持久性有機(jī)污染物在水生食物網(wǎng)中的傳遞行為

馮秋園,萬(wàn) 祎,劉學(xué)勤,劉 永,*

1 北京大學(xué)環(huán)境科學(xué)與工程學(xué)院,水沙科學(xué)教育部重點(diǎn)實(shí)驗(yàn)室,北京 100871 2 北京大學(xué)城市與環(huán)境學(xué)院,地球表面過(guò)程教育部重點(diǎn)實(shí)驗(yàn)室,北京 100871 3 中國(guó)科學(xué)院水生生物研究所,武漢 430072

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持久性有機(jī)污染物在水生食物網(wǎng)中的傳遞行為

馮秋園1,萬(wàn) 祎2,劉學(xué)勤3,劉 永1,*

1 北京大學(xué)環(huán)境科學(xué)與工程學(xué)院,水沙科學(xué)教育部重點(diǎn)實(shí)驗(yàn)室,北京 100871 2 北京大學(xué)城市與環(huán)境學(xué)院,地球表面過(guò)程教育部重點(diǎn)實(shí)驗(yàn)室,北京 100871 3 中國(guó)科學(xué)院水生生物研究所,武漢 430072

食物網(wǎng)是持久性有機(jī)污染物(POPs)在水生生態(tài)系統(tǒng)中傳遞的重要途徑,了解其傳遞行為與機(jī)制是POPs生態(tài)暴露風(fēng)險(xiǎn)評(píng)價(jià)的科學(xué)基礎(chǔ)。從4個(gè)方面展開了討論和分析：①食物網(wǎng)主要特征(營(yíng)養(yǎng)級(jí)和食物鏈長(zhǎng)度)與POPs環(huán)境行為的關(guān)系；②POPs在底棲及底棲-浮游耦合食物網(wǎng)中的環(huán)境行為；③微食物網(wǎng)對(duì)POPs環(huán)境行為的作用；④食物網(wǎng)的變化對(duì)POPs環(huán)境行為的影響。主要結(jié)論如下：①已有研究對(duì)水生生物中POPs生物放大作用存在較大爭(zhēng)議。一般營(yíng)養(yǎng)級(jí)越高,POPs生物富集性越強(qiáng),但由于各種生態(tài)和生理性質(zhì)的影響,也存在例外情況。食物鏈長(zhǎng)度與POPs生物富集性呈正相關(guān)。②POPs通過(guò)底棲食物網(wǎng)將沉積物中的POPs向上傳遞,底棲-浮游食物網(wǎng)的耦合提高了高營(yíng)養(yǎng)級(jí)消費(fèi)者的暴露風(fēng)險(xiǎn),目前就POPs在底棲食物網(wǎng)中的生物放大性是否大于浮游食物網(wǎng)存在爭(zhēng)議。③微生物具有較大的比表面積,是吸附POPs的重要載體。另,沉積物中的微生物通過(guò)分解有機(jī)質(zhì),將POPs釋放到水柱中。微生物降解也是環(huán)境中POPs脫離環(huán)境的重要途徑。④在內(nèi)、外壓力下,食物網(wǎng)結(jié)構(gòu)和功能發(fā)生變化,使物質(zhì)和能量的傳遞方向和效率發(fā)生改變,并與環(huán)境理化性質(zhì)的變化互相耦合,影響POPs的環(huán)境行為。當(dāng)前研究的重點(diǎn)多集中在POPs在浮游食物網(wǎng),尤其是高營(yíng)養(yǎng)級(jí)浮游食物網(wǎng)中的環(huán)境行為,對(duì)POPs在底棲及底棲-浮游耦合食物網(wǎng)和微食物網(wǎng)中環(huán)境行為的研究相對(duì)缺乏。有關(guān)POPs在食物網(wǎng)中環(huán)境行為的研究多集中在食物網(wǎng)的某個(gè)部分,時(shí)間尺度較短,缺乏對(duì)POPs環(huán)境行為動(dòng)態(tài)變化的研究,未來(lái)需深入開展多尺度和多角度的POPs在食物網(wǎng)中環(huán)境行為的動(dòng)態(tài)變化研究。新型POPs的生產(chǎn)和使用量不斷增加,但有關(guān)其在食物網(wǎng)中環(huán)境行為的相關(guān)分析還較為匱乏,需加強(qiáng)研究。

持久性有機(jī)污染物；傳遞行為；營(yíng)養(yǎng)級(jí)；食物鏈長(zhǎng)度；浮游-底棲耦合食物網(wǎng)；微食物網(wǎng)；食物網(wǎng)變化

持久性有機(jī)污染物(POPs)具有持久性、生物蓄積性、高毒性、半揮發(fā)性和遠(yuǎn)距離傳輸?shù)奶攸c(diǎn),會(huì)對(duì)人類健康和生態(tài)安全產(chǎn)生嚴(yán)重危害。POPs通過(guò)大氣干濕沉降、地面徑流和水-氣界面直接交換進(jìn)入水生生態(tài)系統(tǒng),是POPs在生態(tài)圈中傳遞的一個(gè)重要的“匯”。POPs進(jìn)入水體后,被生物體通過(guò)被動(dòng)擴(kuò)散和攝食作用吸收,并沿食物網(wǎng)傳遞,成為POPs在水生生態(tài)系統(tǒng)中非常重要的環(huán)境行為,同時(shí)使頂級(jí)消費(fèi)者受到高水平的POPs暴露風(fēng)險(xiǎn),因此研究POPs在水生生態(tài)系統(tǒng)食物網(wǎng)中的環(huán)境行為成為國(guó)內(nèi)外研究的重點(diǎn)和熱點(diǎn)[1- 3],但當(dāng)前研究的重點(diǎn)仍集中在POPs在浮游食物網(wǎng)的傳遞,對(duì)在底棲及底棲-浮游耦合食物網(wǎng)和微食物網(wǎng)的研究仍不夠充分[4- 5],對(duì)各種內(nèi)、外源環(huán)境壓力下,食物網(wǎng)變化對(duì)POPs環(huán)境行為影響的研究也相對(duì)缺乏[6- 9]。鑒于此,本文從食物網(wǎng)的主要特征與POPs環(huán)境行為的關(guān)系,POPs在底棲及底棲-浮游耦合食物網(wǎng)和微食物網(wǎng)中的環(huán)境行為及影響因素,以及目前較為關(guān)注的幾種內(nèi)外源壓力下食物網(wǎng)變化對(duì)POPs環(huán)境行為的影響等4個(gè)方面,闡述了POPs在水生食物網(wǎng)中的環(huán)境行為,為揭示POPs在水生生態(tài)系統(tǒng)中的歸趨機(jī)制,判斷水生生物的POPs暴露風(fēng)險(xiǎn),制定科學(xué)的環(huán)境標(biāo)準(zhǔn)提供參考。

1 食物網(wǎng)主要特征與POPs環(huán)境行為的關(guān)系

生物體所處的營(yíng)養(yǎng)級(jí)和食物鏈長(zhǎng)度是食物網(wǎng)的主要特征,是研究POPs在食物網(wǎng)中生物富集或生物放大時(shí)的重點(diǎn)考慮因素。在不同的生態(tài)系統(tǒng)中,其影響往往是不同的,了解特定生態(tài)系統(tǒng)中營(yíng)養(yǎng)級(jí)和食物鏈長(zhǎng)度與POPs生物富集或生物放大之間的相關(guān)關(guān)系,并探索影響這種相關(guān)關(guān)系的主要因素一直以來(lái)是生態(tài)學(xué)的熱點(diǎn)問(wèn)題。

1.1 營(yíng)養(yǎng)級(jí)與POPs生物富集性的關(guān)系

生物放大作用是營(yíng)養(yǎng)級(jí)對(duì)POPs生物富集性影響的直接體現(xiàn)。一般情況下,營(yíng)養(yǎng)級(jí)越高,POPs生物放大效應(yīng)越明顯。研究發(fā)現(xiàn),除 PFOA 外,爬行動(dòng)物、哺乳動(dòng)物等高營(yíng)養(yǎng)級(jí)水生動(dòng)物對(duì)PFCs(PFNA、PFDA、PFUA、PFDOA和PFOS)的生物富集(BAFs)系數(shù)大于5000,顯著高于無(wú)脊椎動(dòng)物和魚類,存在明顯的生物放大現(xiàn)象[10- 12]。諸多研究發(fā)現(xiàn)[13-14]生物營(yíng)養(yǎng)級(jí)(15N值)和PCBs濃度之間顯著正相關(guān),生物放大因子大于1。通常認(rèn)為POPs生物富集性隨營(yíng)養(yǎng)級(jí)升高而增大的一個(gè)重要原因是脂質(zhì)含量增加,但實(shí)際上并不是所有食物網(wǎng)中生物體的脂質(zhì)含量都隨15N的升高而增加。另有研究表明,單位脂質(zhì)POPs含量與營(yíng)養(yǎng)級(jí)之間仍有非常顯著的正相關(guān)關(guān)系,說(shuō)明營(yíng)養(yǎng)級(jí)增加本身對(duì)POPs生物富集也具有重要影響,但營(yíng)養(yǎng)級(jí)和脂質(zhì)含量的相對(duì)貢獻(xiàn)大小較難評(píng)價(jià)[15- 16]。

