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1 ENGINEERED BIOCHARS FOR THE REMOVAL OF METALLIC, ORGANIC AND EMERGI NG CONTAMINANTS FROM AQUEOUS SOLUTIONS By MANDU IME INYANG A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2013
2 2013 Mandu Ime Inyang
3 To God, and my family, as none of this would have been possibl e without their love and support
4 ACKNOWLEDGMENTS I would like to extend my deepest gratitude toward those individuals who have made an immense contribution to the successful completion of this research and also made m y PhD journey a worthwhile one. Firstly, I am grateful to God for being my anchor and guide, and my famil y, in particular my parents, Mr. and Mrs. Ime Sampson Inyang for all their encouragement and continuous support toward my academic endeavors. Mum and Dad, your encouragement and advice helped me persevere during the challenging times in my doctoral research. Secondly, my sincere appreciation goes to my advisor, Dr. Bin Gao and all my committee members, Dr. Andrew Zimmerman, D r. Pratap Pullammanappalil, Dr. Jean Claude Bonzongo, and Dr. Ben Koopman for their input in improving the quality of my research. I am especially grateful to Dr. Bin Gao for his patience and continuous guidance in ensuring the timely completion of this re search. Your critique and counsel were valuable in improving the quality of this research. For constantly challenging me to think scientifical ly and develop strong research H ypothes e s, I am grateful to you Dr. Zimmerman. Thank you for serving as Co chair o n my committee. I am also thankful to Dr. Pratap Pullammanappallil for introducing me to a career of research in academia and for serving on my committee. To Dr. Koopman and Dr. Bonzongo, I am grateful for your critique and insightful suggestions that were valuable in the interpretation of my experimental data. Thirdly, special thanks go to my friends: Samriddhi Buxy, Zhouli Tian, Gayathri Ramohan, Samuel Aso and Congrong Yu as well as my Environmental Nanotechnology group members: Ying Yao, Lei Wu, Yuan Ti an, Lin Liu, Yining Sun, Zhou Yanmei, June
5 Fang, Chen Hao, Ming Zhang, Dr. Yu Wang and Dr. Xue Yingwen. Your friendship, co operation and support made my research experience a pleasurable one. I also appreciate the technical input and assistance rendered t o me by the technical staff of the Particle Engineering and Research Center, and Agricultural and Biological Engineering Department in the design of my experiments. Finally, I save my last and special thanks for my c hurch family: the Falade s Maryann and Melissa King; my mentor, Abhay Koppar; my U S moms: Sus ie Studstill and Donna Rowland. I am so grateful for all the love, support and encouragement you have shown me over the years. Thank you for making my stay in Gainesville, a memorable one.
6 TABLE OF C ONTENTS page ACKNOWLEDGMENTS ................................ ................................ ................................ .. 4 LIST OF TABLES ................................ ................................ ................................ ............ 9 LIST OF FIGURES ................................ ................................ ................................ ........ 10 LIST OF ABBREVIATIONS ................................ ................................ ........................... 12 ABSTRACT ................................ ................................ ................................ ................... 14 CHAPTER 1 ENGINEERED BIOCHARS FOR THE REMOVAL OF METALLIC, ORGANIC AND EMERGING CONTAMINANTS FROM AQUEOUS SOLUTIONS .................. 16 Introduction ................................ ................................ ................................ ............. 16 Biochar Engineering: Existing Research and Limitations ................................ ........ 19 Physical Engineering ................................ ................................ ........................ 19 Chemical Engineering ................................ ................................ ...................... 20 Biological Engineering ................................ ................................ ...................... 21 Research Objectives ................................ ................................ ............................... 22 Hypothesis 1 ................................ ................................ ................................ ..... 22 Objective 1 ................................ ................................ ................................ ....... 22 Hypothesis 2 ................................ ................................ ................................ ..... 23 Objective 2 ................................ ................................ ................................ ....... 23 Hypothesis 3 ................................ ................................ ................................ ..... 23 Objective 3 ................................ ................................ ................................ ....... 23 Hypothesis 4 ................................ ................................ ................................ ..... 24 Objective 4 ................................ ................................ ................................ ....... 24 Organization of Dissertation ................................ ................................ .................... 25 2 REMOVAL OF HEAVY METALS FROM AQUEOUS SOLUTIONS BY BIOCHARS DERIVED FROM ANAEROBICALLY DIGESTED BIOMASS ............ 26 Introduction ................................ ................................ ................................ ............. 26 Materials and Me thods ................................ ................................ ............................ 29 Materials ................................ ................................ ................................ ........... 29 Biochar Properties ................................ ................................ ............................ 30 Sorption of Heavy Metals ................................ ................................ ................. 31 Sorption of Lead ................................ ................................ ............................... 31 Post sorptio n Characterizations ................................ ................................ ....... 32 Results and Discussion ................................ ................................ ........................... 33 Biochar Properties ................................ ................................ ............................ 33 Sorption o f Mixed Heavy Metals ................................ ................................ ....... 34 Lead Sorption Kinetics ................................ ................................ ..................... 35
7 Lead Sorption Isotherms ................................ ................................ .................. 36 Post Sorption Characteristics and Sorption Mechanisms ................................ 39 Conclusion ................................ ................................ ................................ .............. 41 3 FILTRATION OF ENGINEERED NANOPARTICLES IN CARBON BASED FIXED BED COLUMNS ................................ ................................ ......................... 54 Introduction ................................ ................................ ................................ ............. 54 Materials and Methods ................................ ................................ ............................ 57 Materials ................................ ................................ ................................ ........... 57 Batch Sorption ................................ ................................ ................................ .. 58 Column filtration ................................ ................................ ............................... 59 Characterizations ................................ ................................ ............................. 60 Mathematical Models ................................ ................................ ....................... 60 Results and Discussion ................................ ................................ ........................... 62 Properties of ENPs and Carbons ................................ ................................ ..... 62 Batch Sorption ................................ ................................ ................................ .. 63 ENP Filtration and Transport in Fixed Bed Columns ................................ ........ 63 Modeling of ENPs Filtration and Transport ................................ ....................... 65 Conclusion ................................ ................................ ................................ .............. 68 4 SYNTHESIS, CHARACTERIZATION AND DYE SORPTION ABILITY OF CARBON NANOTUBES COATED BIOCHAR COMPOSITES ............................... 75 Introduction ................................ ................................ ................................ ............. 75 Materials and Me thods ................................ ................................ ............................ 77 Materials ................................ ................................ ................................ ........... 77 Preparation of CNT biochar Nanocomposite ................................ .................... 77 Characterization ................................ ................................ ............................... 78 Sorption of Methylene Blue ................................ ................................ .............. 79 Effect o f pH and Ionic Strength ................................ ................................ ......... 80 Results and Discussion ................................ ................................ ........................... 80 Biochar Properties ................................ ................................ ............................ 80 Methylene Blue Removal Efficiency of Biochars ................................ .............. 81 Sorption Kinetics ................................ ................................ .............................. 82 Sorption Isotherms ................................ ................................ ........................... 83 Effect of pH ................................ ................................ ................................ ....... 84 Effect of Ionic Strength ................................ ................................ ..................... 85 Conclusion ................................ ................................ ................................ .............. 85 5 SIMULTANEOUS SORPTION OF SULFAPYRIDINE AND LEAD BY BIOCHAR MODIFIED WITH SURFACTANT DISPERSED CARBON NANOTUBES .............. 99 Introduction ................................ ................................ ................................ ............. 99 Materials and Methods ................................ ................................ .......................... 101 Materials ................................ ................................ ................................ ......... 101 Preparation of Surfactant dispersed CNT Biochar Nanocomposites .............. 102
8 Characterization ................................ ................................ ............................. 103 Sorption of Lead and Sulfapyridine in Single Solute System .......................... 103 Co sorption of Lead and Sulfapyridine in B inary Solute System .................... 104 Results and Discussion ................................ ................................ ......................... 105 Biochar Properties ................................ ................................ .......................... 105 Prelimin ary Biochar Assessments ................................ ................................ .. 106 Sorption Kinetics ................................ ................................ ............................ 106 Sorption Isotherms ................................ ................................ ......................... 107 Co sorption of Lead and Sulfapyridine in Binary Solute Systems .................. 108 Conclusions ................................ ................................ ................................ .......... 109 6 ENGINEERED BIOCHARS FOR THE REMOVAL OF METALLIC, ORGANIC, AND EMERGING CONTA MINANTS FROM AQUEOUS SOLUTIONS ................ 118 Conclusions ................................ ................................ ................................ .......... 118 Recommendations ................................ ................................ ................................ 120 LIST OF REFE RENCES ................................ ................................ ............................. 122 BIOGRAPHICAL SKETCH ................................ ................................ .......................... 137
9 LIST OF TABLES Table page 2 1 Summary of sources, health effects, and maximum contaminant levels of selected heavy metals in water (USEPA). ................................ .......................... 42 2 2 Elemental composition (%, mass based) of biochars used in this study. ............ 43 2 3 Relevant physiochemical properties of biochars used in this study .................... 43 2 4 Best fit model parameters of lead removal from aqueous solutions on DAWC and DWSBC ................................ ................................ ................................ ....... 44 3 1 Elemental composition of carbon materials. ................................ ....................... 69 3 2 Physiochemical pro perties of filter materials and engineered nanoparticles (ENPs). ................................ ................................ ................................ ............... 69 3 3 Best fit model parameters for ENP transport in various filter media. .................. 70 4 1 Structural and physiochemical properties of the biochar based sorbents. .......... 87 4 2 Summary of models and best fit parameters of the sorption kinetics and isotherms. ................................ ................................ ................................ ........... 88 4 3 Comparison of methylene blue sorption capacities by various sorbents. ........... 90 5 1 Physiochemical properties of carbons used in this study. ................................ 111 5 2 Best fit model parameters for sorption kinetics and isotherms. ........................ 112
10 LIST OF FIGURES Figure page 2 1 Removal of heavy metals from aqueous solution by the two biochars converted from anaerobically digested biomass. ................................ ................ 45 2 2 Kinetics of lead removal from solution by the two biochars converted from anaerobically digested biomass. ................................ ................................ ......... 46 2 3 Relation between the amounts of Pb removed by the two biochars converted from anaerobically digested biomass and square root of time befor e equilibrium. ................................ ................................ ................................ ......... 47 2 4 Isotherms of lead removal from solution by the two biochars converted from anaerobically digested biomass ................................ ................................ .......... 48 2 5 Changes in solution pH during lead removal from solution by the two biochars converted from anaerobically digested biomass ................................ .. 49 2 6 SEM image and c orresponding EDS spectra of post sorption l ead loaded digested biochars at 10,000X ................................ ................................ ............. 50 2 7 SEM image and c orresponding EDS spectra of pre sorption digested biochars at 5000X ................................ ................................ ............................... 51 2 8 XRD spectra of pre and post sorption digested biochars. ................................ 52 2 9 FTIR spectra of pre and post sorption digested biochars.. ................................ 53 3 1 Removal efficiency of ENPs in batch sorption study ................................ ........... 71 3 2 Filtration and transport of ENPs i n fixed bed columns.. ................................ ...... 72 3 3 Derjaguin Landau Verwey Overbeek (DLVO) energy interactions between filter media and ENPs. ................................ ................................ ........................ 73 3 4 XRD patterns for raw and post filtration carbons loaded with AgNP and raw and post filtration carbons loaded with NTiO 2 ................................ .................... 74 4 1 Thermogravimetric analysis prof iles of biochar based sorbents. ....................... 91 4 2 Raman spectra of biochar based sorben ts. ................................ ........................ 92 4 3 Transmission electron micrographs for samples at 50000X magnification. ........ 93 4 4 Methylene blue removal efficiencies of biochar based sorbents. ........................ 94 4 5 Sorption kinetics plots of biochar based sorbents. ................................ .............. 95
11 4 6 Intraparticle diffusion kinetics plots for biochar based sorbents. ........................ 96 4 7 Sorption isotherms of biochar based sorbents. ................................ .................. 97 4 8 Effects of solution chemistry on methylene sorption on biochar based sorbents. ................................ ................................ ................................ ............. 98 5 1 Thermogravimetric an alysis of pristine and modified biochar based sorbents. 114 5 2 Preliminary assessments for sorption of SPY and Pb onto biochars ................ 114 5 3 K inetic plots for sorption of SPY and Pb onto surfactant CNT modified biochars. ................................ ................................ ................................ ........... 115 5 4 Intra particle dif fusion plots for sorption of SPY and Pb onto surfactant CNT modified biochars ................................ ................................ ............................. 115 5 5 Isotherms for sorption of Pb and SPY onto surfactant CNT modified biochars. ................................ ................................ ................................ ........... 116 5 6 Co sorptio n of Pb and SPY in binary solute system ................................ ........ 116 5 7 Fourier transform infra red analysis of SDBS CNT biochars and SPY laden SDBS CNT biochars. ................................ ................................ ........................ 117 5 8 Possible sorption mechanisms for SPY and Pb sorption on SDBS CNT modified biochars. ................................ ................................ ............................ 117
12 LIST OF ABBREVIATIONS AC Activated carbon AC F E Iron modified activated carbon A G NP Silver nanoparticles BC Bagasse biochar BC CNT Carbo n nanotube coated bagasse biochar BC SDBS C NT Sodium dodecylbenzene sulfonate dispersed carbon nanotubes bagasse biochar nanocomposites BET Brunauer Emmett Teller CNT Functionalized or non functionalized multi walled carbon nanotubes DAWC Digested animal waste biochar DLVO Derjaguin Landau Verwey Overbeek DWSBC Digested whole sugar beet biochar EDL Electrostatic double layer interactions EDS Energy dispersive spectroscopy ENP S Engineered nanoparticles F Freundlich FTIR Fourier transforms infra red spectroscopy HC Hickory biochar HC CNT Carbon nanotube coated hickory biochar HC F E Iron modified hickory biochar HC SDBS CNT Sodium dodecylbenzene sulfonate dispersed carbon nanotubes hickory biochar nanocomposites L Langmuir LL Langmuir Langmuir LW Lifshitz Vander waals attract ion energy
13 MB Methylene blue NT I O2 Nano titanium dioxide SDBS Sodium dodecylbenzene sulfonate SEM Scanning electron microscope SPY Sulfapyridine USEPA United States Environmental Protection Agency UV VIS Ultra violet visible spectroscopy XRD X ray diffraction
14 Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy ENGINEERED BIOCHARS FOR THE REMOVAL OF METALLIC, ORGANIC AND EMERGING CONTAMINANTS FROM AQUEOUS SOLUTIONS By Mandu Ime Inyang August 2013 Chair: Bin Gao Cochair: Andrew Zimmerman Major: Agricultural and Biological Engineering Engineered b iochars combining high tech materials with low cost biomass derived materials hold great promise for the removal of traditional contaminants like heavy metals and organic compounds as well as emerging contaminants (e.g., pharmaceutical residues and nanoparticles ) from wastewater In this work biochars were synthesized and activated using various techniques, and the resulting modified biochars we re evaluated for their abilities to sorb various heavy metals, dyes, engineered nanoparticles (ENPs) or antibiotics. First, anaerobically di gested whole sugar beet and dige sted animal waste residues w ere pyrolyzed into biochar. The sorption capacity (200 mmol kg 1 ) o f these two biochars for Pb was comparable to that of commercial activated carbons and the removal of Pb was mainly controlled by precipitation Next, elemental iron (Fe) was impregnated onto raw hickory biochar (HC) and activated carbons (AC) These c arbon based sorbents could sorb and retain ENPs from aqueous solutions and the iron modification of these sorbents improved their sor ption ability by reducing electrostatic repulsions between the negatively charged ENPs and carbons Third carbon nanotubes ( CNT s ) biochar nanocomposites were
15 synthesized by pyrolyzing dried mixture s of CNT s uspensions sorbed to raw biomass of different types The addition of CNTs significantly enhanced the physiochemical and sorptive properties of the biochars with CNT h ickory and sugarcane b iochars made using 1% CNT suspensions, exhibit ing the g reatest thermal stabilities surface areas an d MB sorption capacit ies. MB sorption onto the raw and modified biochars was predominantly influenced by electrostatic attraction of positively charged MB to negatively charged biochars. Last ly surfactant dispersed CNT biochar nanocomposites were pro duced by dip coating hickory or bagasse biomass in 1% sodium dodecylbenzenesulfo nate (SDBS) dispersed CNT solutions and then pyrolyzing the coated biomass T he S DBS CNT chars had the highest removal of Pb and SPY than the unmodified or CNT biochars in both single and binary solute systems via multiple sorption mechanisms In summary, the modification techniques presented here are time and cost efficient and have shown beneficial results for the treatment of a wide array of contaminants.
