外文翻譯---利用冶金渣副產(chǎn)物去除酸性礦山廢水中的污染物

1、<p>　　Removal of pollutants from acid mine wastewater</p><p>　　using metallurgical by-product slags</p><p>　　D. Feng a,?, J.S.J. van Deventer a, C. Aldrich b</p><p>　　a Departm

2、ent of Chemical and Biomolecular Engineering, The University of Melbourne, Melbourne, Vic., 3010, Australia</p><p>　　b Department of Chemical Engineering, University of Stellenbosch, Private Bag X1, Matielan

3、d, 7602, Stellenbosch, South Africa</p><p>　　Received in revised form 8 January 2004; accepted 12 January 2004</p><p><b>　　Abstract</b></p><p>　　The removal of pollutant

4、s from acid mine drainage using metallurgical by-product slags was studied in laboratory scale. Metallurgical by-product furnace slags were used as sorbents for metal ions and dispersed air column flotation was employed

5、for the solid/liquid separation of the loaded slags. Batch sorption/pH/kinetic studies were conducted using simulated Cu and Pb bearing wastewater. The calcium glass type of slags had high surface area and porosity. Prom

6、ising result was succeeded from the </p><p>　　© 2004 Elsevier B.V. All rights reserved.</p><p>　　Keywords: Furnace slag; Sorption; Flotation; Wastewater treatment; Acid mine drainage</p&

7、gt;<p>　　1. Introduction</p><p>　　Various methods exist for the removal of toxic metal ions from aqueous solution, viz. ion exchange, reverse osmosis, precipitation and adsorption, among others. Adsor

8、ption is by far the most versatile and widely used process.Activated carbon has been the standard adsorbent for the reclamation of municipal and industrial wastewaters. Owing to the high-cost of activated carbon, product

9、ion of its low-cost alternatives has been the focus of research in this area for years. These sorbents for the hea</p><p>　　removal efficiency of mixed metals cannot be achieved at a single precipitation pH

10、level. Conventional sorbents are not acceptable in such a mal-condition as acidic high-turbidity mine drainage. Slags can be used as low-cost adsorbents and neutralising agents and viable alternatives to the combination

11、of much more expensive activated carbon or ion exchange resins and lime. Slags exist often in a powdered form and are mainly applied as dispersions. Downstream of the reaction tank, a suitable soli</p><p>　　

12、2. Experimental work</p><p>　　Two furnace slags, viz. iron making slag and steel making slag, were obtained from Saldanha Steel South Africa in the form of powder with a mean particle size of 24.5 and 24.1 m

13、, respectively. The size distribution was: 100% and100m; 90% and 45 m; 22% and 10 m; and 1.2% and 1m for the iron slag, compared to 100% and 100m; 90% and 45m; 23% and 10m; and 1.5% and 1m for the steel slag. The chemica

14、l composition of the slags expressed as oxides in mass percentage is shown in Table 1.The XRD spectra o</p><p>　　Experiments. HCl and NaOH (analytical grade from Merck) were used to adjust the solution pH. T

15、he Cu and Pb stock solutions were prepared by dissolving their corresponding chloride or nitrate salts (CuCl2 · 5H2O and Pb(NO3)2, analytical grade from Merck) in distilled water. The ion concentrations in stock sol

16、utions were about 5000 mg/l. A cationic flocculant (polyamine type) was obtained from Monica, South Africa.</p><p>　　Sodium dodecyl sulphate (analytical grade from Sigma) was used as a collector in flotatio

17、n. Batch sorption experiments were carried out at an ambient temperature of about 18 ?C on a roller (60 min?1) using 1 l screwed cap plastic bottles. The sorption isotherm studies were conducted by varying the initial io

18、n concentrations. After a contact time of 24 h, the reaction mixture was filtered through a 0.45 m membrane filter (Millipore) and the filtrate was analysed for ion contents. The loading of t</p><p>　　The me

19、tal ion contents in solutions were determined by Varian Inductively Coupled Plasma (ICP). The solution pH values were detected by CRISON Micro pH 2000. Electrokinetic measurements for the determination of point of zero c

20、harge (PZC) were carried out on a Laser Zeta meter (Rank Brothers, Cambridge). The surface area of the sample was measured by BET method on Micropores (Model ASAP 2010, Micromeritics Instrument Corporation, Norcross,GA).

