黃 輝,呂雨薇,梁 敏,朱詠莉,梁 文,蔣亞輝,張增強(qiáng),李榮華*

碳化溫度對牛糞銅和鋅形態(tài)及生態(tài)毒性的影響

黃 輝1,2,呂雨薇1,梁 敏3,朱詠莉1,梁 文2,蔣亞輝2,張增強(qiáng)2,李榮華2*

(1.南京林業(yè)大學(xué)生物與環(huán)境學(xué)院,江蘇 南京 210037；2.西北農(nóng)林科技大學(xué)資源環(huán)境學(xué)院,陜西 楊凌 712100；3.南京農(nóng)業(yè)大學(xué)資源與環(huán)境科學(xué)學(xué)院,江蘇 南京 210095)

通過設(shè)置不同的熱解溫度(350,550和750℃)對牛糞廢棄物進(jìn)行碳化處理,并使用光譜技術(shù)手段對牛糞炭的微觀特點及Cu、Zn賦存形態(tài)進(jìn)行了分析表征,同時結(jié)合淋溶和毒性實驗探究了熱解溫度對牛糞炭生態(tài)毒性的影響.結(jié)果表明,高溫碳化明顯改善牛糞孔隙結(jié)構(gòu),使其比表面積從牛糞原料的1.15m2/g提高至牛糞炭的5.51(350℃)～195.90m2/g(750℃).隨著熱解溫度的提高,牛糞炭pH值從8.18(350℃)提高到了10.14(750℃);牛糞炭中Cu、Zn含量則從牛糞原料中的1.22和1.23mg/g分別升高至18.29～35.11和18.58～31.24mg/g.透射電鏡-選區(qū)衍射以及X射線能譜分析表明,熱解處理可使牛糞中Cu、Zn離子分別轉(zhuǎn)化為副黑銅礦(Cu4O3)和紅鋅礦(ZnO)等金屬氧化物,從而明顯降低了牛糞炭中水溶態(tài)、DTPA提取態(tài)以及HNO3-H2SO4提取態(tài)的Cu、Zn離子濃度;此外,FTIR分析及混合有機(jī)酸浸提實驗結(jié)果也表明,350℃牛糞炭中酚羥基、烷烴基、羧基、酰胺類等有機(jī)官能團(tuán)通過吸附和絡(luò)合作用固定未完全轉(zhuǎn)化的Cu離子,而升高熱解溫度會使得這些官能團(tuán)顯著減少、促進(jìn)Cu離子的完全轉(zhuǎn)化以及無機(jī)物與Cu、Zn離子之間穩(wěn)定金屬氧化物化合鍵的形成.淋溶和生態(tài)毒性實驗表明,高于550℃的熱解溫度能夠顯著降低牛糞炭中Cu、Zn的溶出率以及生態(tài)毒性,是高Cu、Zn含量牛糞廢棄物無害化處理的一種推薦優(yōu)選技術(shù).

熱解溫度；牛糞；Cu；Zn；環(huán)境行為；生態(tài)毒性；影響因素

據(jù)估算,我國每年的畜禽糞便產(chǎn)生量高達(dá)38億t[1].畜禽糞便因含有豐富的有機(jī)質(zhì)及氮磷鉀等養(yǎng)分,是農(nóng)業(yè)種植過程的優(yōu)質(zhì)有機(jī)肥資源[2-3].然而,當(dāng)前我國規(guī)?；B(yǎng)殖場的畜禽糞便中重金屬含量普遍較高[4-7],尤其是Cu和Zn,其含量甚至可高達(dá)1622.81和14679.8mg/kg[5-6].較高Cu和Zn含量的畜禽糞便被長期當(dāng)作有機(jī)肥施入農(nóng)田,存在土壤和農(nóng)作物重金屬污染的環(huán)境風(fēng)險[8-9].

有機(jī)固體廢物高溫?zé)崽幚砑夹g(shù)因具有工藝簡單、運行經(jīng)濟(jì)、能顯著減少有機(jī)固體廢物的排放體積、促使物質(zhì)的穩(wěn)定化和無害化并產(chǎn)生高附加值資源(可燃?xì)?、生物油和生物?的優(yōu)點而廣受關(guān)注[10-13],其中生物炭具有鈍化土壤重金屬、調(diào)控生物多樣性及改良土壤鹽堿化等多重效果[14-17].例如Méndez等[18]和Yuan等[19]在利用含Cu和Zn污泥制備生物炭時發(fā)現(xiàn),熱解處理雖然顯著促進(jìn)了Cu和Zn元素的濃縮,導(dǎo)致總量提高,但卻明顯抑制了污泥生物炭中Cu和Zn離子的溶出量.Devi等[14]在研究含Cu和Zn紙漿經(jīng)高溫?zé)峤夂笊锾康男再|(zhì)時指出,經(jīng)炭化處理后重金屬離子的遷移能力會受到抑制.由此表明,熱解處理能有效降低有機(jī)固體廢物中的重金屬溶出風(fēng)險.但有機(jī)固體廢物經(jīng)高溫?zé)峤猥@得生物炭的環(huán)境穩(wěn)定性受控于熱解溫度、生物質(zhì)廢物原料類型和化學(xué)組成等諸多因素,要推進(jìn)有機(jī)固體廢物的熱處理技術(shù),仍需進(jìn)一步開展深入廣泛的研究[20].

