of organophosphate flame retardants and plasticizers in sediment from Taihu Lake, China [J]. Environmental Toxicology and Chemistry, 2012, 31(7):1478-1484.

[36] Liu Y E, Luo X J, Zapata Corella P, et al. Organophosphorus flame retardants in a typical freshwater food web: Bioaccumulation factors, tissue distribution, and trophic transfer [J]. Environmental Pollution, 2019,255:113286.

[37] Tan X X, Luo X J, Zheng X B, et al. Distribution of organophosphorus flame retardants in sediments from the Pearl River Delta in South China [J]. Science of the Total Environment, 2016,544:77-84.

[38] Zhang R, Yu K, Li A, et al. Occurrence, phase distribution, and bioaccumulation of organophosphate esters (OPEs) in mariculture farms of the Beibu Gulf, China: A health risk assessment through seafood consumption[J]. Environmental Pollution, 2020,263:114426.

[39] Wang X, Zhong W, Xiao B, et al. Bioavailability and biomagnification of organophosphate esters in the food web of Taihu Lake, China: Impacts of chemical properties and metabolism [J]. Environment International, 2019,125:25-32.

[40] Xing L, Tao M, Zhang Q, et al. Occurrence, spatial distribution and risk assessment of organophosphate esters in surface water from the lower Yangtze River Basin [J]. Science of the Total Environment, 2020, 734:139380.

[41] Luo Q, Gu L, Wu Z, et al. Distribution, source apportionment and ecological risks of organophosphate esters in surface sediments from the Liao River, Northeast China [J]. Chemosphere, 2020,250:126297.

[42] Wang L, Huang X, Laserna A K C, et al. Metabolomics reveals that tris(1,3-dichloro-2-propyl)phosphate (TDCPP) causes disruption of membrane lipids in microalga[J]. Science of the Total Environment, 2020,708:134498.

[43] Liu Q, Tang X, Jian X, et al. Toxic effect and mechanism of tris (1,3-dichloro-2-propyl)phosphate (TDCPP) on the marine alga[J]. Chemosphere, 2020,252:126467.

[44] Gill S S, Tuteja N. Reactive oxygen species and antioxidant machinery in abiotic stress tolerance in crop plants [J]. Plant Physiology and Biochemistry, 2010,48:909-930.

[45] Wang L, Huang X, Lim D J, et al. Uptake and toxic effects of triphenyl phosphate on freshwater microalgaeand: Insights from untargeted metabolomics [J]. Science of the Total Environment, 2019,650:1239-1249.

[46] Liu Q, Tang X, Wang Y, et al. ROS changes are responsible for tributyl phosphate (TBP)-induced toxicity in the alga[J]. Aquatic Toxicology, 2019,208:168-178.

[47] Giraudo M, Dube M, Lepine M, et al. Multigenerational effects evaluation of the flame retardant tris(2-butoxyethyl) phosphate (TBOEP) using[J]. Aquatic Toxicology, 2017,190: 142-149.

[48] Morgan M A, Griffith C M, Volz D C, et al. TDCIPP exposure affectsgrowth and osmoregulation [J]. Science of the Total Environment, 2019,694:133486.

[49] Yuan S, Li H, Dang Y, et al. Effects of triphenyl phosphate on growth, reproduction and transcription of genes of[J]. Aquatic Toxicology, 2018,195:58-66.

[50] Kovacevic V, Simpson A J, Simpson M J. Investigation ofSub-Lethal Exposure to Organophosphate Esters in the Presence of Dissolved Organic Matter Using1H NMR-Based Metabolomics [J]. Metabolites, 2018,8:16.

[51] Choi Y, Jeon J, Choi Y, et al. Characterizing biotransformation products and pathways of the flame retardant triphenyl phosphate inusing non-target screening [J]. Science of the Total Environment , 2020,708:135106.

[52] Hong X, Chen R, Hou R, et al. Triphenyl phosphate (TPHP)-induced neurotoxicity in adult male Chinese rare minnows () [J]. Environmental Science & Technology, 2018,52:11895-11903.