21、 The common inorganic anions were determined by a DION</p><p>　　3. Results and discussion</p><p>　　3.1. Sorption equilibria</p><p>　　Two important physico–chemical aspects for the e

22、valuation of the sorption process as a unit operation are the equilibria of sorption and the kinetics. Sorption equilibrium is established when the concentration of metal in a bulk solution is in dynamic balance with tha

23、t of the interface. Fig. 1 shows typical sorption isotherms of Cu2+ and Pb2+ on the two slags, respectively. The slurry pH was maintained at 5.5 and the slag doses were 2 g/l.</p><p>　　Fig. 1. Sorption isoth

24、erms of Cu2+ and Pb2+ on the slags. IS represents</p><p>　　iron slag and SS steel slag. The maximum standard deviation was within</p><p><b>　　3%.</b></p><p>　　As can be

25、seen from Fig. 1, the iron slag had a much higher sorption capacity for heavy metals than the steel slag and Cu2+ had a higher loading on the slags than Pb2+ in terms of molar numbers. The sorption isotherms of uptake me

26、tals by the slags could be expressed as Langmuir isotherms. The metal sorption constants for the Cu2+ and Pb2+ on the slags were calculated from Langmuir plots and are given in Table 2. The Langmuir parameters, Qmax (mg/

27、g) and K (l/mg), are related to saturation capacit</p><p>　　Fig. 2. Dependence of the slag loading with Cu2+ and Pb2+ on equilibriumsolution pH. IS represents iron slag and SS steel slag.</p><p>

28、;　　3.2. Effect of pH on sorption</p><p>　　The original solution concentrations were about 200 mg/l, the slag doses were 2 g/l and the contact time was 24 h. Fig. 2 illustrates the pH dependence of the slag l

29、oading with Cu2+ and Pb2+, respectively. Clearly, the metal uptake was quite low at low pH levels. However, with an increase in solution pH, a significant enhancement in sorption was recorded for both slags, with the opt

30、imum pH values for the metal sorption by the iron and steel slags being 3.5–8.5 and 5.2–8.5, respectively. This beh</p><p>　　3.3. Sorption kinetics</p><p>　　The slag doses were 2 g/l and the slu

31、rry pH was maintained at 5.5. The initial concentrations for Cu2+ and Pb2+ solutions were around 200 mg/l, respectively. The sorption kinetic result is shown in Fig. 3. The sorption equilibrium for the steel slag could r

32、each within 1 h, while the sorption equilibrium for the iron slag could only reach till 24 h.</p><p>　　Fig. 3. Sorption kinetics of Cu2+ and Pb2+ on the slags. IS represents</p><p>　　iron slag a

33、nd SS steel slag.</p><p>　　3.4. Sorption mechanism</p><p>　　In view of the nature of the slags, an exchange interaction of the slag glass with the solution can be expressed as follows</p>

34、<p><b>　　[11]:</b></p><p>　　=SiOCa + 2HOH → =SiOH2 + Ca2+ + 2OH? (1)</p><p>　　It can be expected that in acid environment the above reaction would shift to the left due to the

35、high concentration of hydrogen ions. The basic slags had neutralising effect following the above route. Different amounts of slags were added into 1 l distilled water and stirred at 600 min?1 for 24 h. Table 3 shows the

36、changes of the solution pH and Ca2+ concentrations. Clearly, the calcium ion exchanged with the hydrogen ion and was released from the slags into the solutions, while the solution pH</p><p>　　=SiOCa +M2+ → =

37、SiOM + Ca2+ (2)</p><p>　　The slags had high surface area and porosity as indicated in Table 1, which provided the basis for metal sorption as well. As indicated by the microstructures of the slags in</p&

38、gt;<p>　　Fig. 4. Typical topographic image for the original iron slag by SEM. Bar</p><p>　　length is 1 m and the magnification 10000×.</p><p>　　Fig. 5. Typical topographic image for

39、the original steel slag by SEM. Bar</p><p>　　length is 1 m and the magnification 10000×.</p><p>　　Figs. 4–6, both slags revealed porous structures and the iron slag had smaller pores than t