16 CHAPTER 1 EN GINEERED BIOCHARS FOR THE REMOVAL OF METALLIC, ORGANIC AND EMERGING CONTAMINANTS FROM AQUEOUS SOLUTIONS Introduction W astewater contamination by toxic organic chemicals, heavy metals a nd other emerging pollutants has become a world wide environmental conce rn ( Wang et al., 2010b ) T he undes irable effects of these contaminants to human health and aquatic ecosystems have further necessitated stringent regulations on their discharge levels to environmental waters Today, municipal wastewater treatment plants play a major role in limiting the pollution of these contaminants in aquatic environments ( Li et al., 2013 ) But recent research studies ( Gardner et al., 2013 ; Katsou et al., 2012 ; Lou & Lin, 2008 ) have shown that conventional waste water treatment technologies no longer suffic e in completely eliminating these contaminants from treated waters Moreover, the fate of some contaminants (e.g., emerging pharmaceutical and nanoparticle products ) in wastewater systems is not yet fully understood. Thus environmental remediation studies developing new water treatment technologies for these cont aminants ar e increas ing T he presence of elevated concentrations of heavy metals from point and non point sources in aqueous streams continues to pose challenges in environmental remediation due to their non biodegradable nature. Therefore, m aximum contaminant levels of many heavy metals are set by the United States Environmental Protect ion Agency ( US E PA), close to 0 ppm. In addition to aqueous streams, heavy metals such as cadmium, copper, lead and nickel occur in contaminated soils which mak e s their mobility of great concern ( Uch imiya et al., 2010 ) For instance lead can complex with
17 organic compounds in soil organic matter to form organo metallic complexes, and further increase its environmental persisten ce Unlike heavy metal s, most organic contaminants are biodegradable and often found in trace concentration s in aquatic systems. R emediation studies on organic contaminants ( Es'haghi et al., 2011 ; Lin & Xing, 2008 ; Mishra et al., 2010 ; Zhang et al., 2012c ) however, indicate that even at trace concentrations, certain organic compounds such as phenol, dyes, and dioxins could be bio persistent, and pose severe health problems to living organisms ( Fu et al., 2003 ) Specifically dyes are a class of colored chemicals used in various industries including textile, leather, paper, and plastic industries. But, some w ater soluble aromatic dyes (e.g., az o dye, methylene blue and congo red) are suspected carcinogens that also induce chronic effects on exposed microorganisms ( Yu & Fugetsu, 2010 ) W astewater containing dyes are difficult to treat due to their s tab ility to light and resistan ce to aerobic digesti on. Because t he presence of dyes also produce aesthetic problems their removal from wastewater systems is pertinent to improv ing the q uality of treated water Emerging contaminants typify a class of potentially toxic pollutants including natural or synthetic chemicals whose effects or presence is largely unknown because they have been recently introduced in the environment ( Smital, 2008 ) P harmaceutical residues, personal care products, nanomaterials and per fluoro chemicals are examples of emerging contaminants Among these named pollutants pharmaceutical residues from widely used human and veterinary drugs are considered one of the mo st frequently detected contaminants in wastewaters ( Jesus Garcia Galan et al., 2011 ; Radke et al., 2009 ) Like phar maceutical residues, there have also been increased concentrations of
18 enginee red nanoparticles in wastewater systems due to their widespread applications in many household and industrial products ( Nowack & Bucheli, 2007 ) Also, w hile engineered nanoparticles have high mobility in soil and water and are easily eluted from water treatment systems due to their small size ( Christian et al., 2008 ) ; pha rmaceutical residues are bio persistent and not wholly degraded by anaerobic microbes in many treatment plants ( Radjenovic et al., 2009 ; Radke et al., 2009 ) But the uptake of these emerging materials from wastewater is advantageous because their sorbed products can be further used as purification and disinfection agents in water Biochar is a porous, environmental, and ubiquitous carbon sorbent der ived from the thermal treatment of carbonaceous materials in a closed system under anaerobic conditions ( Uchimiya et al., 2010 ) Today, research on eco friendly biochar is increasing due to its many applications in carbon sequestration, soil fertility enhancements, energy generation, and en vironmental remediation ( Crombie et al., 2013 ; Inyang et al., 2011b ; Liu et al., 2013 ; Namgay et al., 2010 ; Spokas et al., 2012 ; Yao et al., 2011b ) In particular, the application of biochar to environmental r emediation is attractive due to the abundance and low cost of waste biomass that can be used for biochar production. Several research studies ( Inyang et al., 2011b ; Ippolito et al., 2012a ; Ko et al., 2004 ; Liao et al., 2012 ; Yao et al., 2013a ; Yao et al., 2012b ; Zhang et al., 2012a ) have investigated the removal of various contaminants, particularly, dyes, pharmaceutical residues, and heavy metals by biochars produced from a variety of materials In m any cases, the effective removal of these contaminants by biochars was attributed to the modification of their physiochemical properties. But, since no known stud y exist s for the
19 sorption of nano particles on biochar, engineered biochar s may also be co nsidered for immobilizing nanoparticles on biochar surface. Biochar Engineering: Existing Research and Limitations Biochar engineering or modification can be defined as the application of processing techniques to improve the sorptive properties of biochar. Engineering these methods are discussed in great er detail in the following section: Physica l E ngineering Physical engineering of biochar is performed to increase its surface ar ea and pore volume and may include mechanical processes such as milling of the raw biomass (prior to pyrolysis), or finished biochar material ( Peterson et al., 2012 ) For instance t he pulverization of peanut, soybean and canola straws prior to pyrolysis resulted in higher sorption of copper for the respective pulverized s traw biochars than the non pulverized chars Biochars can also be physically modified by manipulating pyrolytic conditions to improve the porosity or oxygen functionalities o f biochar surface Specifically p hysical modificat ion of chars by the passage of oxidizing gases such as steam and CO 2 or a combination of both in pyrolytic reactors are known to burn off loose, dangling carbon atoms and widen existing pores o n the chars making them more accessible to the molecules of adsorbed contaminants ( Lu et al., 2008 ) Moreover, depending on the nature of the feedstock, steam a pplication could reduc e highly acidic functional groups (e.g. c arboxylic and lactonic groups), and enlarg e weakly acidic groups (e.g. phenolic g roups) ( Borchard et al., 2012 ) H eavy metals and cationic dyes in aqueous solutions
20 have been shown to easily sorb onto these negatively charged sites on biochar surfaces ( Carrier et al., 2012 ; Cheng et al., 2008 ) One draw back to th is method of physical modification however, is the additional energy input required for the generation of steam duri ng the modification process. The cost of the ad ditional energy employed could increase th e cost of the engineered biochar, making it a less cost effective treatment option. Chemical Engineering Che mical engineering of biochar involves the incorporation of alkaline (e.g., KOH, NaOH), acidic (e.g., ZnCl 2 H 3 PO 4 H 2 O 2 ), or organic materials (e.g., hydrogel and aerogel), a nd recently, metallic oxides and nanoparticles in bio char to produce active carbons with enhanced surface chemistries and functionalities ( Ippolito et al., 2012b ) ( Karakoyun et al., 2011 ; Li et al., 2012 ; Xue et al., 2012 ; Zhang et al. ) Generally, t he use of KOH or ZnCl 2 and steam activation has been industrially employed in producing commercial activated carbons ( Azargohar & Dalai, 2008 ) This is because the presence of Zn or K constituents intercalated within the C structure of biochar during modification forces apar and the subsequent washing of the modified char, to remove some of these constituents, frees up the interlayer spaces containing these elements, yielding more porous chars ( Marsh, 1987 ) But, the removal of Cu by KOH steam modified pecan shell biochar was found to occur by the interaction of Cu with oxygen functional groups on the surface sites of the char ( Ippolito et al., 2012b ) This suggests tha t modification by KOH also increase s oxygen functionalities (O H groups) on the surface of modified biochars, in addition to increasing the porosity of the chars. In contrast, p(acrylamide)
21 chicken b iochar activat ed with HCl and hydrogel was suggested to i mprove the sorption of phenol by increased hydrophobic interactions. Two limitation s to the s e method s of modification however, are the laborious techniques employed and increased processing cost from the use of many cross linkers and binders to improve th e bonding of chemical reagents to biochar Bio logical Engineering Biological engineering has been recently proposed as another modification method for improving the sorptive properties of biochar ( Inyang et al., 2010 ) T he degradation of biomass substrates by microbes during aerobic or anaerobic treatment processes c ould yield more porous carbon structures when the resulting sludge residuals are pyrolyz ed For instance, the sorption of fluoroquinolone antibiotics on wastewater sludge biochar was attributed to the enhanced porosity of the sludge residual material ( Yao et al., 2013a ) Typically, t he process of anaerobic digestion involves the microbial uptake and assimilation of organic carbon (e.g., hemice llulose and cell ulose portions) which would (depending on the feedstock) result in higher conce ntration of inorganic, cationic (P, K, Ca, Mg, N, and S) and anionic species (CO 3 2 PO 4 3 MgO) on the digested residuals The removal of Pb on digested bagasse biochar which was 20 times better than undigested bagasse biochar ( Inyang et al., 2011a ) was attributed to the precipitation of cerrusite (PbCO 3 ) mineral from slowly released carbonate species interacting with Pb. These results confirm the poss ibility of biologically engineered biochar s to be used in the uptake of contaminants. Despite these results however, the practical application of digested biochars is limited by a paucity of studies utilizing digested biochars i n sorption studies
22 Research Objectives In order to overcome the limitations associated with existing biochar activation techniques t he overarching objective of this study was to develop biochar s using cost effective modification techniques requiring less labor ; yet achieving high removal efficiencies of metal lic emerging and organic contaminants Th e central H ypothesis of this study wa s that the modificati on techniques employed could enhance the physi ochemi cal properties or functionalities of biochar and increase their sorption c apacities for selected contaminants : (a) n anoparticles, (b) heavy metals, (c) cationic dyes, and (d) pharmaceutical residues T he t h re e methods employed for the engineer ing of the biochars were : (1) a naerobic digestion, (2) c hemical modification of biochars by iron impregnation and immobilization of filtered nanoparticles, and (3) pyrolysis of nano particle coated biomass composites. P ostulated H ypotheses and specific research objectives are discussed further in the following section: Hypothesis 1 Th e process of anaerobic digestion does not convert all of the biomass feedstock to methane, but could concentrate inorganic components (e.g., phosphates, carbonates, and oxides) in the residues that when converted to biochar can be slowly released to bind w ith dissolved heavy metals Objective 1 D etermine whether biochars converted from anaerobically digested biomass other than sugarcane bagasse (from previous work) can be used as effective sorbents to remove heavy metals from water The specific o bjectives of this study were to: E valuate the removal efficiency of lead, copper, nickel and cadmium from aqueous solution by two digested biochars.
23 D etermine the removal characte ristics of lead from solution on the digested biochars Investigate the mechanisms of lead removal on the digested biochars. Hypothesis 2 Chemically impregnated iron on negatively charged biochar surfaces can reduce electrostatic repulsions, and fav or the attachment of negatively charged engineered nanoparticles (ENPs) durin g filtration. Objective 2 Evaluate and compare the effectiveness of several carbon based materials including biochar, in retaining three types of ENPs: nano titanium dioxide (NTiO 2 ), silver nanoparticle (AgNP), and multi walled carbon nanotubes (CNT). The s pecific objectives of this study were to: E valuate and compare the ability of the carbon materia ls to sorb and filter the ENPs C ompare the mobility of the th ree ENPs in the carbon filters. D etermine whether biochar can be used as an effective filter media for the ENPs. Hypothesis 3 The carboxyl group incorporated in biochars by pyrolyzing carboxyl f unctionalized CNT coated biomass could provide extra high affinity sorption sites to bind cationic dyes like methylene blue on biochars Objective 3 I denti fy a simple, synthesis method for producing hybrid CNT biochar composite material s and test the sorption potential of the produced hybrid biochar materials for methylene blue removal The specific objectives of this study were to:
24 C haracterize the hybrid CNT biochar material s to investigate the effect of CNT on the physiochemic al properties of the chars. E xamine the influence of pH, contact time, and ionic strength conditions on the sorption capacity of the hybrid sorbents. E lucidate and understand the int eraction mechanisms governing the sorption of MB onto hybrid CNT biochar sorbents. Hypothesis 4 The dispersion of CNT by the surfactant will increase individual CNT threads in suspens ions that can be anchored to the biomass prior to pyrolysis and increas e high affinity sorption sites o n the produced surfactant modified biochar nanocomposites that can bind sulfapyridine and lead Objective 4 Modify hickory and bagasse biochars using sodium dodecylbenzenesulfonate (SDBS) s urfactant dispersed C NT and examine their removal ability for sulfapyridine (SPY) and lead (Pb) The specific objectives of this study were to: Exami ne the effect of SDBS dispersed CNT on the properties of SDBS CNT coated hickory and bagasse biochar nanocomposites Determine the sorption capacity of Pb and SPY on SDBS dispersed CNT hickory and bagasse biochars in a single solute system Investigate co sorption interaction mechanisms between SPY and Pb in a binary solute system f or both SDBS CNT biochars Elucidate and differentiate sorption mechanisms controlling the sorption of Pb and SPY on hickory and bagasse SDBS CNT coated biochars respectively.
25 Organization of Dissertation To achieve the stated objectives of this research as outlined in C hapter 1 this dissertation was organized into six C hapters. Chapter 2 examine d the use of anaerobic activation method for biochar and also test ed the efficiency of digested bio chars to sorb h eavy metals. Based on findings presented in C hapter 2 it was established that d igested biochars had the potential to sorb heavy metals via precipitation mechanism Thus C hapter 3 focus ed on the modification of biochar surface by iron impregnation /precipitation as well as examining how effective iro n modified c arbon s were in filtering nanoparticles But, overall results presented in C hapter 3 pointed to a low retention of CNTs even with these iron modified carb ons compared to other nanoparticles Accordingly, C hapter 4 examined the possibility of incorporating CNT s in the biochars by a dip coating procedure and evaluating the potential of these CNT modified biochar nanocomposites to remove MB from aqueous solutions Results from C hapt er 4, confirmed that incorporating CNT in biochars improve d the ir sorption capacity for MB Next C hapter 5 further ex amine d the effect of increasing the CNT content s of biochars by using SDBS in dispersi ng CNT. SDBS d ispersed CNT biochar composites were then used to sorb Pb and SPY in single and binary solute system s T h is dissertation culminate d in C hapter 6, where relevant conclusion s were drawn from C hapters 1 to 5 and plausible recommendations for future work were presented
26 CHAPTER 2 REMOVAL OF HEAVY METALS FROM AQUEOUS SOLUTIONS BY BIOCHARS DERIVED FROM ANA E RO BICALLY DIGESTED BIOMASS 1 Introduction Heavy metals pose a risk to public he alth because of their toxic, non biodegradab le nature and widespread occurrence in natural and human altered environments. They are mainly introduced into the environment from point sources such as discharges from mining, metal plating, battery, and paper industries. Lead, copper, cadmium, and nickel are among the most toxic and carcinogenic heavy metals that could cause serious environmental and health problems. The United Stat es Environmental Protection Agency (USEPA), therefore, has established very strict maximum contaminant level goals for these heavy metals in natural waters (Table 2 1). Many methods have been developed to address these stringent environmental regulations w hich necessitate removal of heavy me tal compounds from waste water. Traditional water treatment technologies, such as preci pitation, ion exchange, electro coagulation, membrane filtration, and packed bed filtration have been found to be effective in reducin g heavy metal concentrations ( Akbal & Camci, 2011 ; Boudrahem et al., 2011 ; Malamis et al., 2011 ) Most of these technologies however, may be associated with high operation cost and/or sludge disposal problems ( Sud et al., 2008 ) These disadvantages have increased the need of developing alternative and low cost water treatment technologies for heavy metal contaminants. Biosorbents therefore have been suggested to be a potential candidate to satisfy this need to remove toxic metals 1 Reprint with permission from Inyang, M., Gao, B., Yao, Y., Xue, Y., Zimmerman, A.R., Pullammanappallil, P., Cao, X. 2012. Removal of heavy metals from aqueous solution by biochars derived from anaerobically digested biomass. Bioresource Technology, 110, 50 56.