[53] Yuan L, Li J, Zha J, et al. Targeting neurotrophic factors and their receptors, but not cholinesterase or neurotransmitter, in the neurotoxicity of TDCPP in Chinese rare minnow adults () [J]. Environmental Pollution, 2016,208:670-677.

[54] Chen R, Hou R, Hong X, et al. Organophosphate flame retardants (OPFRs) induce genotoxicity in vivo: A survey on apoptosis, DNA methylation, DNA oxidative damage, liver metabolites, and transcriptomics [J]. Environment International, 2019,130:104914.

[55] Hong X, Chen R, Yuan L, et al. Global microRNA and isomiR expression associated with liver metabolism is induced by organophosphorus flame retardant exposure in male Chinese rare minnow () [J]. Science of the Total Environment, 2019,649:829-838.

[56] Zeng X, Sun H, Huang Y, et al. Effects of environmentally relevant concentrations of tris (2-butoxyethyl) phosphate on growth and transcription of genes involved in the GH/IGF and HPT axes in zebrafish () [J]. Chemosphere, 2018,212:376-384.

[57] Huang Y, Liu J, Yu L, et al. Gonadal impairment and parental transfer of tris (2-butoxyethyl) phosphate in zebrafish after long-term exposure to environmentally relevant concentrations [J]. Chemosphere, 2019,218:449-457.

[58] Liu X, Ji K, Choi K. Endocrine disruption potentials of organophosphate flame retardants and related mechanisms in H295R and MVLN cell lines and in zebrafish [J]. Aquatic Toxicology, 2012, 114-115:173-181.

[59] Liu X, Cai Y, Wang Y, et al. Effects of tris(1,3-dichloro-2-propyl) phosphate (TDCPP) and triphenyl phosphate (TPP) on sex-dependent alterations of thyroid hormones in adult zebrafish [J]. Ecotoxicology and Environmental Safety, 2019,170:25-32.

[60] Rhyu D, Lee H, Tanguay R L, et al. Tris(1,3-dichloro-2-propyl) phosphate (TDCIPP) disrupts zebrafish tail fin development [J]. Ecotoxicology and Environmental Safety, 2019,182:109449.

[61] Dishaw L V, Hunter D L, Padnos B, et al. Developmental exposure to organophosphate flame retardants elicits overt toxicity and alters behavior in early life stage zebrafish () [J]. Toxicological Sciences, 2014,142(2):445-454.

[62] Li R, Wang H, Mi C, et al. The adverse effect of TCPP and TCEP on neurodevelopment of zebrafish embryos/larvae [J]. Chemosphere, 2019,220:811-817.

[63] Shi Q, Tsui M M P, Hu C, et al. Acute exposure to triphenyl phosphate (TPhP) disturbs ocular development and muscular organization in zebrafish larvae [J]. Ecotoxicology and Environmental Safety, 2019, 179:119-126.

[64] 張杏麗,鄒 威,周啟星.基于代謝組學(xué)技術(shù)分析磷酸三苯酯誘導(dǎo)斑馬魚胚胎發(fā)育毒性的分子機(jī)制 [J]. 生態(tài)毒理學(xué)報(bào), 2019,14:79-89.

Zhang X L, Zou W, Zhou Q X．Molecular mechanisms of developmental toxicity of triphenyl phosphate on zebrafish embryo revealed by metabonomics [J]. Asian Journal of Ecotoxicology, 2019, 14:79-89.

[65] Li D, Wang P, Wang X, et al. Elucidating multilevel toxicity response differences between tris(1,3-dichloro-2-propyl) phosphate and its primary metabolite in[J]. Science of the Total Environment, 2020,749:142049.

[66] Wu H, Zhong M, Lu Z, et al. Biological effects of tris (1-chloro- 2-propyl) phosphate (TCPP) on immunity in mussel[J]. Environmental Toxicology and Pharmacology, 2018,61:102-106.

