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量是关键因素，其含量越高越利于磷的吸附。
（4）Dubinin-Radushkevich拟合的吸附自由能范围为5.85~7.29kJ·mol-1，说明各类生物炭对磷的吸附作用以物理吸附为主。
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Circumstantiae ,2023,43（5）：261-270.[37]朱艳,肖清波,奚永兰,等.改性生物炭制备条件对磷吸附性能的影响[J].生态环境学报,2020,29（9）：1897-1903.ZHU Y,XIAOQ B,XI Y L,et al.Effect of preparation conditions on the phosphorus adsorption capacities of modified biochar[J].Ecology and Environmen⁃tal Sciences ,2020,29（9）：1897-1903.耿仕呈，张伟涛，顾文源，等.漏缝地板发酵床对育肥羊养殖气体排放的影响及微生物学机理[J].农业环境科学学报,2024,43（4）：926-936.GENG S C,ZHANG W T,GU W Y,et al.Effects of slatted floor fermentation bed on gas emissions and microbiological mechanism during fattening lamb breeding[J].Journal of Agro-Environment Science ,2024,43（4）：926-936.漏缝地板发酵床对育肥羊养殖气体排放的影响及微生物学机理耿仕呈1，张伟涛2，顾文源3，姚惠娇2，高志岭1，何旭4，刘春敬1*，范玉婧1，代宇菲1（1.河北农业大学资源与环境科学学院/河北省农田生态环境重点实验室，河北保定071000；2.河北省畜牧总站，石家庄050000；3.河北省动物疫病预防控制中心，石家庄050000；4.河北省畜牧良种工作总站，石家庄050000）Effects of slatted floor fermentation bed on gas emissions and microbiological mechanism during fatteninglamb breedingGENG Shicheng 1,ZHANG Weitao 2,GU Wenyuan 3,YAO Huijiao 2,GAO Zhiling 1,HE Xu 4,LIU Chunjing 1*,FAN Yujing 1,DAI Yufei 1（1.College of Resources and Environmental Sciences /Key Laboratory of Farmland Ecological Environment of Hebei Province,HebeiAgricultural University,Baoding 071000,China ；2.Hebei Animal Husbandry Station,Shijiazhuang 050000,China ；3.Hebei Animal Disease Prevention and Control Center,Shijiazhuang 050000,China;4.Hebei Provincial Livestock Breeding Work Station,Shijiazhuang 050000,China ）Abstract ：In this study,we aimed at investigating how the slatted fermentation bed impacts ammonia （NH 3）and greenhouse gas emissions,we thus constructed two animal pen types ：ground floor and slatted floor and fermentation bed,then investigated the NH 3,N 2O,CO 2,andCH 4emission characteristics as well as the microbiological mechanisms during fattening sheep breeding using pared to收稿日期：2023-12-05录用日期：2024-01-12作者简介：耿仕呈（1998—），女，河北石家庄人，硕士研究生，研究方向为农业环境保护。
第36卷 第3期 无 机 材 料 学 报Vol. 36No. 32021年3月Journal of Inorganic Materials Mar., 2021收稿日期: 2020-06-20; 收到修改稿日期: 2020-08-27; 网络出版日期: 2020-09-09基金项目: 国家自然科学基金(21776191, 21706181); 上饶师范学院校级自选课题(202028); 山西省重点研发计划(201803D421094); 国家重点研发计划政府间国际科技创新合作重点专项项目(2017YFE0129200)National Natural Science Foundation of China (21776191, 21706181); Project of Shangrao Normal University (202028); Key R&D project of Shanxi Province (201803D421094); National Key R&D Program of China (2017YFE0129200)作者简介: 杨言言(1983–), 女, 博士, 讲师.E-mail:******************YANGYanyan(1983–),female,PhD,lecturer.E-mail:******************通信作者: 郝晓刚, 教授.E-mail:**************.cn;*********************文章编号: 1000-324X(2021)03-0292-07 DOI: 10.15541/jim20200340电活性镍钴双金属氧化物高选择性去除/回收水中磷酸盐离子杨言言1,2, 李永国3, 祝小雯1, 杜 晓2, 马旭莉2, 郝晓刚2(1. 上饶师范学院 化学与环境科学学院, 上饶 334001; 2. 太原理工大学 化学化工学院, 太原 030024; 3. 中国辐射防护研究院, 太原 030006)摘 要: 磷是植物体生长的重要营养素, 也是引发水体富营养化的重要因素, 因此废水中磷酸盐的去除与回收均至关重要。
IFA化肥术语（中英对照）
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磷酸铵镁除磷脱氮技术摘要：目前，生物脱氮除磷常采用A2O工艺，但其流程长且成本高，对进水氨氮浓度变化的适应性及抗负荷冲击的能力较差。
本文介绍一种化学沉淀法，即MAP（Magnesium Ammonium Phosphate）脱氮除磷法。
关键词：磷酸铵镁除磷脱氮MAP 化学沉淀法1 MAP除磷脱氮的基本原理向含NH4+和PO43-的废水中添加镁盐，发生的主要化学反应如下：Mg2++HPO42-+NH4++6H2O→MgNH4PO4·6H2O↓+H+（1）Mg2++PO43-+NH4++6H2O→MgNH4PO4·6H2O↓ （2）Mg2++H2PO4-+NH4++6H2O→MgNH4PO4·6H2O↓+2H+（3）再经重力沉淀或过滤，就得到MAP。
其化学分子式是MgNH4PO4·6H2O，俗称鸟粪石；它的溶度积为2.5×10-13。
因为它的养分比其它可溶肥的释放速率慢，可以作缓释肥（SRFs）；肥效利用率高，施肥次数少；同时不会出现化肥灼烧的情况。
2 MAP除磷脱氮的影响因素和沉淀物组成分析2.1 Mg2+，NH4+，PO43-三者在反应过程中的比例在处理氨氮废水方面，将H3PO4加入到含有MgO的固体粉末中制成一种乳状液，对2.47×10-3mol/L氨氮废水进行处理，得出H3PO4与MgO的物质的量之比大于1.5时，氨氮去除率最高（90％以上），当进水氨氮质量浓度为42mg/L，在最佳条件下，氨氮质量浓度可降到0.5mg/L以下[1]。
赵庆良[2]等人对5618mg/L氨氮的垃圾渗滤液进行处理，按n(Mg2+):(NH4+):n（PO43-)=1:1:1投加氯化镁和磷酸氢二钠，废水中氨氮质量浓度降为172mg/L，过量投加10％的镁盐或磷酸盐，氨氮质量浓度可分别降为112mg/L和158mg/L，继续提高镁盐或磷酸盐的量，废水中剩余氨氮质量浓度处在100mg/L左右，很难进一步降低。
Journal of Hazardous Materials 182 (2010) 156–161Contents lists available at ScienceDirectJournal of HazardousMaterialsj o u r n a l h o m e p a g e :w w w.e l s e v i e r.c o m /l o c a t e /j h a z m atAs(III)removal using an iron-impregnated chitosan sorbentDaniel Dianchen Gang a ,∗,Baolin Deng b ,LianShin Lin caDepartment of Civil Engineering,University of Louisiana at Lafayette,Lafayette,LA 70504,USAbDepartment of Civil and Environmental Engineering,University of Missouri,Columbia,MO 65211,USA cDepartment of Civil and Environmental Engineering,West Virginia University,Morgantown,WV 26506,USAa r t i c l e i n f o Article history:Received 18December 2009Received in revised form 28May 2010Accepted 1June 2010Available online 9 June 2010Keywords:Trivalent arsenic Iron-chitosan AdsorptionAs(III)adsorption kinetics Adsorption isotherma b s t r a c tAn iron-impregnated chitosan granular adsorbent was newly developed to evaluate its ability to remove arsenic from water.Since most existing arsenic removal technologies are effective in removing As(V)(arsenate),this study focused on As(III).The adsorption behavior of As(III)onto the iron-impregnated chi-tosan absorbent was examined by conducting batch and column studies.Maximum adsorption capacity reached 6.48mg g −1at pH =8with initial As(III)concentration of 1007g L −1.The adsorption isotherm data fit well with the Freundlich model.Seven hundred and sixty eight (768)empty bed volumes (EBV)of 308g L −1of As(III)solution were treated in column experiments.These are higher than the empty bed volumes (EBV)treated using iron-chitosan composites as reported by previous researchers.The investi-gation has indicated that the iron-impregnated chitosan is a very promising material for As(III)removal from water.© 2010 Elsevier B.V. All rights reserved.1.IntroductionArsenic,resulting from industrial and mine waste discharges or from natural erosion of arsenic containing rocks,is found in many surface and ground waters [1].Common chemical forms of arsenic in the environment include arsenate (As(V)),arsenite (As(III)),dimethylarsinic acid (DMA),and monomethylarsenic acid (MMA).Inorganic forms of arsenic (As(V)and As(III))are more toxic than the organic forms [2].Arsenite can be predominant in ground-water with low oxygen levels and is generally more difficult to be removed than arsenate [3].Due to the negative impacts of arsenic on human health that range from acute lethality to chronic and car-cinogenic effects,the U.S.Environmental Protection Agency revised the maximum contaminant level (MCL)of arsenic in drinking water from 50to 10g L −1[4].This new regulation has posed a chal-lenge for the research of new technologies capable of selectively removing low levels of arsenic.Existing technologies that are being used for arsenic removal include precipitation [5],membrane separation,ion exchange,and adsorption [6–9].While these approaches can remove arsenic to below 10g L −1under optimal conditions,most of the systems are expensive,not suitable for small communities with limited resources.Of these methods,much work has been done on arsenic removal through adsorption because it is one of the most effec-∗Corresponding author.Tel.:+13374825184;fax:+13374826688.E-mail addresses:ddgang@ ,digang@ ,Gang@ (D.D.Gang).tive and inexpensive methods for arsenic treatment [7].Therefore,development of highly effective adsorbents is a key for adsorption-based technologies.Several iron(III)oxides,such as amorphous hydrous ferric oxide [5]and crystalline hydrous ferric oxide [10]are well known for their ability to remove both As(V)and As(III)from aqueous solutions.In general,arsenate is more readily removed by ferric (hydr)oxides than arsenite [11].Reported mechanisms for arsenic removal include adsorption onto the hydroxide surfaces,entrapment of adsorbed arsenic in the flocculants,and formation of complexes and ferric arsenate (FeAsO 4)[12].The presence of other anions such as sulfate,chloride,and in particular,silicates,phosphate,and natural organic matters,can significantly affect arsenic adsorption [13–15].The use of iron (hydr)oxides in fine powdered or amor-phous forms was found to be effective for arsenic removal,but the process requires follow-up solid/water separation.For packed-bed adsorption systems,high-efficient granular forms of adsorbent are essential.Recently,several iron based granular materials and processes have been developed for arsenic removal.Dong et al.