40、he steel slag. After sorption, some loose flocs appeared at the slag surface as indicated in</p><p>　　Fig. 6. These flocs could be alumino silicates of copper or copper hydroxide complexes depositing on the

41、porous slag surface.</p><p>　　Fig. 6. Typical topographic image for the copper loaded iron slag by</p><p>　　SEM. Bar length is 1 m and the magnification 10000×.</p><p>　　3.5. S

42、lag flotation</p><p>　　Some loose and fairly small flocs appeared in the slurry after 24 h slag adsorption. The solid/liquid separation by simple sedimentation was not complete in short times (up to 10 h) an

43、d required a further stage of filtration. Both ion-exchange and metal precipitation contributed to the removal of metal ions from wastewater by the slags. The hydroxide species of the metal ions increased the turbidity o

44、f solutions after the adsorption. In addition, the presence of some fine slag particles could al</p><p>　　solids, dispersed air column flotation was employed for the subsequent separation of loaded slags fro

45、m solutions. The initial Cu and Pb concentrations were 200 mg/l and the slag doses were 2 g/l. The solution pH was maintained at 5.5. After 24 h contact, the slurry was subjected to flotation at an air flowrate of 200 ml

46、/min. The average bubble size in the flotation column was around 0.62 mm, based on froth image analysis. The flotation result is shown in Table 4. As can be seen from Table 4, at </p><p>　　3.6. Treatment of

47、an acid mine drainage by flotation ofadsorptive slags</p><p>　　The acid mine water used in the experiments was sampled from the drainage of a South African gold mine. The acid mine water had a very low pH of

48、 2.03 and a very high sulphate concentration, as indicated in Table 5. The acid mine water is neutralised with lime in site, resulting in concomitant precipitation of iron, aluminium and other metal hydroxides. Treatment

49、 with lime requires a short reaction period, owing to its high solubility (0.15%). However, since the minimum solubilities for the diffe</p><p>　　a Ion concentration in mg/l.</p><p>　　Direct flo

50、tation by SDS was not very effective for the solid/liquid separation. A cationic flocculant polyamine at an optimised concentration of 100 mg/l was added into the slurry and conditioned for 3 min prior to the addition of

51、 SDS. The fine flocs and slags formed large and well-structured aggregates with the addition of polyamine, which were readily removed by the subsequent flotation process. The flotation result is shown in Table 6. Obvious

52、ly, the residual concentrations of all the base me</p><p>　　a Ion concentration in mg/l.</p><p>　　solution turbidity was under 1.0 NTU (tap water) after the slag neutralisation and sorption, flo

53、cculation and flotation process. The cationic flocculant not only flocculated the soft fine particles to form large aggregates, but also served as a carrier for the removal of anions such as anionic precious metal comple

54、xes due to the electrostatic attraction. After flocculation, the aggregates were more resistant to shearing and suitable for the subsequent flotation. The iron slag had a better perfor</p><p>　　4. Conclusion

55、s</p><p>　　The iron and steel making slags were appropriate sorbents for heavy metal removal from aqueous solutions. The slags combined ion-exchange and sorption properties with an acid-neutralising ability.

56、 The iron slag had a much higher sorption capacity for metals than the steel slag owing to its higher surface area, higher porosity and higher ion-exchangeeffectively separate the slags, yielding very low solution turbid

57、ities. Advantages when compared with settling were observed in terms of process kine</p><p>　　References</p><p>　　[1] V.K. Gupta, S.K. Srivastava, D. Mohan, Ind. Eng. Chem. Res. 36</p>&l

58、t;p>　　(1997) 2207.</p><p>　　[2] K.R. Ramakrishna, T. Viraraghavan, Waste Manage. 17 (8) (1997)</p><p><b>　　483.</b></p><p>　　[3] S.K. Srivastava, V.K. Gupta, D. Mohan

59、, J. Environ. Eng. Div. Am.</p><p>　　Soc. Civ. Eng. 123 (1997) 461.</p><p>　　[4] V.K. Gupta, A. Rastogi, M.K. Dwivedi, D. Mohan, Sep. Sci. Technol.</p><p>　　32 (1997) 2883.</p>

60、;<p>　　[5] K.A. Matis, A.I. Zouboulis, I.C. Hancock, Sep. Sci. Technol. 29</p><p>　　(1994) 1055.</p><p>　　[6] Th.F. Zabel, in: P. Mavros, K.A. Matis (Eds.), Innovations in Flotation</p

61、><p>　　Technology, 1992, Kluwer Academic, Dordrecht, The Netherlands,</p><p><b>　　p. 431.</b></p><p>　　[7] A.I. Zouboulis, K.A. Matis, K.A. Stalidis, G.A. in: P. Mavros, K.