但污染物并不總是隨營(yíng)養(yǎng)級(jí)的升高而顯示生物放大效應(yīng)。Neff等[17]、Varanasi等[18]和Broman等[19]在波羅的海的食物鏈中沒(méi)有發(fā)現(xiàn)PAHs的生物放大現(xiàn)象,推測(cè)是由于營(yíng)養(yǎng)級(jí)越高,生物體對(duì)PAHs的分解活性越高,與其它地區(qū)的類似研究結(jié)果一致[20-22]。Mizukawa等[23]通過(guò)污染物濃度與生物體15N之間的相關(guān)關(guān)系,分析了日本東京灣中20種PBEDs和PCBs的生物放大現(xiàn)象,發(fā)現(xiàn)PCBs和部分PBDEs存在生物放大現(xiàn)象,另幾種PBDE同系物的濃度則隨營(yíng)養(yǎng)級(jí)升高而下降,其原因可能是在生物體內(nèi)被轉(zhuǎn)化或代謝分解了。Nfon等[24]分析了具有不同理化性質(zhì)的POPs在波羅的海浮游和底棲食物網(wǎng)中的傳遞,發(fā)現(xiàn)所有PCBs同系物和OCPs存在顯著的食物鏈放大作用(FCMFs>1),但由于生物體的代謝分解作用,PAHs和PCNs的食物鏈放大系數(shù)(FCMFs)小于1,是營(yíng)養(yǎng)稀釋作用。上述研究說(shuō)明,POPs的食物鏈放大作用受到生物轉(zhuǎn)化和代謝分解的限制,特別是在較高營(yíng)養(yǎng)級(jí)中[25- 26]。另有研究發(fā)現(xiàn),一般低營(yíng)養(yǎng)級(jí)浮游食物網(wǎng)不存在污染物的生物放大或很低,在高營(yíng)養(yǎng)級(jí)消費(fèi)者體內(nèi)存在明顯的生物放大,因?yàn)榈蜖I(yíng)養(yǎng)級(jí)生物對(duì)能量的需求量較低,因此食物的攝入量相對(duì)更低,而食物攝入是POPs富集的一個(gè)重要途徑[27-29]。Lundgren等[30]發(fā)現(xiàn)在亞北極地區(qū)波的尼亞灣的底棲食物網(wǎng)中大多數(shù)高氯代PCNs沒(méi)有生物放大現(xiàn)象,推測(cè)其原因是由于高氯代PCNs體積較大,阻礙了其在生物體內(nèi)的擴(kuò)散和吸收。Evenest研究了幾種POPs在斯的巴特爾群島的孔斯峽灣底棲食物網(wǎng)中的生物富集情況,發(fā)現(xiàn)除了順式九氯,沒(méi)有觀測(cè)到POPs的生物放大現(xiàn)象,幾種無(wú)脊椎動(dòng)物體內(nèi)的POPs濃度等于甚至超過(guò)了魚類或絨鴨體內(nèi)的POPs濃度,原因是這些底棲無(wú)脊椎動(dòng)物是食腐性的,POPs的生物轉(zhuǎn)化和排出能力低,且直接暴露在POPs濃度較高的沉積物間隙水中[31]。

營(yíng)養(yǎng)級(jí)對(duì)POPs生物富集性的影響通常是通過(guò)生物放大或營(yíng)養(yǎng)級(jí)放大來(lái)分析,一般情況下,POPs的生物富集性隨營(yíng)養(yǎng)級(jí)的升高而增強(qiáng),但也有例外,因?yàn)槠渌鞣N生態(tài)和生理性質(zhì),如獵物的豐度、生長(zhǎng)速率、生物與沉積物的接觸程度、攝食POM的量、微生物環(huán)導(dǎo)致的高15N和低POPs濃度等都不同程度地控制著POPs的生物富集,使POPs的實(shí)際濃度偏離了根據(jù)營(yíng)養(yǎng)級(jí)預(yù)測(cè)得到的值[31-33]。另外,目前生物所處營(yíng)養(yǎng)級(jí)一般用15N來(lái)表示,生物放大性通過(guò)15N與POPs濃度的相關(guān)關(guān)系來(lái)分析,但是對(duì)于底棲食物網(wǎng),單純地利用15N有時(shí)并不能很好地指示底棲動(dòng)物所處的實(shí)際營(yíng)養(yǎng)級(jí),因?yàn)槟承┥锘蛩鶖z食的生物具有同位素分餾作用,例如,一些攝食POM的生物,由于細(xì)菌環(huán)(bacteria loop)的分解和同位素分餾作用,POPs暴露風(fēng)險(xiǎn)降低,但是15N的值卻提高了[34]；其次,15N基準(zhǔn)值的季節(jié)性變化也會(huì)影響15N 和POPs相關(guān)關(guān)系的斜率；再者,很多底棲無(wú)脊椎動(dòng)物是雜食性/食腐性,因此導(dǎo)致POPs傳遞的再循環(huán),使POPs的負(fù)荷較高,但是單純地分析15N可能并不能反映出這種循環(huán)。

目前,有關(guān)POPs在水生食物網(wǎng)中的生物放大作用具有很大的爭(zhēng)議,如Bruner等[35]、Losser等[36]和Gobas等[37]等認(rèn)為生物放大在整個(gè)食物網(wǎng)都存在,但Borga等[38]認(rèn)為在低營(yíng)養(yǎng)級(jí)中,由于生物體攝食吸收的POPs可以忽略,所以不存在生物放大作用。Guruge等[39]認(rèn)為生物放大只存在于直接呼吸空氣的生物體內(nèi)。當(dāng)前,對(duì)營(yíng)養(yǎng)級(jí)升高造成的POPs生物富集性增強(qiáng),主要還是從脂質(zhì)含量增加來(lái)解釋[40],但實(shí)際上,營(yíng)養(yǎng)級(jí)升高本身也是一個(gè)重要的影響因素,但是其具體影響機(jī)制紛繁復(fù)雜,存在很多疑問(wèn)和不確定性。Gobas等[31]推測(cè)一是因?yàn)榇蟛糠稚锪吭谘貭I(yíng)養(yǎng)級(jí)傳遞過(guò)程中轉(zhuǎn)化成了能量被耗散,造成污染物的濃縮,二是由于食物在胃腸道中被消化,提高了化學(xué)物質(zhì)的逸度。綜上發(fā)現(xiàn),營(yíng)養(yǎng)級(jí)與POPs在食物網(wǎng)中生物富集性之間的關(guān)系,及潛在的具體機(jī)制仍存在很多疑問(wèn)。在具體的研究中,需要根據(jù)生物體的生理特征,結(jié)合POPs理化性質(zhì)和生態(tài)系統(tǒng)的特點(diǎn)綜合考慮,才能得到較為準(zhǔn)確的預(yù)測(cè)結(jié)果。

1.2 食物鏈長(zhǎng)度與POPs生物放大效應(yīng)的關(guān)系

目前研究認(rèn)為,食物鏈延長(zhǎng)會(huì)增加頂級(jí)消費(fèi)者體內(nèi)的污染物濃度[2,41- 42]。Kidd等[43]發(fā)現(xiàn)加拿大拉柏吉湖中,魚體內(nèi)的毒殺酚濃度異常高,明顯高于其它湖泊的同種魚類。通過(guò)對(duì)比湖泊的污染輸入、營(yíng)養(yǎng)級(jí)結(jié)構(gòu)等發(fā)現(xiàn),拉柏吉湖比其它湖泊的食物鏈更長(zhǎng),同種魚在拉柏吉湖中的N15要顯著高于其他湖泊,從而導(dǎo)致其體內(nèi)有異常高的污染物濃度,由此揭開了食物鏈長(zhǎng)度會(huì)提高POPs生物放大效應(yīng)的認(rèn)識(shí)。Whittle等[44]通過(guò)研究蘇必利爾湖、安大略湖、休倫湖、伊利湖的食物鏈長(zhǎng)度和生物放大系數(shù)發(fā)現(xiàn)食物鏈長(zhǎng)度和生物放大系數(shù)呈正相關(guān),進(jìn)一步證明食物鏈延長(zhǎng)會(huì)提高POPs的生物放大作用。另有諸多研究發(fā)現(xiàn)食物鏈延長(zhǎng)會(huì)導(dǎo)致生物體脂質(zhì)含量增加,那POPs生物富集的提高是否只是由于食物鏈延長(zhǎng)引起的脂質(zhì)含量增加而導(dǎo)致的？根據(jù)是否存在糠蝦和草食性魚類,Rasmussen等[2]對(duì)北美五大湖區(qū)的81個(gè)湖泊劃分為3類,通過(guò)對(duì)比頂級(jí)消費(fèi)者湖紅點(diǎn)鮭(Salvelinusnamaycush) 體內(nèi)的PCBs 濃度發(fā)現(xiàn),增加1個(gè)營(yíng)養(yǎng)級(jí),湖紅點(diǎn)鮭(Salvelinusnamaycush)體內(nèi)的PCBs濃度增加3.5倍,但脂質(zhì)含量只增加了1.5倍,因此認(rèn)為食物鏈延長(zhǎng)本身也會(huì)導(dǎo)致POPs在頂級(jí)消費(fèi)者體內(nèi)生物富集性的提高。Bentzen等[40]進(jìn)一步分析其原因得到：①糠蝦具有晝夜遷移行為,會(huì)攝食部分沉積物,導(dǎo)致其體內(nèi)的POPs濃度比其它同營(yíng)養(yǎng)級(jí)浮游生物更高；②捕食浮游動(dòng)物的魚類通過(guò)捕食降低了浮游動(dòng)物的生物量,降低了生物量稀釋作用,提高了POPs濃度；③頂級(jí)消費(fèi)者由捕食浮游動(dòng)物改為捕食體型較大且營(yíng)養(yǎng)級(jí)相對(duì)較高的糠蝦,攝食能耗降低,脂質(zhì)含量增加[45],POPs濃度升高。

2 底棲及底棲-浮游耦合食物網(wǎng)中POPs的環(huán)境行為

目前,有關(guān)POPs在食物網(wǎng)中傳遞的研究多集中在浮游食物網(wǎng),但底棲食物網(wǎng)是生態(tài)系統(tǒng)中物質(zhì)循環(huán)和能量流動(dòng)的重要環(huán)節(jié),并通過(guò)與浮游食物網(wǎng)相耦合,共同影響POPs在水生生態(tài)系統(tǒng)中的傳遞。據(jù)研究：①底棲初級(jí)和次級(jí)生產(chǎn)力的貢獻(xiàn)非常大,大多魚類對(duì)底棲初級(jí)、次級(jí)生產(chǎn)力的依賴程度甚至高于浮游初級(jí)和次級(jí)生產(chǎn)力[46- 48]。②底棲大型無(wú)脊椎動(dòng)物在營(yíng)養(yǎng)鹽的循環(huán)和輸入方面起到非常重要的作用[49]。③底棲細(xì)菌的生產(chǎn)力往往高于浮游細(xì)菌[50]。另外,底棲生境是POPs在水生生態(tài)系統(tǒng)中循環(huán)流動(dòng)的“匯”,底棲生物通過(guò)攝食和被動(dòng)吸收從沉積物中獲取POPs,并通過(guò)底棲食物網(wǎng)向上傳遞到頂級(jí)消費(fèi)者體內(nèi),是POPs循環(huán)的關(guān)鍵部分。但POPs在浮游和底棲食物網(wǎng)中的傳遞特征具有一定的差異。Kidd等[16]在馬拉維湖研究發(fā)現(xiàn),同一污染物在浮游食物網(wǎng)中生物富集程度大于底棲食物網(wǎng),其原因是同一營(yíng)養(yǎng)級(jí)的底棲(藻類和腹足類等)生物比浮游動(dòng)物、毛翅類昆蟲、腹足類生物的脂質(zhì)含量低,且底棲藻類的生長(zhǎng)率更高,產(chǎn)生“生長(zhǎng)稀釋”作用。Nfon等[24]發(fā)現(xiàn)PCBs在底棲食物鏈中的食物鏈放大系數(shù)要低于浮游食物鏈,推測(cè)是由于浮游生物缺少分解PCBs的酶。但Campfens等[51]認(rèn)為沉積物中POPs的逸度通常遠(yuǎn)大于水體,底棲生物受到的污染會(huì)更嚴(yán)重。目前,已普遍認(rèn)識(shí)到底棲食物網(wǎng)對(duì)POPs環(huán)境行為的重要影響,但底棲食物網(wǎng)在POPs傳遞過(guò)程中所起的作用仍莫衷一是,傳遞機(jī)制尚不清晰,未來(lái)還需要進(jìn)一步的深入研究。