本研究以牛糞為原材料,通過高溫缺氧熱解技術(shù)對牛糞進(jìn)行熱解處理,探討不同熱解溫度(350,550和750℃)對牛糞中Cu和Zn環(huán)境行為和潛在生態(tài)毒性的影響,為高重金屬含量畜禽糞便的無害化處理和資源化利用提供理論參考.

1 材料與方法

1.1 實驗材料

牛糞樣品采自陜西楊凌西北農(nóng)林科技大學(xué)農(nóng)作三站肉牛養(yǎng)殖場;試驗所需試劑包括Cu(NO3)2?3H2O,Zn(NO3)2?6H2O,CaCl2,二乙基三胺五乙酸(DTPA),三乙醇胺(TEA),甲酸,乙酸,乳酸,蘋果酸,檸檬酸,濃硝酸,濃鹽酸,濃高氯酸,過氧化氫,鹽酸羥胺、乙酸銨等,均為購自西隴化工股份有限公司的分析純試劑.

1.2 牛糞的熱解處理

將采集的新鮮牛糞自然風(fēng)干,使用研缽將風(fēng)干牛糞研磨并過2mm篩,裝入自封袋中備用.將300g牛糞干粉裝入1L燒杯,并分別加入800mL Cu(NO3)2和Zn(NO3)2溶液(其中Cu、Zn濃度均為0.5mg/mL),機(jī)械混勻攪拌4h后靜置2d.此后將上述混合物轉(zhuǎn)入80℃烘箱中進(jìn)行烘干處理,并將固體殘渣充分研磨并過0.15mm篩,之后將固體殘渣粉裝入事先稱重的150mL瓷坩堝,壓實并蓋上蓋子后稱重,計算裝填物質(zhì)重量.然后,將坩堝轉(zhuǎn)入馬弗爐并于氮氣保護(hù)條件下以10℃/min速率分別升溫至350,550和750℃,并在設(shè)定溫度下維持2h.此后關(guān)閉馬弗爐,繼續(xù)通入氮氣2h,待自然降溫到室溫后,取出坩堝中的熱解牛糞殘渣(生物炭),小心研磨并過0.15mm篩后,裝入自封袋備用.牛糞原料及生物炭樣品經(jīng)HCl-HNO3- HClO4消解后,用Hitachi Z-2000型原子吸收分光光度計測定CuZn總量[21].

1.3 牛糞炭的表征測試

牛糞炭的比表面積(BET)通過N2-吸附-脫附比表面積儀(V-Sorb 2800P,GoldAPP,中國)測定,利用氮氣吸附等溫線計算牛糞炭吸附累積孔內(nèi)表面積(csa)以及總孔體積(TPV)和平均孔徑(AVP) (AVP=4TPV/csa);用元素分析儀(Vario EL cube CHONS, Elementar,德國)測定C、H、O、N等元素含量;微觀形貌及金屬礦物晶型采用場發(fā)射高分辨透射電鏡-選區(qū)衍射(TEM-SEAD,JEM-1230, 日本)進(jìn)行分析;重金屬Cu和Zn的化合價態(tài)變化通過X射線光電子能譜分析儀(XPS,AXIS Ultra DLD,Shimadzu,日本)分析測定;牛糞炭的表面官能團(tuán)變化通過傅里葉紅外光譜儀(FTIR,Nicolet NEXUS 470,Thermo Nicolet,美國)測定,其中波數(shù)范圍為4000～400cm-1.

1.4 銅、鋅的淋溶能力及牛糞炭的生態(tài)毒性測試

在室溫條件下,將4份0.5g生物炭樣品分別用10mL的去離子水(pH值 6.9)[22]、有機(jī)混合酸(總濃度10mmol/L, pH值 5.60,甲酸、乙酸、乳酸、蘋果酸和檸檬酸的物質(zhì)的量比為1:4:2:2:1)、DTPA- TEA-CaCl2(0.005mol/L DTPA+0.01mol/L CaCl2+ 0.1mol/L TEA, pH值 7.30)及HNO3-H2SO4(體積比2:1, pH值 3.18)[23]振蕩提取24h,懸液經(jīng)0.45μm濾膜過濾后,以電感耦合等離子體質(zhì)譜(ICP-MS)測定濾液中Cu和Zn的含量.每個處理設(shè)置3個平行實驗.