27 from wastewater ( Demirbas, 2009 ) For example, Demirbas ( 2008 ) indicated that agricultural by products and in some cases appropriately modified could be used to develop cost effect ive technologies to treat heavy metals in both industrial and municipal wastewater. Bioc har is a pyrogenic carbon rich material, derived from thermal decomposition of biomass in a closed system with little or no oxygen ( Das et al., 2008 ; Lehmann et al., 2006 ; Van Zwieten et al., 2010 ) When cheap biomass, particularly agricultural by products, is used for biochar produce, the cost of biochar produc tion is mainly associated with the machinery and heating, which is only about $4 per gigajoule ( Lehmann, 2007 ) The use of biochar as a low cost sorbent to remove metallic contaminants from aqueous solutions is an emerging and promising wastewater treatment technology, which has already been demonstrated in previous studies ( Beesley & Marmiroli, 2011 ; Liu & Zhang, 2009 ; Uchimiya et al., 2010 ) Biochars converted from agricultural residues, animal waste, and woody materials have been tested for their ability to sorb various heavy metals, including lead, copper, nickel, and cadmium ( Cao et al., 2009a ; Uchimiya et al., 2 011 ; Uchimiya et al., 2010 ) In addition, anaerobically digested biomass has been found to be a good feedstock to produce biochars with suitable physicochemical properties to serve as a low cost sorbent ( Inyang et al., 2010 ; Yao et al., 2011a ) A recent study indicated that biochar converted from anaerobically digested sugarcane bagasse is a far more effective sorbent of lead than biochar from undigested bagasse and even more effective than commercial activated carbon ( Inyang et al., 2011b ) It is suggested that anaerobic digestion could be used as a
28 high efficiency c arbon based sorbents for heavy metals ( Inyang et al., 2011b ) In addition, this method may also provide other benefits, such as producing renewable bioenergy through anaerobic digestion and pyrolysis and reducing waste management cost. However there is still a paucity of data showing the universal applicability of biochars converted from other digested biomass types also have superior ability to remov e heavy metals from water ( Inyang et al., 2011b ) Sugar beets and dairy manure are two of the most common biomass types used in anaerobic digesters to produce bioenergy. Sugar beets are traditionally used for sugar production; however, they require rapi d processing to maximize sugar extraction and minimize spoilage. Traditionally, dairy waste could be applied directly to agricultural lands as amendment for soils, but there are increasing concerns over the potential risk of surface and ground water contam ination ( Hooda et al., 2000 ) Recent studies suggest that anaerobic digestion could be an effective waste management strategy to reduce the volume of sugar beets and dairy waste as well as to generate bioenergy ( Brooks et al., 2008 ; Fang et al., 2011 ; Wang et al., 2010a ) Because most bacterial digestion processes cannot utilize all the feedstock material s, it is therefore important to develop methods to handle the residuals. To our knowledge, however, little research has been conducted to develop methods to process anaerobic digestion residuals, particularly with respect to using the digested biomass to m ake biochar based sorbents. The overarching objective of this work was to determine whether biochars converted from anaerobically digested biomass other than sugarcane bagasse can be used as effective sorbents to remove heavy metals from water. Two biocha rs were
29 produced from anaerobically digested dairy waste and whole sugar beets in the laboratory through slow pyrolysis. Batch sorption experiments were used to examine the sorption behaviors of heavy metals on the biochars and the physicochemical properti es of the pre and post sorption biochars were determined. Mathematical models were used to help data analysis and interpretation of sorption mechanism. The specific ob jectives of this work were to: (1) evaluate the removal efficiency of lead, copper, nick el and cadmium from aqueous solution by the two biochars; (2) determine the sorption characteristics of lead on the biochars; and (3) understand the sorption mech anisms of lead on the biochars. Materials and Methods Materials Digested dairy waste residue w as produced by a single stage, thermophilic, anaerobic digester at the Dairy Research Unit of the Animal Science Department, University of Florida (UF) in Gainesville, FL. Digested whole sugar beet residue was obtained from a two stage, thermophilic, high solids sequencing, anaerobic digester in the Sequential Batch Anaerobic Composting (SBAC) pilot plant at UF. The residues were pressed, de watered, then stored in air tight plastic bags, and refrigerated prior to use. To make the biochars, the residue materials were first dried at 80 o C. About 500 g of the dried feedstocks were converted into biochar through slow pyrolysis at 600 o C for 2 h in a N 2 environment in a furnace (Olympic 1823HE) following the procedures of ( Yao et al., 2011a ) The resulting biochars are referred to as DAWC (digested animal waste char) and DWSBC (digested whole sugar beet char). The biochar samples were ground and sieved to 0.5 1 mm sized particles. After several rins es with deionized (DI)
30 water to remove impurities such as ash, both DAWC and DWSBC samples were dried at 80 o C for further testing. All chemical reagents used were of high purity grades from Fisher Scientific (Suwanee, Georgia). Stock solutions of 1000 ppm lead (II) nitrate, cadmium (II) nitrate tetrahydrate, nickel (II) nitrate hexahydrate, and copper (II) nitrate trihydrate were prepared by dissolving appropriate amount of chemicals in DI water. Biochar P roperties Carbon, hydrogen, and nitrogen contents of the biochars were determined using a CHN Elemental Analyzer (Carlo Erba NA 1500) via high temperature catalyzed combustion followed by infrared detection of resulting CO 2 H 2 and NO 2 gases. Major inorganic elemental constituents of the biochars were det ermined using the EPA 200.7 method of acid digestion followed by analysis by inductively coupled plasma with atomic em ission spectroscopy (ICP AES). The pH of the biochar samples was measured by combining biochar with DI water in a mass ratio of 1:20. The solution was then hand stirred and allowed to stand for 5 min before measurement with a pH meter (Fisher Scientific Accumet Basic AB15). colloidal biochar suspensions obtained through sonication according to the procedure of ( Johnson et al., 1996 ) Charge mobility of each biocha r suspension was determined using a Brookhaven Zeta Plus (Brookhaven Instruments, Holtsville, NY) and Specific surface areas of the biochars were determined on a Quantach rome Autosorb1 surface area analyzer. N 2 adsorption isotherms measured at 77 K and interpreted using Brunauer, Emmet, and Teller (BET) theory yielded mesoporous
31 surface area (pores > 1.5 nm) and CO 2 adsorption isotherms at 273K were interpreted using Monte Carlo simulations of the non local density functional theory and yielded microporous surface area (pores < 1.5 nm). Sorption of Heavy Metals An initial evaluation of the sorption ability of DAWC and DWSBC was performed using a mixed heavy metal solution c ontaining Pb 2+ Cu 2+ Cd 2+ and Ni 2+ The concentration of each metal in th e solution was adjusted to be 0. 1 mmol L 1 About 0.1 g of the test biochar was added into 68 mL digestion vessels (Environmental Express) and mixed with 50 mL of the heavy metal solution at room temperature (22 0.5 o C). After shaking in a reciprocating shaker for 24 h, the vessels were withdrawn and filtered immediately through 0.1 m pore size nylon membranes (GE cellulose nylon membranes). The Ni, Cu, Cd a nd Pb concentrations in the filtrates were determined using ICP AES (Perkin Elmer Plasma 3200RL). The sorbed heavy metal concentrations were calculated based on the difference between the initial and final metal concentrations in the supernatant. Vessels w ithout the sorbent (biochar) or the sorbates (metals) were included as experimental controls. Sorption of Lead Sorption kinetics of lead on DAWC and DWSBC were determined by mixing 50 mL of 200 ppm Pb solution with 0.1 g of each sorben t in the at room temperature, and shaking over the course of a 24 h period. Sample solutions with their corresponding blank controls were withdrawn at specific time intervals to examine sorption kinetics. The mixtures were immediately fil tered and the filtrates we re stored for further analysis.
32 Sorption isotherms were obtained by adding 0.1g of each biochar to 50 mL of Pb 2+ solutions with varying concentrations (5 to 600 ppm) in each vessel and shaken for 24 h. The solutions were then filt ered and pH values of the filtrates were recorded. Both the filtrates and lead laden biochars were collected for further analysis. The lead laden biochars were washed with DI water several times and oven dried before analysis. For all lead sorption experim ents, blank experiments without the sorbent or sorbate were included as experimental controls, which indicated that there was no addition or loss of lead in the experiments. Lead concentrations in the filtrates were determined with the ICP AES. Lead concentrations on the biochars were calculated based on the differences between initial and final aqueous solutions. All the sorption experiments (mixed heavy metals and lead) were performed in duplicate and the average values are reported here. Additional analyses were conducted whenever two measurements showed a difference larger than 5%. Post sorption Characterizations Scanning electron microscopy (SEM) coupled with energy dispersive spectroscopy (EDS) was used to examine surface morphology and elemental composition of both pre sorption and post sorption (lead laden) DAWC and DWSBC. The tested samples were mounted on carbon stubs using carbon conductive paint. The samples were placed under a JEOL JSM 6330F field emission SEM equipped with an Oxford EDS. D uring operation, the accelerating voltage of the instrument was maintained at 10 kV and varying magnifications were used. X ray diffraction (XRD) analysis was conducted on both pre and post sorption biochar samples to identify possible crystalline structu res. A computer controlled X ray diffractometer (Philips Electronic Instruments) equipped with a stepping motor and
33 graphite crystal monochromator was used. Fourier transform infra red (FTIR) analysis was used to characterize functional groups present on t he biochar surfaces. The biochar samples were directly mounted on the diamond base of a Nicolet 6700 FTIR (Thermo Scientific) and a transparent polyethylene film was used to cover the samples for the FTIR analysis. Results and Discussion Biochar P roperties CHN analysis revealed that DAWC contained much more carbon than DWSBC (Table 2 2 ), probably because the feedstocks were from different types of anaerobic digesters. DAWC also had higher nitrogen content but slight ly lower hydrogen content than DWSBC (Tab le 2 2 ). Elemental analysis showed that the two biochars had various amounts of inorganic elements and slightly higher amounts of Ca (5.5%) and K (2.3%) were present in DWSBC and DAWC, respectively (Table 2 2 ). These inorganic elements originated from nutr ients rich in plant and animal residues, but the most pre dominant component in both biochars was carbon. Previou s studies have shown that physi ochemical properties of biochars, such as pH, surfac e potential, and surface area, are important factors control ling their environmental applications ( Inyang et al., 2010 ) Both DAWC and DWSBC were alkaline with a relatively high pH (Table 2 3 ), which is similar to the biochars obtained from anaerobically digested sugarcane bagasse ( Inyang et al., 2010 ) This suggests that the two biochars could be goo d conditio ners for acid soils. The surface potential measurements indicated that DAWC had more negative surface charge than DWSBC which may be related to its higher surface area and pore volume (Table 2 3 ). All of
34 these data seem to suggest a greater potential for D AWC to sorb abundant posit ively charged heavy metals. Sorption of Mixed Heavy M etals Both DWSBC and DAWC showed good ability to remove the mixture of four heavy metals from aqueous phase (Figure 2 1 ) The removal efficiency of the four metals by DWSBC was higher than 97%, indicating this biochar has a strong affinity for all the tested heavy metals. DAWC also showed high removal efficiency for Pb 2+ (99%) and Cu 2+ (98%), but relatively low removal efficiency for Cd 2+ (57%) and Ni 2+ (26%). Previous studies indicated that the effectiveness of biochar in the immobilization of heavy metals strongly depends on the metal contaminant type ( Uchimiya et al., 2010 ) The two biochars, however, showed different trends in removal rates (ability) for the four metals in the mixed solution, indicating both m etal and biochar type played important roles. Although the differences were very small, the removal ability of DWSBC for the metals followed the order of Cd>Ni>Pb>Cu. In contrast, DAWC followed the order of Pb>Cu>Cd>Ni, which is consistent with the removal efficiency trend of biochars converted from poultry litter ( Uchimiya et al., 2010 ) A ccording to Shi et al. (2009 ) high sorption of Pb 2+ from solution on sorbents through surface electrostatic attraction could be attributed to its high electronegativity constant (2.33), which results in a high tendency for specific adsorption. However, the electronegativity constants of Ni 2+ Cu 2+ and Cd 2 + are 1.93, 1.90, and 0.69, respectively, is not consistent with the heavy metal removal trends for either DAWC or DWSBC. Surface electrostatic interaction, therefore, might not be a dominant heavy metal removal mechanism for these biochars. Other mechanis ms, such as precipitation and surface complexation, should also be considered ( Cao et al., 2009a ; Inyang et al.,
35 2011b ; Uchimiya et al., 2011 ) In the following sections, sorption of lead on DAWC and DWSBC was examined in greater detail to improve current mechanistic understanding of heavy metal removal by biochars fro m anaerobically digested biomass. Lead Sorption Kinetics DAWC and DWSBC sorbents showed similar lead sorption kinetics and reached apparent sorption equilibrium after about 24 h (Figure 2 2 ). There was an initial rapid increase in lead removal followed by a slow down as sorption approached equilibrium. Pseudo first order, pseudo second order, and Elovich models were used to simulate the sorption kinetics data collected. The governing equations can be written as ( Gerente et al., 2007 ; Yao et al., 2011 b ) First order (2 1 ) Second order (2 2) Elovich (2 3) W here q t (mmol kg 1 ) and q e (mmol kg 1 ) are the amounts of lead sorbed at time t and at equilibrium respectively; k 1 (h 1 ) and k 2 (kg mmol 1 h 1 ) are the first order and second order apparent sorption rate constants, respectively; and 1 h 1 mmol 1 ) are the initial Elovich sorption and desorption rate constant at time t, respectively. The first order, second order and Elovich models reproduced the sorption data closely for both biochars with coefficients of correla tion all above 0.94 (Figure 2 2 ). While the Elovich model had the best fit for DWSBC with an R of 0.97, the second order model was a better fit for DAWC with an R of 0.95 (Table 2 4 ). Fittings of the three models to DWSBC Pb sorption were slightly better than that of DAWC (Table 2 4 ),
36 suggesting that sorption of lead on DAWC was more energetically heterogeneous. Although those models assume different mechanisms ( Gerente et al., 2007 ) comparisons of the fittings did not help reveal the governing mechanisms of lead sorption on the two biochars because there were only slight differences among the simulated results (Figure 2 2 ). A plot of the pre equilibrium sorbed Pb amounts against the square root of contact times for both biochar sorbents showed a strong linear dependency with an R of 0.90 and 0.98 for DAWC and DWSBC, respectively (Figure 2 3 ). This strong linear relationship, along with the slow sorption rate prior to reaching equilibrium for both sorbents, suggests diffusion controlled removal of Pb 2+ by the two biochars. This mensions, either for a surface adsorption or precipitation removal mechanism. The more rapid approach to Pb sorption equilibrium by DAWC and its larger pore volume is consistent with this interpretation. Lead Sorption Isotherms The lead sorption isotherms on the two biochars showed a very rapid increase in solid phase concentrations, removing close to all the lead at low equilibrium solution concentrations (Figure 2 4 ). Above lead equilibrium solution concentrations of 0.5 mmol L 1 there was very little additional lead removal. Similar phenomena was observed for lead removal by biochars made from anaerobically digested bagasse, in which precipitation was shown to be the dominant sorption mechanism ( Inyang et al., 2011b ) Previous studies have demonstrat ed that slow release of negatively charged ions, such as carbonate and phosphate, from biochars can precipitate heavy metal ions,
37 particularly lead ( Cao et al., 2009a ; Inyang et al., 2011b ) Because Pb 2+ has strong chemical affinity with those ions, biochar sorbents may completely remove lead from aqueous solutions when its initial concentration is low. When the initial lead concentration is high, however, Pb 2+ can cons ume all the available anions in solution and thus the isotherms will reach a plateau. Comparisons between solution pH before (i.e., before adding biochar) and after sorption for different initial lead concentrations in the isotherm experiment (Figure 2 5 ) support the surface precipitation mechanism of lead removal by the two biochars. When the initial lead concentrations were low, solution pH increased from about 5 to about 10 after the lead was removed, which is consistent with the alkali nature of the bio chars (Table 2 3). This was because initial lead concentration of lead was not high enough to consume all the carbonate and/or phosphate ions released by the biochars. Because both carbonate and phosphate have strong buffer capacities, their release from b iochars could increase solution pH even if lead precipitation on biochar surfaces might release some H + under certain conditions. When the amount of lead in the solution was greater, the solution pH stayed unchanged or became lower because of the full consumption of these alkali ions (Figure 2 5). The fact that final pH reached the minimum at the same solvent concentration at which isotherm curve reached plateau further confirmed the importance of the surface precipitation mechanism (Figures 2 4 and 2 5 ). The Langmuir (L), Freundlich (F) and Langmuir Langmuir (LL) models were used to simulate the sorption isotherms of lead on the two biochars. These governing
38 equations (Equations 2 4 2.6) can be written as ( Cao et al., 2009a ; Gerente et al., 2007 ) : Langmuir (2 4) Freundlich (2 5) Langmuir Langmuir (2 6 ) W here S max (mmol kg 1 ) is the maximum amount of Pb sorbed ; K (L mmol 1 ) and K f (mmol (1 n) L n kg 1 ) are the Langmuir adsorption constant related to the interaction bonding energies and the Freundlich equilibrium constant, respectively; C (mmol L 1 ) is the equilibrium solution concentration of the sorbate; and n is the Freundlich linearity constant. Th e models fit the experimental sorption data of DWSBC well, but failed to describe that of DAWC (Figure 2 3 ). The best fit model for DAWC isotherm was the Freundlich model, but the r 2 was only 0.416 (Table 2 4 ). The failure of these models was probably due to the higher heterogeneity of DAWC, which was converted from digested dairy waste, a highly heterogeneous feedstock, a mixture of components with different compositions. For lead sorption on DWSBC, the two Langmuir based models (i.e., L and LL models) sim ulated the data better (r 2 > 0.930) than the Freundlich model (r 2 = 0.808) (Table 2 4 ). Although Langmuir based models are developed for weak physical sorption, L or LL model can be used to describe the sorption of metals on biochars through precipitation ( Cao et al., 2009a ; Inyang et al., 2011b ) Inyang et al. (2011b ) indicated that precipitation of lead on biochar derived from anaerobically digested
39 bagasse could be modeled with the L model with a large bonding constant (K = 189 L mmol 1 ). The Langmuir sorption constant for the DWSBC tested in this study was also very high (K = 266 L mmol 1 ), confirming a s trong affinity of lead. The Langmuir sorption capacity of lead sorption on the DWSBC was around 197 mmol kg 1 which is comparable to that of commercial activated carbons (101 395 mmol kg 1 ) and other biochar sorbents (11 680 mmol kg 1 ) ( Beesley & Marmiroli, 2011 ; Cao et al., 2009a ; Inyang et al., 2011b ; Liu & Zhang, 2009 ; Mohan et al., 2007 ; Uchimiya et al., 2010 ) Although the Langmuir model could not be used, a rough estimation of the lead sorption capacity of DAWC directly from the plateau of the isotherm indicated that it shoul d be even higher (> 200 L mmol 1 ) than that of DWSBC (Figure 2 4 ). This confirms that biochars converted from anaerobically digested biomass can be used as effective sorbents to remove lead from aqueous solutions. Post Sorption Characteristics and Sorption Mechanisms SEM image analysis of the post sorption (Pb laden) DAWC and DWSBC revealed the presence of many hexagonal and prismatic crystalline structures on their surfaces (Figure 2 6 ). The corresponding EDS spectra of the SEM image focusing area showed very high peaks of lead element, which demonstrated the presence of lead on surfaces of the post sorption biochars. This strongly suggests the precipitation of lead mineral(s) from aqueous solution onto the biochar surfaces, because both the elemental analysis (Table 2 2 ) and the SEM EDS analysis of pre sorption biochars (Figure 2 7 ) showed no lead or crystals of this t ype in the original biochars. Compared to pre sorption biochars, XRD spectra of post sorption biochars showed several new peaks at specific d values associated with lead minerals, further supporting the precipitation mechanism of lead remo val by the two bi ochars (Figure 2
40 8 ). As discussed above, carbonate and/or phosphate released from the biochars can react with lead in aqueous solution to form stable minerals on biochar surfaces through following reactions in Equations 2 7 2.9: ce rrusite ( 2 7 ) hydrocerrusite ( 2 8 ) pyromorphite ( 2 9 ) W here X can be either F Cl Br or OH Three types of lead minerals, cerrusite (PbCO 3 ), hydrocerrusite (Pb 3 (CO 3 ) 2 (OH) 2 ), and pyromorphite (Pb 5 (PO 4 ) 3 Cl), were identified in the post sorption DAWC, indicating lead removal by DAWC could be controlled by all the three precipitation mechanisms ( E quations 2.7 2.9 ). This was probably because DAWC was converted from a complicated feedstock (digested d airy waste) and thus could release both carbonate and phosphate to react with heavy metals in solution. Because of this heterogeneity, the ability of biochar from digested manure to sorb aqueous heavy metals might tend to fluctuate among samples, which may explain why the three sorption models failed to describe the isotherms of lead sorption on DAWC. In spite of the fluctuation, biochar converted from anaerobically digested dair y waste still has an unexceptionable ability to remove heav y metals from water (Figures 2 4 ). Only one lead precipitate (hydrocerrusite), however, was found on the post sorption DWSBC. Similarly, hydrocerrusite was also found in a recent study of lead sorption through precipitation on biochar converted from anaerobically digested sugarcane bagasse ( Inyang et al., 2011b ) XRD spectra of the pre sorption biochars indicated the existence of calcite ( CaCO 3 ) in both DAWC and DWSBC, which could be
41 the source of carbon ate s release d into the solution. Carbonate minerals such as calcite h ave also been found in other biochars converted from other anaerobically digested biomass ( Yao et al., 2011a ) Anaerobic digestion may concentrate exchangeable cations, such as Ca 2+ K + Na + into residue materials ( Gu & Wong, 2004 ; Hanay et al., 2008 ) T hose cations could react with the dissolved CO 2 during anaerobic digestion to form carbonate minerals during slow pyrolysis ( Inyang et al., 2010 ) Surface organic functional groups of the two biochars converted from anaerobically digested biomass were characterized using FTIR spectroscopy (Figure 2 9 ). Functional group distributions of DAWC and DWSBC were similar to biochars made from other types of digested biomass ( Inyang et al., 2010 ; Yao et al., 2011a ) Although the two biochars had relat ively high surface area s (Table 2 3 ), comparisons of pre a nd post sorption FTIR spectra for both DAWC and DWSBC showed high similarity, which provides no evidence of lead adsorption on biochar through interacting with the surface functional groups. Conclusion This study indicat e d that biochars produced from anaerobically digested biomass ( dairy waste and sugar beet s ) can effectively re move heavy metals from aqueous solutions. The lead sorption capacity of the two biochars used in this study is comparable to that of commercial activated carbons. Thus, biochar converted from anaerobically digested biomass can be used as an alternative sorbent for ac tivated carbon or other water purifiers to tr eat heavy metals in wastewater. High metal removal efficiency of biochars from digested feedstock suggest s that anaerobic digestion could be used as a means of to make high quality biochar based sorbents.