62、A.</p><p>　　Matis (Eds.), Innovations in Flotation Technology, 1992, Kluwer</p><p>　　Academic, Dordrecht, The Netherlands, p. 475.</p><p>　　[8] C. Aldrich, D. Feng, Minerals Eng. 13

63、 (10–11) (2000) 1129.</p><p>　　[9] J. Rubio, F. Tessele, Minerals Eng. 10 (7) (1997) 671.</p><p>　　[10] A.I. Zouboulis, K.A. Kydros, K.A. Matis, Water Res. 29 (1995) 1755.</p><p>　　

64、[11] K.K. Panday, G. Prasad, V.N. Singh, Water Res. 19 (1985) 869.</p><p>　　[12] S.V. Dimitrova, D.R. Mehandgiev, Water Res. 32 (11) (1998) 3289.</p><p>　　利用冶金渣副產(chǎn)物去除酸性礦山廢水中的污染物</p><p&

65、gt;<b>　　摘要</b></p><p>　　在實驗室規(guī)模，研究利用冶金副產(chǎn)物渣料去除酸性礦物排泄的污染物。作為吸附劑用于金屬離子和分散空氣柱浮選的冶金學(xué)副產(chǎn)物還原渣被用作固體/液體的分離。利用模擬的Cu和Pb的軸承廢水，實施了批量吸附/PH/動力學(xué)研究。鈣玻璃型渣具有較高的表面積和孔隙度。在處理南非金礦酸性礦山廢水的渣的吸附/浮選的聯(lián)合過程中成功地獲得了預(yù)期結(jié)果。</p>

66、<p><b>　　引言</b></p><p>　　從水溶液中去除有毒金屬離子有許多方法，離子交換、反滲透、沉淀、吸附及其他。吸附是目前為止最廣泛和應(yīng)用最普遍的工藝?；钚蕴渴鞘姓凸I(yè)廢水再利用的標(biāo)準(zhǔn)吸附劑。由于活性炭的高成本，數(shù)年來，廉價的替代產(chǎn)品成為這一領(lǐng)域研究的焦點。用于重金屬吸附的吸附劑從自然材料到工農(nóng)業(yè)副產(chǎn)物，如粉煤灰、碳材料、金屬氧化物、沸石、泥沼、苔蘚、氫氧化

67、物、木質(zhì)素、粘土、生物質(zhì)、花生殼、硫鐵礦骨料、針鐵礦、珊瑚沙子。</p><p>　　還原渣作為冶金副產(chǎn)物用作填充物或者礦渣水泥產(chǎn)品中。據(jù)報道顆粒狀還原渣能夠轉(zhuǎn)變成有效的吸附劑，可以用于去除染色劑和金屬離子。以堿性為主的作為各種重金屬離子非常規(guī)吸附劑的渣利用酸中和能力使離子交換和吸附性能結(jié)合。酸性采礦水是采礦和礦物工業(yè)不可避免的副產(chǎn)物，特別是就氧化的硫化礦物而言，令人擔(dān)憂。酸性采礦水往往含有高濃度的可溶解重金屬和

68、硫酸鹽，而且可能有較高的渾濁度和低于2 的PH 。這些條件可能阻止 未經(jīng)處理的酸性礦水排入公用的河流，因為他們對水生植物和魚的生命有有害的影響。類似地，由酸性礦水排放造成地地下水污染是一個同等麻煩的問題。傳統(tǒng)上，通過加入石灰處理，酸性礦水被中和，導(dǎo)致離子共沉淀，鋁和其它金屬的氫氧化物。然而，發(fā)現(xiàn)受污染水的不同金屬的最小溶解度出現(xiàn)在不同的PH值，而且在自然界中的氫氧化物是兩性的，在單一的沉淀PH水平混合金屬的最大去除效率不能夠獲得。作為酸