早期的研究中,底棲食物網(wǎng)和浮游食物網(wǎng)被認(rèn)為是互相獨(dú)立的,但是隨著研究的深入,逐漸認(rèn)識(shí)到二者是通過(guò)各種方式耦合在一起的,通過(guò)物質(zhì)和能量的互相傳遞和交換共同支持生態(tài)系統(tǒng)的功能[52],POPs在此過(guò)程中隨物質(zhì)循環(huán)、能流流動(dòng)出現(xiàn)耦合。一般情況下,浮游-底棲食物網(wǎng)主要通過(guò)以下方式耦合：懸浮顆粒從水柱中沉降,實(shí)現(xiàn)浮游生境對(duì)底棲動(dòng)物的營(yíng)養(yǎng)供給[53- 56]；底棲生物降解顆粒有機(jī)質(zhì)為營(yíng)養(yǎng)鹽,并再循環(huán)進(jìn)入水柱[57- 58],底棲沉積物的再懸浮加強(qiáng)浮游-底棲耦合過(guò)程[59]；雜食性魚類捕食生境多變,能廣泛捕食浮游、底棲生物及一些陸地的無(wú)脊椎動(dòng)物[47],從而實(shí)現(xiàn)不同生境之間物質(zhì)和能量的傳遞與交換[60- 61]。在上述過(guò)程中,POPs隨之一起實(shí)現(xiàn)浮游-底棲食物網(wǎng)的耦合,因此只分析一種POPs傳遞路徑,不能準(zhǔn)確地判斷頂級(jí)捕食者體內(nèi)POPs的來(lái)源和暴露風(fēng)險(xiǎn)。例如,通過(guò)底棲-浮游食物網(wǎng)的耦合,沉積物中的有機(jī)碳可以再循環(huán)進(jìn)入到浮游生物體內(nèi),同時(shí)將沉積物中與有機(jī)碳相絡(luò)合的POPs通過(guò)食物網(wǎng)向上傳遞,沉積物中積累多年的POPs使得水柱中捕食者體內(nèi)的污染物濃度維持在一定的水平[15,62]?？肺r等白天在湖泊底部攝食底棲生物,晚上遷移到上層水柱中,捕食浮游生物,實(shí)現(xiàn)能量和POPs的浮游-底棲耦合。通過(guò)模擬糠蝦的3種生活情景：①只生活在浮游生境；②部分時(shí)間生活在浮游生境,部分時(shí)間生活在底棲生境；③只生活在底棲生境。模擬結(jié)果顯示情景①中糠蝦體內(nèi)的PCBs濃度最小,情景③最大,情景②介于兩者之間,且其模擬結(jié)果與實(shí)測(cè)值最相符[92]。這說(shuō)明了浮游-底棲食物網(wǎng)的耦合,及耦合提高了浮游動(dòng)物和魚類的POPs暴露風(fēng)險(xiǎn)。

在過(guò)去,底棲生境往往只作為污染物在水生生態(tài)系統(tǒng)中傳遞的 “匯”,且由于采樣困難等原因,POPs在底棲食物網(wǎng)中環(huán)境行為的研究相對(duì)較少,但了解POPs在底棲食物網(wǎng)中的環(huán)境行為是解析POPs在水生生態(tài)系統(tǒng)中歸趨機(jī)制必不可少的內(nèi)容,需進(jìn)一步加強(qiáng)研究。另外,底棲食物網(wǎng)通過(guò)多種過(guò)程和浮游食物網(wǎng)耦合,對(duì)POPs的環(huán)境行為產(chǎn)生復(fù)雜影響,但目前對(duì)POPs在底棲-浮游耦合食物網(wǎng)中的歸趨機(jī)制知之甚少,是未來(lái)探索的一個(gè)重點(diǎn)。

3 微食物網(wǎng)對(duì)POPs環(huán)境行為的作用

微生物,包括病毒、細(xì)菌、鞭毛蟲、纖毛蟲、浮游植物和微型的浮游動(dòng)物,在生態(tài)系統(tǒng)中具有非常重要的作用,這些生物形成一個(gè)非常復(fù)雜的微食物網(wǎng)。微食物網(wǎng)通過(guò)以下機(jī)制影響食物網(wǎng)的結(jié)構(gòu)、功能：①礦化分解有機(jī)質(zhì)為浮游植物提供營(yíng)養(yǎng)鹽；②有機(jī)碎屑消耗殆盡時(shí),細(xì)菌會(huì)與浮游植物競(jìng)爭(zhēng)無(wú)機(jī)營(yíng)養(yǎng)鹽[63- 64]。③細(xì)菌可以替代小型浮游動(dòng)物,緩解初級(jí)生產(chǎn)者的被捕食壓力。④沉積物中的細(xì)菌是大型無(wú)脊椎動(dòng)物的食物來(lái)源,但有關(guān)這一點(diǎn)存在較大的爭(zhēng)議[65- 66]。研究表明,微食物網(wǎng)在POPs的傳遞過(guò)程中也起到重要作用,微生物吸附是POPs傳遞的一個(gè)重要途徑,可以通過(guò)微食物網(wǎng)向高營(yíng)養(yǎng)級(jí)生物傳遞。微生物豐度高、體型小、周轉(zhuǎn)速率快、具有最大的比表面積,在水生生態(tài)系統(tǒng)中異養(yǎng)細(xì)菌通常占到了生物表面積的80%,因此成為一個(gè)非常重要的POPs吸附體。且由于微生物體型較小,POPs在微生物與水相之間很快達(dá)到平衡,幾乎不受“生長(zhǎng)稀釋”的影響,因此微生物成為比浮游植物更重要的POPs吸收介質(zhì)[67]。另外,顆粒物沉降過(guò)程中有機(jī)質(zhì)被不斷分解,在一個(gè)較短的時(shí)間內(nèi),細(xì)菌分解不會(huì)影響PCBs在顆粒物與水之間的分配,導(dǎo)致PCBs濃度不斷升高[68]。同時(shí),礦化分解提高了顆粒物的表面積,增加了PCBs的吸附[69- 70]。當(dāng)顆粒物沉降到沉積物表層時(shí),微生物豐度較大,有機(jī)質(zhì)被大量分解,然后大部分PCBs再次進(jìn)入到水柱中,進(jìn)行再循環(huán)。由此可見,微食物網(wǎng)對(duì)POPs在水生生態(tài)系統(tǒng)中的傳遞和再循環(huán)起到了非常重要的作用。另外,眾所周知,微生物分解是POPs離開水生生態(tài)系統(tǒng)的重要途徑,但由于微食物網(wǎng)采集及分析存在諸多困難,其相關(guān)研究還非常缺乏。

4 主要內(nèi)外源壓力下的食物網(wǎng)變化及對(duì)POPs環(huán)境行為的影響

食物網(wǎng)是POPs在水生生態(tài)系統(tǒng)中傳遞的關(guān)鍵路徑,在受到一系列的內(nèi)源驅(qū)動(dòng)和外源壓力時(shí),食物網(wǎng)的組成、結(jié)構(gòu)和功能會(huì)發(fā)生很大的變化,從而使食物網(wǎng)的動(dòng)力學(xué)特征發(fā)生較大的變化。其中,季節(jié)變化會(huì)使水生生態(tài)系統(tǒng)發(fā)生規(guī)律性變化,探索由于季節(jié)演替導(dǎo)致的POPs在環(huán)境介質(zhì)與食物網(wǎng)之間,及食物網(wǎng)內(nèi)部環(huán)境行為的規(guī)律性變化是生態(tài)學(xué)研究的重點(diǎn)之一；另外,富營(yíng)養(yǎng)化和外來(lái)物種入侵或引進(jìn)導(dǎo)致的生態(tài)和環(huán)境災(zāi)變是目前國(guó)際上水生生態(tài)系統(tǒng)普遍存在的兩個(gè)問(wèn)題,在這種外源壓力下,水生生態(tài)系統(tǒng)的結(jié)構(gòu)和功能會(huì)發(fā)生顯著變化,甚至發(fā)生穩(wěn)態(tài)轉(zhuǎn)換,從而使POPs在食物網(wǎng)中的環(huán)境行為和歸趨機(jī)制發(fā)生變化。因此,本文從季節(jié)變化、富營(yíng)養(yǎng)化和外來(lái)物種入侵或引進(jìn)這3個(gè)較為常見和普遍關(guān)注的內(nèi)、外源壓力入手,探討了食物網(wǎng)變化對(duì)POPs環(huán)境行為的影響。

4.1 食物網(wǎng)的季節(jié)演替對(duì)POPs傳遞的影響

光照、溫度、風(fēng)和徑流輸入等的季節(jié)變化會(huì)導(dǎo)致水生生態(tài)系統(tǒng)的理化性質(zhì),生物體的生理特征和種群結(jié)構(gòu)發(fā)生周期性變化,進(jìn)而引起食物網(wǎng)結(jié)構(gòu)和功能的變化,使POPs的環(huán)境行為發(fā)生變化。