42 Table 2 1 Summary of sources, health effects, and maximum contaminant levels of sel ected heavy metals in water (USEPA). Heavy Metal Source Potential health effect Maximum contaminant level (mg/L) Lead Corrosion of household plumbing, erosion of natural deposits Delays in physical or mental development in children, kidney problems and high blood pressure in adults 0 .0 Copper Corrosion of household plumbing, erosion of natural deposits Gastrointestinal distress over short term, liver and kidney damage over long term 1.3 Nickel Corrosion of steel pipes and metal fittings, erosion of rocks with nickel ore deposits, discharge from electroplating and battery industries. Decreased body and organ weight. < 0.01 Cadmium Corrosion of galvanized pipes, erosion of natural deposits, discharge from metal refineries, runoff from waste batteries Itai itai disease, renal and kidney damage, emphysema, hypertension and testicular atrophy 0.005
43 Table 2 2 Elemental composition (%, mass based) of biochars used in this study Biochar C H N O P K Ca Mg Zn Mn Cu Fe Al Pb DAWC 65.42 0.68 3.63 24.35 0.36 2.33 1.89 0.55 0.02 0.02 a 0.09 0.16 a DWSBC 66.67 1.07 0.43 20.15 0.54 1.51 0.64 5.5 1.08 0.04 0.01 1.54 1.3 a a:<0.01 Table 2 3 Relevant physiochemical properties of biochars used in this study Biochar pH CO 2 Surface Area (m 2 /g) N 2 Surface Area (m 2 /g) Pore Volume (cc/g) Zeta Potential (mv) DAWC 10 .0 555.2 161.2 0.147 29.18 DWSBC 9 .0 128.5 48.6 0.034 15.85
44 Table 2 4 Best fit model parameters of lead removal from aqueous solutions on DAWC and DWSBC Biochar Model Parameter 1 Parameter 2 Parameter 3 Parameter 4 R 2 DAWC First order k 1 = 0.280 q e = 266 0.94 Second order k 2 = 0.00100 q e = 311 0.95 Elovich 174 0.94 Langmuir K = 928 S max = 248 0.0 6 Freundlich K f = 248 n = 0.0619 0.4 2 Double Langmuir K 1 = 1120 S max1 = 234 K 2 = 0.0023 S max2 = 1.90 0.12 DWSBC First order k 1 = 0.181 q e = 203 0.95 Second order k 2 = 0.000760 q e = 250 0.96 Elovich 0.9 7 Langmuir K = 266 S max = 197 0.93 Freundlich K f = 189 n = 0.140 0.8 1 Double Langmuir K 1 = 351 S max1 = 172 K 2 = 2.87 S max2 = 43.9 0 0.94
45 Figure 2 1 Removal of h eavy metals from aqueous solution by the two biochar s converted from anaerobically digested biomass
46 Figure 2 2 Kinetics of lead removal from solution by the two biochar s converted from anaerobically digested biomass A) DAWC and B) DWSBC
47 Figure 2 3 Relation between the amounts of Pb removed by the two biochar s converted from anaerobically digested biomass A) DAWC and B) DWSBC and square root of time before equilibrium.
48 Figure 2 4 Isotherms of lead removal from solution by the two biochar s converted from anaerobically digested biomass A) DAWC and B) DWSBC
49 Figure 2 5 Changes in solution pH during lead removal from solution by the two biochar s converted from anaerobically digested biomass A) DAWC and B) DWSBC
50 Figure 2 6 SEM image (left) and corresponding EDS spectra (right) of post sorption lead loaded digested biochars. A) DAWC and B) DWSBC at 10,000 X. The EDS spectra were recorded at the location shown in the SEM image
51 Figure 2 7 SEM image (left) and corresponding EDS spectra (right) of pre sorp tion digested biochars. A) DAWC and B) DWSBC at 500 0 X. The EDS spectra were recorded at the same location shown in the SEM image.
52 Figure 2 8 XRD spectra of pre and post sorption digested biochars. A) DAWC and B) DWSBC
53 Figure 2 9 FTIR spectra of pre and post sorption digested biochars. A) DAWC and B) DW SBC
54 CHAPTER 3 FILTRATION OF ENGINEERED NANOPARTICLES IN CARBON BASED FIXED BED COLUMNS 1 Introduction Engineered nanoparticles (ENPs) have become the foundation of a novel br and of technology, impacting consumer products, manufacturing techniques, and material usages ( Albrecht et al., 2006 ) This can be attributed to their intrinsic properties, such as high surface area to volume ratio, small size (1 100nm), and unsaturated surface atoms that readily bind to other atoms ( Christian et al., 2008 ; Ghaedi et al., 2012b ) In addition, most ENPs exhibit various quantum effects such as resonance, optical properties, mechanical strength, thermal and electrical conductivity that can be exploited in the development of various household and industrial applications ( Nowack & Bucheli, 2007 ) The projected global demand for nano material products is expected to reach $1 trillion dollars in 2015 ( Eckelman et al., 2008 ; Nowack & Bucheli, 2007 ) which will increase loadings of ENPs to the environment and further pose a risk to soil and groundwater systems. ENPs may be released into the environment from both point sources (e.g., production facilities, landfills, and wastewater treatment pla nts) and non point sources (e.g., accidental spills and wear from ENP products) ( Nowack & Bucheli, 2007 ) As a result, the occurrences of ENPs in aquatic systems, parti cularly from municipal discharges and wastewater treatment plants, have been reported in several recent studies ( Baun et al., 2008 ; Isaacson et al., 2009 ; Upadhyayula et al., 2012 ) Because of 1 R eprint with permission from Inyang, M., Gao, B., Wu, L., Yao, Y., Zhang, M., Liu, L. 2013. Filtration of engineered nanoparticles in carbon based fixed bed columns. Chemical Engine ering Journal, 220(0), 221 227.
55 the potential toxic effect of ENPs to aquatic ecosystems ( Petersen et al., 2010 ; Tervonen et al., 2009 ; Wiesner et al., 2009 ) it is crucial to develop cost effective treatment technologies to remove them from water systems. To our knowledge, however, only little resea rch has been conducted to study the removal of aqueous ENPs, particularly with respect to evaluating the removal ability of carbon based filters to ENPs in aqueous solutions. Biochar is pyrogenic black carbon derived from the thermal degradation of carbon rich biomass in an oxygen limited environment. Recent studies have demonstrated that biochars can be used as low cost adsorbent s to remove various contaminants from water ( Inyang et al., 2011b ; Inyang et al., 2012 ; Xue et al., 2012 ; Yao et al., 2011a ; Yao et al., 2011b ) It is estimated that the production cost of biochar from a typical biomass is around 0.076 U.S. dollar per kilogram ( Yoder et al., 2011 ) much lower than other commercial carbon based adsorbents including activated carbon (1.44 2.93 U.S. dollar per kilo gram) ( Lima et al., 2008 ) This promising new carbon material (particularly after further modifications) could be used as poten tial filter media to remove ENPs from water, although fu rther testing is still needed. Most biochars are predominantly negatively charged ( Inyang et al., 2010 ; Yao et al., 2012c ) and may readily bind positively charged ENPs through electrostatic attractions. Plant and animal derived biochars produced at relatively high temperatures (> 400 o C) may also contain disordered electrons for electrostatic attractions and bonding of ENPs on oxidized graphene sheets in biochars ( Keiluweit & Kleber, 2009 ) The possibili ty of retaining ENPs on biochars is further enhanced by the presence of functional groups of different surface charges that co exist
56 within the outer surface and pores of biochars These functional groups could coordinate, or complex with ENPs to sequester them on biochar surfaces. Today, several modification/engineering methods have been recently developed to improve the retention of biochar based sorbents to aqueous contaminants ( Xue et al., 2012 ; Yao et al., 2011a ) which could also be ap plied for the removal of ENPs. Nevertheless, most current studies have focused on the use of activated carbons and other high tech sorbents to remove ENPs from water ( Ghaedi, 2012 ; Ghaedi et al., 2012a ; Ghaedi et al., 2012b ; Marahel et al., 2012 ; Yao et al., 2012a ) and, there is no research effort devoted to the development of a biochar based technol ogy for the removal of aqueous ENPs. Filtration and transport of ENPs in porous media, particularly in artificial soil columns packed with quartz sand, have been recently investigated ( Lecoanet & Wiesner, 2004 ; Tian et al., 2010 ; Tian et al., 2012c ; Tian et al., 2011 ; Wang et al., 2012a ) Findings from those studies have demonstrated that the retention and transport of ENPs in porous media are controlled by several factors, such as surface charge, particle shape and size, and solution chemistry ( Lecoanet & Wiesner, 2004 ; Tian et al., 2012d ; Tian et al., 2011 ) The Derjaguin Landau Verwey Overbeek (DLVO) theory has been applied to quantify the attractive and repulsive interaction forces between ENPs and porous medium grains ( Tian et al., 2010 ) In addition, it has been reported that theories and models developed for simulating the filtration and transport behaviors of colloidal particles in porous media can be modified or applied directly to that of ENPs ( Tian et al., 2010 ) Thus, it is anticipated that those t heories and models could also be used to describe the filtration of ENPs in carbon base d filters (porous media).
57 The overarching objective of this study was to evaluate and compare the effectiveness of several carbon based materials, including biochar, in retaining three types of ENPs: nano titanium dioxide (NTiO 2 ), silver nanoparticle (AgNP), and multi walled carbon nanotubes (CNT). These nanoparticles are among the most popular ENPs employed in industrial and household applications ( El Sheikh et al., 2007 ; Ghaedi, 2012 ; Leonard & Setiono, 1999 ; Li et al., 2011a ; Lu et al., 2009 ; Sumesh et al., 2011 ) A biochar produced from hickory wood (HC), an activated carbon (AC), and Fe modified HC and AC were used as carbon sorbents in both batch and fixed bed settings to test their sorption of the three ENPs. Simulations of mathematical models (i.e., DLVO theory and colloid transport model) were used to help data analysis and aid in the interpretation of experimental results. The specific objectives of this study were to: (a) evaluate and compare the ability of the carbon materials to sorb and filter the ENPs, (b) compare the mobility of the three ENPs in the carbon filters, and (c) determine whether biochar can be used as an effective filter media for the ENPs. Materials and Methods Materials NTiO 2 AgNP, and CNT nanoparticle powders were obtained from Sinonano Company (China), Particle Engineering and Research Center (University of Florida), and Shenzhen Nanotech Port Co. (China), respectively. Their solutions were prepared by adding 20 mg each to 200 mL of de ionized (DI) water and sonicated in an ultrasound homogeniz er (Model 300 V/T, Biologics, Inc.) for 1 h at pulse intervals of 12 min s. The resulting suspensions were used as stock solutions for subsequent batch sorption and filtration studies. Zeta potential and the effective diameters (i.e., hydrodynamic diameters ) of the nanoparticles were measured with a 10 ppm
58 suspension diluted from the stock (same solution chemistry) using a Brookhaven Zeta Plus (Brookhav en Instruments, Holtsville, NY) and follow ed the procedures of previous studies ( Tian et al., 2010 ; Wang et al., 2012a ) Hickory chips w ere obtained from the UF North Florida Research and Education Center as feedstock for biochar production. About 500 g of the dried hickory chips were converted into HC through slow pyrolysis at 600 o C for 2 h in a nitrogen environment in a furnace (Olympic 1823HE), following the procedures of ( Inyang et al., 2012 ) Granular AC (coconut shell, steam activated) was purchased from Fisher Scientific (Suwanee, Georgia). All other chemical reagents employed in this study were of high purity grade, from Fisher Scientific, Suwanee, Georgia. The HC and AC samples were ground and sieved to 0.5 1 mm sized particles. After several rinses with deionized dis tilled water, both HC and AC were dried at 80 o C for further testing and Fe modification. The iron modified carbons (i.e., HC Fe and A C Fe) were produced using previously reported method ( Chen et a l., 2007 ; Thirunavukkarasu et al., 2003 ) Briefly, 2g each of HC and AC were added to 8 mL of 2 M Fe(NO 3 ) 3 solution, about 3ml of 10 M NaOH was then added to create an iron precipitate on the carbon surfaces. The mixtures were dried at 105 o C and then rinsed with DI water before use. Quartz sand (1.3mm sized) was obtained from Standard Sand and Silica Co as reference filter media. The sand was washed sequentially with t ap water, 10% nitric acid and deionized water, and baked at 550 o C to remove metal oxides and organic impurities before use ( Tian et al., 2010 ) Batch Sorption An initial evaluation of the sorption ability of HC, HC Fe, AC, AC Fe and sand to the nanoparticles was performed in batch experiments. About 0.1 g of each sorbent was added into 68ml digestion vess els (Environmental Express), and mixed with 50ml of 10
59 ppm ENP solutions at room temperature of (22 0.5 o C). The sample solutions with their corresponding blanks and experimental controls (without sorbent or sorbate) were agitated for 3 h on a reciprocat ing shaker, and withdrawn at the end of 3 h to examine their sorption capacities on a UV VIS spectrophotometer. Measurements of NTiO2, AgNP, and CNT concentrations on the UV VIS spectrophotometer were conducted at wavelengths of 655 nm, 645 nm, and 255 nm, respectively ( Tian et al., 2010 ; Wang et al., 2012a ) ENP concentrations on the sorbents were calculated based on the differences between i nitial and final aqueous solutions. The experime nts were performed in duplicate and average values were used in the analysis. Column filtration The carbon sorbents were wet packed with sand in laboratory columns measuring 1.56 cm in diameter and 5.6 cm in height following the procedures of Xue et al. ( Xue et al., 2012 ) About 5.5 g of sand was used at each end of the col umn to help distribute the flow, and the carbon sorbent was sandwiched between the sand inside the column. The heights of the lower sand, carbon and upper sand layers in the column were 2.34 cm, 1.28 cm, and 2.50 cm respectively. Columns packed with sand only were also used in the experiment. For each experiment, the column was first flushed with DI water for 2 h to equilibrate it. A peristaltic pump (Masterflex L/S, Cole Parmer Instrument, Vernon Hills, IL) was then connected to the influent (bottom) of the column to maint ain an upward flow rate of 1 ml min 1 The filtration experiment was initiated by switching the influent to a 10 ppm ENP solut ion for 3 h followed by 2 h of DI water flushing. Effluent samples from the columns were collected with a fraction collector (IS 95 Interval Sampler, Spectrum Chromatography, Houston, TX) during the experiment to determine the ENP concentrations with the U V VIS spectrophotometer.