為了解牛糞炭的生態(tài)毒性,試驗中將牛糞炭和去離子水以1:10(/)的比例加入到50mL離心管中,于25℃振蕩1h后靜置0.5h,用0.45μm濾膜過濾上清液,濾液保存?zhèn)溆?之后,在事先滅菌的潔凈培養(yǎng)皿中放入一張無菌濾紙,加入上述濾液5mL,并均勻分布播入10顆中國小白菜種子(品種:秦都96-12),蓋上培養(yǎng)皿蓋后放入25℃培養(yǎng)箱進(jìn)行培養(yǎng).每個處理設(shè)置4個平行,并設(shè)置去離子水處理為對照,培養(yǎng)48h后取出培養(yǎng)皿,計算種子發(fā)芽率和平均根長,并按公式(1)進(jìn)行種子發(fā)芽指數(shù)(GI)的計算.

GI=處理種子發(fā)芽率×處理平均根長/

(對照種子發(fā)芽率×對照平均根長) (1)

2 結(jié)果與討論

2.1 熱解溫度對牛糞炭理化性質(zhì)的影響

如表1所示,隨著熱解溫度的升高,牛糞炭對應(yīng)的BET、csa和TPV逐漸增大而AVP逐漸減小.例如,隨著熱解溫度從350℃增加到550和750℃,對應(yīng)的牛糞炭BET從5.51m2/g提高到81.19和195.90m2/ g,csa從4.55m2/g提高到了36.53和68.35m2/g,TPV從0.03cm3/g增加到0.06和0.12cm3/g,AVP則從20.73nm逐漸減小到6.46～ 6.67nm左右.這表明在缺氧熱解過程中牛糞發(fā)生了劇烈的物質(zhì)分解和炭化,促使生物炭的孔隙結(jié)構(gòu)形成[21,24-26].本研究中, 750℃的牛糞炭的AVP為6.67nm,稍大于550℃牛糞炭的6.46nm,說明過高的熱解溫度會導(dǎo)致炭的孔隙進(jìn)一步發(fā)生崩塌從而產(chǎn)生較大的孔隙[20].與原始牛糞相比,350℃熱解使牛糞有機(jī)質(zhì)發(fā)生分解并保留了一定量的酸性有機(jī)物質(zhì),從而導(dǎo)致牛糞炭pH值有所減小;此后,隨著溫度升高至550和750℃,酸性有機(jī)化合物在高溫下被分解并生成堿性物質(zhì)[12],從而導(dǎo)致牛糞炭pH值從8.18升高到9.95和10.14.但不同熱解溫度下生物炭的EC值不存在顯著差異(介于0.95～ 1.21mS/cm),暗示了高溫?zé)峤鈼l件對原材料組分中可溶性鹽成分的影響較小[27].隨著裂解溫度的提高,灰分產(chǎn)率從350℃的30.08%提高至750℃的49.70%,而炭產(chǎn)率則由46.43%逐漸降低至27.99%,這與Chen等[24]在熱解污泥研究中的觀察結(jié)果相類似;此外,隨著熱解溫度的升高,牛糞炭(350～750℃)的C/N比逐漸增大而H/C比和O/C比降低,表明升高熱解溫度會使牛糞炭中的芳香環(huán)結(jié)構(gòu)含量增加[28-29],且牛糞炭中的碳原子活性降低,滲碳能力受到抑制,暗示了牛糞炭施入土壤后的礦化速度會降低[18,30],這均表明了高溫?zé)峤饽茉鰪?qiáng)生物炭的環(huán)境穩(wěn)定性[27,31].此外,與牛糞原料(Cu 1.22mg/g, Zn 1.23mg/g)相比,熱解導(dǎo)致Cu和Zn發(fā)生濃縮使其含量分別從18.29和18.58mg/ g(350℃)升高到35.11和31.24mg/g(750℃),這與Jin等[21]和Zhang等[26]對城市污泥的炭化處理結(jié)果相類似.

表1 牛糞及牛糞炭的基本理化性質(zhì)

注:—為未檢驗;表格同一行中a、b、c、d等小寫字母表示不同處理之間數(shù)據(jù)在<0.05下存在顯著差異,數(shù)據(jù)顯著性分析軟件為IBM SPSS Statistics (Version 26.0).

2.2 熱解溫度對牛糞中Cu和Zn賦存形態(tài)變化的影響

如圖1所示,牛糞中Cu和Zn主要以自由金屬離子形式存在.但隨著熱解過程的進(jìn)行,350℃熱解的牛糞炭中分布著尺寸約為200nm大小的礦物顆粒,對其進(jìn)一步經(jīng)SEAD衍射分析后發(fā)現(xiàn),這些礦物組分有副黑銅礦(Cu4O3)、紅鋅礦(ZnO)、黑銅礦(CuO)等金屬氧化物礦物(圖1a);此后,當(dāng)熱解溫度進(jìn)一步提高到550和750℃時,由于濃縮效應(yīng)而使牛糞炭中包裹的金屬礦物明顯增多[30],且此時牛糞炭中金屬礦物種類主要以副黑銅礦和紅鋅礦為主(圖1b和1c).這表明富含高濃度Cu和Zn的牛糞經(jīng)高溫裂解后,Cu和Zn離子會分別相應(yīng)地轉(zhuǎn)化為Cu4O3及ZnO等金屬氧化物,這可能是有機(jī)質(zhì)炭化后強(qiáng)烈固定原物質(zhì)中金屬離子的直接原因[21,26,32].