[16]devel-oped iron coated pottery granules (ICPG)for both As(III)and As(V)removal from drinking water.The column tests showed that ICPG consistently removed total arsenic from test water to below 5g L −1level.In another study,Gu et al.[17]used iron-containing granular activated carbon for arsenic adsorption.This iron-containing granular activated carbon was shown to remove arsenic most efficiently when the iron content was approximately 6%.Viraraghavan et al.[18]reported a green sand filtration process and found a strong correlation between influent Fe(II)concen-0304-3894/$–see front matter © 2010 Elsevier B.V. All rights reserved.doi:10.1016/j.jhazmat.2010.06.008D.D.Gang et al./Journal of Hazardous Materials182 (2010) 156–161157tration and arsenic removal percentage.The removal percentage increased from41%to above80%as the ratio of Fe/As was increased from0to20.Granular ferric hydroxide(GFH),another iron based granular material,showed a high treatment capacity for arsenic removal in a column setting before the breakthrough concentration reached10g L−1[19].It was found that complexes were formed upon the adsorption of arsenate on GFH[20].Selvin et al.[21]con-ducted laboratory-scale tests over50different media for arsenic removal and found GFH with a particle size of0.8–2.0mm was the most effective one among the tested media.However,some disad-vantages with GFH exist,including quick head loss buildup within 2days because of thefine particle size,and significant reduction (50%)in adsorption capacity with larger sized media(1.0–2.0mm).Chitin and its deacetylated product,chitosan,are the world’s second most abundant natural polymers after cellulose.These polymers contain primary amino groups,which are useful for chemical modifications and can be used as potential separa-tors in water treatment and other industrial applications.Many researchers focused on chitosan as an adsorbent because of its non-toxicity,chelating ability with metals,and biodegradability[22]. Several studies have demonstrated that chitosan and its deriva-tives could be used to remove arsenic from aqueous solutions [23,24].Based on the fact that both iron(III)oxides and chitosan exhib-ited high affinity for arsenic,this study focused on examining the effectiveness of an iron-impregnated chitosan granular adsorbent for arsenic removal.Most arsenic removal technologies are more effective for removing arsenate than for arsenite[12].We found in this study that the iron-impregnated chitosan was effective for arsenite removal from experiments in both batch and column set-tings.2.Experimental2.1.Preparation of iron-chitosan beadsThe experimental procedure for the preparation of iron-chitosan beads was described in detail by Vasireddy[25].To summarize, approximately10g of medium molecular weight chitosan(Aldrich Chemical Corporation,Wisconsin,USA)was added to0.5L of0.01N Fe(NO3)3·9H2O solution under continuous stirring at60◦C for2h to form a viscous gel.The beads were formed by drop-wise addition of chitosan gel into a0.5M NaOH precipitation bath under room temperature.Maintaining this concentration of NaOH was critical for forming spherically shaped beads[25].The beads were then separated from the0.5M NaOH solution and washed several times with deionized water to a neutral pH.The wet beads were then dried in an oven under vacuum and in air.Thefinal iron content of the chitosan bead was about8.4%.2.2.Arsenic measurementAn atomic absorption spectrometer(AAS)(Thermo Electron Corporation)equipped with an arsenic hollow cathode lamp was employed to measure arsenic concentration.An automatic inter-mittent hydride generation device was used to convert arsenic in water samples to arsenic hydride.The hydrides were then purged continuously by argon gas into the atomizer of an atomic absorption spectrometer for concentration measurements.As(III)stock solution(1000mg L−1)was prepared by dissolving 1.32g of As2O3(obtained from J.T.Baker)in distilled water con-taining4g NaOH,which was then neutralized to pH about7with 1%HCl and diluted to1L with distilled water.All the working solu-tions were prepared with standard stock solution.To50mL of each sample solution(i.e.,reagent blank,standard solutions,and water samples),5mL1%HCl and5mL of100g L−1NaI solution were used to convert arsenic in water samples to arsenic hydride.2.3.Arsenic adsorption experimentsEach arsenic solution(100mL)of desired concentration was mixed with the iron-chitosan beads in a250mL conicalflask.The solution pH was adjusted with0.1M HCl or0.1M NaOH to obtain the desired pHs.A pH buffer was not used to avoid potential com-petition of buffer with As(III)sorption.One sample of the same concentration solution without adsorbent(blank),used to estab-lish the initial concentration of the samples,was also treated under same conditions as the samples containing the adsorbent.The solu-tions were placed in a shaker for afixed amount time,followed by filtration to remove the adsorbent.Thefiltrate was then analyzed for thefinal concentration of arsenic using the atomic absorption spectrometer.The solid phase concentration was calculated using the following formula:q=(C i−C f)VM(1) where,q(g g−1)is the solid phase concentration,C i(g L−1)is the initial concentration of arsenic in solution,C f(g L−1)is thefinal concentration of arsenic in treated solution;V(L)is the volume of the solution,and M(g)is the weight of the iron-chitosan adsorbent.2.4.Kinetic experimentsAdsorption kinetics was examined with various initial concen-trations at25◦C.The pH of the solutions was chosen at8.0for optimal adsorption.The adsorbent loading for three different ini-tial concentrations of306,584,and994g L−1was all0.2g L−1.A predetermined quantity of iron-chitosan adsorbent(20mg)was placed in separate conicalflasks with pH-adjusted As(III)solution. The conicalflasks were covered with parafilm and placed in a shaker (150rpm),and sub-samples of the solutions were then removed periodically andfiltered prior to arsenic analysis.To determine the reaction rate constants of arsenic adsorption onto iron-chitosan,both the pseudo-first-order and pseudo-second-order models were used.Kinetics of the pseudo-first-order model can be expressed as[26]:ln(q e−q t)=ln q e−k1t(2) where,k1(min−1)is the rate constant of pseudo-first-order adsorp-tion,q t(mg g−1)is the amount of As(III)adsorbed at time t(min), and q e(mg g−1)is the amount of adsorption at equilibrium.The model parameters k1and q e can be estimated from the slope and intercept of the plot of ln(q e−q t)vs t.The pseudo-second-order model can be expressed as follow[27]:tq t=tq e+1k2q2e(3)where,k2(g mg−1min−1)is the pseudo-second-order reaction rate. Parameters k2and q e can be estimated from the intercept and slope of the plot of(t/q t)vs t.2.5.Isotherm modelsAdsorption isotherms such as the Freundlich or Langmuir mod-els are commonly utilized to describe adsorption equilibrium.The Freundlich isotherm model is represented mathematically as:q e=k f C1/ne(4) where,q e(mg g−1)is the amount of As(III)adsorbed,C e(g L−1) is the concentration of arsenite in solution(g L−1),k f and1/n158 D.D.Gang et al./Journal of Hazardous Materials182 (2010) 156–161Fig.1.Scanning electron micrograph(SEM)of iron-chitosan bead.are parameters of the Freundlich isotherm,denoting a distribu-tion coefficient(L g−1)and intensity of adsorption,respectively.The