60 Characterizations Carbon, hydrogen, and nitrogen contents of the carbon sorbents were determined using a CHN elemental analyzer (Carlo Erba NA 1500) via high temperature catalyzed combustion followed by infrared detection of the re sulting CO 2 H 2 and NO 2 gases. Major inorganic elemental constituents, pH, zeta potential, and surface area of all the sorbents were determined using previously reported methods ( Inyang et al., 2012 ; Yao et al., 2011a ) Bulk density of the carbon sorbents and sand materials were determined using the tap and fill method reported previously ( Abdullah & Wu, 20 09 ) X ray diffraction (XRD) analysis was conducted on these HC, HC Fe, AC, AC Fe samples to identify possible crystalline structures. A computer controlled X ray diffractometer (Philips Electronic Instruments) equipped with a stepping motor and graphit e crystal monochromator was used to obtain diffraction patterns. Mathematical Models The filtration and transport of the ENPs in fixed bed columns was described by the advection dispersion equation (ADE) based on the colloid filtration theory ( Yao et al., 1971 ) : ( 3 1) ( 3 2) W here C is the sorbate concentration in po re water (mg L 1 ), t is the time (min), b is the medium bulk density (g L 1 ), is the dimensionless volumetric moisture content (porosity), S is the adsorbed particle concentration (mg g 1 ) z is the distance travelled in the direction of the flow (cm), D is the dispersion coefficient (cm 2 min 1 ), v is the average
61 linear pore water velocity (cm min 1 ), and k d is the removal or deposition rate constant (min 1 ). Equation ( 3 1) was solved numerically with a zero initial concentration, pulse input and a zero concentration gradient boundary conditions for the carbon layer. The Levenberg Marquardt algorithm was used to estimate the value of the model parameters by minimizing the sum of the squared differences between model calculated and measured effluent ENPs concentrations over multiple calculation interactions. The classic DLVO theory was used to determine the interaction energies between the ENP and filter media ( Shen et al., 2007 ) The Lifshitz van der Waals LW ) (Equation 3 3) and the electric double layer repulsion energy EDL ) for a sph ere plate system (Equation 3 4) were used to determine the total DLVO energy between the ENPs and sorbents ( Tian et al., 2010 ; Vanoss et al., 1990 ) : (3 3 ) ( 3 4 ) ( 3 5 ) W here A is the Hamaker constant, h is the separation distance, r is the radius of the permittivity (8.854*10 12 C 2 N 1 m 2 23 C 2 J K 1 ), T is the temperature, z is the valence electrolyte, e is the electron charge (1.602*10 19
62 sand surface) length. Results and Discussion Properties of ENPs and Carbons Elemental analysis of the carbon sorbents (Table 3 1) showed that AC, HC, AC Fe, and HC Fe had varied amounts of inorganic constituents with more predominant amounts of carbon observed. In particular, elemental iron content in the raw carbon materials (AC and HC) was observed to significantly increase by a 100 fold after impregnation with iron in AC Fe and HC Fe. Slightly high er am ounts of iron were noted in AC Fe (2.33 %) than in HC Fe (1.03 %), probably because AC (pH 7.1) is less basic than the HC (pH 8.5). Recently, Nieto Delgado and Rangel Mendez ( Nieto Delgado & Rene Rangel Mendez, 2012 ; Ofir et al., 2007 ) suggested that irons may be more effectively anchored on acidic carbons than on basic ones Typically, the precipitation of Fe (III) on carbon surfaces is known to enhance their interaction with negatively charged contaminant species, such as phosphate, arsenic, and colloids (including ENPs) ( Chen et al., 2007 ; Nieto Delgado & Rene Rangel Mendez, 2012 ; Ofir et al., 2007 ; Wang et al., 2012b ) Moreover, the precipitation may promote the electron transfer of O 2 from the aqueo iron complex to the metal cation (i .e., Fe), weakening the O H bond of the complex, to release protons into the reaction media ( Nieto Delgado & Rene Rangel Mendez, 2012 ; Ofir et al., 2007 ) As a result, the pH of AC Fe (4.7) and HC Fe (4.9) was observed to be lower than that of their original carbons. The zeta potential values of AC Fe and HC Fe were also less negative than the AC and HC (Table 3 2), confirming the presence of iron particles on carbon surfaces. All ENPs used in this study were predominantly negatively charged at
63 their original pH (Table 3 2). Surface areas and pore volumes of the ENPs were much smaller than those of the carbon materials, suggesting that the ENPs could also be physically sorbed on the surface sites within the carbon matrix. Batch S orption Batch experiments demonstrated that the carbons were better sorbents than the pure sand (Figure 3 1). In addition, iron impregnation improved the sorption of the ENPs on the c arbon sorbents. This improvement was more significant in the removal of CNT (Figure 3 1 a) and NTiO 2 (Figure 3 1 b) than in the removal of AgNPs by the iron modified carbons (Figure 3 1 c). The raw carbons showed lower removal of the ENPs and the lowest re moval was observed for CNT, possibly because of the surface similarities between the carbon sorbents and the CNT (both surfaces are highly negatively charged). The removal of the three ENPs by the Fe modified biochar (HC Fe) (with lower loadings of Fe) was the highest and was even higher than that of Fe modified activated carbon (AC Fe). This indicated that the biochar based sorbent could be a better (more cost effective) option to be used in the filter for ENPs. Although the removal efficiency trends of EN Ps by carbon materials can be evaluated from batch sorption studies, batch experiments usually are conducted under optimized conditions, For example, although the so lution chemistry of the batch and column experiments was identical, the contact time between ENPs and adsorbents of the adsorption experiments ( i.e. 3 h) was much longer than that of the filtration tests (less than 3 min s). ENP Filtration and Transport in Fixed Bed Columns The transport of the ENPs in the fixed bed columns (Figure 3 2 a c ) showed rapid breakthrough responses and the effluent concentrations reached a plateau about
64 30 mins after the ENP injections. The breakthrough curves returned to the b aseline level after the columns were flushed with DI water, reflecting the completion of the breakthrough process. P eak concentrations of ENPs in the columns packed with unmodified carbons (i.e. HC or AC) were similar to or even higher than that in the ref erence sand columns, indicating the HC or AC could not improve the ENP filtration in the fixed bed columns. This result is different from findings in the batch sorption experiments, probably due to difference in sorption dynamics of the t wo systems as disc ussed above. Except the AC Fe filters for AgNP, performances of the iron modified carbon (i.e., AC Fe or HC Fe) filters were better and the peak breakthrough concentrations of the ENPs were lower than that of other sorbents. This further confirmed that iro n modification can improve the removal of ENPs by the filters Previous studies have reported that the interaction between impregnated me tal oxyhydroxides, including Fe (OH) 3 and colloidal/nanosized particles can increase intra particle bridging and reduce the electron double layer repulsions to facilitate particle deposition in filter media ( Ofir et al., 2007 ; Yao et al., 1971 ) Among all the filters, HC Fe s howed the best filtration performance for all the ENPs, which is consistent with the batch sorption experimental data. T he iron modified biochar is a bett er ENP filter material than AC Fe and can be used to remove ENPs from water. Among the three ENPs, CNT showed the highest peak breakthrough concentrations for all the tested experimental conditions, which corresponded to the findings of the batch study. The poor interactions between CNT and pyrolyzed carbons (AC and HC) may have resulted from high amorphic ity and relatively few graphitic sites on the carbons that could have ( Keiluweit & Kleber, 2009 ) In
65 addition, previous studies have also indi cated that tubular CNTs may have higher mobility in porous media than spherical ENPs because they can orient parallel to the streamlines in the flow to reduce their retention ( Tian et al., 2012a ; Tian et al., 2011 ) Under the tested experimental conditions, surfaces of the three ENPs and carbon materials were all negatively charged (T able 3 2). The solution chemistry therefore was unfavorable for the attachment of the nanoparticles to the carbon surfaces ( Tian et al., 2012d ; Wang et al., 2012a ) which explains why the HC and AC enabled columns showed no difference to t he reference sand column. Although, the presence of iron hydroxides on the carbon surfaces did not alter the overall surface charge to positive (Table 3 2), it greatly reduced the surface potential, to promote the deposition of nanoparticles on carbon surf aces (enhance attachment efficiency) ( Morales et al., 2011 ; Wu et al., 2012b ) Because, iron hydroxides are positively charged under mos t practical circumstances, they may also introduce charge heterogeneity to the carbon surfaces, which could serve as the sorption sites dominating the filtration and transport of nanoparticles in the fixed bed columns ( Tian et al., 2010 ; Tian et al., 2012c ) Modeling of ENPs Filtration and Transport The ADE model reproduced the e xperimental data closely with good coefficients of correlation (R 2 > 0.90) (Table 3 3 ). The model estimated removal rate constants (k d ) for the ENPs in various filter media as ranging between 0.04 0.07 min 1 for NTiO 2 and 0.05 0.07 min 1 for AgNP, with least retention rates for CNT ranging from 0.005 0.02 min 1 further supporting the dis cussion above. As shown in T able 3 3 the kd values of the iron modified carbons were generally much higher than that of the other sorbents, except for the kd of AgNP removal by the AC Fe. The kd values of HC Fe were the
66 highest for each nanoparticle, further confirming that the iron modified biochar can be used as an alternative, low cost adsorbent for the treatment of ENPs in wastewater. The optimized filter length (L o = ln where v (cm min 1) is the fluid pore velocity and C/C o is 0.001), at which 99.9% of the ENPs are filtered from the solution ( Wang et al., 2008 ) was calculated using the model estimated kd. Compared to other ENPs, the estimated L o values were highest (422 1450 cm) for CNT (Table 3 3), indicating it requires more filter material for its complete removal, especially for the unmodified carbons. The L o values for NTiO 2 and AgNP ranged between 141 217 cm and 118 170 cm, respectivel y, indicating these two ENPs require less amount of filter media to remove them from solution. Although the HC Fe was the most effective, it still requires the filter to be designed at 116, 118, and 422 cm to remove NTiO 2 AgNP, and CNT, respectively. DLVO interaction energy profiles were calculated to evaluate the relative contributions of van der Waals and electrostatic interactions to the interactions between the ENPs and carbon or sand surfaces. The Hamaker constants of the van der Waals interactions be tween the ENPs and the filter media in water were determined from previously reported individual Hamaker value of pyrolyzed carbon (6x10 20 J) ( Maurer et al., 2001 ) sand (8.8x10 20 J) ( Tian et al., 2010 ) CNT (8.2x10 20 J) ( Tian et al., 2010 ) NTiO 2 (6x10 20 J) ( Butt et al., 2005 ) and for AgNP (38.5x10 20 J) ( Butt et al., 2005 ) Electr olyte concentrations of the filtrate solutions were assumed as 0.001 M for raw carbons and quartz sand, while 0.01 M was assumed fo r the iron impregnated carbons. For all the ENPs, the interaction energy profiles were characterized by the absence of an attractive primary minimum and the presence of high energy barriers in
67 both sand and carbon media (Figure 3 3). In particular, the highest energy barriers were observ ed for CNT transport in both sand and carbon media ranging between 135 235 KT (Figure 3 3 a), with no obvious secondary minimums observed for the unmodified carbons. The presence of such high energy barrier would limit the deposition of the CNT on the un modified carbon surface, which is consistent with the batch and column experimental data. A deep secondary minimum well for CNT was observed for the iron modified carbons (HC Fe and AC Fe) at separation distance of 7 nm (Figure 3 3 a), suggesting that the CNT can attached to the iron modified carbon surfaces through secondary minimum deposition ( Tufenkji & Elimelech, 2005 ) Because effective diameter (hydrod ynamic diameter) was used in the calculations to determine the DLVO interaction between tubular CNTs and the filter media, the results may not reflect the actual interactions and may overestimate the repulsive forces ( Tian et al., 2012d ; Wang et al., 2008 ) Previous studies of CNT transport in porous media, however, suggested that this approach might be used as exploratory estimations ( Tian et al., 2010 ; Tian et al., 2012c ; Tian et al., 2011 ; Wang et al., 2008 ) A new or modified DLVO theory is thus necessary to better describe the filtration of CNTs in carbonaceous filters. Much lower energy barriers were o bserved for both NTiO 2 (15 40 KT, Figure 3 3 b) and AgNP (15 60 KT, Figure 3 3 c) to the sorbents, which correlates with earlier findings of higher filtration of NTiO 2 and AgNP in the columns. Similarly, the DLVO energy profiles for AgNP and NTiO 2 (Fig ure 3 3 b and c) did not show any obvious secondary minimums for the deposition of AgNP and NTiO 2 on the raw carbons (HC and AC), but showed shallow secondary minimum wells at separation distance of ~20 nm for the iron modified carbons. This also suggested that AgNP and NTiO 2 can attach
68 to the iron modified carbon surfaces through secondary minimum deposition ( Tufenkji & Elimelech, 2005 ) In addition to electrostatic interactions, the existence of other possible mechanisms was further probed using XRD analysis ( Figures 3 4 ). XRD patterns, however, did not show any crystalline structures on the post filtration samples (Figures 3 10 and 3 11). This result c ould rule out the possibility of any precipitation mechanisms for the attachment of the ENPs on the carbon surfaces. Conclusion The removal efficiencies of unmodified and iron impregnated carbons to three ENPs were evaluated with both experimental and mode ling investigations. The results indicated that iron impregnation improved the removal ability of the c arbons for the ENPs. Among all the carbon sorbents, the iron modified biochar was the best filter material, suggesting that biochar based sorbents can be used in low cost filters for ENP removal. Because CNT showed high mobility in all the carbon enabled filters, additional investigations are still need ed to further modify the biochar to enhance its ability to remove CNTs from water.
69 Table 3 1 Elemental composition of carbon materials Sample P% K% Ca% Mg% Zn% Mn% Cu% Fe% Al% Ni% Pb% C% H% N% O% HC 0.02 0.24 0.82 0.13 a a a 0.01 0.06 a a 81.81 2.17 0.73 14.02 HC Fe 0.02 0.03 0.43 0.16 a 0.02 a 1.03 0.03 a a 80.54 1.65 1.34 16.48 AC 0.01 0.09 0.17 0.08 a a a 0.02 a a a 86.07 0.12 1.17 17.05 AC Fe a 0.02 0.14 0.04 a a a 2.33 0.03 a a 75.06 0.61 0.57 23.77 a < 0.01% Table 3 2 Physiochemical properties of filter materials and engineered nanoparticles (ENPs) BET N2 Sample pH Surface area (m 2 /g) Pore Volume (cc/g) Zeta Potential (mv) Effective diameter (nm) Bulk density (g/cm 3 ) HC 8.5 431 .0 0.2 0 43.7 n.d 0.4 HC Fe 4.9 12.5 0.00 18.8 n.d 0.4 AC 7.1 956.2 0.3 0 28.9 n.d 0.6 AC Fe 4.7 1090 .0 0.03 19.8 n.d 0.5 Sand 6.5 nd nd 40.4 n.d 1.4 CNT 6.6 142 .0 0.1 0 46.3 140 .0 n.d TiO2 6.8 58.8 0 .00 11 .0 90 .0 n.d AgNP 7.9 30.7 0 .00 35.1 52.4 n.d nd not determined
70 Table 3 3 Best fit model parameters for ENP transport in various filter media Filter Media Nanoparticles K d (min 1 ) R 2 Maximum column length L max (cm) Sand NTiO 2 0.045 0.992 173 HC NTiO 2 0.038 0.961 206 HC Fe NTiO 2 0.067 0.995 116 AC NTiO 2 0.036 0.996 217 AC Fe NTiO 2 0.055 0.991 141 Sand CNT 0.006 0.965 1430 HC CNT 0.008 0.908 1045 HC Fe CNT 0.019 0.933 422 AC CNT 0.005 0.997 1450 AC Fe CNT 0.014 0.994 567 Sand AgNP 0.046 0.930 170 HC AgNP 0.062 0.989 129 HC Fe AgNP 0.066 0.985 118 AC AgNP 0.053 0.993 147 AC Fe AgNP 0.048 0.993 166
71 Figure 3 1 Removal efficiency of EN Ps in batch sorption study. A) CNT, B) NTiO2, and C) AgNP
72 Figure 3 2 Filtration and transport o f ENPs in fixed bed columns. A) CNT, B) NTiO 2 and C) AgNP
73 Figure 3 3 Derjaguin Landau Verwey Overbeek (DLVO) energy interactions b etween filter media and ENPs A) CNT, B ) NTiO 2 and C) AgNP.
74 Figure 3 4 XRD patterns for A) raw and post filtration carbons loaded with AgNP and B) raw and post filtration carbons loaded with NTiO 2 Minerals detected were peak labeled as C for calcite (CaCO 3 ), A for anatase (TiO 2 ), Q for quartz (SiO 2 ), and Ag for metallic silver (Ag)
75 CHAPTER 4 SYNTHESIS, CHARACTERI Z ATION AND DYE SORPTIO N ABILITY OF CARBON N ANOTUBES COATE D BIOCHAR COMPOSITES Introduction The use and release of organic dyes in many industrial products are a threat to water systems ( Iriarte Velasco et al., 2011 ) The complex aromatic structure of dyes makes them of low biodegradability and stable toward light and chemical treatments ( Ai & Jiang, 2012 ) Methylene blue (3,7 bis(Dimethylamino) phenothiain 5 ium chloride) is a cationic dye found in many industrial effluents that may induce aesthetic, and more importantly health problems such as cancers, reproductive and neurological disorders in humans and aquatic organisms ( Yan et al., 2011 ) A number of treatment techniques including ionic exchange, adsorption, coagulation, membrane filtration and photo catalysis have been extensively tes ted for the removal of dyes from wastewater ( Cheng et al., 2012 ; Kannan & Sundaram, 2001 ; Lee et al., 1999 ; Malakootian & Fatehizadeh, 2010 ; Ramkumar et al., 2010 ) Among these methods, adsorption is known to be a more economical and simple treatment approach ( Ai & Jiang, 2012 ) Thus, research on low cost, high capacity adsorbents for organic dyes is increasing ( Ma et al., 2012 ) Biochar is a low cost, porous, carbon rich product derived from the thermal degradation of organic matter in an oxygen limited environment ( Lehmann, 2007 ) The benefits of employing eco friendly biochar in wastewater treatment technologies have already been established ( Inyang et al., 2012 ; Inyang et al., 2011c ; Kasozi et al., 2010 ; Xue et al., 2012 ; Yao et al., 2011c ; Zhang & Gao, 2013 ) In addition, a recent study showed that ENPs may bind to biochar surfaces (particularly after modification) to a greater extent than to commercial activated carbons ( Inyang et al., 2013 ) Thus, marrying existing b iochar technology with emerging nanotechnology to create hybrid
76 biochar nano composites, has great potential to create a new class of environmentally friendly and cost effective sorbents to treat a wide array of contaminants ( Yao et al., 2011a ; Yao et al., 2013b ; Zhang et al., 2013a ; Zhang et al., 2013b ; Zhang et al., 2012a ; Zhang et al., 2012b ) Carbon nanotubes (CNTs) are cylindrical tubes of graphene material that exhibit exceptional properties su ch as ultra low weight, high mechanical strength, and thermal and chemical stability ( Zhang et al., 2009 ) The potential use of CNTs as adsorbents has generated much interest ( Ai & Jiang, 2012 ; Li et al., 2011b ; Tian et al., 2013a ; Ti an et al., 2013b ; Tian et al., 2012b ) because the hollow, l ayered structure of CNTs endows them with characteristically high specific surface areas and correspondingly high sorption capacities for various contaminants ( Ma et al., 20 11 ; Tian et al., 2012b ) Moreover, chemically functionalized CNT surfaces, grafted with specific functional groups (carbox yl, hydroxyl, amine, fluorine) provide high affinity sorption sites for increased electron bonding ( Ma et al., 2012 ; Theodore et al., 2011 ; Wang, 2009 ) Despite these sorptive properties, practical applicati on of CNTs remains limited by its poor solubility, and rapid aggregation in its native state ( Lee et al., 2008 ) Several research efforts have been made to overcome these limitations, by loading CNTs on sorptive supports using sol gel ( Es'haghi et al., 2011 ) crosslinking agents ( Salipira et al., 2008 ) and carbon vapor deposition (CVD) growth techniques ( Huang et al., 2012 ; Zhang et al., 2009 ) However, the high cost and formation of by products with many of these methods has made it necessa ry to consider other supports. Thus, bio char is examined here as such a potential, low cost support for CNTs.