圖1 不同溫度下牛糞炭的高分辨透射電鏡(TEM)照片及選區(qū)衍射(SEAD)光斑圖

如圖2所示,隨著裂解溫度從350℃提高到750℃,Zn 2p的譜圖并未發(fā)生明顯變化,說明了Zn元素在熱解過程中未發(fā)生價態(tài)變化(圖2a).在結(jié)合能1022和1045eV附近出現(xiàn)的Zn 2p 3/2和 Zn 2p 1/2的特征峰也表明了缺氧熱解過程中Zn(II)大部分以ZnO形式存在于牛糞生物炭中.然而與Zn不同,牛糞熱解過程明顯改變了Cu元素的價態(tài).Cu(II)峰的電子結(jié)合能與Cu 2p 3/2激發(fā)能的差值20.0eV是Cu(II)不同于Cu(I)及Cu(0)的主要區(qū)分標(biāo)志[33],350℃牛糞炭Cu(II)和Cu(0)峰對應(yīng)電子結(jié)合能差值為19.7eV (圖2b),這表明了熱解過程中Cu(II)→Cu(I)→Cu(0)還原過程的可能發(fā)生[34-35],即部分Cu(II)被缺氧還原成了Cu(0),從而促使副黑銅礦(Cu4O3)和黑銅礦(CuO)的形成;當(dāng)熱解溫度提高到550和750℃時,電子結(jié)合能952和933eV處兩個峰仍然存在,但在電子結(jié)合能963和943eV處出現(xiàn)兩個新峰,表明了熱解溫度超過550℃能使更低價態(tài)的Cu化合物形成,這也證實了高溫?zé)峤馀＜S過程中Cu(II)會被還原為Cu(I)并進(jìn)一步被還原為Cu(0)[33-34].這些結(jié)果證明了,高Cu、Zn含量的牛糞經(jīng)過不同溫度熱解處理后,Zn元素的價態(tài)未發(fā)生改變,主要以ZnO形式存在于牛糞炭中;而Cu元素則以Cu(II)、Cu(I)和Cu(0)等多種價態(tài)并存于牛糞炭中,在低溫(350℃)環(huán)境下形成Cu4O3和CuO,而在高溫(550℃及以上)下主要以Cu4O3形式存在于牛糞炭中.

圖2 不同熱解溫度下牛糞炭中Zn (2p)、Cu (2p)的XPS光譜圖

2.3 熱解溫度對牛糞炭Cu和Zn環(huán)境淋溶行為的影響

從圖3可知,不同熱解溫度下牛糞炭的Cu和Zn浸提濃度存在顯著差異.用去離子水浸提時,350℃牛糞炭中Cu和Zn浸提濃度分別為0.16和0.11mg/g;當(dāng)炭化裂解溫度升高到550℃時,Cu和Zn浸提濃度分別顯著降低至7.07和0.40μg/g;當(dāng)炭化裂解溫度進(jìn)一步升高到750℃時,Cu和Zn浸提濃度分別進(jìn)一步減少至6.21和0.10μg/g(圖3a),這些結(jié)果表明牛糞炭中Cu要比Zn更容易溶出,而熱解溫度高于550℃則對牛糞炭中Cu和Zn的溶出具有抑制作用. DTPA-TEA-CaCl2和HNO3-H2SO4浸提牛糞炭中Cu和Zn元素的淋溶特征與用去離子水浸提基本相似,即隨著熱解溫度從350℃升高到750℃,浸提液中Cu溶出濃度呈降低趨勢但仍然處于較高濃度水平(15.3～19.9mg/g),而Zn溶出濃度則從3.2mg/g (350℃)顯著減少至低于0.05mg/g(3550℃)(圖3b),這表明牛糞炭中具有濃度水平相對較高的螯合態(tài)Cu和Zn離子.隨著裂解溫度升高至750℃,螯合態(tài)Cu濃度仍然超過15mg/g,但螯合態(tài)Zn離子則顯著降低至0.05mg/g以下(圖3b),這一結(jié)果與XPS分析牛糞炭孔結(jié)構(gòu)中觀察到的Cu發(fā)生價態(tài)變化而Zn的價態(tài)不變有關(guān)的規(guī)律相印證[26,35],即Cu4O3(Cu2O·2CuO)礦物結(jié)構(gòu)容易被螯合劑破壞,而ZnO則較為穩(wěn)定.HNO3-H2SO4浸提的不同牛糞炭中Cu溶出濃度高達(dá)31.1(350℃)、1.1mg/g(550℃)和0.2mg/ g(750℃)(圖3c),表明350℃牛糞炭中Cu的生態(tài)毒性可能較大,而提高生物炭熱解溫度則能顯著降低Cu的淋溶風(fēng)險從而減低其生態(tài)環(huán)境風(fēng)險,這表明提高熱解溫度是降低有機(jī)固體廢棄物中重金屬元素淋溶風(fēng)險和生態(tài)毒性的有效手段[36-37].此外,隨炭化熱解溫度的升高,混合有機(jī)酸浸提的Zn元素濃度從0.7mg/g(350℃)顯著降低到0.05mg/g以下(3550℃) (圖3d),這表明高Zn含量的牛糞廢棄物經(jīng)高溫(3550℃)炭化后能夠有效降低其生態(tài)環(huán)境風(fēng)險[21].然而,混合有機(jī)酸浸提的Cu濃度呈現(xiàn)先顯著升高后顯著下降的趨勢,這可能與牛糞炭中官能團(tuán)的種類及含量變化有關(guān)[36-37].