69、性高濁度礦物排水在如此惡劣的條件下常規(guī)吸附劑是不可接受的。礦渣可用于低成本的吸附劑和中和劑、可行的替代物來結(jié)合更加昂貴的活性炭或者離子交換樹脂和石灰。 </p><p>　　渣通常是以粉狀存在的，主要用作分散劑。反應(yīng)槽的下流，合適的固液分離是普遍需要的。與其它工藝，例如，過濾、沉淀、離心分離、廢水中眾所周知的方法和水處理相比，浮選為分離范圍提供各種優(yōu)點。聯(lián)合吸附和后續(xù)的浮選被證明是一

70、種從溪流中除去重金屬的有效的方法。</p><p>　　本研究涉及到兩種渣對去除廢水溪流中Cu和Pb的吸附性能測試。用模擬Cu和Pb的軸承廢水在實驗室水平上實施了批量吸附/PH/動力學(xué)研究。離子吸附后的渣浮選給固液分離提供了有效的方法。也有報道成功應(yīng)用新技術(shù)處理金礦酸性排棄物。</p><p><b>　　2實驗研究</b></p><p>　

71、　兩種還原渣，即鐵渣和鋼渣，從南非Saldanha Steel以粉體的形式獲得，平均粒徑分別為24.5和24.1μm。粒徑分布是：鐵渣100μm100%，45μm45%，10μm23%，1μm1.2%：相比之下，鋼渣100μm100%，45μm90%，10μm23%，1μm1.2%。表1為渣的化學(xué)組成以氧化物質(zhì)量百分數(shù)算。DRON X-ray衍射儀獲得的XRD圖譜表明鈣玻璃是兩種渣的主要相。渣經(jīng)過三次蒸餾水清洗來去除粘附的雜質(zhì)，然后20

72、0℃干燥。干燥的渣儲存在干燥箱中用于實驗。</p><p>　　HCl和NaOH（分析純，Merck）用來調(diào)節(jié)溶液的PH。用蒸餾水溶解他們相應(yīng)的氯化物或者硝酸鹽（CuCl2?5H2O或者Pb(NO3)2,分析純，Merck）來制備Cu和Pb的常備溶液。在儲備溶液中的離子濃度為5000㎎/L。從南非莫尼卡獲得陽離子絮凝劑(聚胺型)。十二烷基硫酸鈉（分析純，Sigma）用作浮選過程中的收集器。</p>

73、<p>　　用11個帶有螺絲帽的塑料瓶在滾筒上18℃恒溫條件下實施定量吸附實驗。通過改變原始離子濃度，來實施等溫吸附實驗。接觸24小時后，通過0.45μm的薄膜過濾器（Millipore）過濾反應(yīng)混合物，然后分析濾液離子含量。渣的運載能力首先由不同離子含量決定，其次是吸附平衡。所有的動力學(xué)實驗都是在18℃恒溫水浴條件下，11個圓底透明塑料容器中實施的。容器中的溶液由玻璃攪拌器以600轉(zhuǎn)每分的速度攪動。所有的實驗僅重復(fù)已報道的平

74、均值。</p><p>　　溶液中的金屬離子含量由電感耦合等離子矩決定。由CRISON Micro pH 2000檢測溶液PH。在Laser Zeta儀表上實施電荷為零的動電學(xué)測試。樣品表面區(qū)域通過BET方法在Micropores (ModelASAP 2010, Micromeritics Instrument Corporation, Norcross,GA)上測試。常見無機陰離子通過傳感器由DIONEX A

75、I450離子色譜儀決定。濁度（NTU units）測量，采用一種高頻濁度計儀器。測試浮選柱由玻璃制成，周長和高度分別為35㎜和300㎜。水力旋流器通過燒結(jié)的具有小孔尺寸為4的玻璃進行分散。實驗在18℃恒溫條件下進行。泡沫自動溢出。在一個由磁力攪拌器攪拌的燒杯中，一定數(shù)量的表面活性劑添加到礦渣泥漿中和允許先于浮選2min的條件。在酸性礦物廢水處理中，渣漿中添加陽離子絮凝劑和先于收集器3min的條件。通過改變?nèi)芤簼岫?，浮選效率由通過渣的去除