Langmuir equation is another widely used equilibrium adsorption model.It has the advantage of providing a maximum adsorption capacity q max(mg g−1)that can be correlated to adsorption proper-ties.The Langmuir model can be represented as:q e=q maxK L C e1+K L C e(5)where,q max(mg g−1)and K L(L mg−1)are Langmuir constants representing maximum adsorption capacity and binding energy, respectively.2.6.Column studyColumn study was conducted to investigate the use of iron-chitosan as a low-cost treatment technology for arsenite removal. Experiments were conducted with a12-mm-ID glass column packed with1.5g iron-chitosan as afixed bed.The influent solu-tion had an inlet As(III)concentration of308g L−1at pH8,and was passed the column at aflow rate of25mL h−1.Effluent solu-tion samples were collected and analyzed for arsenic concentration during the column test.3.Results and discussion3.1.Structure characterization of iron-chitosan beadsThe prepared iron-chitosan beads were examined by scanning electron microscope(SEM)(AMRAY1600)for the surface morphol-ogy.A working distance of5–10mm,spot size of2–3,secondary electron(SE)mode,and accelerating voltage of20keV were used to view the samples.It can be seen from Fig.1that the beads are porous in structure.X-ray Photoelectron Spectroscopy(XPS),a sur-face sensitive analytic tool to determine the surface composition and electronic state of a sample,was used in this study.In XPS analysis,a survey scan was used to determine the elements exist-ing on the surface.The high resolution utility scans were then used to measure the atomic concentrations of Fe,C,N and O in the sam-ple.Fig.2shows the peak positions of carbon,nitrogen,oxygen,and iron obtained by the XPS for iron-chitosan beads.In Fig.2,the car-bon1s peak was observed at283.0eV with a FWHM(full width at maximum height)of2.015.The Fe peak was observed at730.0eV. The N-1s peak for iron-chitosan bead was found at398.0eV(FWHM 2.00eV),which can be attributed to the amino groups inchitosan.Fig.2.XPS spectrum of iron-chitosan bead.3.2.Effect of pHThe effect of pH on arsenite removal with the iron-chitosan adsorbent was examined using100mL As(III)solution with an initial concentration of314g L−1and a solid loading rate of 0.15g L−1.The solution pH was adjusted with0.1M HCl or0.1M NaOH to obtain pHs ranging from4to12.Lower pHs were avoided because the acid environments could lead to partial dissolution of the chitosan polymer and make the beads unstable[25,28]. The solutions were placed in a shaker(150rpm)for20h at room temperature(25◦C),followed byfiltration to remove the adsor-bent.The amounts of As(III)adsorbed,calculated using Eq.(1),are present in Fig.3.Under the experimental conditions,approximately 2.0mg g−1of As(III)was adsorbed and that amount did not change significantly in the pH range4–9.However,when pH was higher than9.2,arsenite removal decreased dramatically with increasing pH.The results can be explained using arsenic chemical speciation in different pH ranges[29].Arsenite remains mostly as a neutral molecule for pH<9.2,and negatively charged at pH>9.2.So at pH>9.2,arsenite sorption is less because of the unfavorable electro-static interaction with negatively charged surfaces.This adsorptive behavior is common for arsenite with other adsorbents[17,30].Gu et al.[17]reported that pH had no obvious effect on As(III)removal in the range of4.4–9.0,with removal efficiency above95%.Another study indicated that the uptake of As(III)by fresh andimmobi-Fig.3.Arsenite removal of the iron-chitosan adsorbent(0.15g L−1)as a function of pH for initial arsenite concentration of314g L−1at T=25◦C.D.D.Gang et al./Journal of Hazardous Materials182 (2010) 156–161159Fig.4.Adsorption kinetics for different initial arsenite concentrations with iron-chitosan adsorbent loading of0.2g L−1at pH=8and T=25◦C.lized biomass was not greatly affected by solution pH with optimal biosorption occurring at around pH6–8[30].Raven et al.[11] reported that a maximum adsorption of arsenite on ferrihydrite was observed at approximately pH9.3.3.Kinetics of adsorptionFig.4illustrates the adsorption kinetics for three different ini-tial arsenite concentrations.More than60%of the arsenite was adsorbed by iron-chitosan within thefirst30min,then adsorption leveled off after2h.Given the initial concentrations and adsorbent loading,equilibrium was reached after about2h.The adsorption capacity increased from1.51to4.60mg g−1as the initial arsen-ite concentration was increased from306to994g L−1.The rapid adsorption in the beginning can be attributed to the greater con-centration gradient and more available sites for adsorption.This is a common behavior with adsorption processes and has been reported in other studies[31].The sorption rate of As(III)on nat-urally available red soil was initially rapid in thefirst2h and slowed down thereafter[32].Elkhatib et al.[33]reported that the initial adsorption was rapid,with more than50%of As(III) adsorbed during thefirst0.5h in an arsenite adsorption study. Fuller et al.[34]reported that As(V)adsorption onto synthesized ferrihydrite had a rapid initial phase(<5min)and adsorption con-tinued for182h.Raven et al.[11]studied the kinetics of As(V) and As(III)adsorption on ferrihydrite and found that most of the adsorption occurred within thefirst2h.It has been reported that arsenite forms both inner-and outer-sphere surface complexes on amorphous Fe oxide[35].Another possible adsorption mech-anism is hydrogen bond formation between As(III)and chitosan bead[24].Figs.5and6illustrate modelfits of the kinetic data for the pseudo-first-order and pseudo-second-order kinetic models. In general,the pseudo-second-order characterized the kinetic data better than the pseudo-first-order model.Table1summa-Fig.5.Adsorption kinetics of the iron-chitosan adsorbent(0.2g L−1)for three initial arsenite concentrations at pH=8and T=25◦C,and corresponding pseudo-first-ordermodels.Fig.6.Adsorption kinetics of the iron-chitosan adsorbent(0.2g L−1)for three initial arsenite concentrations at pH=8and T=25◦C,and corresponding pseudo-second-order models.rizes adsorption capacities determined from the modelfits.It is noted that the second order rate constant(k2)decreased from 3.19×10−2to 1.15×10−2g mg−1min−1as the initial concen-tration increased from306to994g L−1.The initial rate(k2q2e) increased from8.48×10−2to27.97×10−2with increasing initial As(III)concentration.Because as initial concentration increased,the concentration difference between the adsorbent surface and bulk solution increased.Jimenez-Cedillo et al.[36]investigated arsenic adsorp-tion kinetics on iron,manganese and iron-manganese-modified clinoptilolite-rich tuffs and concluded that the adsorption pro-cesses could be described by the pseudo-second-order model.Table1Adsorption capacities and parameter values of kinetic models for three initial arsenite concentrations and iron-chitosan loading of0.2g L−1at pH=8.Initial conc.(g L−1)Pseudo-first order Pseudo-second orderk1×102(min−1)R2q e,exp(mg g−1)q e,col(mg g−1)k1×102(g mg−1min−1)R2q e,exp(mg g−1)q e,col(mg g−1)k2q2e×102306 2.630.98 1.51 1.24 3.190.99 1.51 1.638.48584 2.380.96 2.90 2.30 1.310.99 2.90 3.1913.28994 2.370.93 4.60 3.26 1.150.99 4.60 4.9327.97160 D.D.Gang et al./Journal of Hazardous Materials182 (2010) 156–161Fig.7.Adsorption isotherms of the iron-chitosan adsorbent (0.2g L −1)for three initial arsenite concentrations at pH =8,and corresponding isotherm models.Thirunavukkarasu et al.