圖3 不同浸提劑提取的不同裂解溫度得到的牛糞炭中Cu和Zn含量

圖中不同小寫字母表示數(shù)據(jù)在<0.05下存在顯著差異

如圖4所示, 隨著熱解溫度的升高,牛糞炭中官能團(tuán)的種類和數(shù)量均呈現(xiàn)顯著減少的變化.在4000～400cm-1波長范圍內(nèi),350℃牛糞炭中官能團(tuán)種類與數(shù)量最多,分別存在酚羥基(3391cm-1)、烷烴基(2926和1437cm-1)、羧基(1601cm-1)以及酰胺類(1601, 1096和781cm-1)等官能團(tuán)化合物[25,38]及600cm-1附近或以下范圍內(nèi)的有機(jī)物或無機(jī)物與重金屬離子之間的配位鍵等官能團(tuán)[21].然而,隨著碳化裂解溫度的提高,牛糞炭中官能團(tuán)的種類及相對含量均顯著減少,與350℃牛糞炭相比,550℃牛糞炭中僅存在少量的磷酸鍵官能團(tuán)(1028cm-1)以及明顯減少的有機(jī)物與重金屬離子之間的配位鍵.當(dāng)裂解溫度提高到750℃時,牛糞炭中僅存在少量的有機(jī)物與重金屬離子之間配位鍵,且配位鍵逐漸發(fā)展為無機(jī)物與金屬離子之間形成的金屬氧化物化合鍵(圖4).導(dǎo)致上述官能團(tuán)變化的原因可能是高溫(3550℃)裂解使得牛糞中烴類物質(zhì)缺氧轉(zhuǎn)化為生物氣(例如CO2、CH4或其他有機(jī)分子氣體)、生物油(芳香結(jié)構(gòu)物質(zhì))等附屬產(chǎn)品,從而導(dǎo)致有機(jī)官能團(tuán)種類及數(shù)量顯著減少,并促進(jìn)金屬氧化物化合鍵的形成[21,26,37].這一結(jié)果也與本研究中TEM及SEAD分析的金屬氧化物礦物晶型演化結(jié)果相印證,即當(dāng)熱解溫度從350℃升高到550℃時,大量存在的有機(jī)官能團(tuán)發(fā)生缺氧轉(zhuǎn)化,同時結(jié)合的Cu離子被未完全轉(zhuǎn)化,使得混合有機(jī)酸從牛糞炭中提取的Cu元素濃度較高;而熱解溫度提高至750℃時,未完全轉(zhuǎn)化的Cu離子進(jìn)一步形成更加穩(wěn)定的金屬氧化物,從而導(dǎo)致混合有機(jī)酸浸提的750℃牛糞炭中Cu濃度顯著降低(圖3d,圖4).

圖4 不同熱解溫度下牛糞炭傅里葉紅外光譜圖

2.4 牛糞炭的生態(tài)毒性

如圖5所示,在對照去離子水中,小白菜種子基本能全部發(fā)芽(發(fā)芽指數(shù)為98.7%),其平均根長為19.1mm;與之相比,350℃牛糞炭浸出液中,近70%的小白菜種子受到重金屬Cu、Zn毒害而不能發(fā)芽(圖5a),其平均根長顯著受到抑制,僅為12.3mm(圖5b),這與低溫裂解生物炭中存在大量小分子有機(jī)物和較高的活性Cu和Zn有關(guān)[36-37].而當(dāng)熱解溫度升高至550和750℃時,可使小白菜種子發(fā)芽指數(shù)明顯提高至350℃牛糞炭處理的2.4～3.1倍(圖5a),并使小白菜平均根長增長至15.5～16.5mm(圖5b),說明提高熱解溫度至550℃以上,能夠減少小分子有機(jī)物含量并使生物活性較高的Cu、Zn轉(zhuǎn)化為穩(wěn)定的低毒性金屬氧化物,從而顯著降低牛糞炭的生態(tài)毒性,并明顯緩解這些生物炭對植物生長的抑制效應(yīng).

圖5 不同熱解溫度的牛糞炭浸出液中大白菜種子發(fā)芽指數(shù)及發(fā)芽平均根長

生物炭與去離子水比率為1:10(/)

3 結(jié)論

3.1 高溫?zé)峤饽軐⑴＜S轉(zhuǎn)化為富含穩(wěn)定孔隙結(jié)構(gòu)的牛糞生物炭.