77 The overarching objective of this study was to develop a simple method to synthesize hybrid CNT biochar nanocomposite material s and test their potential applications. Our specific objectives were to: (1) characterize the CNT biochar nanocomposite, (2) determine the effects of CNT hybridization on the physiochemical properties of the biochars, ( 3) examine the influence of pH and ionic strength conditions on the sorption of MB on the CNT biochar nanocomposite, and (4) elucidate and understand the interaction mechanisms governing the sorption of MB on the C NT biochar nanocomposite Materials and Methods Materials Carboxylic acid functionalized multi walled CNTs with diameters ranging 1 0 20 nm were purchased from the Sinonano Company (P. R. China). Hickory chips and sugarcane bagasse biomass were obtained from the North Florida Research and Education Center of the University of Florida. The biomass feedstocks were dried and milled to 500 m size fraction. Methylene blue (C 16 H 18 ClN 3 S, molecular weight, 319.86 g) and other chemicals employed in this study were of analytical grade and obtained from Fisher Scientific, Georgia. Preparation of CNT biochar N anocomposite CNT suspensions were prep ared by adding either 20 mg (0.01% by weight) or 2 g (1% by weight) of CNT powder to 200 ml of deionized (DI) water. The CNT suspensions were sonicated in an ultrasound homogenizer (Model 300 V/T, Biologics, Inc.) for 1 h at pulse intervals of 12 min. The resulting suspensions were designated as CNT 0.01% and CNT 1%, respectively, and used for the preparation of CNT biochar nanocomposite.
78 Milled hickory chips and sugarcane bagasse biomass (feedstocks) were converted to CNT biochar nanocomposite following a dip coating procedure ( Schoen et al., 2010a ; Zhang et al., 2012a ) Specifically, 10 g of each feedstock were placed in 100 ml of the CNT suspensions and stirred for 1 h using a magnetic stirre r at 500 rpm, after which, the dip coated CNT treated feedstocks were removed and oven dried at 105 o C. Next, the dried CNT treated feedstock were each placed in a quartz tube, inside a tubular furnace (MTI, Richmond, CA) and pyrolyzed at 600 o C for 1 h in a flowing N 2 environment. In addition, untreated feedstocks were also converted into biochars using the same pyrolysis conditions. The resulting biochars produced were designated as hickory chips (HC), CNT modified hickory chips (HC CNT 0.01% and HC CNT 1 %), sugarcane bagasse (BC), and CNT modified sugarcane bagasse (BC CNT 0.01% and BC CNT 1%). All biochars were rinsed with distilled, de ionized water several times; oven dried, and sealed in glass containers for subsequent testing. Characterization Elemen tal carbon, hydrogen, nitrogen and oxygen content (C, H, N, and O); zeta potential, pH, and surface areas of the sorbents were determined using previously reported methods ( Inyang et al., 2012 ; Yao et al., 2011a ) Thermogravimetric analysis (TGA) was performed in a stream of air at a heating rate of 10 o C min 1 with a Mettler TGA/D SC1 analyzer (Columbus, OH) to test the t hermal stability of the samples T he morphology of modified CNT biochar composites was examined by transmission electron microscopy (JEOL 2010F TEM). Physiochemical features of the samples were investigated by Raman spectroscopy (Renishaw Bio Raman).
79 Sorption of Methylene B lue An initial evaluation of the sorption ability of the chars was conducted using MB solution in batch sorption experiments. About 25 mg of each test biochar was mixed in 50 ml digestion vessels ( Environmental Express) with 12.5 ml of 20 mg L 1 MB solution at room temperature (22 0.5 o C). The sample solutions and their corresponding blanks and experimental controls (without either sorbent or sorbate) were agitated for 24 h on a reciprocating shak er, then filtered through 0.22 m pore size nylon membrane (GE cellulose nylon membranes). Measurements of MB concentrations in the filtrates were determined using a Thermo Scientific EVO 60 UV VIS spectrophotometer at a wavelength of 665nm. Sorbed amounts of MB on test biochars were calculated as the difference between the initial and final aqueous MB solution concentrations. Sorption experiments were conducted in duplicate and the average values are reported here. Following the initial evaluation experime nts, sorption kinetics and isotherm studies of MB sorption on unmodified biochar (HC and BC) and CNT biochar nanocomposites (HC CNT 1% and BC CNT 1%) were conducted. To examine sorption kinetics, 25 mg of each sorbent was mixed with 12.5 ml of 20 mg L 1 MB solution in 50 ml digestion vessels at room temperature. The sample solutions and their corresponding controls were withdrawn from the agitator at time intervals of about 1 h up to 24 h and filtered through 0.22 m pore size nylon membranes for measuremen ts. The pH of the filtered sample solutions were noted prior to and after sorption experiments. Sorption isotherms were obtained by adding 25 mg of each biochar to 12.5 ml, MB solutions of varying concentrations (5 80 mg L 1 ). Sorption kinetics and isoth erm experiments were performed in triplicate and the average results are presented with standard deviations.
80 Effect of pH and I onic S trength The effect of pH on MB sorption by the CNT modified and unmodified biochar sorbents was evaluated by adding 25 mg of each biochar to 12.5 ml of 20 mg L 1 MB solutions in 50ml digestion ves sels with pH condition ranging 2 10, adjusted by adding aqueous solutions of either 0.1 M NaOH or 0.1mM HCl. To study the effect of ionic strength, pre determined amounts of NaCl were added to obtain 0.01 M, 0.05 M, 0.1 M, and 0.5 M ionic strength solutions. The sample mixtures and their corresponding blanks were agitated for 24 h, and then filtered and treated as described above. Sorption experiments at different pH or ionic stren gth were conducted in triplicate. Results and Discussion Biochar P roperties The properties of both hickory (HC) and bagasse biochars (BC) were generally improved by the addition of 1% CNTs (Table 4 1). In particular, the surface areas of HC CN T 1% and BC C NT 1% were about 1.2 and 40 times greater than their unmodified control bioc hars (HC and BC) surface areas respectively, suggestin g more CNTs were anchored to BC than to HC. In addition, the pore volume of the sorbents also increased with the addition of the CNTs (Table 4 1 ), indicating the CNT pretreatment could potentially increase the porosity of the biochars. Results of the zeta potential measurements showed that the surfaces of the hybrid biochar sorbents also became increasingly negatively charged wi th increasing amounts of CNTs added, probably because the CNTs used in this work are negatively charged (zeta potential 46.3 mv). TGA profiles of HC and BC samples exhibited a slightly higher thermal stability with increasin g introduction of CNTs (Figure 4 1 a and b), though, the difference in stability was more obvious for BC CNT nanocomposites. The thermal degradation of
81 pyrolyzed carbon materials, typically show loss of moisture (50 100 o C), followed by the disappearance of transformation carbon (e.g. aliphatic C C groups) from 100 350 o C, and finally, the formation of graphitic chars beyond 350 o C ( Chen et al., 2008 ; Zhang et al., 2012a ) Weight losses in both modified and unmodified biochars were insignificant until 350 o C, thereafter, comparatively g reater weight losses (~80 %) were observed in the unmo dified chars from about 350 5 00 o C than modified chars (~70 %). The thermal decomposition behavior of the CNT 1% biochars, particularly, BC CNT 1% closely r esembled the CNT thermal curve, degrading at about 400 o C. Features of the Raman spectra (Figure 4 2 a and b) include the disorder mode D band (~1350 cm 1 ) induced by sp 3 hybridization and the tangential mode G band, representing crystalline graphitic/sp 2 carbon stretching vibrations (1500 1600 cm 1 ) ( Rong et al., 2013 ; Theodore et al., 2011 ) ( Osswald et al., 2005 ) and increased I D /I G indicates higher defect concentration (increased functional groups) on the sorbents surface ( Theodore et al., 2011 ) Unmodified HC and BC biochars had lower I D /I G ratios (1.12 and 1.11 respectively), than CNT (1.78) (Table 4 1). But, after incorporating CNTs onto the biochars, I D /I G ratios of the hybrid HC CNT and BC CNT sam ples increased, indicating increased functionalities on the hybrid sorbents. TEM images of HC CNT 1% and BC CNT 1% ( Fi gure 4 3 ) showed the presence of tubular CNT bundles (diameter, 15 25 nm) on the char surfaces which further demonstrate d the incorporat ion of CNT in the biochar nanocomposites Methylene Blue Removal Efficiency of Biochars Though MB was sorbed by both modified and unmodified biochars, MB removal efficiencies increased with increasing CNT additions (Figure 4 4 ). The highest removal
82 of MB b y the sorbents was observed for HC CNT 1% (47% removal) and BC CNT 1% (64% removal). These findings are consistent with characterization results which showed the most improvements in specific surface area/pore volume and surface chemistry properties (e.g., zeta potential and Raman functionality) of biochars with additions of 1% CNT. Several mechanisms have been proposed for the interaction of organic contaminants with pristine and functionalized CNT including hydrophobic interaction, hydrogen bonding, elect graphitic surfaces of CNT and organic molecules containing C=C bonds ( Ai & Jiang, 2012 ; Ma et al., 2011 ; Mishra et al., 2010 ) To elucidate the sorption interaction between the hybrid CNT sorbents and MB, further testing under varied conditions was conducted using the HC CNT 1% and BC CNT 1%. Sorption Kinetics Sorption versus time profiles for the sorbents showed different sorption behaviors for the biochars (Figure 4 5 a d). For instance, pseudo equilibrium times for MB sorption were reached in the order of BC CNT 1% < HC < BC < HC CNT 1%. The sorption kinetic data were simulated with the pseudo first order, pseudo second order, and Elovich models. The CNT biochar nanocomposites exhibited much better correlation of the experimental data with first and second order models (R 2 > 0.75) than HC and BC (R 2 < 0.59, Table 2). The first (k 1 ) and second order (k 2 ) sorption rate constants were higher for HC (1.08 and 1.90 h 1 ) and BC (18.94 and 19.44 h 1 ) than for their respective modified chars (HC CNT 1% (0.97 and 0.67 h 1 ) and B C CNT 1% (13.08 and 5.16 h 1 )), suggesting faster initial uptake of MB on the unmodified biochars. This could be attributed to the fact that the CNT biochar nanocomposites have larger porosities (pore volume), which could increase the diffusive interaction time to delay the
83 initial MB sorption. Intraparticle diffusion kinetic plots were linear and well correlated with MB sorption (R 2 > 0.75) for all sorbents ( Figure 4 6 ), confirming the importance of the diffusive interaction between the MB and the sorbents The pseudo second order model, which predominantly describes chemisorption processes and interaction of functional groups on sorbents with contaminants ( Gerente et al., 2007 ) was the best fit for BC CNT 1% sorption kinetics data (R 2 = 0.96). This suggests stronger affiliations of MB to the COOH functionality on BC CNT 1% than present for the unmodif ied biochars. The Elovich model, which also evaluates chemisorption mechanisms ( Chien & Clayton, 1980 ) did very well modeling MB sorption by HC CNT 1% (R 2 = 0.98) and may also due to the involvement of carboxyl groups in MB sorption. Sorption Isotherms Equilibrium isotherms of HC, BC, HC CNT 1%, and BC CNT 1% (Figure 4 7 a d) showed increasing uptake of MB with increasing concentrations of aqueous MB until apparent maximum sorption capacity was reached. Sorption capacities of HC CNT 1% and BC CNT 1% i ndicated via Langmuir modeling (2.4 and 5.5 mg g 1 respectively), were almost twice those of their respective unmodified biochar, (1.3 and 2.2 mg g 1 respectively). Similar sorption capacities were reported for some carbonaceous waste materials but the o bserved values are much lower than those of some other hybrid CNT materials (Table 4 3 ), suggesting further investigations are still need ed to optimize the synthesis of the CNT biochar nanocomposites to improve their sorption capacities to MB. While both L angmuir and Freundlich models reproduced the sorption isotherm data reasonably well, the Langmuir Freundlich (L F) model fit both unmodified and
84 modified biochars (R 2 > 0.90) best (Table 4 2 and Figure 4 7 ), indicating the sorption of MB on the biochar bas e sorbents could be controlled by multiple mechanisms. In literature, the L F model is often used to describe the sorption of chemicals by heterogeneous materials, including biochars, through multiple processes ( Jeppu & Clement, 2012 ; Kasozi et al., 2010 ) In this work, the sorption of MB on the CNT biochar nanocomposites could be controlled by two processe s: 1) MB sorbed onto high affinity binding sites within CNT and 2) MB sorbs to biochar itself as the CNT sites become filled. Effect of pH Solution pH can influence both the surface charge of a sorbent as well as the degree of ionization and conformation o f a sorbate (such as MB) ( Dias et al., 2002 ; Ma et al., 2012 ; Pavan et al., 2008 ) With increasing pH, there was an increase in MB sorption by all the biochars, until pH of 7, above which, no significant pH dependence was observed (Figure 4 8 a ). Similar trends have been reported in the literature for MB sorption on other sorbents ( Iriarte Velasco et al., 2011 ; Pavan et al., 2008 ) Al l the biochars used in this study were predominantly negatively charged in DI water (Table 1), which would promote the sorption of positively charged MB (pKa 3.8) by electrostatic attraction ( Dias et al., 2002 ) In particular, the electrostatic attraction of positively charged MB to the CNT biochar n anocomposites should increase with increasing pH (below 7) because of the increasing deprotonation of the functional (e.g., carboxyl and hydroxyl) groups of the CNTs within the biochar matrix. On the other hand, because MB consists of benzene rings, which can participate in electron donor acceptor reactions with graphene in promoting the sorption of MB on the CNT biochar nanocomposites. In this case, a low
85 pH should favor don or compounds ( Iriarte Velasco et al., 2011 ; Ji et al., 2009 ; Zhu et al., 2004 ) Effect of Ionic Strength The presence of cations such as Na + have been shown previously to reduce the sorption of MB onto fungus ( Maurya et al., 2006 ) In theory, when electrostatic forces between sorbent surfaces and sorbate ions are attractive, an in crease in ionic strength will decrease the sorption capacity of the sorbate due to competition of Na+ ions with positively charged MB for sorption sites ( Ma et al., 2012 ) Here, MB sorption onto all the biochar based sorbents decreased somewhat as concentrations of NaCl increased from 0.01 to 0.1 M (Figure 4 8 b and c), confirming involvement of electrostatic interaction in MB sorption. With 0.5 M of NaCl, however, there was no significant decreas e, and even a slight increase in MB sorption by some of the sorbents. This effect has been reported previously and may arise from dimerization or aggregation of dye molecules at very high salt concentrations ( Ma et al., 2012 ; Mukerjee & Ghosh, 1970 ) and is independent of MB interactions with the sorbent. Conclusion Characterization of the sorbents indicated that the physiochemical properties (e.g., surface area, porosity, and thermal stability) of the biochars were enhanced by additions of CNTs. The BC CNT 1% char had the hi ghest sorption capacity to MB among all the sorbents, likely because it may have anchored more CNTs judging from its better thermal stability, higher surface area, and larger pore volume. The data collected suggests that electrostatic attraction was the do minant mechanism for sorption of MB onto the chars, but chemisorption such as pi pi bonding should not be ruled out
8 6 as a contributing sorption mechanism. Intrapore diffusion was also likely to control the rate at which MB was removed from solution onto the biochar based sorbents. Though the overall sorption capacity of the CNT hybridized chars studied were lower than those of some other hybrid sorbents reported in literature, the synthesis procedure employed here is simple, inexpensive, and can be further optimized. The results presented have established the potential of biochar CNT hybrid sorbents to be used for environmental remediation of dyes and possibly oth er organic pollutants. In addition to their low cost, they may provide additional environmental benefits, such as carbon sequestration and soil amelioration.