[37]examined As(III)adsorption kinet-ics with granular ferric hydroxide (GFH)and found that most of As(III)adsorption onto GFH occurred at pH 7.6,with 68%of As(III)removed within 1h and 97%removed at the equilibrium time of 6h.Kinetic data fitted the pseudo-second-order kinetic model well with a kinetic rate constant of 0.003g GFH h −1g −1As,which is equivalent to 5.0×10−2g mg −1min −1[37].In our study,the kinetic rate constants were from 3.19×10−2to 1.15×10−2g mg −1min −1,which were smaller than using GFH.This could be attributed to the differences in adsorbent parti-cle size and initial arsenic concentrations between these two studies.3.4.Adsorption isothermsFig.7presents the adsorption isotherm data and two isotherm models at pH 8.The maximum adsorption capacity was found to increase from 1.97to 6.48mg g −1as the initial concentration of As(III)increased from 295to 1007g L −1.Maximum adsorp-tion capacity reached 6.48mg g −1with initial As(III)concentration of 1007g L −1.Chen and Chung [24]reported that the adsorp-tion capacity of As(III)was 1.83mg As g −1for pure chitosan bead.This study confirmed that impregnating iron into chitosan could significantly increase the As(III)adsorption capacity of the chi-tosan bead.In another study,Driehaus et al.[19]reported that the adsorption capacity could reach 8.5mg As g −1of granular fer-ric hydroxide (GFH).Model parameters and regression coefficients are listed in Table 2.The Freundlich model agreed better with the experimental data compared to the Langmuir model.The adsorp-tion intensity (1/n )and the distribution coefficient (k f )increased as the initial arsenite concentration increased.This indicated the dependence of adsorption on initial concentration.Low 1/n values (<1)of the Freundlich isotherm suggested that any large change in the equilibrium concentration of arsenic would not result in a significant change in the amount of arsenic adsorbed.Selim and Zhang [38]reported that adsorption isotherms of three differ-ent soils for As(V)were better fit to the Freundlich modelandFig.8.Breakthrough curve for an inlet arsenite concentration of 308g L −1at pH =8for a column reactor packed with the iron-chitosan adsorbent.adsorption intensity values ranged from 0.270to 0.340.Salim and Munekage [39]found that adsorptions of As(III)onto silica ceramic were well fit by the Freundlich isotherm.Similarly low 1/n values for As(V)adsorption have been reported by others [40].3.5.Column studyFig.8shows a breakthrough curve for an inlet arsenite con-centration of 308g L −1at pH 8.The break point was observed after 768empty bed volumes (EBV)and adsorbent was exhausted at 1400bed volumes.In comparison,Boddu et al.[23]reported that the break through point was about 40and 120EBV for As(III)and As(V),respectively using chitosan-coated biosorbent.Gupta et al.[41]conducted column tests using iron-chitosan compos-ites for removal of As(III)and As(V)from arsenic contaminated real life groundwater.Their result showed that the iron-chitosan flakes (ICF)could treat 147EBV of As(III)and 112EBV of As(V)spiked groundwater with an As(III)or As(V)concentration of 0.5mg L −1.Given the difference of the initial concentrations between the two studies,the numbers of EBV were lower than what we found in this study.This can be partially attributed to the difference of the water constituents in the real grounder water used in the previous study [41].Gu et al.[17]examined the arsenic breakthrough behaviors for an As-GAC sample prepared from Dacro 20×40LI with an inlet concentration of 56.1g L −1As(III).Their results demonstrated that the adsorbent could effectively remove arsenic from ground-water in a column setting.Dong et al.[16]also reported that average removal efficiencies for total arsenic,As(III),and As(V)for a 2-week test period were 98%,97%,and 99%,respectively,at an average flow rate of 4.1L h −1and Empty Bed Contact Time (EBCT)>3min.Table 2Values of the Freundlich and Langmuir isotherm model parameters for three arsenite concentrations with iron-chitosan loading of 0.2g L −1at pH 8.Initial concentration(g L −1)Freundlich parameters Longmuir constants k f (L g −1)1/n R 2q max (mg g −1)K L (L mg −1)R 22950.590.240.98 2.000.120.985960.640.260.95 2.820.070.9410070.740.330.996.820.010.95D.D.Gang et al./Journal of Hazardous Materials182 (2010) 156–1611614.ConclusionsOverall,the study has demonstrated that iron-impregnated chi-tosan can effectively remove As(III)from aqueous solutions under a wide range of experimental conditions and removal efficiency depends on various factors including pH,adsorption time,adsor-bent loading,and initial concentration of As(III)in the solution. Results from the kinetic batch experiments indicated that more than60%of the arsenic was adsorbed by the iron-chitosan within 30min of adsorption.Kinetic resultsfit the pseudo-second-order model well.The second order reaction rate constants were found to decrease from3.19×10−2to1.15×10−2g mg−1min−1as the initial As(III)concentration increased from306to994g L−1.Adsorp-tion isotherm results indicated that maximum adsorption capacity increased from1.97to6.48mg g−1at pH=8as the initial concen-tration of As(III)increased from0.3to1mg L−1.The adsorption isotherm datafit well to the Freundlich model.Column experi-ments of As(III)removal were conducted using12-mm-ID column at aflow rate of25mL h−1with an initial As(III)concentration of 308g L−1.This study corroborates that impregnating iron into chitosan can significantly increase As(III)adsorption capacity of the chitosan bead.Advantages of using the iron-impregnated chitosan include its high efficiency for As(III)treatment and low cost compared with the pure chitosan bead.We expect that the iron-impregnated chi-tosan is a useful adsorbent for As(III)and could be used both in conventional packed-bedfiltration tower and Point of Use(POU) systems.The possible concerns include the physicochemical sta-bility of the adsorbent because of the biodegradable nature of the chitosan material.Further research is underway to examine the adsorbent stability and whether the iron-impregnated chitosan can maintain its capability after several regeneration andCompeting adsorption of other ions will also be AcknowledgmentsThe authors would like to thank Mr.Ravi K.Kadari and Ms. Dhanarekha Vasireddy for conducting the laboratory experiments. 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氧化锆-磁性壳聚糖材料对磷酸盐的吸附研究安风霞;刘景亮【摘要】制备了氧化锆负载的磁性壳聚糖复合材料,并研究了其对水体中磷酸盐的吸附行为.XRD表征结果表明负载在磁性壳聚糖上的氧化锆纳米颗粒并未影响材料的晶型结构.吸附实验结果表明负载氧化锆后磁性材料对磷酸盐表现出了良好的吸附性能,其吸附机理包括离子交换和静电作用.该磁性材料对磷酸盐的吸附量随溶液pH值的增大而减小;随温度的升高而升高;阴离子对水中磷酸盐吸附效果影响的次序为:Cl->NO-3>SO2-4.【期刊名称】《广州化工》【年(卷),期】2017(045)020【总页数】4页(P29-32)【关键词】磁性壳聚糖;氧化锆;吸附;磷【作者】安风霞;刘景亮【作者单位】国电环境保护研究院, 江苏南京 210031;南京晓庄学院, 江苏南京211171【正文语种】中文【中图分类】O647.3磷是水体富营养化的限制性元素之一，过量磷的引入能引起水体的富营养化，因此除磷对控制和防治水体富营养化有重要的意义[1]。
除磷的方法主要有生物法、化学沉淀法和吸附法。
吸附法除磷由于成本低、吸附效率高、可再生循环等优点，受到广泛关注[2-3]。
常用的吸附除磷材料如工业废渣[4]、天然材料[5-6]等对磷的吸附容量有限。
近年来，将金属氧化物或氢氧化物作为除磷吸附剂取得了显著效果，特别是铝[7-8]、钙[9]、铁[10-11]、锆[12-13]的氧化物对磷酸盐均有良好的吸附性能。
研究表明氧化锆化学性质稳定，与其他阴离子相比对磷酸盐具有较高的吸附选择性[14-15]。
Chitrakar[12]合成了无定形的氢氧化锆，对磷酸盐的最大吸附量为10 mg·g-1；Chubar 等[16]用溶胶-凝胶法合成的ZrO2·xH2O 对磷酸盐的最大吸附量约40 m g·g-1，结果表明，ZrO2 对磷酸盐的吸附机制在于 ZrOH 吸附位点与磷酸根离子的配位作用，吸附过程中-OH 被置换到水体中。
利用啤酒工业废水养殖小球藻的研究摘要：本研究用啤酒工业废水和不同碳源培养小球藻，主要目的是初步探讨小球藻对不同碳源的偏好和小球藻对啤酒工业废水的处理效果。
为我们达到利用各种有机废水作为替代性小球藻培养基，同时用小球藻处理各种有机废水的双重目的提供可能。
选取5种碳源分别加入培养基，来研究不同碳源对小球藻生长的影响，确定小球藻生长所需的最适合的碳源。
另外，用啤酒废水代替蒸馏水配制小球藻培养基，来研究小球藻对啤酒废水的处理效果。
啤酒废水中的重要环境污染物在培养小球藻的过程中可以得到有效地清除,并从中可以获得具有商业价值的小球藻细胞。
关键词：小球藻；碳源；啤酒废水前言：小球藻(Chlorella)是高价值微藻,具有在一般水域甚至在废水中快速生长的特性,目前普遍采用的开放式自养生产方式成本较高,为了提高生产效益,许多学者把小球藻的生产纳入综合利用和环境治理之中[1]。
同时,随着人口增长和资源短缺的矛盾不断加剧,水资源已面临短缺危机,水污染问题也已成为一个非常严重的全球问题。
加强水资源的循环利用是减轻污染和实现节能减排的重要途径。
啤酒废水COD、BOD高达1000 mg/L以上,BOD:COD值一般都在0.15以上,可生化性好,目前常用的处理方法是活性污泥法及其改进形式和生物接触氧化法但传统活性污泥法曝气动力消耗大,单位废水处理费用高,同时易发生污泥膨胀和产生大量污泥,使许多用户在经济和管理上难以承受[2],而且处理后水中的氮、磷含量仍然很高,大量排放易造成水体富营养化,同时也是一种资源浪费。
过去近60年的研究表明,微藻能有效地去除废水的氮、磷等营养物,藻类的强化培养已作为一种废水二级处理工艺,用于去除废水中残留的无机化合物。
不同藻类对氮、磷的净化效率是不同的,通过多种藻类的比较研究,Ganter等认为小球藻和栅藻是对这两种元素去除率最高的藻类。
藻类能有效地富集和降解多种有机化合物,如有机氯化合物、有机氮化合物、金属有机化合物等;Hosehis等对11个属的微生物去除废水BOD的比较实验表明,单种藻类对BOD的去除比单种细菌或原生动物更有效,其中普通小球藻对BOD的最大去除率可达到83%[3]。