研究認(rèn)為,溫度、光照、營(yíng)養(yǎng)鹽和水動(dòng)力條件的變化使浮游植物的生物量及物種組成發(fā)生季節(jié)演替,浮游植物是食物網(wǎng)中物質(zhì)循環(huán)和能量流動(dòng)的起點(diǎn),通過(guò)“上行控制”作用使高營(yíng)養(yǎng)級(jí)消費(fèi)者的棲息環(huán)境、生長(zhǎng)率、食性等發(fā)生較大的變化,進(jìn)而引起整個(gè)食物網(wǎng)的變化。在諸多地區(qū)的浮游和底棲食物網(wǎng)中都觀察到了明顯的季節(jié)變化[31，71- 73]。研究表明,水生生態(tài)系統(tǒng)中POPs濃度具有明顯的季節(jié)變化,7月份最低、9月底最高,之后又不斷下降,其原因主要是由于藻類生產(chǎn)力的季節(jié)變化造成的[74- 75]。Nizzetto[8]研究發(fā)現(xiàn),浮游動(dòng)物體內(nèi)的PCBs濃度在水華后期劇烈下降,生物富集系數(shù)(BAF)最低,不存在生物放大作用,在水華爆發(fā)前期和爆發(fā)期存在明顯的生物放大作用。因?yàn)樵谒A期間,浮游動(dòng)、植物快速生長(zhǎng),生物量劇烈變化,通過(guò)“生長(zhǎng)稀釋”和 “生物量稀釋”改變了PCBs的暴露風(fēng)險(xiǎn)。另外,浮游動(dòng)物物種組成、捕食者和生理學(xué)特征的季節(jié)性變化也會(huì)影響POPs的生物富集性。魚類的生長(zhǎng)率和種群結(jié)構(gòu)與初級(jí)生產(chǎn)力和浮游生物的密度密切相關(guān),浮游動(dòng)、植物的季節(jié)性演替,使得魚類的攝食習(xí)性也有周期性的變化,進(jìn)而導(dǎo)致魚類POPs生物富集性的季節(jié)變化[44]。另外,各種不同魚類其單位脂質(zhì)POPs濃度的季節(jié)變化不一樣,其原因目前還不是十分明確,推測(cè)可能是食性不同和遷移方式不同造成的[27,31]。

此外,浮游植物的季節(jié)性生長(zhǎng)會(huì)影響有機(jī)質(zhì)的沉降,有研究發(fā)現(xiàn)一年中大約有70%脂質(zhì)沉降發(fā)生在春季水華時(shí)期,與此同時(shí)POPs沉降量也最大,底棲生境的POPs暴露風(fēng)險(xiǎn)增加[76],而在冬季,脂質(zhì)的沉降量較低,POPs主要在浮游食物網(wǎng)中富集[67]。底棲群落的能量很大程度上依靠浮游生境的浮游碎屑[53],同位素分析顯示浮游植物是底棲無(wú)脊椎動(dòng)物的主要食物來(lái)源[77]。浮游碎屑沉降量的顯著季節(jié)變化,使底棲群落的生存環(huán)境和食物資源發(fā)生顯著的季節(jié)變化[65],進(jìn)而使底棲大型無(wú)脊椎動(dòng)物的脂質(zhì)含量發(fā)生季節(jié)變化[78],而脂質(zhì)含量是影響POPs生物富集的主要因素；另外,攝食作用是生物體POPs的主要來(lái)源之一,因此浮游生態(tài)系統(tǒng)中污染物的季節(jié)性變化會(huì)反映到底棲群落中,但其變化趨勢(shì)和幅度可能會(huì)存在差異,因?yàn)楹芏嗟讞珓?dòng)物是雜食性的,能夠緩解季節(jié)變化的影響,而且新沉降的物質(zhì)能夠很快地與老的沉積物混合(生物擾動(dòng)),使得POPs分布特征發(fā)生混合?？傮w上,底棲食物網(wǎng)生物富集性的季節(jié)性變化不如浮游食物網(wǎng)顯著。另外,不同底棲種群之間的季節(jié)變化趨勢(shì)和幅度也各不相同,因?yàn)樗鼈兊纳鷳B(tài)特征(攝食模式、死亡率)和生理特征(氧吸收速率、生物轉(zhuǎn)化能力、新陳代謝等)不同,這些都會(huì)影響污染物吸收、傳遞、排泄[31]。Hummel等[79]和Capuzzo等[80]的研究發(fā)現(xiàn),底棲軟體動(dòng)物體內(nèi)的PCBs濃度受到季節(jié)性產(chǎn)卵和生殖的影響,會(huì)發(fā)生明顯的季節(jié)變化,生殖會(huì)減少軟體動(dòng)物體內(nèi)50%—66%的PCBs含量[81]。

環(huán)境理化性質(zhì)和生物因素的季節(jié)變化造成POPs在環(huán)境和食物網(wǎng)中的分配和行為發(fā)生變化。如,夏季水溫升高,底棲生物和微生物活性增強(qiáng),表層水流變慢和熱分層出現(xiàn),加劇了底部的缺氧,促進(jìn)DDT厭氧分解,DDD濃度升高[82]。另外,夏季水溫升高改變生物的呼吸率和膜的通透性,POPs吸收速率加快,生物富集性增強(qiáng),POPs暴露風(fēng)險(xiǎn)升高[83]。

季節(jié)演替通過(guò)溫度、風(fēng)速、光照、DO等因素的周期性變化,驅(qū)動(dòng)食物網(wǎng)發(fā)生規(guī)律性變動(dòng),造成POPs在食物網(wǎng)中的環(huán)境行為的季節(jié)性變化。環(huán)境介質(zhì)、POPs的性質(zhì)與食物網(wǎng)變化互相影響,共同作用于POPs的環(huán)境行為,但目前往往只考慮某一方面對(duì)POPs環(huán)境行為的影響,對(duì)三者之間的聯(lián)系關(guān)注不夠。另外,大多數(shù)研究只集中在食物網(wǎng)的某一環(huán)節(jié),如低營(yíng)養(yǎng)級(jí)浮游食物網(wǎng),缺乏從整體角度的探索,且研究的時(shí)間尺度較短,多集中在某年的特定季節(jié),難以發(fā)掘其規(guī)律性變化。因此,未來(lái)的研究需要從多角度、整體性和長(zhǎng)時(shí)間尺度上深入研究POPs在水生食物網(wǎng)中環(huán)境行為的季節(jié)性變化。

4.2 富營(yíng)養(yǎng)化條件下食物網(wǎng)變化對(duì)POPs環(huán)境行為的影響

富營(yíng)養(yǎng)化和POPs污染是水生態(tài)系統(tǒng)目前遇到的兩個(gè)重要環(huán)境問(wèn)題,這兩者之間往往是同時(shí)發(fā)生、互相聯(lián)系的。傳統(tǒng)上,富營(yíng)養(yǎng)化和POPs污染的研究通常是分開的,但后來(lái)的研究發(fā)現(xiàn),營(yíng)養(yǎng)水平能夠影響污染物的循環(huán)和可生物利用水平[8](圖1)。瑞典環(huán)境保護(hù)局于1995年開展了揭示白令海附近海域及湖泊的富營(yíng)養(yǎng)與污染物(POPs,微量金屬)之間互相聯(lián)系的5年研究計(jì)劃(EUCON)[6]。

富營(yíng)養(yǎng)化會(huì)促進(jìn)水-氣界面的POPs輸入[74],或減少揮發(fā)[4],改變食物網(wǎng)組成、多樣性和優(yōu)勢(shì)種以及生態(tài)作用機(jī)制和能流傳遞效率、方向,直接或間接地影響POPs的環(huán)境行為和傳遞機(jī)制[6,84- 85]。對(duì)于初級(jí)生產(chǎn)者,富營(yíng)養(yǎng)化促使浮游植物的優(yōu)勢(shì)種由藍(lán)、綠藻取代硅藻、金藻等,初級(jí)生產(chǎn)力大大提高,生物量增加,大型水生植物和底棲藻類死亡,初級(jí)生產(chǎn)力從底棲為主導(dǎo)變?yōu)楦∮握贾鲗?dǎo)。由于藍(lán)藻的食物品質(zhì)較差,使消費(fèi)者的食物類型從浮游向底棲轉(zhuǎn)變[86- 87],從而改變了食物網(wǎng)中物質(zhì)流的傳遞方向,同時(shí)改變了POPs沿食物網(wǎng)的流動(dòng)路徑。浮游生物POPs濃度降低,其原因在于：相較于硅藻和金藻等,藍(lán)藻的脂質(zhì)含量較低[88- 89]；且生物量大大增加導(dǎo)致“生物量稀釋”,生長(zhǎng)速率較快,發(fā)生“生長(zhǎng)稀釋”[90- 93]；另外,在富營(yíng)養(yǎng)條件下,藻類的脂質(zhì)含量相較于貧營(yíng)養(yǎng)條件下更低[16,67]。

對(duì)于浮游和底棲動(dòng)物,富營(yíng)養(yǎng)化導(dǎo)致透光度下降,附生藻類和大型水生植物死亡,①使得浮游和底棲動(dòng)物的棲息、避難生境減少,增強(qiáng)了魚類的捕食作用[94- 95],大型無(wú)脊椎動(dòng)物減少,且小型化；②藍(lán)藻爆發(fā),可攝食的食物資源減少,質(zhì)量變差,導(dǎo)致初級(jí)消費(fèi)者豐度下降、優(yōu)勢(shì)種改變,且食性發(fā)生變化。因此POPs在食物網(wǎng)中的傳遞路徑發(fā)生變化；③藻類爆發(fā)導(dǎo)致有機(jī)質(zhì)的沉降量增加。首先,有機(jī)質(zhì)大量分解造成沉積物-水界面缺氧[97],底棲生物大量死亡,優(yōu)勢(shì)種變?yōu)槟褪艿脱醴N[6,87,95- 96],使得底棲食物網(wǎng)發(fā)生較大的變化,POPs沿食物網(wǎng)的傳遞發(fā)生改變。其次,POPs隨顆粒物沉降的通量增加[67,97],大部分POPs從水柱中移除,降低了浮游生境的POPs暴露風(fēng)險(xiǎn)[98- 99],多項(xiàng)研究表明浮游食物鏈中POPs濃度與富營(yíng)養(yǎng)化水平呈負(fù)相關(guān)[92,100]。同時(shí)沉積物中POPs濃度升高[4,89],底棲食物網(wǎng)的暴露風(fēng)險(xiǎn)提高[75,101- 102]。

對(duì)于魚類等高營(yíng)養(yǎng)級(jí)生物,富營(yíng)養(yǎng)化會(huì)減少物種進(jìn)化過(guò)程中對(duì)環(huán)境變化的生態(tài)響應(yīng),導(dǎo)致物種滅絕,降低多樣性[103],另外浮游和底棲群落的變化,使魚類的攝食食性、生長(zhǎng)率和脂質(zhì)含量都會(huì)發(fā)生較大的,影響POPs的生物富集和傳遞效率[93，104- 105]。例如,Lasson等[98]對(duì)瑞典的61個(gè)湖泊研究發(fā)現(xiàn),魚類體內(nèi)PCBs生物富集系數(shù)(BAF)與湖泊生產(chǎn)力呈顯著負(fù)相關(guān),因?yàn)楦粻I(yíng)養(yǎng)化條件下生產(chǎn)力提高,魚類生長(zhǎng)加快,發(fā)生“生長(zhǎng)稀釋”。另外,魚類等高營(yíng)養(yǎng)級(jí)的生物,其攝食吸收的POPs所占比例較大,低營(yíng)養(yǎng)級(jí)POPs濃度的變化會(huì)沿食物鏈影響魚類體內(nèi)的POPs濃度[93,106]。但是上述幾個(gè)不同的影響因素,在不同的生態(tài)系統(tǒng)中或是不同的條件下,起主導(dǎo)作用的因素不同。