87 Table 4 1 Structural and physiochemical properties of the biochar based sorbents. Sample pH Zeta potential (mV) BET N2 surface area (m 2 /g) Pore volume (cc/g) C % H % N % O % Raman intensity (disorder/order ratio) I D /I G HC 7.3 28.8 289 0.001 81.8 2.2 0.7 15.3 1.11 HC CNT 0.01% 7.4 33.2 257 0.003 84.1 2.5 0.4 13 1.14 HC CNT 1% 7.5 41.4 352 0.138 80.3 2.1 0.2 17.4 1.30 BC 6.9 32.7 9 0.000 76.4 2.9 0.8 19.9 1.12 BC CNT 0.01% 7.0 34.1 120 0.008 79.3 2.2 0.8 17.7 1.15 BC CNT 1% 7.3 44.6 390 0.220 85.7 1.7 0.7 11.9 1.28
88 Table 4 2 Summary of models and best fit parameters of the sorption kinetics and isotherms Sorbents Model* Parameter 1 Parameter 2 Parameter 3 R 2 HC First order ( ) k 1 = 1.08 q e1 = 0.80 0.464 Second order ( ) k 2 = 1.90 q e2 = 0.89 0.588 Elovich ( ) 0.761 Langmuir ( ) K = 1.10 S max = 1.28 0.930 Freundlich ( ) K F = 0.62 n = 0.19 0.748 L F ( K = 1.16 S max = 1.24 n = 9.23 0.940 HC CNT 1% First order k 1 = 0.97 q e1 = 2.10 0.759 Second order k 2 = 0.67 q e2 = 2.28 0.869 Elovich 0.978 Langmuir K = 17.92 S max = 2.40 0.851 Freundlich K F = 1.44 n = 0.14 0.764 L F K = 40.47 S max = 2.40 n = 27.86 0.987 BC First order k 1 = 18.94 q e1 = 1.90 0.265 Second order k 2 = 19.44 q e2 = 1.92 0.336 Elovich 0.764 Langmuir K = 2.09 S max = 2.20 0.917 Freundlich K F = 1.07 n = 0.20 0.797 L F K = 2.24 S max = 2.20 n = 1.81 0.920
89 Table 4 3 Summary of models and best fit parameters of the sorption kinetics and isotherms continued Sorbents Model* Parameter 1 Parameter 2 Parameter 3 R 2 BC CNT 1% First order k 1 = 13.08 q e1 = 4.30 0.935 Second order k 2 = 5.16 q e2 = 4.37 0.956 Elovich 0.799 Langmuir K = 4.84 S max = 5.50 0.961 Freundlich K F = 3.02 n = 0.18 0.757 L F K = 5.12 S max = 5.50 n = 1.08 0.961 : q t and q e are the amount of sorbate removed at time t and at equilibrium, respectively (mg g 1 ), and k 1 and k 2 are the first order and second order sorption rate constants (h 1 ), respectively, is the initial sorption rate (mg g 1 ) and is the desorption constant (g mg 1 ), K and K f are the Langmuir bonding term related to interaction energies (L mg 1 ) and the Freundlich affinity coefficient (mg (1 n) L n g 1 ), respectively, S max is the Langmuir m aximum capacity (mg g 1 ), C e is the equilibrium solution concentration (mg L 1 ) of the sorbate, and n is the Freundlich linearity constant.
90 Table 4 4 Comparison of methylene blue sorption capacities by various sorbents. Adsorbents Methylene blue adsorption capacity (mg g 1 ) Reference HC CNT 1% 2.6 This work BC CNT 1% 6.2 This work Microwaved modified bamboo biochar 35.3 ( Liao et al., 2012 ) Fly ash stabilized hydrogen titanate nanosheets 0.6 ( Hareesh et al., 2012 ) Blast furnace sludge 6.4 ( Wang et al., 2005 ) Activate coir pith carbon 5.9 ( Kavitha & Namasivayam, 2007 ) Graphene carbon nanotube hybrid 82.0 ( Ai & Jiang, 2012 ) Magnetite loaded multi walled carbon nanotube 48.1 ( Ai et al., 2011 ) Powdered activated carbon 91.0 ( Yener et al., 2008 ) Alkali activated carbon nanotubes 400.0 ( Ma et al., 2012 ) Carbon nanotubes 46.2 ( Yao et al., 2010 )
91 Figure 4 1 Thermogravimetric analysis profile s of biochar based sorbents. A) HC and HC CNT composites and B) BC and BC CNT composites
92 Figure 4 2 Raman spectra of biochar based sorbents. A) HC and HC CNT composites and B) BC and BC CNT composites
93 Figure 4 3 Transmissi on electron micrographs for A) HC CNT 1% and B) BC CNT 1% samples at 50000X magnification
94 Figure 4 4 Methylene blue removal efficiencies of biochar based sorbents
95 Figure 4 5 Sorption kinetics plots of biochar based sorbents. A) HC, B) HC CNT 1%, C) BC and D) BC CNT 1%
96 Figure 4 6 Intraparticle diffus ion kinetics plots for A) HC and B) BC sorbents.
97 Figure 4 7 Sorption isotherms of biochar based sorbents. A) HC, B) HC CNT 1%, C) BC, and D) BC CNT 1%
98 Figure 4 8 Effects of s olution chemistry on methylene sorption on biochar based sorbents. A) pH effects, B) ionic strength effec ts on HC and HC CNT 1%, and C) ionic strength effects on BC and BC CNT 1%
99 CHAPTER 5 SIMUL TANEOUS SORPTION OF SULFAPYRIDINE AND LEAD BY BIOCHAR MODIFIED WITH SURFACTANT DISPERSED CARBON NANOTUBES Introduction Human and veterinary pharmaceutical antibiotics in water systems have been and remain a public concern due to chronic toxic effects and potential development of antibiotic resistance in microbial populations (Challis et al., 2013; Ji et al., 2009; Schwarz et al., 2012). Sulfon amide antibiotics (SA) are a popular class of broad spectrum antibiotics whose metabolites are not wholly digested in animal systems, but can be leached into water bodies when sulfonamide contaminated animal manure are applied to soils (Kurwadkar et al., 2 007). Research studies (Diaz Cruz et al., 2008; Gobel et al., 2004; Radke et al., 2009) have reported high concentrations of SA and their metabolites in waterways, and their total elimination in conventional water treatment plants has yet to be demonstrate d (Jesus Garcia Galan et al., 2012). Sulfapyridine (SPY) is a fairly water soluble SA often detected at high concentrations in wastewaters (70 227 ng L 1 ) (Gobel et al., 2004; Jesus Garcia Galan et al., 2011; Thiele Bruhn et al., 2004). Typically, wastew ater treatment plants are operated at short hydraulic residence times (~ 40 h), whereas residues of SPY and their metabolites require longer residence times (32 62 days) to be completely degraded (Gros et al., 2010; Radjenovic et al., 2009).Thus, the pot ential contamination of environmental waters by SPY in treated wastewater effluents is likely (Gros et al., 2010). Lead (Pb) is a long standing toxic heavy metal pollutant found in wastewater and industrial effluents (Muhammad et al., 2006). The effects of heavy metals, like Pb in the environment can be severe because it is non biodegradable and environmentally persistent, even at low concentrations. Specifically, bioaccumulation of Pb in human
100 and animal tissues has been linked to hypertension, mental reta rdation, renal impairment, and reproductive disorders (Pokras & Kneeland, 2008). Among several treatment options aimed at minimizing Pb concentrations in wastewater, sorption has been generally recognized as an effective and relatively economical treatment approach (Huang et al., 2012). Moreover, previous research on the application of low cost biosorbents (after modification) have reported higher uptake of Pb compared to some commercial activated carbons (Inyang et al., 2012; Yao et al., 2011c). Biochar is a porous, carbon rich material derived from the thermal treatment of biomass in a closed system under anaerobic conditions. The properties and constituents of biochar are heterogeneous and vary extensively with thermal treatment conditions (Chen et al., 2 012). Also, recent studies (Chen et al., 2008; Chun et al., comprising of highly carbonized and less carbonized organic matter. While, the carbonized matter with condense d aromatic compounds, can act as an adsorbent; the less carbonized, amorphous fraction acts as a partition/absorption phase (Amymarie & Gschwend, 2002; Chen et al., 2008; Keiluweit et al., 2010). These two phases may well sorb Pb and SPY. Like biochar, car bon nanotubes (CNTs) are a promising, new class of sorbents composed of covalently bonded carbon sheets with high specific surface areas that can effectively sorb heavy metals and organic compounds (Huang et al., 2012; Ji et al., 2009; Lin & Xing, 2008; Ti an et al., 2012b). Yet, practical application of CNTs is limited by its poor solubility and propensity to aggregate into bundles (Matarredona et al., 2003). Functionalization of CNTs and the use of surfactants in CNTs dispersion are two
101 suggested means of overcoming these limitations (Clark et al., 2011; Lin & Xing, 2008). Also, our previous study (Inyang et al., 2013) had demonstrated that biochar could serve as a low cost support for CNTs, and the incorporation of CNTs into biochar significantly improved its sorption ability for methylene blue. The combination of surfactant dispersed CNTs with biochar can be suggested as a means of optimizing the pro cess of producing CNT biochars because the surfactant dispersion of CNT w ould increase the amount of individ ual CNT threads to be anchored onto biochar for the sorption of Pb and SPY. Thus, the overarching objective of this study wa s to synthesize sodium dodecyl benzene sulfonate (SDBS) dispersed CNT biochar nanocomposites from hickory chips and sugarcane bagass e biomass (HC SDBS CNT and BC SDBS CNT, respectively), and determine their sorption potential for SPY and Pb in aqueous solutions. Our specific objectives were to: (1) examine the effects of SDBS dispersed CNT on the properties of SDBS CNT biochar nanocomp osites, (2) determine the sorption capacity of Pb and SPY on SDBS CNT biochar nanocomposites in a single solute system, (3) examine interactions between SPY and Pb in a binary solute system on SDBS CNT biochar nanocomposites, and (4) elucidate and differen tiate sorption mechanisms controlling the sorption of Pb and SPY on SDBS CNT biochar nanocomposites. Materials and Methods Materials Hickory chips and sugarcane bagasse biomass were obtained from UF North Florida Research and Education Center, and dried an d milled to < 500 m size fractions. Carboxylic acid functionalized multi walled carbon nanotubes (CNTs) with diameter, 10 20 nm and purity > 95 % was purchased from Sinonano Company (P.R.
102 China). All reagents used in this study were of high purity grade. Sulfapyridine (99% Sigma, mol. wt. 249.29 g mol 1 ), lead nitrate and sodium dodecylbenzene sulfonate (SDBS, mol. wt. 348.48 g mol 1 ) surfactant were obtained from Sigma Aldrich Co. (St. Louis, MO), and Fisher Scientific, Georgia, respectively. The chemical properties of SPY used in this study are presented in Table 5 1. SPY solutions (5 60 mg L 1 ) were prepared by sonicating pre determined amounts of SPY in known volumes of deionized, distilled water for 1 h, and the solutions left to stand for 1 3 days until complete dissolution Preparation of S urfactant dispersed C NT Biochar N anocomposites Surfactant based CNT suspensions were prepared by adding 2 g each of CNTs and SDBS pow der (1% by weight) to 200ml of deionized, distilled water. Surfactant CNT sus pensions and CNT suspensions containing 2 g of CNTs powder without SDBS were each sonicated in an ultrasound homogenizer (Model 300 V/T, Biologics, Inc.) for 1 h at pulse intervals of 12 min. The resulting suspensions, designated as CNT (without surfactant ), and SDBS CNT (with surfactant) were used for the preparation of CNT biochar nanocomposites. SDBS dispersed CNT biochars and CNT biochars (without SDBS) were produced by dip coating milled sugarcane bagasse and hickory chips biomass (feedstocks) each in SDBS CNT and CNT suspensions respectively, before converting the coated biomass to SDBS CNT biochars and CNT biochars following previously reported procedure (Schoen et al., 2010b; Zhang et al., 2012a). Specifically, 10 g of milled sugarcane bagasse and h ickory chips biomass were each dip coated in 100 ml of SDBS CNT or CNT suspensions, and stirred for 1 h on a magnetic stirrer at 500 rpm, after which, the dip coated CNTs treated feedstock were removed and oven dried at
103 105 o C. Next, dried SDBS CNT and CNT feedstocks along with untreated feedstocks were each placed in a quartz tube and slowly pyrolyzed at 600 o C for 1 h inside a tubular furnace (MTI, Richmond, CA) under a flowing N2 environment. The resulting biochars were tagged as SDBS dispersed CNT modif ied hickory and bagasse biochars (HC SDBS CNT and BC SDBS CNT), CNT modified hickory and bagasse biochars without SDBS (HC CNT and BC CNT), and pristine hickory and bagasse biochars (HC and BC). All samples were stored in plastic vials for further testing. Characterization Thermogravimetric analysis (TGA) of the biochars was performed in a stream of air at a heating rate of 10 o C min 1 with a Mettler TGA/DSC1 analyzer (Columbus, OH) to test the thermal stability of the chars. Elemental carbon, hydrogen, nit rogen and oxygen content (C, H, N, and O); zeta potential, pH, and surface areas of the sorbents were determined using previously reported methods (Inyang et al., 2012; Yao et al., 2011a). Fourier transform infra red analysis of pre and post sorption SDBS CNT biochars loaded with SPY was conducted to elucidate the interaction mechanisms during sorption. Sorption of Lead and Sulfapyridine in Single Solute S ystem The modified and pristine biochars were preliminary assessed for their sorption ability of lead and SPY by mixing 25 mg of each biochar with 12.5 ml of 40 mg L 1 Pb(NO 3 ) 2 or 20 mg L 1 SPY in 50 ml digestion vessels (Environmental Express) at room temperature. The sample solutions and their corresponding blank controls were agitated for 24 h on a reciprocating shaker and withdrawn at specific time intervals and filtered through 0.22 m p ore size nylon membranes (GE cellulose nylon membranes). Concentrations of SPY and Pb concentrations in filtrates were determined using a
104 Thermo Scientific EVO 60 UV VIS spectrophotometer at a wavelength of 260 nm, and an inductively coupled plasma spectro meter (ICP AES), respectively. The sorbed amounts of Pb and SPY on test biochars were each calculated as the difference between their initial and final aqueous solution concentrations. All sorption experiments were conducted in triplicate and the average v alues are presented with standard deviations. Following preliminary sorption assessments, Pb and SPY sorption kinetics and isotherm studies were conducted on surfactant CNT modified biochars (HC SDBS CNT and BC SDBS CNT) in a single solute system using sim ilar procedures. Sorption kinetics were examined by mixing 25 mg of each biochar with 12.5 ml of 40 mg L 1 Pb(NO 3 ) 2 or 20 mg L 1 SPY respectively, in 50 ml digestion vessels at specific intervals from 1 to 24 h and filtered to 0.22 m. The pH of sample sol utions was monitored prior to and after sorption experiments. Pb and SPY sorption isotherms were also determined by mixing 25 mg of each biochar with 12.5 ml of varying concentrations (5 100 mg L 1 ) of Pb(NO 3 ) 2 or (10 60 mg L 1 ) of SPY in 50 ml digesti on vessels. Co sorption of Lead and Sulfapyridine in Binary Solute S ystem Co sorption of Pb and SPY by SDBS CNT biochars in a binary solute system was examined. The effect of Pb on SPY sorption was determined by mixing 25 mg of each biochar with 12.5 ml of 20 mg L 1 SPY and different concentrations of Pb (0.05 0.5 mM). Likewise, the effect of SPY on Pb sorption was tested by mixing 25 mg of each biochar with 12.5 ml of 40 mg L 1 Pb and different concentrations of SPY (0.02 0.2 mM). Binary solute sample mixtures of Pb and SPY and their corresponding blank controls were agitated on a reciprocating shaker for 24 h, withdrawn and treated as
105 described above for the single solute system. Co sorption experiments were also conducted in triplicate. Results and Di scussion Biochar Properties The physiochemical properties of pristine hickory (HC) and bagasse (BC) biochars were generally improved by the addition of CNT, but no obvious changes in physiochemical properties were observed after the addition of SDBS CNT to the biochars (Table 5 1 ). Although, there was a slight increase in the surface area (from 351 to 359 m 2 g 1 ) and pore volume (from 0.14 to 0.27 cc g 1 ) of HC SDBS CNT, there was no corresponding improvement in BC SDBS CNT. T ypically, the sonication of surfactant CNT suspensions would debundle CNT aggregates either by steric or electrostatic repulsions ( Bandyopadhyaya et al., 2002 ; Clark et al., 2011 ; Ham et al., 2005 ; Matarredona et al., 2003 ; Vaisman et al., 2006 ) which sh ould increase the amount of CNT threads in solu tion that can be anchored to the coated biomass and produced bio char ( Vaisman et al., 2006 ) but the corresponding improvements in the physiochemical properties of the biochars were not obvious. TGA profiles of the SDB S CNT biochar nanocomposites (Figure 5 1) showed higher stability than pristine biochars, and in some cases, greater thermal stability than CNT biochars. All the carbons showed insignificant weight losses during thermal treatment until 350 o C, after which CNT, or biochars began to degrade. However, unlike BC SDBS CNT, slightly higher stability was observed in HC SDBS CNT profile than HC CNT, which is consistent with the small improvements in its physiochemical properties. This could also indicate that more CNT may have anchored to the surface of HC SDBS CNT biochar than BC SDBS CNT But the higher thermal stability of HC
106 SDBS CNT could also be attributed to the presence of more aromatic, lignin component s in hickory compared to the bagasse biomass. Prelimin ary Biochar A ssessments Lead and SPY were bot h sorbed onto pristine and CNT biochars (Figure 5 2), but SDBS CNT biochars removed the most SPY and Pb. Specifically, the amounts of SPY sorbed increased in the order of pristine biochar < CNT biochar < SDBS CN T biochar, with highest removal efficiencies of 86 % and 56 % SPY for HC SDBS CNT and BC SDBS CNT, respectively. Distribution coefficients, Kd for SPY on the modified SDBS CNT chars were 3111 and 655 L kg 1 for HC SDBS CNT and BC SDBS CNT, respectively which are much higher than previously reported Kd values for sulfamethaxole ( 2 104 L kg 1 ) sorption on biochar (Yao et al., 2012b), and SPY ( 1.1 5.6 L kg 1 ) sorption on soil organic matter (Haham et al., 2012). Likewise, HC SDBS CNT and BC SDBS CNT biochars had the most Pb removal efficiency (71 % and 53 %, respectively), than CNT biochars, or pristine biochars. Given the heterogeneity of the synthesized SDBS CNT biochars, the possibilit y of multiple mech anisms participating in Pb or SPY s orption on to SDBS CNT chars cannot be ruled out. Further sorption and characterization studies were conducted on SDBS CNT biochar to investigate these possibilities. Sorption K inetics Sorption kinetics of SPY or Pb in a s ingle solute system is shown in sorption versus time profiles (Figures 5 3 a and b). Both Pb and SPY sorption onto SDBS CNT biochars occurred at an initially fast rate with 30 50 % of SPY or Pb sorbed during the first 1 h, thereafter, the rate of Pb and SPY gradually slowed down until equilibrium states were reached. Rate limited pseudo first order, pseudo second order, and Elovich