第49卷第12期2021年6月广州化工Guangzhou Chemical IndustryVol.49No.12Jun.2021溶解氧对亚硫酸盐型自养反硝化的影响连雨露，倪伟敏，李可，刘月瞳(杭州师范大学生态系统保护与恢复杭州市重点实验室，浙江杭州311121)摘要：亚硫酸盐可代替碳源作为电子供体被脱氮硫杆菌利用进行反硝化，去除水中硝酸盐。
亚硫酸钠可作为脱氧剂，在30t,10.17mmol/L的亚硫酸钠可在4min内将水中DO降至0mg/L。
添加相同量亚硫酸钠，半封闭状态下，DO容易上升，稳定周期短，一个周期内硝态氮去除率仅为46.74%；而全封闭状态下，DO可长期稳定在0mg/L,—个周期内硝态氮的去除率可达到 97.74%o溶解氧会对亚硫酸盐型自氧反硝化产生抑制，通过改变亚硫酸钠投加量与采用全封闭状态可以消除抑制。
关键词：自养反硝化；亚硫酸盐；溶解氧抑制；pH影响中图分类号：X703.1文献标志码：A文章编号：1001-9677(2021)012-0079-03 Effect of Dissolved Oxygen on Sulfite-type Autotrophic Denitrification*LIAN Yu-lu,NI Wei-min,LI Ke,LIU Yue-tong(Hangzhou Normal University Key Laboratory of Ecosystem Conservation and Restoration，Zhejiang Hangzhou311121,China)Abstract:Sulfite can replace the carbon source as an electron donor and be used by Thiobacillus denitrificans to denitrify and remove nitrate from water.Sodium sulfite can also be used as a deoxidizer.At30t,10.17mmol/L sodium sulfite can reduce DO in water to0mg/L within4min.Adding the same amount of sodium sulfite,the DO was easy to rise in the semi-closed state,the stable period was short,and the nitrate nitrogen removal rate in one cycle was only 46.74%.While in the fully closed state,DO can be stabilized at0mg/L for a long time,and the removal rate of nitrate nitrogen in one cycle can reach97.74%.Dissolved oxygen can inhibit sulfite-type auto-oxygen denitrification，which can be eliminated by changing the dosage of sodium sulfite and adopting a fully enclosed state.Key words:autotrophic denitrification；sulfite；dissolved oxygen inhibition；pH effect机械、化工、电镀、光伏等行业常大量使用硝酸或硝酸盐作为原材料或辅助剂，其排放的废水含有大量硝酸盐且有机物含量低。
Phosphate removal from water using an iron oxide impregnated strongbase anion exchange resinT.Nur,M.A.H.Johir,P.Loganathan,T.Nguyen,S.Vigneswaran*,J.KandasamyFaculty of Engineering and Information Technology,University of Technology,Broadway,Sydney,NSW2007,Australia1.IntroductionPhosphorus(P)is an essential nutrient for the growth of plantsand microorganisms.It is also an important chemical element formany industries and a major nutrient contaminant in water.It isreleased to water bodies in the form of organic and inorganicphosphates by domestic,mining,industrial and agriculturalactivities and municipal discharges.Excessive concentration of Pin water causes eutrophication.Eutrophication is a majorenvironmental problem as it can lead to abundant developmentof aquatic plants,including algae,threatenfish and other aquaticlife,and disturb the ecological balance of organisms present in thewater.The excessive P in water should be removed for controllingeutrophication and maintaining a sustainable green environmentfor future generations.To control eutrophication,environmentalprotection agencies in many countries have recommended thattotal P should not exceed0.005–0.1mg P/L in natural water bodies[1,2].Several physical,biological and chemical processes have beeninvestigated for the removal of dissolved phosphates in water andwastewaters.These processes include adsorption/ion exchange,chemical precipitation/coagulation,crystallization and membranefiltration/reverse osmosis.Of the various methods of P removal,adsorption/ion exchange methods are promising,because they aresimple and economical,result in less sludge production andtherefore experience minimal disposal problems[3].Furthermore,these methods seem to be the most suitable for small watersupplies contaminated by P because of its simplicity,effectiveness,selective removal in the presence of other ions,easy recovery of Pand relatively low cost[4].These methods are also able to handleshock loadings and operate over a wide range of temperatures.The performances of the adsorbents with reference tophosphate removal have been reported in literature with varyingdegrees of success[3].Generally,ion exchangers are able toeffectively remove nitrate from water,but they have not beensuccessful in removing phosphate.For example,Gupta et al.[5]reported that Purolite500A anion exchange resin had a Langmuiradsorption capacity of64mg N/g for nitrate adsorption whereas ithad only7mg P/g for phosphate adsorption.Therefore,they usedtwo columns in series,one with Purolite for removing nitratefollowed by the other with hydrous ferric oxide for the removal ofphosphate from water containing both these anions.Others usedJournal of Industrial and Engineering Chemistry20(2014)1301–1307A R T I C L E I N F OArticle history:Received3May2013Accepted4July2013Available online12July2013Keywords:Anion exchange resinFixed-bed columnBreakthrough curvePhosphorusPhosphateAdsorptionA B S T R A C TRemoving phosphate from water is important as it causes eutrophication,which in turn has a harmfuleffect on aquatic life,resulting in a reduction in biodiversity.On the other hand,recovery of phosphatefrom phosphorus containing wastewater is essential for developing an alternative source of phosphorusto overcome the global challenge of phosphorus scarcity.Phosphate removal from aqueous solutionswas studied using an iron oxide impregnated strong base anion exchange resin,Purolite FerrIX A33E inbatch andfixed-bed column experiments.Phosphate adsorption in the batch study satisfactorilyfitted tothe Langmuir isotherm with a maximum adsorption capacity of48mg P/g.In the column study,increasein inlet phosphate concentration(5–30mg P/L),andfiltration velocity(2.5–10m/h)resulted in fasterbreakthrough times and increase in breakthrough adsorption capacities.Increase in bed height(3–19cm)also increased adsorption capacity but the breakthrough time was slower.The breakthrough datawere reasonably well described using the empirical models of Bohart–Adams,Thomas,and Yoon–Nelson,except for high bed heights.Phosphate adsorbed was effectively desorbed using1M NaOH andthe adsorbent was regenerated after each of three adsorption/desorption cycles by maintaining theadsorption capacity at>90%of the original value.Greater than99.5%of the desorbed P was recovered byprecipitation using CaCl2.ß2013The Korean Society of Industrial and Engineering Chemistry.Published by Elsevier B.V.All rightsreserved.*Corresponding author at:P.O.Box123,Broadway,NSW2007,Australia.Tel.:+61295142641;fax:+61295142633.E-mail addresses:s.vigneswaran@.au,Saravanamuth.Vigneswaran@.au(S.Vigneswaran).Contents lists available at SciVerse ScienceDirectJournal of Industrial and Engineering Chemistryj ou r n al h o m e p a g e:w w w.e l se v i e r.co m/l oc a t e/j i e c1226-086X/$–see front matterß2013The Korean Society of Industrial and Engineering Chemistry.Published by Elsevier B.V.All rights reserved./10.1016/j.jiec.2013.07.009polymeric anion exchanger bound with ferric oxide nanoparticles to successfully remove phosphate[4,6].These studies showed that adsorbents containing ferric oxide are required for effectively remove phosphate.Purolite FerrIX A33E media is another nanoparticle derived resin designed to selectively remove arsenic (arsenate and arsenite)from water supply[7].This resin unites a unique blend of hydrous iron oxide nanoparticles that have a very high attraction for arsenic with a durable,non-friable,spherical polymer substrate.As this resin was found to be effective in removing arsenic anions,it is expected to have strong affinity for phosphate anions as well,because phosphate like arsenic is specifically adsorbed on iron oxides[3].However,no detailed study has been conducted on phosphate adsorption by this ion exchange resin.The aim of the present research is to study and model the removal of phosphate from synthetic wastewater employing the ion exchange resin,Purolite FerrIX A33E in batch andfixed-bed column systems.The effects of bed height,initial phosphate concentration andfiltration velocity were investigated in the column study.The empirical models of Bohart–Adams,Thomas and Yoon–Nelson were tested for their ability to describe the column adsorption data.A study was also conducted to develop a suitable method to regenerate the adsorbent for reuse,as well as recovery of the desorbed P for beneficial use.2.Materials and methods2.1.Ion exchange resinA commercially available and reasonably economical iron oxide impregnated Type II hybrid strong base anion exchange resin, Purolite FerrIX A33E obtained from the Purolite Company,U.S.A.