對(duì)于微生物,富營(yíng)養(yǎng)化會(huì)直接或通過(guò)改變棲息生境的理化性質(zhì)間接地影響微生物對(duì)POPs的分解速率,Graham等[107]研究認(rèn)為富營(yíng)養(yǎng)條件下微生物降解POPs的速率要高于寡營(yíng)養(yǎng)水平,但是目前這方面的研究相對(duì)較少。另如前文所述,富營(yíng)養(yǎng)化會(huì)擴(kuò)大和加強(qiáng)底棲生境的缺氧程度,促使微生物將DDT轉(zhuǎn)化為DDD[82- 83]。

圖1 富營(yíng)養(yǎng)化條件下食物網(wǎng)變化對(duì)POPs環(huán)境行為的影響Fig.1 Influences of foodweb changes caused by eutrophication on environmental behaviors of POPs

4.3 物種入侵或引進(jìn)對(duì)POPs環(huán)境行為的影響

外來(lái)物種的入侵或引進(jìn)是目前水生生態(tài)系統(tǒng)遇到的危機(jī)之一,對(duì)食物網(wǎng)的組成、結(jié)構(gòu)和動(dòng)態(tài)變化產(chǎn)生了重要的影響,例如,生物體的食性、生長(zhǎng)率、營(yíng)養(yǎng)位等,進(jìn)而影響POPs的暴露風(fēng)險(xiǎn),及在食物網(wǎng)中的環(huán)境行為和傳遞機(jī)制。研究發(fā)現(xiàn),中營(yíng)養(yǎng)級(jí)物種入侵會(huì)導(dǎo)致浮游動(dòng)物和攝食浮游動(dòng)物魚類的營(yíng)養(yǎng)位(trophic position)增加,提高其捕食者體內(nèi)的污染物濃度,例如,尾突蚤或胡瓜魚等中營(yíng)養(yǎng)級(jí)物種入侵后,改變了浮游動(dòng)物的群落結(jié)構(gòu),優(yōu)勢(shì)種由枝角類變?yōu)闃镒泐?浮游動(dòng)物生物量降低[108- 109],提高了浮游動(dòng)物和魚類的營(yíng)養(yǎng)位,進(jìn)而導(dǎo)致其體內(nèi)的污染物濃度升高[9,110]。另外,在食物網(wǎng)中引入糠蝦等新物種,會(huì)延長(zhǎng)食物鏈長(zhǎng)度,且由于糠蝦的晝夜遷移行為,會(huì)增加浮游-底棲食物網(wǎng)的耦合程度,導(dǎo)致POPs在頂級(jí)消費(fèi)者體內(nèi)的濃度升高[2]。

但其它研究發(fā)現(xiàn),外來(lái)種入侵并不總是引起較高營(yíng)養(yǎng)級(jí)生物污染物暴露風(fēng)險(xiǎn)的增加。在加拿大實(shí)驗(yàn)湖L227和L110中引入梭子魚或肉食性白斑狗魚后,生態(tài)系統(tǒng)的群落結(jié)構(gòu)、物種豐度和優(yōu)勢(shì)種都發(fā)生了改變,很多魚類的攝食習(xí)慣發(fā)生變化,從主要攝食浮游動(dòng)物轉(zhuǎn)變?yōu)閿z食底棲動(dòng)物,提高了營(yíng)養(yǎng)位,但魚類體內(nèi)的PCBs、DDT、HCH和Hg等污染物的濃度或是沒(méi)有變化,或是下降了,推測(cè)這可能是由于營(yíng)養(yǎng)級(jí)之間復(fù)雜的級(jí)聯(lián)作用導(dǎo)致的[111]。例如,中營(yíng)養(yǎng)級(jí)物種入侵,導(dǎo)致浮游動(dòng)物減少,浮游植物的生物量增加,產(chǎn)生“生物量稀釋”[109,112- 113],或因?yàn)槿肭值奈锓N比傳統(tǒng)獵物的營(yíng)養(yǎng)位更高[110,114]、密度更大[115],消費(fèi)者的捕食偏好改變,覓食效率提高,生長(zhǎng)率更高[9],從而導(dǎo)致“生長(zhǎng)稀釋”,使污染物濃度下降。

物種入侵或是人工引入會(huì)通過(guò)營(yíng)養(yǎng)級(jí)聯(lián)作用對(duì)多個(gè)營(yíng)養(yǎng)級(jí)產(chǎn)生不同程度的影響,從而對(duì)食物網(wǎng)的結(jié)構(gòu)、功能和生態(tài)過(guò)程產(chǎn)生影響,進(jìn)而導(dǎo)致POPs沿食物網(wǎng)的傳遞發(fā)生很大的變化,但目前對(duì)這方面的研究還相對(duì)較少,作用機(jī)制還不甚清晰。

5 結(jié)論與展望

營(yíng)養(yǎng)級(jí)和食物鏈長(zhǎng)度是食物網(wǎng)的2個(gè)主要特征,是研究POPs生物富集或生物放大時(shí)的重點(diǎn)考慮因素。一般情況下,POPs生物富集與營(yíng)養(yǎng)級(jí)呈正相關(guān),但受到生物體各種生理、生態(tài)因素和POPs理化性質(zhì)的影響,有時(shí)會(huì)出現(xiàn)例外。各種影響因素紛繁復(fù)雜,如何互相聯(lián)系作用于營(yíng)養(yǎng)級(jí)與POPs生物富集性之間的相關(guān)關(guān)系尚不十分明確,需進(jìn)一步探索。食物鏈長(zhǎng)度與生物富集呈正相關(guān),除脂質(zhì)的影響外,其它具體作用機(jī)制還需深入研究。底棲食物網(wǎng)是POPs在水生生態(tài)系統(tǒng)中傳遞的一個(gè)重要途徑,但目前對(duì)POPs在底棲食物網(wǎng)中環(huán)境行為的認(rèn)識(shí)莫衷一是,尚不明確。另外,底棲食物網(wǎng)通過(guò)各種過(guò)程與浮游食物網(wǎng)耦合,對(duì)POPs的環(huán)境行為產(chǎn)生復(fù)雜的影響,但相關(guān)研究甚少,需要加強(qiáng)關(guān)注。微食物網(wǎng)對(duì)POPs環(huán)境行為的影響逐漸被關(guān)注,目前相關(guān)研究主要集中在細(xì)菌對(duì)POPs的分解和吸附作用,由于微生物體型小,分析困難,微食物網(wǎng)中POPs環(huán)境行為的動(dòng)態(tài)研究和定量研究還相對(duì)缺乏。在內(nèi)源驅(qū)動(dòng)和外源壓力的作用下,比如季節(jié)演替、富營(yíng)養(yǎng)化、物種入侵等,使環(huán)境介質(zhì)、食物網(wǎng)和POPs的結(jié)構(gòu)、性質(zhì)發(fā)生復(fù)雜的變化,三者之間互相反饋導(dǎo)致POPs沿食物網(wǎng)的環(huán)境行為發(fā)生更為錯(cuò)綜復(fù)雜的變化,但是目前這方面的研究還不夠全面和系統(tǒng),是研究的重點(diǎn)和難點(diǎn)。

目前,POPs在食物網(wǎng)中環(huán)境行為的研究主要集中在“classic” POPs,包括PCBs、OCPs和PAHs等,但在傳統(tǒng)POPs被禁用多年后,很多替代性的新型POPs被越來(lái)越多地使用,并在世界范圍內(nèi)的不同環(huán)境中都有檢出。例如,在中國(guó),海洋哺乳動(dòng)物和沉積物中傳統(tǒng)PBDE開始逐漸被新型阻燃劑BFRs替代[116- 117],渤海地區(qū)更是遭到了新型POPs的嚴(yán)重污染[118]。在南、北極地區(qū)也分別都檢測(cè)到了新型POPs,并認(rèn)為存在生物放大的可能。目前對(duì)于新型POPs可以獲取的數(shù)據(jù)還很少[119- 121],對(duì)于其生物富集性的研究也多集中在少數(shù)地區(qū)的極少數(shù)大型哺乳動(dòng)物和魚類中。未來(lái),需要加大對(duì)新型POPs在水生食物網(wǎng)中的研究。

[1] Kelly B C, Ikonomou M G, Blair J D, Morin A E, Gobas F A P C. Food web-specific biomagnification of persistent organic pollutants. Science, 2007, 317(5835): 236- 239.

[2] Rasmussen J B, Rowan D J, Lean D R S, Carey J H. Food chain structure in Ontario Lakes determines PCB levels in Lake Trout (salvelinusnamaycush) and other pelagic fish. Canadian Journal of Fisheries and Aquatic Sciences, 1990, 47(10): 2030- 2038.

[3] Wan Y, Zhang K, Dong Z M, Hu J Y. Distribution is a major factor affecting bioaccumulation of decabrominated diphenyl ether: Chinese sturgeon (Acipensersinensis) as an example. Environmental Science and Technology, 2013, 47(5): 2279- 2286.

[4] Jeremiason J D, Eisenreich S J, Paterson M J, Beatly K G, Hecky R, Elser J J. Biogeochemical cycling of PCBs in lakes of variable trophic status: a paired-lake experiment. Limnology and Oceanography, 1999, 44(3): 889- 902.

[5] 李雪梅, 張慶華, 甘一萍, 周軍, 戴家銀, 曹宏, 許木啟. 持久性有機(jī)污染物在食物鏈中積累與放大研究進(jìn)展. 應(yīng)用與環(huán)境生物學(xué)報(bào), 2008, 13(6): 901- 905.

[6] Skei J, Larsson P, Rosenberg R, Jonsson P, Olsson M, Broman D. Eutrophication and contaminants in aquatic ecosystem. AMBIO, 2000, 29(4- 5): 184- 194.

[7] Gunnarsson J, Bj?rk M, Gilek M, Granberg M, Rosenberg R. Effects of eutrophication on contaminant cycling in marine benthic systems. AMBIO, 2000, 29(4- 5): 252- 259.