3.2 在熱解過程中,Cu和Zn會發(fā)生濃縮效應(yīng)而導(dǎo)致總量顯著升高;牛糞中的Cu和Zn離子會隨熱解的溫度升高而發(fā)生形態(tài)轉(zhuǎn)化,缺氧熱解使得部分Cu(II)被還原為Cu(I)和Cu(0)并促使副黑銅礦(Cu4O3)和黑銅礦(CuO)的形成,同時會促使Zn(II)轉(zhuǎn)化為穩(wěn)定的紅鋅礦(ZnO),從而對Cu和Zn產(chǎn)生強(qiáng)烈的固定作用.

3.3 升高熱解溫度會使牛糞炭中酚羥基、烷烴基、羧基、酰胺類以及磷酸鍵等官能團(tuán)種類及含量均顯著減少,同時促使無機(jī)物與Cu和Zn離子之間形成更加穩(wěn)定的金屬氧化物化合鍵,從而明顯降低牛糞炭中Cu和Zn的淋溶風(fēng)險,進(jìn)而有效降低牛糞炭的生態(tài)毒性.

3.4 將溫度提升到550℃以上進(jìn)行安全熱解處理,有利于降低高Cu和Zn含量牛糞的環(huán)境生態(tài)風(fēng)險.

[1] 葛勉慎,周海賓,沈玉君,等.添加劑對牛糞堆肥不同階段真菌群落演替的影響[J]. 中國環(huán)境科學(xué), 2019,39(12):5173-5181.

Ge M S, Zhou H B, Shen Y J, et al. Effect of additives on the succession of fungal community in different phases of cattle manure composting [J]. China Environmental Science, 2019,39(12):5173- 5181.

[2] 吳浩瑋,孫小淇,梁博文,等.我國畜禽糞便污染現(xiàn)狀及處理與資源化利用分析[J]. 農(nóng)業(yè)環(huán)境科學(xué)學(xué)報, 2020,39(6):1168-1176.

Wu H W, Sun X Q, Liang B W, et al. Analysis of livestock and poultry manure pollution in China and its treatment and resource utilization [J]. Journal of Agro-Environment Science, 2020,39(6):1168-1176.

[3] 岳 霞,陳德珍,安 青,等.生物炭對污泥熱解液與牛糞共厭氧發(fā)酵的影響[J]. 中國環(huán)境科學(xué), 2022,42(3):1267-1277.

Yue X, Chen D Z, An Q, et al. Biochars enhancing anaerobic co-digestion of sewage sludge pyrolysis liquid and cow dung: influences of inorganics in biochar raw [J]. China Environmental Science, 2022,42(3):1267-1277.

[4] Ahadi N, Sharifi Z, Hossaini S M T, et al. Remediation of heavy metals and enhancement of fertilizing potential of a sewage sludge by the synergistic interaction of woodlice and earthworms [J]. Journal of Hazardous Materials, 2020,385:121573.

[5] Li Y X, Xiong X, Lin C Y, et al. Cadmium in animal production and its potential hazard on Beijing and Fuxin farmlands [J]. Journal of Hazardous Materials, 2010,177(1):475-480.

[6] 潘 尋,韓 哲,賁偉偉.山東省規(guī)?；i場豬糞及配合飼料中重金屬含量研究[J]. 農(nóng)業(yè)環(huán)境科學(xué)學(xué)報, 2013,32(1):160-165.

Pan X, Han Z, Ben W W. Heavy metal contents in pig manure and pig feeds from intensive pig farms in Shandong Province, China [J]. Journal of Agro-Environment Science, 2013,32(1):160-165.

[7] 朱建春,李榮華,張增強(qiáng),等.陜西規(guī)模化豬場豬糞與飼料重金屬含量研究[J]. 農(nóng)業(yè)機(jī)械學(xué)報, 2013,44(11):98-104.

Zhu J C, Li R H, Zhang Z Q, et al. Heavy metal contents in pig manure and feeds under intensive farming and potential hazard on farmlands in Shananxi Province, China [J]. Transactions of the Chinese Society for Agricultural Machinery, 2013,44(11):98-104.

[8] Chen H, Yuan X, Li T, et al. Characteristics of heavy metal transfer and their influencing factors in different soil–crop systems of the industrialization region, China [J]. Ecotoxicology and Environmental Safety, 2016,126:193-201.

[9] Liu Z, Tran K Q. A review on disposal and utilization of phytoremediation plants containing heavy metals [J]. Ecotoxicology and Environmental Safety, 2021,226:112821.

[10] Gholizadeh M, Hu X, Liu Q. Progress of using biochar as a catalyst in thermal conversion of biomass [J]. Reviews in Chemical Engineering, 2021,37(2):229-258.

[11] Guo H N, Wu S B, Tian Y J, et al. Application of machine learning methods for the prediction of organic solid waste treatment and recycling processes: A review [J]. Bioresource Technology, 2021,319: 124114.