[7]was used as the adsorbent in this study.This adsorbent is a blend of hydrous iron oxide nanoparticles and highly porous polystyrene cross-linked with divinylbenzene polymer having an arsenic operating capacity of0.5–4g As/L and available in the form of spherical beads(0.3–1.2mm diameter).The highly porous nature of the resin bead allows for maximum utilization of the impregnated iron oxide.The polymer component of the adsorbent had positively charged quarternary ammonium functional groups with chloride as counter ion[8].The resin contained13%Fe.2.2.Feed solutionsThe feed solutions were prepared using KH2PO4with distilled water spiked with different concentrations of phosphate(5–30mg P/L).The pH of the solutions ranged from7.2to7.6.2.3.Chemical analysisThe analysis of phosphate ion was carried out using a Metrohm ion chromatograph(model790Personal IC)equipped with an auto sampler and conductivity cell detector.The separation was achieved using an A SUPP column3(150mmÂ4mm).Na2CO3 (3.2mmol/L)and NaHCO3(1.0mmol/L)were used as mobile phase with aflow rate of0.9mL/min.2.4.Batch adsorption experimentEquilibrium adsorption experiments were conducted in a set of glassflasks with100mL solutions spiked with phosphate(10mg P/L)and ion exchange resin concentrations of0.1–10g/L at room temperature(24Æ18C).The suspensions were agitated in aflat shaker at a shaking speed of120rpm for72h to ensure that the adsorption equilibrium is reached.However,preliminary experi-ments showed that the adsorption equilibrium was achieved within 48h.The experiments were duplicated and the average values were taken for data analysis.The difference between duplicate values was withinÆ2%.The amount of phosphate adsorption at equilibrium,q e (mg P/g),was calculated using Eq.(1),q e¼ðC0ÀC eÞVM(1) where C0=initial concentration of phosphate(mg P/L);C e=equi-librium concentration of the phosphate(mg P/L);V=volume of the solution(L)and M=mass of adsorbent(g).2.5.Column mode experimentsThefixed-bed column was made of2.0cm inner diameter pyrex glass tube.At the bottom of the column,a stainless steel sieve was attached followed by a layer of glass beeds in order to provide a uniformflow of the solution through the column.A known quantity(12–86g)of the resin was packed in the column to yield the desired bed height(3–19cm)of the adsorbent.Phosphate solutions of known concentrations(5,10,15,20and30mg P/L) was pumped upward through the column at a desiredfiltration velocity(2.5,5.0and10.0m/h)controlled by a peristaltic pump. The effluents at the outlet of the column were collected at regular time intervals and the phosphate concentration was measured using ion-chromatograph.The breakthrough curves show the loading behaviour of phosphate to be removed from solution in afixed-bed column and are usually expressed in terms of adsorbed phosphate-P concentration(C ad),inlet phosphate-P concentration(C o),outlet phosphate-P concentration(C)or normalized concentration defined as the ratio of outlet phosphate concentration to inlet phosphate concentration(C/C o)as a function of time.The maximum column capacity,q total(mg P),for a given feed concentration andfiltration velocity is equal to the area under the plot of the adsorbed phosphate-P concentration,C ad (C ad=C oÀC)(mg/L)versus effluent time(t,min)and is calculated from Eq.(2).q total¼Q1000Z tÀtotaltÀ0C ad dt(2)Equilibrium uptake q eq(mg P/g)or maximum capacity of the column is defined by Eq.(3)as the total amount of adsorbed phosphate-P concentration(q total)per g of adsorbent(M)at the end of the totalflow time:q eq¼q totalM(3)The detention times of Purolite FerrIX A33E duringfluidization are calculated using Eq.(4).t¼Hv(4) where t is the detention time(h),H is the column bed height(m) and v is the upflowfiltration velocity(m/h).2.6.Batch adsorption isotherm modellingBatch adsorption data were modelled using Langmuir adsorp-tion isotherm equation[9]expressed as follows:Q e¼q max K L C e1þK L C e(5)where q max=the maximum amount of the phosphate-P concen-tration per unit weight of the adsorbent(mg P/g),K L=Langmuir adsorption constant(L/mg).T.Nur et al./Journal of Industrial and Engineering Chemistry20(2014)1301–1307 1302This model can be linearized as follows:C e Q e ¼1q max K L þC eq max(6)2.7.Fixed bed column modelling2.7.1.The Bohart–Adams modelThe Bohart–Adams model [10]established the fundamental equations describing the relationship between C t /C o and t in a continuous flow system.The model expression is shown below:lnC t C o¼k AB C o t Àk AB N o ÁZ F (7)where,C o and C t (mg P/L)are the inlet and effluent phosphateconcentration,k AB (L/mg min)is the kinetic constant,F (cm/min)is the linear velocity calculated by dividing the filtration velocity by the column section area,Z (cm)is the bed depth of column and N o (mg P/L)is the saturation concentration.The values for k AB and N o were determined from the intercept and slope of the linear plot of ln(C t /C o )against time (t ).The Bohart–Adams adsorption model was applied to experimental data to describe the initial part of the breakthrough curve.2.7.2.The Thomas modelThe Thomas model [11]assumes plug flow behaviour in the bed,and uses the Langmuir isotherm for equilibrium and second order reversible reaction for kinetics.This model is suitable for adsorption processes where the external and internal diffusion limitations are absent.The linearized form of the Thomas model can be expressed as follows:lnC o C tÀ1 ¼k Th q o MQ Àk Th C o t (8)where,k Th (mL/min mg)is the Thomas rate constant;q o (mg P/g)is the equilibrium phosphate uptake per g of the Purolite resin;C o (mg/L)is the inlet phosphate concentration;C t (mg P/L)is the outlet phosphate concentration at time t ;M (g)is the mass of adsorbent,Q (mL/min)is the filtration velocity and t (min)stands for filtration time.A linear plot of ln[(C o /C t )À1]against time (t )was employed to determine values of k Th and q o from the intercept and slope of the plot.2.7.3.The Yoon–Nelson modelYoon and Nelson [12]developed a model based on the assumption that the rate of decrease in the probability of adsorption of adsorbate molecule is proportional to the probability of the adsorbate adsorption and the adsorbate breakthrough on the adsorbent.The linearized form of the Yoon–Nelson model for a single component system is expressed as:ln C to t¼k YN t Àt k YN(9)where,k YN (1/min)is the rate velocity constant,t (min)is the time required for 50%phosphate breakthrough.A linear plot of ln [C t /(C o ÀC t )]against sampling time (t )was used to determine values of k YN and t from the intercept and slope of the plot.2.8.Regeneration of adsorbent and phosphate recoveryWhen the adsorbent Purolite FerrIX A33E was saturated with phosphate ions in the column,it was important to regenerate the adsorbent for the recovery of phosphate ions as well as the reuse of the adsorbent for further adsorption of phosphate.In this study,the column regeneration studies were carried out using differentdesorbing agents such as 1M NaCl,1M Na 2SO 4,1M NaOH,and MQ water.The regeneration was performed by leaching the resin containing the adsorbed phosphate with the leaching solution at a filtration velocity 10m/h for 30min.Since phosphate was enriched through the adsorption process,the enriched phosphate was desorbed by the regeneration solutions and recovered as calcium phosphate by adding different concentrations of CaCl 2to the leachate.3.Results and discussion3.1.Batch adsorption experimentThe results of the batch