[8] Nizzetto L, Gioia R, Li J, Borga K, Pomati F, Bettinetti R, Dachs J, Jones K C. Biological pump control of the fate and distribution of hydrophobic organic pollutants in water and plankton. Environmental Science and Technology, 2012, 46(6): 3204- 3211.

[9] Rennie M D, Strecker A L, Palmer M E.Bythotrephesinvasion elevates trophic position of zooplankton and fish: implications for contaminant biomagnification. Biological Invasions, 2011, 13(11): 2621- 2634.

[10] Houde M, Pacepavicius G, Wells R S, Fair P A, Letcher R J, Alaee M, Bossart G D, Hohn A A, Sweeney J, Solomon K R, Derek C G. Polychlorinated bipphenyls and hydroxylated polychlorinated biphenyls in plasma of bottlenose dolphins (Tursiopstruncatus) from the Western Atlantic and the Gulf of Mexico. Environmental Science and Technology, 2006, 40(19): 5860- 5866.

[11] Kannan K, Tao L, Sinclair E, Pastva S D, Jude D J, Giesy J P. Perfluorinated compounds in aquatic organisms at various trophic levels in a Great Lakes food chain. Archives of Environmental contamination and Toxicology, 2005, 48(4): 559- 566.

[12] Martin J W, Mabury S A, Solomon K R, Muir D C G. Dietary accumulation of perfluorinated acids in juvenile rainbow trout (Oncorhynchusmykiss). Environmental Toxicology and Chemistry, 2003, 22(1): 189- 195.

[13] Nfon E, Cousins I T. Interpreting time trends and biomagnification of PCBs in the Baltic region using the equilibrium lipid partitioning approach. Environmental Pollution, 2006, 144(3): 994- 1000.

[14] Hobson K A, Fisk A, Karnovsky N, Holst M, Gagnon J M, Fortier M. A stable isotope (δ13C, δ15N) model for the North Water food web: implications for evaluating trophodynamics and the flow of energy and contaminants. Deep Sea Research Part Ⅱ: Topical Studies in Oceanography, 2002, 49(22- 23): 5131- 5150.

[15] Kucklick J R, Baker J E. Organochlorines in Lake Superior′s food web. Environmental Science and Technology, 1998, 32(9): 1192- 1198.

[16] Kidd K A, Bootsma H A, Hesslein R H, Muir D C G, Hecky R E. Biomagnification of DDT through the benthic and pelagic food webs of Lake Malawi, East Africa: importance of trophic level and carbon source. Environmental Science and Technology, 2001, 35(1): 14- 20.

[17] Neff J M. Bioaccumulation of organic micropollutants from sediments and suspended particulates by aquatic animals. Fresenius′ Zeitschrift für Analytische Chemie, 1984, 319(2): 132- 136.

[18] Varanasi U, Reichert W L, Stein J E, Brown D W, Sanborn H R. Bioavailability and biotransformation of aromatic hydrocarbons in benthic organisms exposed to sediment from an urban estuary. Environmental Science and Technology, 1985, 19(9): 836- 841.

[19] Broman D, N?uf C, Lundbergh I, Zebühr Y. An in situ study on the distribution, biotransformation and flux of polycyclic aromatic hydrocarbons (PAHs) in an aquatic food chain (Seston-MytilusedulisL.-SomateriamollissimaL.) from the baltic: an ecotoxicological perspective. Environmental Toxicology and Chemistry, 1990, 9(4): 429- 442.

[20] Niimi A J, Oliver B G. Influence of molecular weight and molecular volume on dietary absorption efficiency of chemicals by fishes. Canadian Journal of Fisheries and Aquatic Sciences, 1988, 45(2): 222- 227.

[21] Selck H, Palmqvist A, Forbes V E. Biotransformation of dissolved and sediment-bound fluoranthene in the polychaete,CapitellaSp. I. Environmental Toxicology and Chemistry, 2003, 22(10): 2364- 2374.

[22] Gewurtz S B, Lazar R, Haffner G D. Comparison of polycyclic aromatic hydrocarbon and polychlorinated biphenyl dynamics in benthic invertebrates of Lake Erie, USA. Environmental Toxicology and Chemistry, 2000, 19(12): 2943- 2950.

[23] Mizukawa K, Takada H, Takeuchi I, Ikemoto T, Omori K, Tsuchiya K. Bioconcentration and biomagnification of polybrominated diphenyl ethers (PBDEs) through lower-trophic-level coastal marine food web. Marine Pollution Bulletin, 2009, 58(8): 1217- 1224.

[24] Nfon E, Cousins I T, Broman D. Biomagnification of organic pollutants in benthic and pelagic marine food chains from the Baltic Sea. Science of the Total Environment, 2008, 397(1/3): 190- 204.

[25] Hanari N, Kannan K, Horii Y, Taniyasu S, Yamashita N, Jude D J, Berg M B. Polychlorinated naphthalenes and polychlorinated biphenyls in benthic organisms of a Great Lakes food chain. Archives of Environmental Contamination and Toxicology, 2004, 47(1): 84- 93.

[26] Corsilini S, Kannan K, Imagawa T, Focardi S, Giesy J P. Polychloronaphthalenes and other dioxin-like compounds in Arctic and Antarctic marine food webs. Environmental Science and Technology, 2002, 36(16): 3490- 3496.

[27] Hallanger I G, Warner N A, Ruus A, Evenset A, Christensen G, Herzke D, Gabrielsen G W, Borg? K. Seasonality in contaminant accumulation in Arctic marine pelagic food webs using trophic magnification factor as a measure of bioaccumulation. Environmental Toxicology and Chemistry, 2011, 30(5): 1026- 1035

[28] Borg? K, Kidd K A, Muir D C G, Berglund O, Conder J M, Gobas F A P C, Kucklick J, Malm O, Powell D. Trophic magnification factors: considerations of ecology, ecosystems and study design. Integrated Environmental Assessment and Management, 2012, 8(1): 64- 84.

[29] Binelli A, Provini A. The PCB pollution of Lake Iseo (N. Italy) and the role of biomagnification in the pelagic food web. Chemosphere, 2003, 53(2): 143- 151.

[30] Lundgren K, Tysklind M, Ishaq R, Broman D, Van Bavel B. Polychlorinated naphthalene levels, distribution, and biomagnification in a benthic food chain in the Baltic Sea. Environmental Science and Technology, 2002, 36(23): 5005- 5013.

[31] Evenset A, Hallanger I G, Tessmann M, Warner N, Ruus A, Borg? K, Gabrielsen G W, Christensen G, Renaud P E. Seasonal variation in accumulation of persistent organic pollutants in an Arctic marine benthic food web. Science of the Total Environment, 2016, 542: 108- 120.

[32] Paterson G, Ryder M, Drouillard K G, Haffner G D. Contrasting PCB bioaccumulation patterns among Lake Huron lake trout reflect basin-specific ecology. Environmental Toxicology and Chemistry, 2016, 35(1): 65- 73.

[33] McLeod A M, Paterson G, Drouillard K G, Haffner G D. Ecological factors contributing to variability of persistent organic pollutant bioaccumulation within forage fish communities of the Detroit River, Ontario, Canada. Environmental Toxicology and Chemistry, 2014, 33(8): 1825- 1831.

[34] Jardine T D, Kidd K A, Fisk A T. Applications, considerations, and sources of uncertainty when using stable isotope analysis in ecotoxicology. Environmental Science and Technology, 2006, 40(24): 7501- 7511.

[35] Bruner K A, Fisher S W, Landrum P F. The role of the zebra mussel,Dreissenapolymorpha, in contaminant cycling: Ⅱ. Zebra Mussel contaminant accumulation from algae and suspended particles, and transfer to the benthic invertebrate,Gammarusfasciatus. Journal of Great Lakes Research, 1994, 20(4): 735- 750.

[36] Looser R, Ballschmiter K. Biomagnification of polychlorinated biphenyls (PCBs) in freshwater fish. Fresenius′ Journal of Analytical Chemistry, 1998, 360(7/8): 816- 819.

[37] Gobas F A P C, Wilcockson J B, Russell R W, Haffner G D. Mechanism of biomagnification in fish under laboratory and field conditions. Environmental Science and Technology, 1999, 33(1): 133- 141.

[38] Borg? K, Gabrielsen G W, Skaare J U. Biomagnification of organochlorines along a Barents sea food chain. Environmental Pollution, 2001, 113(2): 187- 198.

[39] Guruge K S, Tanabe S, Iwata H, Taksukawa R, Yamagishi S. Distribution, biomagnification, and elimination of butyltin compound residues in common cormorants (Phalacrocoraxcarbo) from Lake Biwa, Japan. Archives of Environmental Contamination and Toxicology, 1996, 31(2): 210- 217.

[40] Bentzen E, Lean D R S, Taylor W D, Mackay D. Role of food web structure on lipid bioaccumulation of organic contaminants by lake trout (Salvelinusnamaycush). Canadian Journal of Fisheries and Aquatic Sciences, 1996, 53(11): 2397- 2407.

[41] Oliver B G, Niimi A J. Trophodynamic analysis of polychlorinated biphenyl congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environmental Science and Technology, 1988, 22(4): 388- 397.

[42] Evans M S, Noguchi G E, Rice C P. The biomagnification of polychlorinated biphenyls, toxaphene, and DDT compounds in a Lake Michigan offshore food web. Archives of Environmental Contamination and Toxicology, 1991, 20(1): 87- 93.

[43] Kidd K A, Schindler D W, Muir D C G, Lockhart W L, Hesslein R H. High concentrations of toxaphene in fishes from a subarctic lake. Science, 1995, 269(5221): 240- 242.

[44] Whittle D M, Kiriluk R M, Carswell A A, Keir M J, MacEachen D C. Toxaphene congeners in the Canadian Great Lakes basin: temporal and spatial food web dynamics. Chemosphere, 2000, 40(9- 11): 1221- 1226.

[45] Kerr S R. A simulation model of lake trout growth. Journal of the Fisheries Research Board of Canada, 1971, 28(6): 815- 819.

[46] Hecky R E, Hesslein R H. Contributions of benthic algae to lake food webs as revealed by stable isotope analysis. Journal of the North American Benthological Society, 1995, 14(4): 631- 653.

[47] Vander Zanden M J, Vadeboncoeur Y. Fishes as integrators of benthic and pelagic food webs in lakes. Ecology, 2002, 83(8): 2152- 2161.

[48] Strayer D, Likens G E. An energy budget for the zoobenthos of Mirror Lake, New Hampshire. Ecology, 1986, 67(2): 303- 313.

[49] Covich A P, Palmer M A, Crowl T A. The Role of benthic invertebrate species in freshwater ecosystems: zoobenthic species influence energy flows and nutrient cycling. BioScience, 1999, 49(2): 119- 127.