[12] Mahima J, Sundaresh R K, Gopinath K P, et al. Effect of algae (Scenedesmus obliquus) biomass pre-treatment on bio-oil production in hydrothermal liquefaction (HTL): Biochar and aqueous phase utilization studies [J]. Science of the Total Environment, 2021,778: 146262.

[13] Nanda S, Berruti F. A technical review of bioenergy and resource recovery from municipal solid waste [J]. Journal of Hazardous Materials, 2021,403:123970.

[14] Devi P, Saroha A K. Risk analysis of pyrolyzed biochar made from paper mill effluent treatment plant sludge for bioavailability and eco-toxicity of heavy metals [J]. Bioresource Technology, 2014,162: 308-315.

[15] 龍秋寧,王潤松,徐涵湄,等.沼液與生物炭聯(lián)合施用對楊樹人工林土壤甲螨密度的影響[J]. 南京林業(yè)大學(xué)學(xué)報(自然科學(xué)版), 2021, 44(3):211-215.

Long Q N, Wang R S, Xu H M, et al. Effects of biogas slurry and biochar on oribatida density in poplar plantation [J]. Journal of Nanjing Forestry University (Natural Science Edition), 2021,44(3): 211-215.

[16] 姚晶晶,馮象千,肖 賀,等.不同固廢及其處理產(chǎn)物對黃驊港鹽堿土的改良效果[J]. 南京林業(yè)大學(xué)學(xué)報(自然科學(xué)版), 2021,45(3):45- 52.

Yao J J, Feng X Q, Xiao H, et al. Improvement effects of different solid waste and their disposal by products on saline-alkali soil in Huanghua Port [J]. Journal of Nanjing Forestry University (Natural Science Edition), 2021,45(3):45-52.

[17] 馬鋒鋒,趙保衛(wèi),鐘金魁,等.牛糞生物炭對磷的吸附特性及其影響因素研究[J]. 中國環(huán)境科學(xué), 2015,35(4):1156-1163.

Ma F F, Zhao B W, Zhong J K, et al. Characteristics phosphate adsorption onto biochars derived from dairy manure and its influencing factor [J]. China Environmental Science, 2015,35(4): 1156-1163.

[18] Mendez A, Terradillos M, Gasco G. Physicochemical and agronomic properties of biochar from sewage sludge pyrolysed at different temperatures [J]. Journal of Analytical and Applied Pyrolysis, 2013, 102:124-130.

[19] Yuan H, Tao L, Huang H, et al. Influence of pyrolysis temperature on physical and chemical properties of biochar made from sewage sludge [J]. Journal of Analytical and Applied Pyrolysis, 2015,112:284-289.

[20] Jeong C Y, Dodla S K, Wang J J. Fundamental and molecular composition characteristics of biochars produced from sugarcane and rice crop residues and by-products [J]. Chemosphere, 2016,142:4-13.

[21] Jin J, Li Y, Zhang J, et al. Influence of pyrolysis temperature on properties and environmental safety of heavy metals in biochars derived from municipal sewage sludge [J]. Journal of Hazardous Materials, 2016,320:417-426.

[22] HJ 557-2009 固體廢物浸出毒性浸出方法水平振蕩法 [S].

HJ 557-2009 Solid waste-Extraction procedure for leaching toxicity- Horizontal vibration method [S].

[23] HJ/T299-2007 固體廢物浸出毒性浸出方法硫酸硝酸法 [S].

HJ/T299-2007 Solid waste-Extraction procedure for leaching toxicity-Sulphuric acid & nitric acid method [S].

[24] Chen T, Zhang Y, Wang H, et al. Influence of pyrolysis temperature on characteristics and heavy metal adsorptive performance of biochar derived from municipal sewage sludge [J]. Bioresource Technology, 2014,164:47-54.

[25] Ma F, Dai J S, Fu Z, et al. Biochar for asphalt modification: A case of high-temperature properties improvement [J]. Science of the Total Environment, 2022,804:150194.

[26] Zhang P Z, Zhang X X, Li Y F, et al. Influence of pyrolysis temperature on chemical speciation, leaching ability, and environmental risk of heavy metals in biochar derived from cow manure [J]. Bioresource Technology, 2020,302:122850.

[27] Zhang H Z, Chen C R, Gray E M, et al. Effect of feedstock and pyrolysis temperature on properties of biochar governing end use efficacy [J]. Biomass & Bioenergy, 2017,105:136-146.

[28] Park J, Hung I, Gan Z H, et al. Activated carbon from biochar: Influence of its physicochemical properties on the sorption characteristics of phenanthrene [J]. Bioresource Technology, 2013,149: 383-389.

[29] Zhang P, Huang P, Xu X J, et al. Spectroscopic and molecular characterization of biochar-derived dissolved organic matter and the associations with soil microbial responses [J]. Science of the Total Environment, 2020,708:134619.

[30] Cheng Y, Luo L, Lv J, et al. Copper speciation evolution in swine manure induced by pyrolysis [J]. Environmental Science & Technology, 2020,54(14):9008-9014.