equilibrium adsorption experimentindicated that the removal efficiency of phosphate improved with increased resin dose due to increased availability of surface area for adsorption (Fig.1a).Approximately 90%removal of phosphate was achieved with a resin dose 1g/L and the removal efficiency reached 97%with resin doses !5g/L.The adsorption isotherm data (Fig.1b)satisfactorily fitted to the Langmuir adsorption model (data not shown)and the value of the maximum adsorption capacity was 48mg P/g.This adsorption capacity is among the highest values reported for most adsorbents in the literature (Table 1).The high adsorption capacity is probably due to the presence of nano-sized iron oxide particles in the adsorbent which can have adsorption capacity in the order of 145mg P/g Fe [28].Based on this adsorption capacity the contribution of the iron oxide to the total adsorption capacity of the Purolite resin is estimated to be 19mg P/g resin (13%Fe in resin Â145mg P/g).This amount of phosphate is adsorbed specifically by inner sphere surface complexation on iron oxide [3,13,14]and the remaining amount of phosphate is adsorbed by coulombic forces (outer-sphere complexation)on the quaternary ammonium functional group positive charges in the polymer component of thePurolite.020*********)%(y c n e i c i f f E l a v o m e R Resin dose (g/L)(a)51015202530246810Q e (m g /g )Ce (mg/L)(b)fit (R2=0.97)Fig.1.(a)Effect of Purolite FerrIX A33E resin dose on the efficiency of P removal from a solution containing PO 4-P concentration of 10mg/L and (b)Langmuir Isotherm model fit for phosphate removal by the resin.T.Nur et al./Journal of Industrial and Engineering Chemistry 20(2014)1301–130713033.2.Fixed-bed column studiesAdsorption of phosphate by Purolite FerrIX A33E is presented in the form of breakthrough curves (Figs.2–4).The breakthrough curves became less sharp when the mass transfer rates were decreased [29].Since mass transfer rates were finite,the break-throughs were diffused and exhibited an S-shape pattern.3.3.Effect of adsorbent bed heightFig.2shows the breakthrough curves obtained for phosphate adsorption on the Purolite resin at bed heights of 3,6,12,14,16,17and 19cm (12,28,56,66,76,80and 86g of resin),at a constant filtration velocity 2.5m/h and inlet concentration 20mg P/L.Based on Fig.2it is evident that at low bed heights,the breakthrough occurred faster than that at high bed heights.This pattern of breakthrough at different bed heights is similar to the findings in other column studies reported on different adsorbents and adsorbates [30,31].The starting time of saturation occurred after 3,9,13,19,23,25and 28h when the bed height was 3,6,12,14,16,17and 19cm,respectively,and 50%saturation was achieved within the interval of 5,13,21,25,31,34and 39h,respectively.As the bed height increased,phosphate had more time to contact with the Purolite ion exchange resin as shown by the higher detentiontime (Table 2),resulting in more efficient removal of phosphate.Thus,the higher bed height resulted in a greater decrease in phosphate concentration in the effluent.The slope of the breakthrough curve decreased with increase in bed height as a result of broadened influent movement zone [31].However,it is found that the complete pattern of adsorption breakthrough curve was formed at all bed heights.The adsorption of phosphate increased when bed height rose from 3to 19cm,because of the increased amount of adsorption sites available at higher bed heights.3.4.Effect of initial phosphate concentrationThe effect of increase in the influent phosphate concentration from 5to 30mg P/L on breakthrough curves is shown in Fig.3.The starting time of saturation occurred at 30,23,17,15and 13h and 50%saturation was achieved after an operation period of 45,28,21,19and 15h for the influent concentrations of 5,10,15,20and 30mg P/L,respectively.The breakthrough time occurred faster and the breakthrough curves were sharper with increasing influent phosphate concentration.These results are consistent with nitrate adsorption on an ion exchange resin [32].These results demon-strate that the change of concentration gradient affects the saturation rate and breakthrough time,or in other words,theTable 1Comparison of Langmuir adsorption capacity for phosphate on Purolite FerrIX A33E with that on various other adsorbents in synthetic solutions.AdsorbentExperimental conditions(E,equilibrium concentration (mg P/L);I,initial concentration (mg P/L))Adsorption capacity (mg P/g)ReferenceAkagenite258C,pH 7,E 1À25060[13]Akagenite granulated 208C,pH 5.5,E 0À417–23[14]Activated alumina 208C,pH 5.5,E 0À412–14[14]Activated aluminaE 0À23.3[15]Iron/zirconium binary oxide 258C,pH 4,E 0À8013.7[16]MgAlLDH (granular)258C,pH 6.9,E 0À15047.3[17]Zeolite208C,E 0À400.13[18]Activated carbonE 0À283.2[19]Activated carbon (granular)208C,E 0À251.2[18]Anion exchange resin from soybean hulls 258C,pH 7,10g resin,I 0À62020[20]BauxsolpH 5.2–6.2,E 0À4314–15[21]Blast furnace slag 258C,E 0À60044.2[22]Activated red mud 308C,pH 5.2,E 0À6222.5[23]Activated red mud 258C,E 0À80054,113[24]Raw red mud 258C,E 0À80038[24]Blast furnace slag308C,E 0À1129[25]Basic oxygen furnace slag 208C,E 0À12011–20[26]Fly ashE 0À90020–26[24]Iron oxide tailings20–218C,pH 6.6–6.8,E 0À1305–8[27]Purolite A500P anion exchange resinResin dose 0.5–10g/L,I 157[6]Purolite FerrIXA33E anion exchange resin248C,pH 7.2–7.6,resin dose 0.1–10g/L,I 1048This study00.20.40.60.810102030405060C t /C 0Time (h)Fig. 2.Breakthrough curves for different bed heights (initial phosphate concentration =20mg P/L and filtration velocity =2.5m/h).0.20.40.60.81C t /C 0Time (h)Fig.3.Breakthrough curves for different inlet concentrations (bed height =12cm and filtration velocity =2.5m/h).T.Nur et al./Journal of Industrial and Engineering Chemistry 20(2014)1301–13071304diffusion process is concentration dependent.As the influent concentration increases,phosphate loading rate increases,so does the driving force for mass transfer,and decreases in the adsorption zone length [33].The extended breakthrough curve at low influent concentration indicates that a higher volume of solution can be treated.3.5.Effect of filtration velocityThe effect of filtration velocity on the adsorption of phosphate on Purolite resin was investigated by varying the filtration velocity (2.5,5.0and 10.0m/h)at a constant adsorbent bed height of 12cm and the inlet concentration of 30mg P/L.The breakthrough generally occurred faster and the breakthrough curve was steeper with higher filtration velocity (Fig.4).The time to reach the plateau of C t /C 0increased significantly with a decrease in the filtration velocity.The plateau of C t /C 0occurred at 9,15,and 23h for the inletfiltration velocity of 10,5.0,and 2.5m/h with the values of C t /C 0of 0.90,0.85,and 0.83,respectively.Faster breakthrough of adsor-bates and steeper breakthrough at higher filtration velocities have also been reported elsewhere for adsorption onto other adsorbents [31,34].At a lower filtration velocity,phosphate had more time to contact with Purolite resin as indicated by the higher detention time (Table 4),which allowed the diffusion of the phosphate ions into the pores of the adsorbent,resulting in a higher proportion of the removal of the influent phosphate ions in the column (lower C t /C o ).However,the quantity of phosphate ions removed was lower at lower filtration rate (Table 4),due to the lower amounts of phosphate ions (lower bed volumes)passing through the column per unit time.Mixed results have been reported in the literature on the effect of filtration velocity on adsorption capacity.Increase of filtration velocity was found to decrease the adsorption of Cu on activated carbon [31],whereas increase of filtration velocity increased adsorption of Pb on oil palm fibre [35].However,Hekmatzadeh et al.