[50] Theil-Nielsen J, S?ndergaard M. Production of epiphytic bacteria and bacterioplankton in three shallow lakes. Oikos, 1999, 86(2): 283- 292.

[51] Campfens J, Mackay D. Fugacity-based model of PCB bioaccumulation in complex aquatic food webs. Environmental Science and Technology, 1997, 31(2): 577- 583.

[52] Vadeboncoeur Y, Vander Zanden M J, Lodge D M. Putting the lake back together: reintegrating benthic pathways into lake food web models. BioScience, 2002, 52(1): 44- 54.

[53] Johnson R K, Wiederholm T. Pelagic-benthic coupling-the importance of diatom interannual variability for population oscillations ofMonoporeiaaffinis. Limnology and Oceanography, 1992, 37(8): 1596- 1607.

[54] Jonasson P M. Ecology and production of the profundal benthos in relation to phytoplankton in Lake Esrom. Oikos, 1972, 14(Supplement): 1- 148.

[55] Johnson R K, Bostr?m B, Van De Bund W. Interactions betweenChironomusplumosus(L.) and the microbial community in surficial sediments of a shallow, eutrophic lake. Limnology and Oceanography, 1989, 34(6): 992- 1003.

[56] Rasmussen J B. Effects of density and microdetritus enrichment on the growth of chironomid larvae in a small pond. Canadian Journal of Fisheries and Aquatic Sciences, 1985, 42(8): 1418- 1422.

[57] Kemp M W, Boynton W R. Spatial and temporal coupling of nutrient inputs to estuarine primary production: the role of particulate transport and decomposition. Bulletin of Marine Science, 1984, 35(3): 522- 535.

[58] Strayer D L. Perspectives on the size structure of lacustrine zoobenthos, its causes, and its consequences. Journal of the North American Benthological Society, 1991, 10(2): 210- 221.

[59] Doering P H, Kelly J R, Oviatt C A, Sowers T. Effect of the hard clamMercenariamerceniaon benthic fluxes of inorganic nutrients and gases. Marine Biology, 1987, 94(3): 377- 383.

[60] Jeppesen E, Jensen J P, S?ndergaard M, Lauridsen T, Pedersen L J, Jensen L. Top-down control in freshwater lakes: the role of nutrient state, submerged macrophytes and water depth. Hydrobiologia, 1997, 342- 343: 151- 164.

[61] Stein R A, DeVries D R, Dettmers J M. Food-web regulation by a planktivore: exploring the generality of the trophic cascade hypothesis. Canadian Journal of Fisheries and Aquatic Sciences, 1995, 52(11): 2518- 2526.

[62] Klump J V, Kaster J L, Sierszen M E.Mysisrelictaassimilation of Hexachlorobiphenyl from sediments. Canadian Journal of Fisheries and Aquatic Sciences, 1991, 48(2): 284- 289.

[63] Danger M, Oumarou C, Benest D, Lacroix G. Bacteria can control stoichiometry and nutrient limitation of phytoplankton. Functional Ecology, 2007, 21(2): 202- 210.

[64] Joint I, Henriksen P, Fonnes G A, Bourne D, Thingstad T F, Riemann B. Competition for inorganic nutrients between phytoplankton and bacterioplankton in nutrient manipulated mesocosms. Aquatic Microbial Ecology, 2002, 29(2): 145- 159.

[65] Goedkoop W, Johnson R K. Pelagic-benthic coupling: profundal benthic community response to spring diatom deposition in mesotrophic Lake Erken. Limnology and Oceanography, 1996, 41(4): 636- 647.

[66] Goedkoop W, Johnson R K. Exploitation of sediment bacterial carbon by juveniles of the amphipodMonoporeiaaffinis. Freshwater Biology, 1994, 32(3): 553- 563.

[67] Larsson P, Andersson A, Broman D, Nordb?ck J, Lundberg E. Persistent organic pollutants (POPs) in pelagic systems. AMBIO, 2000, 29(4- 5): 202- 209.

[68] Axelman J, Broman D, N?f C. Vertical flux and particulate/water dynamics of polychlorinated biphenyls (PCBs) in the open Baltic Sea. AMBIO, 2000, 29(4- 5): 210- 216.

[69] Koelmans A A, Lijklema L, Jiménez C S. Sorption of chlorobenzenes to mineralizing phytoplankton. Environmental Toxicology and Chemistry, 1993, 12(8): 1425- 1439.

[70] Koelmans A A, Anzion S F M, Lijklema L. Dynamics of organic micropollutant biosorption to cyanobacteria and detritus. Environmental Science and Technology, 1995, 29(4): 933- 940.

[71] Fisk A T, Stern G A, Hobson K A, Strachan W J, Loewen M D, Norstrom R J. Persistent organic pollutants (POPs) in a small, herbivorous, Arctic marine zooplankton (Calanushyperboreus): trends from April to July and the influence of lipids and trophic transfer. Marine Pollution Bulletin, 2001, 43(1/6): 93- 101.

[72] Hargrave B T, Phillips G A, Vass W P, Bruecker P, Welch H E, Siferd T D. Seasonality in bioaccumulation of organochlorines in lower trophic level Arctic marine biota. Environmental Science and Technology, 2000, 34(6): 980- 987.

[73] Hallanger I G, Ruus A, Herzke D, Warner N A, Evenset A, Heimstad E S, Gabrielsen G W, Borg? K. Influence of season, location, and feeding strategy on bioaccumulation of halogenated organic contaminants in Arctic marine zooplankton. Environmental Toxicology and Chemistry, 2011, 30(1): 77- 87.

[74] Dachs J, Eisenreich S J, Baker J E, Ko F C, Jeremiason J D. Coupling of phytoplankton uptake and air-water exchange of persistent organic pollutants. Environmental Science and Technology, 1999, 33(20): 3653- 3660.

[75] S?derstr?m M, Nylund K, J?rnberg U, Lithner G, Rosén G, Kylin H. Seasonal variations of DDT compounds and PCB in a eutrophic and an oligotrophic lake in relation to algal biomass. AMBIO, 2000, 29(4- 5): 230- 237.

[76] Chi J, Wang Q Y, Huang J J, Huang G L. Sedimentation and seasonal variation of hexachlorocyclohexanes in sediments in a eutrophic lake, China. Journal of Environmental Science and Health, Part B: Pesticides, Food Contaminants, and Agricultural Wastes, 2008, 43(7): 611- 616.

[77] Renaud P E, Tessmann M, Evenset A, Christensen G N. Benthic food-web structure of an Arctic fjord (Kongsfjorden, Svalbard). Marine Biology Research, 2011, 7(1): 13- 26.

[78] Gardner W S, Nalepa T F, Frez W A, Cichocki E A, Landrum P F. Seasonal patterns in lipid content of lake michigan macroinvertebrates. Canadian Journal of Fisheries and Aquatic Sciences, 1985, 42(11): 1827- 1832.

[79] Hummel H, Bogaards R H, Nieuwenhuize J, De Wolf L, Van Liere J M. Spatial and seasonal differences in the PCB content of the musselMytilusedulis. Science of the Total Environment, 1990, 92: 155- 163.

[80] Capuzzo J M, Farrington J W, Rantamaki P, Clifford C H, Lancaster B A, Leavitt D F, Jia Xiaoping. The relationship between lipid composition and seasonal differences in the distribution of PCBs inMytilusEdulisL. Marine Environmental Research, 1989, 28(1/4): 259- 264.

[81] Person N J, Axelman J, Broman D. Validating possible effects of eutrophication using PCB concentrations in bivalves and sediment of the US musselwatch and benthic surveillance programs. AMBIO, 2000, 29(4- 5): 246- 251.

[82] Kale S P, Murthy N B K, Raghu K, Sherkhane P D, Carvalho F P. Studies on degradation of14C-DDT in the marine environment. Chemosphere, 1999, 39(6): 959- 968.

[83] Sethajintanin D, Anderson K A. Temporal bioavailability of organochlorine pesticides and PCBs. Environmental Science and Technology, 2006, 40(12): 3689- 3695.

[84] Horner-Devine M C, Leibold M A, Smith V H, Bohannan B J M. Bacterial diversity patterns along a gradient of primary productivity. Ecology Letters, 2003, 6(7): 613- 622.

[85] Smith V H. Microbial diversity-productivity relationships in aquatic ecosystems. FEMS Microbiology Ecology, 2007, 62(2): 181- 186.

[86] Carlier A, Riera P, Amouroux J M, Bodiou J Y, Desmalades M, Grémare A. Food web structure of two Mediterranean lagoons under varying degree of eutrophication. Journal of Sea Research, 2008, 60(4): 264- 275.

[87] Vadeboncoeur Y, Jeppesen E, Vander Zanden M J, Schierup H H, Christoffersen K, Lodge D M. From Greenland to green lakes: cultural eutrophication and the loss of benthic pathways in lakes. Limnology and Oceanography, 2003, 48(4): 1408- 1418.

[88] Borg? K, Fisk A T, Hoekstra P F, Muir D C G. Biological and chemical factors of importance in the bioaccumulation and trophic transfer of persistent organochlorine contaminants in arctic marine food webs. Environmental Toxicology and Chemistry, 2004, 23(10): 2367- 2385.

[89] Larsson P, Okla L, Cronberg G. Turnover of polychlorinated biphenyls in an oligotrophic and a eutrophic lake in relation to internal lake processes and atmospheric fallout. Canadian Journal of Fisheries and Aquatic Sciences, 1998, 55(8): 1926- 1937.

[90] Niimi A J, Cho C Y. Elimination of hexachlorobenzene (HCB) by rainbow trout (Salmogairdneri), and an examination of its kinetics in lake Ontario Salmonids. Canadian Journal of Fisheries and Aquatic Sciences, 1981, 38(11): 1350- 1356.

[91] Axelman J, Broman D, N?f C. Field measurements of PCB partitioning between water and planktonic organisms: influence of growth, particle size, and solute-solvent interactions. Environmental Science and Technology, 1997, 31(3): 665- 669.

[92] Taylor W D, Carey J H, Lean D R S, McQueen D J. Organochlorine concentrations in the plankton of lakes in Southern Ontario and their relationship to plankton biomass. Canadian Journal of Fisheries and Aquatic Sciences, 1991, 48(10): 1960- 1966.

[93] Ryan M J, Stern G A, Kidd K A, Croft M V, Gewurtz S, Diamond M, Kinnear L, Roach P. Biotic interactions in temporal trends (1992- 2010) of organochlorine contaminants in the aquatic food web of lake Laberge, Yukon Territory. Science of the Total Environment, 2013, 443: 80- 92.