76、來決定。</p><p><b>　　3結(jié)果與討論</b></p><p><b>　　3.1吸附平衡</b></p><p>　　作為一個單元運算的評估吸附進程的兩個重要的物理化學(xué)方面是吸附平衡和動力學(xué)。當(dāng)金屬體積濃度和表面濃度處于動態(tài)平衡時，估計達到吸附平衡。圖1所示，典型的兩種渣各自的Cu2+ 和 Pb2+的等溫吸附

77、曲線。渣漿PH保持在5.5和劑量保持在2g/L。</p><p>　　由圖1可知，鐵渣比鋼渣對重金屬有更高的吸附能力，就摩爾數(shù)而言Cu2+ 比 Pb2有更高的對渣的運載能力。</p><p>　　由礦渣吸收金屬的等溫吸附曲線能夠由朗繆爾等溫線表示。在礦渣上對Cu2+ 和 Pb2+的金屬吸附常數(shù)是由朗繆爾圖計算而來，見圖2。朗繆爾參數(shù)，Qmax (mg/g) and K (l/mg),分別與

78、飽和量和吸附結(jié)合常數(shù)有關(guān)。渣的吸附數(shù)據(jù)和朗繆爾等溫方程相吻合。關(guān)于兩種金屬的吸附數(shù)據(jù)非常完美地與朗繆爾等溫線相吻合，給礦渣提供了很高的相關(guān)系數(shù)。</p><p>　　見表2， Cu2+對鐵渣和鋼渣的吸附能力是88.50 mg/g和16.21 mg/g。Pb2+對鐵渣和鋼渣的吸附能力是95.24和 32.26 mg/g。鐵渣吸附常數(shù)比鋼渣吸附常數(shù)更高，很好的符合鐵渣比鋼渣有更高的吸附能力這一事實。據(jù)報道，一個大約4

79、0 mgPb2+/g吸附能力的顆粒狀的高爐渣的粒度小于0.25㎜。這次研究中這種渣和鐵渣有相似的組成，包含34% SiO2, 44% CaO, 6.4% Al2O3, 2.45% MgO and 0.5%</p><p>　　Fe2O3。然而，鐵渣比文獻中報道的渣更具吸附能力，可能是因為鐵渣的比表面積更大。</p><p>　　3.2 PH對吸附的影響</p><p&g

80、t;　　原始溶液濃度200㎎/L，渣的量是2g/L，接觸時間為24h。圖2表明PH依賴于各自的Cu2+ 和 Pb2+的運載能力。很明顯，PH低時，金屬吸收也相當(dāng)?shù)?。然而，隨著溶液PH的上升，記錄的兩種渣的吸附都有顯著的增強，鐵渣和鋼渣的金屬吸附最佳PH是3.5–8.5 和5.2–8.5 。考慮渣表面電荷很大程度上依賴于二氧化硅和氧化鋁的電位這種行為，就可以做出解釋。吸附渣的復(fù)合電位，由電泳測量，發(fā)現(xiàn)鋼渣的為3.2，鐵渣的為4.8。對于鐵

81、渣，在PH<4.8的范圍內(nèi)，粒子表面有一個正電荷密度。這種條件下，因為靜電學(xué)排斥，重金屬的吸收將會相當(dāng)?shù)?。隨著PH的上升，也就是說，超過鐵渣電位為零的點（>4.8）,鐵渣表面負電荷密度增加，因此導(dǎo)致在鐵渣表面重金屬吸收顯著增強。而且，PH<3.2鋼渣運輸相當(dāng)?shù)停琍H>3.2鋼渣運輸較高。此外，析出的重金屬陽離子在高溶液PH可能有助于在渣上的金屬運輸。</p><p><b>　　

82、3.3 吸附動力學(xué)</b></p><p>　　礦渣量為2g/L、渣漿PH維持在5.5。Cu2+ 和 Pb2+溶液各自原始濃度大約在200㎎/L。吸附動力學(xué)結(jié)果見圖3.鋼渣吸附平衡在1h內(nèi)達到，然而鐵渣的吸附平衡直到24小時才能達到。</p><p><b>　　3.4吸附機理</b></p><p>　　鑒于渣的特性，礦渣玻璃和溶