[32]reported that filtration velocity had no effect on the adsorption of nitrate on an ion exchange resin.The differences in the effects of filtration velocity may be due to the type of adsorbent and adsorbate,and experimental conditions such as filtration velocities,bed heights and influent concentra-tions used in the different studies.3.6.Empirical modelling3.6.1.Bohart–Adams modelThe model’s fit to the experimental data was good for all the inlet phosphate concentrations (R 2!0.72,Table 3)and filtration velocities (R 2!0.80,Table 4),but only for the low bed heights (3–14cm,R 2!0.80,Table 2).The values of k AB calculated from the model decreased when inlet phosphate-P concentration and bed height increased,but it increased when the filtration velocity increased.The value of N o increased with the increasing inlet concentration and bed height but decreased with increasingC t /C 0Time (h)Fig.4.Breakthrough curves for different filtration velocities (bed height =12cm and initial phosphate concentration =30mg P/L).Table 2The parameters of three models for different bed heights (initial concentration =20mg P/L and filtration velocity =2.5m/h).Bed height (cm)Bohart–Adams modelThomas modelYoon–Nelson modelBreakthrough adsorption capacityDetention time k AB(L/mg.min)x 10À5N o(mg/L)Â103R 2k Th(mL/min mg)Â10À2q o(mg/g)R 2k YN(1/min)Â10À3t (min)Â103R 2q eq (mg/g)t (h)350.311.80.8865.811.40.9815.30.30.9611.40.012617.816.50.8622.812.60.96 5.30.90.9013.50.024128.617.00.9113.212.70.94 3.5 1.50.9311.10.048147.020.70.8012.013.40.91 3.2 1.80.8311.70.05616 5.621.80.6310.913.60.86 1.6 3.10.6312.20.06417 5.023.20.729.613.70.90 1.6 3.20.7512.00.068194.425.00.428.714.50.841.53.40.6112.50.076Table 3The parameters of three models for different initial concentrations (bed height =12cm and filtration velocity =2.5m/h).Influentconcentration (mg P/L)Bohart–Adams modelThomas modelYoon–Nelson modelBreakthrough adsorption capacityDetention time k AB(L/mg min)Â10À5N o (mg/L)Â103R 2k Th(mL/min mg)Â10À2q o(mg/g)R 2k YN (1/min)Â10À3t (min)Â103R 2q eq (mg/g)t (h)514.18.40.8324.8 4.70.880.9 3.30.81 4.10.0481013.511.10.7219.87.40.89 2.1 2.20.76 6.90.0481511.712.80.9318.67.60.89 2.8 1.50.767.00.0482011.215.30.8417.711.30.91 3.3 1.40.8211.40.048309.220.00.9016.112.80.944.11.30.8513.20.048T.Nur et al./Journal of Industrial and Engineering Chemistry 20(2014)1301–13071305filtration velocity.The increase in k AB with filtration velocity shows that the overall system kinetics was dominated by external mass transfer in the initial part of adsorption in the column [36].3.6.2.Thomas modelThe high R 2values (0.80–0.98)obtained for the model fit to experimental data indicate that Thomas model described the column data very well (Tables 2–4).The model prediction of the column adsorption capacity (q o )increased with bed height (Table 2),inlet phosphate concentration (Table 3),and filtration velocity (Table 4)consistence with the observations made earlier based on the breakthrough curves.The values obtained for q o from the model are approximately equal to those calculated from the breakthrough curves.3.6.3.Yoon–Nelson modelThe experimental data fitted satisfactorily to Yoon–Nelson model for all phosphate concentrations and filtration velocities (R 2=0.76–0.99,Tables 3and 4)but,similar to Bohart–Adams model,the fit was good only at low bed heights (3–14cm,R 2!0.83,Table 2).The rate constant k YN increased and the 50%breakthrough time (t )decreased when both filtration velocity and inlet concentration increased (Tables 3and 4),but the opposite trend occurred with increasing bed height (Table 2).With the increase of bed height,t rose while k YN fell,which was also observed for Cu adsorption on rice husk-based activated carbon [31].Considering the values of R 2for the models fits to the data and breakthrough curves,it can be concluded that all three models can be used to reasonably describe the behaviour of the adsorption of phosphate on the Purolite resin in a fixed-bed column.The exception is perhaps the use of the Bohart–Adams and Yoon–Nelson models at high bed heights,where the models’fits to the data were poor.3.6.4.Regeneration of Purolite resinThe phosphate adsorption capacity of the column was estimated by Thomas model to be 16.4mg P/g,and by breakthrough curve calculation to be 12.9mg/g (Table 4).Among the reagents used for the regeneration of Purolite resin,NaOH was found to be effective in desorbing the phosphate from the column.No detectable phosphate was desorbed using 1M NaCl,Na 2SO 4and water.The inability of the high concentrations of Cl Àand SO 42Àin the Na salts in desorbing the phosphate ions from the adsorbent suggests that phosphate ions were strongly adsorbed specifically by inner sphere complexation probably on the Fe oxide component of the adsorbent [3,13,14].This is consistent with the adsorption capacity of the iron oxide in the Purolite resin estimated in Section 3.1(19mg P/g),which is more than the amount of phosphate adsorbed in the column,suggesting that all the phosphate in the column may have adsorbed to the iron oxide component of the resin.Only 60–70%of the adsorbed phosphate was desorbed by 0.5M NaOH whereas 90–95%of the phosphate was desorbed by 1.0M NaOH in 42bed volumes.Approximately 70%phosphate was desorbed within 10min (14bed volumes).The regenerated adsorbent was tested for reuse after every adsorption–desorption cycle using fresh 1M NaOH for each desorption cycle (Fig.5).The results showed that the efficiency of phosphate adsorption–desorption was nearly the same for two cycles and dropped a little in the third cycle.The decline in efficiency of adsorption/desorption was not more than 10%,indicating that the Purolite resin has good potential to adsorb phosphate and is reusable for at least three times.Further studies need to be conducted for higher number of adsorption/desorption cycles to determine the number of such cycles that can be performed without significantly reducing the adsorption capacity.3.6.5.Recovery of desorbed phosphateThe effect of different concentrations of CaCl 2added to the solution containing the desorbed phosphate on the reduction in solution concentration of phosphate is shown in Fig.6.The results showed that increase in CaCl 2concentration decreased solution phosphate concentration,indicating that phosphate has been precipitated probably as calcium phosphate.Kuzawa et al.[17]showed that addition of CaCl 2to desorbed phosphate solution can form calcium phosphate and hydroxyapatite compounds which have good fertiliser value.Midorikawa et al.[37]also proposed a P removal and recovery system with recycling of the alkaline desorbing solution similar to the method reported by Kuzawa et al.[17].Recently this system was successfully tested in pilot plants in Japan and the United States [37,38].If P removed from the water is economically recovered,it can partly overcome the perceived future scarcity of P when natural phosphate rock reserves will be exhausted [3].Table 4The parameters of three models for different filtration velocities (bed height =12cm and initial concentration =30mg P/L).Filtration velocity (m/h)Bohart–Adams modelThomas modelYoon–Nelson modelBreakthrough adsorption capacityDetention time k AB (L/mg.min)Â10À5N o (mg/L)Â103R 2k Th(mL/min mg)Â10À2q o(mg/g)R 2k YN (1/min)Â10À3t (min)Â103R 2q eq(mg/g)t (h)2.5 5.829.70.8011.516.40.91 4.0 1.30.8512.90.0485.019.117.70.9213.619.20.849.00.60.9715.90.02410.033.216.40.9813.922.70.8014.00.30.9916.30.0120.60510152025C t /C 0Time (h)Fig.5.Breakthrough curves for phosphate desorption by 1M NaOH in the column adsorption experiment (bed height 12cm,initial concentration 30mg P/L,filtration velocity 2.5m/h)for three adsorption–desorption cycles of Purolite FerrIX A33E.T.Nur et al./Journal of Industrial and Engineering Chemistry 20(2014)1301–13071306。