[94] Jeppesen E, Jensen J P, S?ndergaard M, Lauridsen T. Trophic dynamics in turbid and clearwater lakes with special emphasis on the role of zooplankton for water clarity. Hydrobiologia, 1999, 408- 409: 217- 231.

[95] Rasmussen J B, Kalff J. Empirical models for zoobenthic biomass in lakes. Canadian Journal of Fisheries and Aquatic Sciences, 1987, 44(5): 990- 1001.

[96] Kemp W M, Sampou P A, Garber J, Tuttle J, Boynton W R. Seasonal depletion of oxygen from bottom waters of Chesapeake Bay: roles of benthic and planktonic respiration and physical exchange processes. Marine Ecology Progress Series, 1992, 85: 137- 152.

[97] Auer B, Elzer U, Arndt H. Comparison of pelagic food webs in lakes along a trophic gradient and with seasonal aspects: influence of resource and predation. Journal of Plankton Research, 2004, 26(6): 697- 709.

[98] Larsson P, Collvin L, Okla L, Meyer G. Lake productivity and water chemistry as governors of the uptake of persistent pollutants in fish. Environmental Science and Technology, 1992, 26(2): 346- 352.

[99] Millard E S, Minns C K, Charlton C C, Halfon E. Effect of primary productivity and vertical mixing on pcb dynamics in planktonic model ecosystems. Environmental Toxicology and Chemistry, 1993, 12(5): 931- 946.

[100] Berglund O, Larsson P, Ewald G, Okla L. The effect of lake trophy on lipid content and PCB concentrations in planktonic food webs. Ecology, 2001, 82(4): 1078- 1088.

[101] Gunnarsson J S, Schaanning M T, Hylland K, Sk?ld M, Eriksen D ?, Berge J A, Skei J. Interactions between eutrophication and contaminants. III. Mobilization and bioaccumulation of benzo(a)pyrene from marine sediments. Marine Pollution Bulletin, 1996, 33(1/6): 80- 89.

[102] Franz J M S, Hauss H, Sommer U, Dittmar T, Riebesell U. Production, partitioning and stoichiometry of organic matter under variable nutrient supply during mesocosm experiments in the tropical Pacific and Atlantic Ocean. Biogeosciences, 2012, 9(11): 4629- 4643.

[103] Vonlanthen P, Bittner D, Hudson A G, Young K A, Müller R, Lundsgaard-Hansen B, Roy D, Di Piazza S, Largiader C R, Seehausen O. Eutrophication causes speciation reversal in whitefish adaptive radiations. Nature, 2012, 482(7385): 357- 362.

[104] Thomann R V. Bioaccumulation model of organic chemical distribution in aquatic food chains. Environmental Science and Technology, 1989, 23(6): 699- 707.

[105] Hammar J, Larsson P, Klavins M. Accumulation of persistent pollutants in normal and dwarfed Arctic Char (Salvelinusalpinussp. complex). Canadian Journal of Fisheries and Aquatic Sciences, 1993, 50(12): 2574- 2580.

[106] Smith D W. Synchronous response of hydrophobic chemicals in herring gull eggs from the Great Lakes. Environmental Science and Technology, 1995, 29(3): 740- 750.

[107] Graham D W, Miles D, De Noyelles F, Smith V H. Development of small outdoor microcosms for studying contaminant transformation rates and mechanisms under various water column conditions. Environmental Toxicology and Chemistry, 1999, 18(6): 1124- 1132.

[108] Fernandez R J, Rennie M D, Sprules W G. Changes in nearshore zooplankton associated with species invasions and potential effects on larval lake whitefish (Coregonusclupeaformis). International Review of Hydrobiology, 2009, 94(2): 226- 243.

[109] Strecker A L, Arnott S E. Invasive predator,Bythotrephes, has varied effects on ecosystem function in freshwater lakes. Ecosystems, 2008, 11(3): 490- 503.

[110] Vander Zanden M J, Rasmussen J B. A trophic position model of pelagic food webs: impact on contaminant bioaccumulation in lake trout. Ecological Monographs, 1996, 66(4): 451- 477.

[111] Kidd K A, Paterson M J, Hesslein R H, Muir D C G, Hecky R E. Effects of northern pike (Esoxlucius) additions on pollutant accumulation and food web structure, as determined by δ13C and δ15N, in a eutrophic and an oligotrophic lake. Canadian Journal of Fisheries and Aquatic Sciences, 1999, 56(11): 2193- 2202.

[112] Swanson H K, Johnston T A, Schindler D W, Bodaly R A, Whittle D W. Mercury bioaccumulation in forage fish communities invaded by rainbow smelt (Osmerusmordax). Environmental Science and Technology, 2006, 40(5): 1439- 1446.

[113] Hogan L S, Marschall E, Folt C, Stein R A. How non-native species in Lake Erie influence trophic transfer of mercury and lead to top predators. Journal of Great Lakes Research, 2007, 33(1): 46- 61.

[114] Campbell L M, Thacker R, Barton D, Muir D C G, Greenwood D, Hecky R E. Re engineering the eastern Lake Erie littoral food web: the trophic function of non-indigenous Ponto-Caspian species. Journal of Great Lakes Research, 2009, 35(2): 224- 231.

[115] Coulas R A, Macisaac H J, Dunlop W. Selective predation on an introduced zooplankter (Bythotrephescederstroemi) by lake herring (Coregonusartedii) in Harp Lake, Ontario. Freshwater Biology, 1998, 40(2): 343- 355.

[116] Chen S J, Feng A H, He M J, Chen M Y, Luo X J, Mai B X. Current levels and composition profiles of PBDEs and alternative flame retardants in surface sediments from the Pearl River Delta, southern China: comparison with historical data. Science of the Total Environment, 2013, 444: 205- 211.

[117] Zhu B Q, Lai N L S, Wai T C, Chan L L, Lam J C W, Lam P K S. Changes of accumulation profiles from PBDEs to brominated and chlorinated alternatives in marine mammals from the South China Sea. Environment International, 2014, 66: 65- 70.

[118]Li X M, Gao Y, Wang Y W, Pan Y Y. Emerging persistent organic pollutants in Chinese Bohai Sea and its coastal regions. The Scientific World Journal, 2014, 2014: 1- 10.

[119] Kim J T, Son M H, Kang J H, Kim J H, Jung J W, Chang Y S. Occurrence of legacy and new persistent organic pollutants in avian tissues from King George Island, Antarctica. Environmental Science and Technology, 2015, 49(22): 13628- 13638.

[120] Vorkamp K, Bossi R, Rigét F F, Skov H, Sonne C, Dietz R. Novel brominated flame retardants and dechlorane plus in Greenland air and biota. Environmental Pollution, 2015, 196: 284- 291.

[121] Wolschke H, Meng X Z, Xie Z Y Ebinghaus R, Cai M H. Novel flame retardants (N-FRs), polybrominated diphenyl ethers (PBDEs) and dioxin-like polychlorinated biphenyls (DL-PCBs) in fish, penguin, and skua from King George Island, Antarctica. Marine Pollution Bulletin, 2015, 96(1/2): 513- 518.

Environmental behavior of persistent organic pollutants in aquatic food web

FENG Qiuyuan1, WAN Yi2, LIU Xueqin3, LIU Yong1,*

1CollegeofEnvironmentalScienceandEngineering,TheKeyLaboratoryofWaterandSedimentSciencesMinistryofEducation,PekingUniversity,Beijing100871,China2CollegeofUrbanandEnvironmentalScience,MOELaboratoryforEarthSurfaceProcess,PekingUniversity,Beijing100871,China3InstituteofHydrobiologyChineseAcademyofSciences,Wuhan430072,China

Food web is an import transfer path in aquatic ecosystem. It is essential to explore the environmental behavior of persistent organic pollutants (POPs) along the food web, which will provides cientific foundations for risk evaluation of ecological exposure. Here, we analyzed and summarized the environmental behavior of POPs in aquatic food web, including (a) relationships between trophic levels and food chain length and biomagnification; (b) POPs transfer along benthic and benthic-pelagic coupling food web; (c) adsorption and degradation of POPs by microbes; (d) influences of food web changes caused by several common environmental pressures on the environmental behavior of POPs. The results showed increased bioaccumulation in higher trophic levels, but exceptions were noted because of ecological and physiological factors, such as diet, prey abundance, POPs degradation and isotope enrichment by microbial loop, growth rate, and lipid content. Food chain length was positively correlated with bioaccumulation. When POPs were transferred from the sediment to top consumers, the coupling of pelagic-benthic food web would enhance the exposure risks of higher trophic level consumers to POPs. Controversies existed whether the biomagnification of POPs along benthic food web was greater than that along the pelagic food web. Microbes could adsorb POPs more efficiently because of having a larger surface area. Microbes in the sediment decomposed organic materials, recycling POPs into the water column. Microbial degradation is an important way for POPs leaving aquatic ecosystem. Food web changes caused by seasonal succession, eutrophication and exotic invasions could change the direction and efficiency of energy transfer, and further resulted in changes of the environmental behavior of POPs. Most current studies mainly focused on the environmental behavior of POPs in pelagic food web, especially on the higher trophic levels, lacking researches on the environmental behavior of POPs in benthic and pelagic-benthic coupling food webs and microbial loops. Furthermore, the studies on environmental behaviors of POPs were always restricted to parts of the food web on small temporal and spatial scales, lacking researches on dynamic changes of the environmental behavior of POPs from multiple perspectives and large scales. New POPs have been increasing in production and usage, but studies about its environmental behavior along food web were still scare. Future studies should be conducted considering the above.

persistent organic pollutants (POPs); environmental behavior; trophic level; length of food chain; pelagic-benthic coupling food web; microbial food web; food web changes

國(guó)家重點(diǎn)基礎(chǔ)研究發(fā)展計(jì)劃(973計(jì)劃)(2015CB458900)

2016- 02- 04; 網(wǎng)絡(luò)出版日期:2016- 12- 19

10.5846/stxb201602040256

*通訊作者Corresponding author.E-mail: yongliu@pku.edu.cn

馮秋園,萬(wàn)祎,劉學(xué)勤,劉永.持久性有機(jī)污染物在水生食物網(wǎng)中的傳遞行為.生態(tài)學(xué)報(bào),2017,37(9):2845- 2857.

Feng Q Y, Wan Y, Liu X Q, Liu Y.Environmental behavior of persistent organic pollutants in aquatic food web.Acta Ecologica Sinica,2017,37(9):2845- 2857.