[31] Mobaraki M, Afshang B, Rahimpour M R, et al. Effect of cracking feedstock on carburization mechanism of cracking furnace tubes [J]. Engineering Failure Analysis, 2020,107:104216.

[32] Huang H, Liang W, Li R, et al. Converting spent battery anode waste into a porous biocomposite with high Pb (II) ion capture capacity from solution [J]. Journal of Cleaner Production, 2018,184:622-631.

[33] Din S U, Awan J M, Imran M, et al. Novel nanocomposite of biochar-zerovalent copper for lead adsorption [J]. Microscopy Research and Technique, 2021,84(11):2598-2606.

[34] Liu W J, Tian K, Jiang H, et al. Harvest of Cu NP anchored magnetic carbon materials from Fe/Cu preloaded biomass: their pyrolysis, characterization, and catalytic activity on aqueous reduction of 4-nitrophenol [J]. Green Chemistry, 2014,16(9):4198-4205.

[35] Liu W J, Tian K, Jiang H, et al. Selectively improving the bio-oil quality by catalytic fast pyrolysis of heavy-metal-polluted biomass: Take copper (Cu) as an example [J]. Environmental Science & Technology, 2012,46(14):7849-7856.

[36] Wang Q, Wang B, Ma Y, et al. Stabilization of heavy metals in biochar derived from plants in antimony mining area and its environmental implications [J]. Environmental Pollution, 2022,300:118902.

[37] Zhang X, Zhao B, Liu H, et al. Effects of pyrolysis temperature on biochar’s characteristics and speciation and environmental risks of heavy metals in sewage sludge biochars [J]. Environmental Technology & Innovation, 2022,26:102288.

[38] Yang G X, Jiang H. Amino modification of biochar for enhanced adsorption of copper ions from synthetic wastewater [J]. Water Research, 2014,48:396-405.

Effects of pyrolysis temperatures on occurrence speciation of copper and zinc in cattle manures and their potential ecotoxicity.

HUANG Hui1,2, Lü Yu-wei1, LIANG Min3, ZHU Yong-li1, LIANG Wen2, JIANG Ya-hui2, ZHANG Zeng-qiang2, LI Rong-hua2*

(1.College of Biology and the Environment, Nanjing Forestry University, Nanjing 210037, China；2.College of Natural Resources and Environment, Northwest A&F University, Yangling 712100, China；3.College of Resources and Environmental Sciences, Nanjing Agricultural University, Nanjing 210095, China)., 2022,42(9)：4240～4247

Pyrolysis to biochar is an efficient technique of disposing organic solid wastes. However, pyrolysis treatment of the cattle manure (CM) containing high concentrations of Cu and Zn is scarcely investigated, and the environmental behaviors of Cu and Zn in the cattle manure biochar (CMB) accompanying with their ecological risk keep unknown. In this study, the occurrence speciation of Cu and Zn in CMB, their leaching characteristics and the ecological risk of CMB from different pyrolysis temperatures (350, 550 and 750oC) were evaluated using spectrum technology (i.e., BET, TEM-SEAD, XPS, and FTIR), the leaching experiments, and the risk tests. Results show that carbonization of CM to CMB with an increasing temperature from 350 to 750oC improves the pore structure of CM and enlarges the specific surface area from 1.15m2/g in raw CM to 5.51～195.90m2/g in CMB. The pH value of the CMB increases from 8.18 (350℃) to pH 10.14 (750℃). Importantly, the Cu concentration increases from 1.22mg/g in raw CM to 18.29～35.11mg/g in CMB while the Zn concentration elevates from 1.23mg/g to 18.58～31.24mg/g. Meanwhile, most of the Cu and Zn are oxidized to paramelaconite (Cu4O3) and zincite (ZnO), which significantly reduces the concentrations of Cu and Zn in forms of extractable ones with deionized water, DTPA-TEA-CaCl2and HNO3-H2SO4as extracting agents, respectively. Furthermore, many functional groups (i.e., phenolic hydroxyl, alkane, carboxyl, amide, etc.) can immobilize labile Cu through adsorption and complexation, but a high pyrolysis temperature (>550oC) tends to arise a significant decrease in species and amounts of functional groups and promote the complete conversion of Cu ions and the formation of stable metal oxide bonds between inorganics and Cu or Zn ions. In summary, the pyrolysis temperature over 550oC could dramatically reduce the leaching rates of Cu and Zn and mitigate the ecotoxicity of CMB, which can be a potential approach to dispose the CM wastes with high concentrations of Cu and Zn.

pyrolysis temperature；cattle manure；copper zinc；environmental behavior；ecological risk；influencing factors

X705,X171

A

1000-6923(2022)09-4240-07

2022-02-23

中央高校基本科研業(yè)務(wù)費專項資金項目(2452015177);南京林業(yè)大學(xué)水杉師資科研啟動項目(163108167)

*責(zé)任作者, 教授, rh.lee@nwsuaf.edu.cn

黃 輝(1993-),男,江蘇連云港人,副教授,博士,主要從事固體廢棄物資源化與重金屬污染控制.發(fā)表論文23篇.