83、液的互相交換作用能夠表示為：</p><p>　　SiOCa + 2HOH → =SiOH2 + Ca2+ + 2OH? (1)</p><p>　　預(yù)期在酸性環(huán)境中上面的反應(yīng)因為高氫離子濃度，能夠轉(zhuǎn)移到左邊。根據(jù)上面的路線，堿性渣具有中和作用。</p><p>　　不同數(shù)量的渣加入到11個蒸餾水中，然后600r/min攪拌24h。表3顯示了溶液PH改變和濃度Ca+

84、改變。明顯地，當(dāng)溶液PH增加，Ca+和H+交換，而且從渣中釋放出來到溶液中。這就證實了化學(xué)反應(yīng)方程式1，在渣和溶液接觸中發(fā)生。接觸24h后，在等量情況下，鋼渣中Ca+濃度比鐵渣中的更低?？梢酝茢喑觯F渣具有比鋼渣更高的離子交換能力，這和等溫吸附曲線一致。在二價重金屬存在的情況下，上述方程可以描述如下：</p><p>　　=SiOCa +M2+ → =SiOM + Ca2+ （2）</p>&

85、lt;p>　　如表1所列，礦渣具有較高的比表面積和孔隙度，同時也提供了重金屬吸附的基礎(chǔ)。礦渣微觀結(jié)構(gòu)圖4-6，兩種渣都揭示了多孔結(jié)構(gòu)和鐵渣具有比鋼渣更小的孔。吸附后，一些松散的絮狀體出現(xiàn)在渣的表面，如圖6.這些絮狀物可能是銅的鋁硅酸鹽或者銅的復(fù)雜配合物沉積于渣的孔上。</p><p><b>　　3.5渣浮選</b></p><p>　　24h渣吸附后，在渣漿上

86、出現(xiàn)一些松散且相當(dāng)小的孔。通過簡單沉淀的固液分離在短時間內(nèi)不會完成（10h以上），而且需要進一步的過濾。通過渣，離子交換和金屬氧化物沉淀都有助于從廢水中去除重金屬離子。吸附后，多種金屬陽離子配合物增加了溶液濁度。除此之外，一些精煉渣粒子的存在能夠引起溶液濁度的增加，然而，這還可以避免使用粗渣粒子。為了去除懸浮固體，分散空氣柱浮選應(yīng)用于后續(xù)的從溶液中分離已負載的渣。原始Cu和Pb濃度為200mg/L，渣的量為2g/L。溶液PH保持在5.5

87、。接觸24h后，在200ml/min的空氣流量中對渣漿進行浮選。根據(jù)前面的圖像分析，空氣柱中，平均氣泡尺寸接近0.62㎜。浮選結(jié)果如表4。</p><p>　　由表4可知，15㎎/l SDS量，對于渣漿體系來說，浮選后的溶液濁度已經(jīng)低于1.00NTU（自來水濁度）。浮選后Cu2+ 和 Pb2+濃度進一步減小，且隨著SDS量的增加，這種影響變得更加顯著。與沒有渣的空白實驗比較，這主要歸因于重金屬離子和SDS的絡(luò)合作

88、用。吸附渣浮選，使用少量渣作為載體，發(fā)現(xiàn)對處理含有重金屬如Cu和Pb的溶液非常有效。</p><p>　　沉淀也是一個有效的過程，在典型絮凝劑的協(xié)助下留下一個清澈的浮層。沉淀簡單且相對成本較低。然而，與沉淀相比，浮選較快且在回收精細粒子方面效率更高。目前研究中，精煉渣粉體用于吸附劑。因此，浮選被選作后續(xù)的固液分離方法。如果粗渣粒子用于廢水處理，沉淀方法是可選的。用粗渣粒子吸附填料柱的污染物是十分有效的，盡管吸附區(qū)

89、域小。這可以避免后續(xù)浮選或者固液分離沉淀。</p><p>　　3.6吸附渣浮選法對礦山酸性廢水的處理</p><p>　　用于實驗的酸性礦山廢水是從南非金礦排泄物中取的樣品。酸性礦山廢水的PH較低，2.03，硫酸鹽濃度很高，如表5.</p><p>　　酸性礦水由石灰原位中和，并伴隨離子、鋁和金屬氫氧化物的沉淀。處理石灰需要短暫反應(yīng)期，由于它的高濁度（0.15%）