[67] Meng X, Li F, Wang X, et al. Combinatorial immune and stress response, cytoskeleton and signal transduction effects of graphene and triphenyl phosphate (TPP) in mussel[J]. Journal of Hazardous Materials, 2019,378:120778.

[68] Meng X, Li F, Wang X, et al. Toxicological effects of graphene on musselhemocytes after individual and combined exposure with triphenyl phosphate[J]. Marine Pollution Bulletin, 2020,151:110838.

[69] Wang Q, Hong X, Chen H, et al. The neuropeptides of Asian freshwater clam () as new molecular biomarker basing on the responses of organophosphate chemicals exposure [7]. Ecotoxicology and Environmental Safety, 2018,160:52-59.

[70] Sanchez-Marin P, Vidal-Linan L, Fernandez-Gonzalez L E, et al. Proteomic analysis and biochemical alterations in marine mussel gills after exposure to the organophosphate flame retardant TDCPP [J]. Aquatic Toxicology, 2021,230:105688.

[71] Yu L, Jia Y, Su G, et al. Parental transfer of tris(1,3-dichloro-2-propyl) phosphate and transgenerational inhibition of growth of zebrafish exposed to environmentally relevant concentrations [J]. Environmental Pollution, 2017,220:196-203.

[72] Ren X, Wang W, Zhao X, et al. Parental exposure to tris(1,3- dichloro-2-propyl) phosphate results in thyroid endocrine disruption and inhibition of growth in zebrafish offspring [J]. Aquatic Toxicology, 2019,209:132-141.

[73] Shi Q, Wang Z, Chen L, et al. Optical toxicity of triphenyl phosphate in zebrafish larvae [J]. Aquatic Toxicology, 2019,210:139-147.

[74] Yan S, Wang Q, Yang L, et al. Comparison of the Toxicity Effects of Tris(1,3-dichloro-2-propyl)phosphate (TDCIPP) with Tributyl Phosphate (TNBP) Reveals the Mechanism of the Apoptosis Pathway in Asian Freshwater Clams () [J]. Environmental Science & Technology, 2020,54:6850-6858.

Pollution status and ecotoxicity of organophosphate esters (OPEs) in aquatic environment.

ZENG Jia-min1, ZHONG Shi-hua2, QIAN Wei1, YUAN Sheng-wu1, ZHU Xiao-shan1,3*

(1.Institute of Marine Engineering, Shenzhen International Graduate School, Tsinghua University, Shenzhen 518055, China；2.Shenzhen Agricultural Products Quality and Safety Inspection and Testing Center, Shenzhen 518055, China；3.Guangdong Laboratory of Southern Ocean Science and Engineering (Zhuhai), Zhuhai 519000, China)., 2021,41(9)：4388～4401

To effectively assess the potential ecological health risks of organophosphate esters (OPEs), this study gave an overview of the global OPE pollution in water and sediment with a focus on their toxic effects on aquatic organisms. The potential toxicity mechanisms were also analyzed in different kinds of biota including plankton, nekton and benthos, and the future research directions and scientific issues in aquatic environmental studies were finally prospected. This study would be helpful for effective assessment of ecological effects and health risks of OPEs. It would also positively promote the food safety monitoring and ecotoxicology research of OPEs, and provide reference for regulating their green applications.

aquatic environment；organophosphate esters (OPEs)；pollution；bioaccumulation；toxicity

X52

A

1000-6923(2021)09-4388-14

曾佳敏(1997-),女,湖南常德人,清華大學(xué)環(huán)境工程專業(yè)碩士研究生,主要從事有機(jī)磷酸酯生態(tài)環(huán)境毒理研究.發(fā)表論文1篇.

2021-02-07

國(guó)家自然科學(xué)基金(41877352,42077227);廣東省基礎(chǔ)與應(yīng)用基礎(chǔ)研究基金項(xiàng)目(2021A1515010158);深圳市基礎(chǔ)研究重點(diǎn)項(xiàng)目(JCYJ20180507182227257)

* 責(zé)任作者, 副研究員, zhu.xiaoshan@sz.tsinghua.edu.cn