107 models (Table 5 2) were used to simulate the sorption kinetic data but, only best fitted models plots are graphed (Figures 5 3 a and b).The Elovich model best fit the kinetic sorption data for SPY on both biochars (R 2 > 0.90), but the simulated Elovich rate constant for HC SDBS CNT was 50 times greater than for BC SDBS CNT. The Elovich model evaluates chemisorption mechanisms, s uch as chemical bo ndings between con taminants and heterogeneous surfaces such as SDBS CNT biochars But, pH of SPY sample solutions remained at pH 6 7 below its pKa 2 value of 8.4, before and after sorption, so that most of the SPY was protonated and neut ral which could encourage hydrophobic interactions between neutral SPY species and hydrophobic CNT/biochar sites. Moreover, the protonation of SPY would increase the electron i nteractions between graphene in CNTs or biochar and SPY (Ji et al., 2009). In the case of Pb, the pseudo second order modeled sorption onto both SDBS CNT chars well, but the Elovich model had a much stronger fit for HC SDBS CNT (R 2 = 0.98) than BC SDBS CN T (R 2 = 0.66) which had a better fit with the second order model (R 2 = 0.82). The interactions of Pb to oxygen containing functional groups in CNT/biochars via complexation or electrostatic attraction are possible sorption mechanisms for Pb sorption on SDB S CNT biochars. Intra particle diffusion plots for pre equilibrium sorption of SPY an d Pb (Figure 5 4 a and b) fit both SDBS CNT biochars which implies that diffusion also influenced the rate of Pb and SPY sorption into the pores of modified biochars Sorption Isotherms Isotherms for the sorption of Pb and SPY onto SDBS CNT chars are presented in Figure 5 5 a and b. Experimental equilibrium data for SPY and Pb were simulated
108 with Langmuir (L), Freundlich (F) and Langmuir Freundlich (L F) models, and the best fit model parameters were presented (Table 5 2). The L F model best reproduced SPY sorption (R 2 > 0.89) on both biochars, which suggests that multiple controlling sorption mechanisms may occur. The L F model has been applied to the sorption of chemic als on heterogeneous materials, including biochars (Jeppu & Clement, 2012; Kas ozi et al., 2010). Pb sorption data was also well fit with the L F model (R 2 > 0.87). The L bonding term, K (L mg 1 ) was 40 times higher for BC SDBS CNT (1.66) than HC SDBS CNT ( 0.04) which suggests that the affiliation of Pb on BC SDBS CNT was more influenced by surface sorption. Previous sorption isotherm study (Inyang et al., 2011) for Pb sorption on pristine, unmodified BC had also shown that Pb bonding on BC was similar to ac tivated carbon (with similar Langmuir bonding term) and mainly surface controlled. Co sorption of Lead and Sulfapyridine in Binary Solute Systems Sorption capacities for SPY for HC SDBS CNT (7.7 8.7 mg g 1 ) and BC SDBS CNT (4.4 4.5 mg g 1 ) in the binary solute system were similar to the single solute system for HC SDBS CNT (8.64 mg g 1 ) and BC SDBS CNT (4.8 mg g 1 ), which showed that there was limited interaction between SPY and Pb (Figure 5 6 a). Likewise, sorption capacities for Pb for HC SDBS CNT (13. 1 13.3 mg g 1 ) and BC SDBS CNT (10.0 10.4 mg g 1 ) in the binary solute system were similar to the single solute system for HC SDBS CNT (14.4 mg g 1 ) and BC SDBS CNT (13.2 mg g 1 ) (Figure 5 6 b). Student t test analysis also showed that there was no sign ificant dependence of the amounts of SPY or Pb sorbed on either biochars with increasing Pb or SPY concentrations. A previous study reported that the presence of exchangeable cations in solution increased the sorption of SPY on soil organic matter via thei r complexation wi th
109 the amine and SO 2 groups in SPY (Haham et al., 2012; Schwarz et al., 2012). But, there have also been previous studies that found no interaction between the heavy metal and organic contaminant during co sorption (Cao et al., 2009b; Wu et al., 2012a). Co sorption studies for Pb and SPY indicate that the sorption of Pb or SPY on the SDBS CNT biochars was site specific, and there was likely no significant competition between SPY and Pb for functional groups on the SDBS CNT biochar matrix. Additionally, complexation of Pb with the amine or SO 2 groups in SPY may not have occurred. Further evidence of insignificant competition for functional groups between Pb and SPY was observed in the FTIR spectra of the SPY loaded SDBS CNT biochars which sh owed no change in the functional chemistry of SDBS CNT chars after SPY sorption (Figure 5 7). The SPY sorption process in single and binary solute systems for either SDBS CNT biochars (Figure 5 8) could have occurred via the or acceptor reactions between graphite in CNT or carbonized biochar matter and the amine group in SPY; (2) hydrophobic interactions between SPY and hydrophobic sites on CNT/biochar sites; and (3) physical sorption of SPY on SDBS CNT biochar surfaces. While Pb sorption on SDBS CNT biochars may have also occurred by: (1) Pb complexing with oxygen containing functional groups in biochar and CNT, (2) physical sorption of Pb on the surface of the modified biochars, and (3) electrostatic interaction between the dissociated carboxyl group in CNT and Pb. Conclusion s The incorporation of SDBS CNT into biochar improved their sorption capacities for Pb and SPY even though, no obvious improvements in the physiochemical properties (e.g., surface area, pore volume, and zeta potential) of the modified SDBS CNT biochars were observed. Despite limited improvements in their properties, both
110 SDBS CNT biochar nanocomposites effectively removed Pb or SPY from single solute and binary solute aqueous solutions. Binary solute systems however showed no significant change in the sorption capacities of the biochar nanocomposites from their sorption in a single solute system, suggesting site specific sorption interactions and no significant competit ion for functional groups between SPY and Pb. The synthesized SDBS CNT biochar nanocomposites show that they have the potential to be employed as alternative treatment technologies for the remediation of metallic and organic contaminants. Furthermore, beca use of the anti microbial properties of SPY, the SPY laden SDBS CNT biochars can be further expl oited for disinfection purposes
111 Table 5 1 Physiochemical p roperties of carbons used in this study Sample pH Zeta potential (mV) Surface area (m 2 /g) Pore volume (cc/g) C % H % N % O % H/C (O+N)/C Pka 1 /Pka 2 Log Kow CNT 8.50 46.34 142 0.115 ND ND ND ND ND ND ND ND HC 7.25 28.84 289.2 0.001 81.81 2.17 0.73 14.02 0.027 0.17 ND ND HC CNT 7.49 41.42 351.5 0.138 80.30 2.08 0.22 17.40 0.026 0.22 ND ND HC SDBS CNT 6.74 42.65 359 0.27 77.69 2.07 0.19 20.05 0.027 0.26 ND ND BC 6.94 32.72 9.3 0 76.44 2.93 0.79 19.84 0.038 0.26 ND ND BC CNT 7.31 44.58 390 0.22 85.73 1.74 0.66 11.88 0.020 0.14 ND ND BC SDBS CNT 6.72 32.21 336 0.167 84.30 1.98 0.63 13.09 0.023 0.15 ND ND SPY ND ND ND ND ND ND ND ND ND ND 2.9/8.4 0.35 ND Not determined
112 Table 5 2 Best fit model parameters for sorption kinetics and isotherms SPY Sorbents Model* Parameter 1 Parameter 2 Parameter 3 R 2 HC SDBS CNT First order ( ) k 1 = 12.94 q e1 = 7.36 0.524 Second order ( ) k 2 = 2.41 q e2 = 7.52 0.638 Elovich ( ) 0.923 Langmuir ( ) K = 0.13 S max = 27.90 0.894 Freundlich ( ) K F = 3.98 n = 0.61 0.891 L F K = 0.10 S max = 31.05 n = 0.93 0.895 BC SDBS CNT First order k 1 = 1.24 q e1 = 3.87 0.664 Second order k 2 = 1.34 q e2 = 3.86 0.811 Elovich 0.910 Langmuir K = 0.02 S max = 19.36 0.889 Freundlich K F = 0.87 n = 0.65 0.908 L F K = 0.00 S max = 122.63 n = 0.69 0.907
113 Table 5 2 Best fit model parameters for sorption kinetics and isotherms continued. Pb Sorbents Model* Parameter 1 Parameter 2 Parameter 3 R 2 HC SDBS CNT First order k 1 = 1.00 q e1 = 11.52 0.847 Second order k 2 = 0.13 q e2 = 12.30 0.932 Elovich 0.979 Langmuir K = 0.81 S max = 15.2 0.819 Freundlich K F = 6.82 n = 0.21 0.873 L F K = 0.04 S max = 28.03 n = 0.33 0.877 BC SDBS CNT First order k 1 = 10.82 q e1 = 9.81 0.786 Second order k 2 = 1.65 q e2 = 9.98 0.816 Elovich 0.657 Langmuir K = 2.16 S max = 13.7 0.935 Freundlich K F = 7.24 n = 0.17 0.898 L F K = 1.66 S max = 14.7 n = 0.62 0.954 *q t and q e are the amount of sorbate removed at time t and at equilibrium, respectively (mg g 1 ), and k 1 and k 2 are the first order and second order sorption rate constants (h 1 ), respectively, is the initial sorption rate (mg g 1 ) and is the desorption constant (g mg 1 ), K and K f are the Langmuir b onding term related to interaction energies (L mg 1 ) and the Freundlich affinity coefficient (mg (1 n) L n g 1 ), respectively, S max is the Langmuir maximum capacity (mg g 1 ), C e is the equilibrium solution concentration (mg L 1 ) of the sorbate, and n is the Freundlich linearity constant.
114 Figure 5 1 Thermogravimetric analysis of pristine and modified. A) HC and B) BC sorbents Figure 5 2 Preliminary assessments for sorption of SPY an d Pb o nto biochars
115 Figure 5 3 K inetic plots for sorption of A) SPY and B) Pb onto surfactant CNT modified biochars Figure 5 4 Intra particl e diffusion plots for sorption of A) SPY and B) Pb onto surfactant CNT modified biochars
116 Figure 5 5 Is otherms for sorption of A) Pb and B) SPY onto surfactant CNT modified biochars Figure 5 6 Co sorption of Pb an d SPY in binary solute system ( a b)
117 Figure 5 7. Fourier transform infra red ana lysis of SDBS CNT biochars and SPY laden SDBS CNT biochars Figure 5 8. Possible sorption mechanisms for SPY and Pb sorption on SDBS CNT modified biochars
118 CHAPTER 6 ENGINEERED BIOCHARS FOR THE REMOVAL OF METALLIC, ORGANIC AND EMERGING CONTAMINANTS FROM AQUEOUS SOLUTIONS Conclusion s The overarching o bjective of this study to develop useful biochars using cost effective activation techniques requiring less cost and labor; yet achiev i ng high removal efficiencies for traditional (e.g., heavy metals and dyes) and emerging contaminants such as nanoparticles and pharmaceutical residues was achieved in most cases The important findings critical to this research work are summarized as foll ows: First, b iochar produced from anaerobically digested dairy waste (DAWC), and sugar beet s (DWSBC) residuals could effectively remove lead cadmium, nickel and copper from aqueous solutions. In particular, their sorption capacity for lead was comparable to that of commercial activated ca rbons and p recipitation was considered the primary mechanism co ntrolling the sorption of lead. Lead was precipitated on digested anaerobically digested biochars because of its reaction with slowly released carbonate or pho sphate ions within DAWC or DWSBC. Moreover, t he prese nce of these carbonate and phosphate deposits within the digested biochar matrix makes them good soil conditi oner s, particularly in acidic soils. Based on the performance and potential applications of an aerobically digested biochars, digested biochars can be competitive water purification products in the market of adsorbents. Second c hemically activated carbon s from iron impregnation/precipitation on carbon surfaces generally improved their retention for CNT, and NTiO 2 but were not effective for improving AgNP retention. In addition, a mong all the ENPs CNT had the highest mobility and the most environmental risk of being released from filter media because over 90% of the CNT was released f rom the columns But i ron modified
119 biochars showed better ret ention of CNT than raw biochars and commercial activated carbons and may be considered as more cost effective option s than activated carbons for mobilizing CNT. Still the re tention of ENPs on the carbon based filter media was only slightly improved by iron impregnation of the carbons and an optimization of p rocess conditions may be required (e.g., reducing flow rate and increasing column length ) to enhance the retention of ENP s on the carbons Third, i ncorporating CNT with biochar to make hybrid CNT biochar nanocomposites was beneficial in improving the physiochemical properties and sorptive properties of hickory (HC) and bagasse (BC) biochars. Specifically the add ition of 1% CNT by weight in BC dramatically increased its surface area by 40 times and also doubled its sorption capacity for MB via electrostatic attractions mechanisms. Additionally, t he effect of solution chemistry on the sorption of MB on biochars was evident for both modified an d unmodified bochars. Thus, i n order t o maximize the sorption of MB on raw and CNT modified biochars, low ionic strength conditions and an optimum pH of 7 should be maintained to reduce the competition effect between MB and other cations. Lastly, d epositin g SDBS dispersed CNT onto biochar matrix (SDBS CNT biochar nanocomposites) generally improved their sorption capacities for Pb and SPY compared to CNT biochars (without surfactants) or pristine hickory (HC) and bagasse (BC) biochars. But, slight improve men ts in the physiochemical properties (e.g., thermal stability, surface area, pore volume, and zeta potential) of the SDBS CNT biochar nano composites were only ob served in HC SDBS CNT biochars, which might be due to higher CNT contents In addition, higher t hermal stability of HC SDBS CNT may be due
120 to more aromatic lignin components of HC SDBS CNT compared to BC SDBS CNT. Sorption of Pb and SPY onto SDBS CNT biochars was likely influenced by multiple mechanisms such as surface adsorption, electrostatic bondings. The use of high temperature biochars (> 700 o C) with significant graphite contents can also be suggested to improve retention of SPY when combined with graphite rich CNT because of enhanced In concl usion, this study has shown that biochars from low cost materials can be sorbing a wider array of contaminants. In some cases, the sorption capacities of these modified biochars were comparable and even surpass ed commercial activated carbons for some c ontaminants. In addition to high sorption capacities, these modified biochars have resulted in the improvement of phys iochemical properties including, higher thermal stabilities comparably better than emerging sorb ents such as carbon nanotubes in some instances Moreover the modification techniques employed here in this study are simple, and non laborious. T hese modified biochars could be potential cost effective technologies for water puri fication, soil fertility enhancements and carbon sequestration. Recommendations The research studies presented here opens exciting avenues to improv e the process of biochar production and activation as well as expand the application of biochar for remediation purposes. Possible a venues to improve the potential of this research for practical applications which can be considered for further studies are o utlined in the following : Firstly, t he process of filtering ENPs on the biochars did not achieve high retention of ENPs on the carbon based filter media. Therefore, m odifying solution
121 chemistries and process conditions, (e.g., i ncreasing the column length, and reducing flow rates and increasing the ionic strength of the filter media) may be considered to maximize the retention of ENPs in the carbon filter media Secondly, SDBS CNT biochars showed great potential for removing SPY and Pb from aqueous solutions, but the improvements in the properties of BC SDBS CNT may have been limited by SDBS dispersion of CNT. The physiochemical p roperties of BC SDBS CNT may be improved by examining the optimum SDBS to CNT ratios t hat can be employed to maximize the sorptive properties of BC SDBS CNT Finally, r ecycling SPY laden SDBS CNT biochar nanocomposites for disinfection purposes can be cons idered to exploit the anti microbial properties of SPY Further study examining the potential of these materials to be used in reducing bacterial or algal populations in environmental systems would be advantageous.
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137 BIOGRAPHICAL SKETCH Mandu Ime Inyang was born in Lagos, Nigeria. She received her Bachelor of Technology degree in chemical e ng i neering from Ladoke Akintola University of Technology, Oyo state, Nigeria in 2005. Prior to her graduation, she served as an Intern in the National Engineering and Technical Company (NETCO), Lagos, Nigeria, where she gained experience in process design. Aft er grad uation, she worked as a c hemistry instructor, teaching chemistry to National Diploma students in the Basic and Applied Science D epartment, Niger state polytechnic, Zungeru, Nigeria before proceeding to the United States for graduate studies. Mandu began he r graduate research as a research scholar/exchange student in the Bioprocess laboratory, Agricultu ral and Biological Engineering D epartment where she was involved in the pilot scale production of biodiese l from waste vegetable oil for six (6) months and ga ined experience in the characterization of biodiesel according to ASTM fuel quality standards. At the end of her exchan ge program, she enrolled in a m aster s program in A gricultu ral and B iological E ngineering D epartment and continued her research i n renewa ble energy. During her m aster s, Mandu served as a research assistant working in the Bioprocess, and Environmental Nanotechnology Laborator ies She gained experience in generating biogas from anaerobic digestion of biomass materials, and converting the dig estion residuals to carbon adsorbents. She immediately proceeded to continue doctoral studies in agricultural and biological e ngineering, in the field o f environmental n anotechnology Her doctoral research has focused on the engi neering of hybrid carbon adsorbents using anaerobic digestion, chemicals and nano materials to improve their sorption capacity for a wider range of contaminants. Mandu attribute s her success to a firm trust, and unwavering confidence in God.