90、。然而，由于發(fā)現(xiàn)在污水中不同金屬最小溶解度通常出現(xiàn)在不同的PH和氫氧化物沉淀實質(zhì)上是兩性的，混合金屬的最大去除效率在單一的沉淀PH水平大約7時，不能夠獲得。渣結(jié)合離子交換和具有酸中和的吸附性能，而且可能提供選擇性的去除酸性排泄物的金屬離子。</p><p>　　定量吸附試驗用11個酸性礦山排泄物實施、接觸時間24h。從酸性礦物中去除陽離子的渣的量的影響見表5。渣漿PH增減到中性PH，且渣的量增加到30g/L。在3

91、0g/L的鐵渣中的大多數(shù)重金屬，如，Cu，Pb,Cr被去除了，殘留濃度能夠達到排泄標(biāo)準(zhǔn)（<0.05㎎/L）。在重金屬離子去除方面鐵渣的性能優(yōu)于鋼渣。貴金屬陽離子如Au,Ag,Pt,Pd,Rh和Ru僅僅去除了少部分。普遍的，在酸性環(huán)境中貴金屬離子以氯的陰離子配合物的方式存在。，如AuCl4?, PdCl62?和PtCl42-</p><p>　　。在PH>ZPC時陰離子配合物不能夠物理上地吸收在帶負電荷

92、的渣的表面，而且不能和渣進行Ca2+交換。由于PH的上升，一些貴金屬離子的去除可以歸因于和Fe(OH)3形成的共絮狀物。由于同樣地陰離子貴金屬配合物的原因，除了PO43?的其它陰離子外，去除也是比較困難。PO43?以低溶解性鈣磷酸鹽沉淀物方式除去。Ca和Mg能夠和F-形成一些難容沉淀物，PO43-和SO42-沉積在渣的表面。</p><p>　　30g/L的渣接觸24h后，渣漿中出現(xiàn)了大量的松散且細小的孔洞。通過

93、SDS直接浮選對固液分離不是很有效。陽離子絮凝劑聚胺以100㎎/L的優(yōu)化濃度加入渣漿中，先于SDS加入前3min加入。添加聚胺后，優(yōu)良的絮凝體和渣形成大而且結(jié)構(gòu)較好的聚合物，這是有準(zhǔn)備去除通過后續(xù)的渣浮選工藝。浮選結(jié)果如表6.明顯地，對于排放到公共溪流中，所有堿性金屬的殘余濃度和在溶液中的貴金屬離子濃度足夠低，在渣中和、吸附、沉淀、浮選過程后溶液PH呈中性和溶液濁度低于1.0NTU（自來水）。由于靜電作用，陽離子絮凝劑不僅絮凝柔軟的精細

94、粒子形成較大的聚合物，而且作為載體去除陰離子，如陰離子貴金屬配合物。絮凝后，聚合物更耐剪切力，且更加適合后續(xù)的浮選。對于礦物排泄物處理，就最后離子殘余濃度而言，鐵渣的性能優(yōu)于鋼渣。</p><p>　　最后，應(yīng)當(dāng)強調(diào)通過采用非傳統(tǒng)的冶金副產(chǎn)物吸附劑和高產(chǎn)能浮選技術(shù)來</p><p>　　相信吸附渣浮選概念。研究的案例僅聚焦在酸性礦山排泄物的處理，并且預(yù)期這種工藝可能成為未來無機和有機污染物

95、去除領(lǐng)域的可選技術(shù)。</p><p><b>　　4結(jié)論</b></p><p>　　鐵渣和鋼渣適合用做從水中去除重金屬的吸附劑。渣利用酸中和能力結(jié)合離子交換和吸附性能。由于鐵渣的高比表面積，高孔隙度，高離子交換能力，鐵渣吸收重金屬比鋼渣有較高的吸附能力。渣吸附后的浮選能夠有效的分離渣，產(chǎn)生非常低的溶液濁度。當(dāng)比較沉淀就動力學(xué)過程和上浮物質(zhì)量的時候察到了優(yōu)點。<