1 IRON (FE) A ND MANGANESE (MN) OXIDE MINERAL MODIFIED BIOCHARS: CHARACTERIZATION AND REMOVAL OF ARSENATE AND LEAD By SHENGSEN WANG 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 201 4
2 Â© 201 4 Shengsen Wang
3 To my parents for their encouragements
4 ACKNOWLEDGMENTS I thank all people who have offered help to my academic study and accommodations in my study at University of Florida. I want to show greatest appreciation to the ch air of my supervisory committee, Dr. Y.C. Li , and my co chair, Dr. B. Gao, for their guidance in experimental design and implement of whole projects, their input in dissertation and manuscript preparation and their greatest contribution to my academic training and encouragement . I would like to thank Dr s . L.Q. Ma, W.G. Harris, and K. W. Migliacco for involving in my supervisory committee and offering valuable professional suggestions and guidance for my research, and their c ontribution to my dissertation. I am very grateful to Dr s . K.R. Reddy and J. Burns for their support for my program. I also want to thank professors in soil and water science department for offering classes to me as an integral part of graduate study. These professors include Dr. S. Grunwald, Dr. P. Inglett, Dr. D. Rhue, Dr. J. Sartain and Dr. P. Nkedi Kizza. I want to thank Mr. Michael J. Sisk as student service for his patience and generous help. People in my lab also deserve great gratitude. I thank Dr. X. Hu for his help with equipment s . Great thanks to Dr. A.K. Alva for his help with experimental materials. Last and most importantly, I thank my parents for their encouragement and support, my wife Xianni Yan g for her encourage ment , support, caring and being with me to allow me to focus on my research, and my son J ames Wang for bringing me joyful ness and for being a considerate kid.
5 TABLE OF CONTENTS page ACKNOWLEDGMENTS ................................ ................................ ................................ .. 4 LIST OF TABLES ................................ ................................ ................................ ............ 7 LIST OF FIGURES ................................ ................................ ................................ .......... 8 ABSTRACT ................................ ................................ ................................ ................... 10 CHAPTER 1 INTRODUCTION AND LITERATURE REVIEW ................................ ..................... 12 Introduction ................................ ................................ ................................ ............. 12 Biochar ................................ ................................ ................................ .................... 13 Feedstock ................................ ................................ ................................ ............... 17 Arsenic (As) ................................ ................................ ................................ ............ 20 Lead (Pb) ................................ ................................ ................................ ................ 23 Hydrous Iron and Manganese Oxides ................................ ................................ ..... 25 Hypotheses and Research Objectives ................................ ................................ .... 30 Hypotheses ................................ ................................ ................................ ...... 30 Objectives ................................ ................................ ................................ ......... 31 2 PYROLYSIS TEMPERATURES AFFECTED CHARACTERISTICS OF BIOCHAR PRODUCED FROM WOODY AND GRASS FEEDSTOCKS AND THEIR SORPTION OF ARSENATE AND LEAD ................................ .................... 32 Introduction ................................ ................................ ................................ ............. 32 Materials and Methods ................................ ................................ ............................ 34 Reagents ................................ ................................ ................................ .......... 34 Biochar Production ................................ ................................ ........................... 34 Biochar Ch aracterization ................................ ................................ .................. 35 As and Pb Sorption ................................ ................................ .......................... 36 Statistical Analysis ................................ ................................ ............................ 36 Results and Discussion ................................ ................................ ........................... 37 Elemental Composition ................................ ................................ .................... 37 Basic Properties ................................ ................................ ............................... 38 As(V) and Pb(II) Sorption ................................ ................................ ................. 39 Summary and Conclusion ................................ ................................ ....................... 41 3 REMO VAL OF ARSENIC BY MAGNETIC BIOCHAR PREPARED FROM PINE WOOD AND NATURAL HEMATITE ................................ ................................ ....... 51 Introduction ................................ ................................ ................................ ............. 51 Materials and Methods ................................ ................................ ............................ 53
6 Reagen ts ................................ ................................ ................................ .......... 53 Biochar Preparation ................................ ................................ .......................... 54 Characte rization ................................ ................................ ............................... 54 Adsorption Kinetics and Isotherm ................................ ................................ ..... 56 Sorption Models ................................ ................................ ............................... 56 Statistical Analysis ................................ ................................ ............................ 58 Results and Discussion ................................ ................................ ........................... 58 Biochars Property ................................ ................................ ............................. 58 Adsorption Kinetics ................................ ................................ ........................... 60 Sorption Isotherm ................................ ................................ ............................. 60 Sorption M echanisms ................................ ................................ ....................... 61 Summary and Conclusion ................................ ................................ ....................... 62 4 MANGANESE OXIDE MODIFIED BIOCHARS: PREPARATION, CHARACTERIZATION, AND SORPTION OF LEAD AND ARSENATE ................. 70 Introduction ................................ ................................ ................................ ............. 70 Materials and Methods ................................ ................................ ............................ 70 Reagents ................................ ................................ ................................ .......... 72 Sorbent Preparation ................................ ................................ ......................... 72 Mn oxide modified pine biochar (MPB) ................................ ...................... 72 Birnessite modified pine biochar (BPB) ................................ ..................... 72 Sorbent Characterization ................................ ................................ .................. 73 Adsorption Kinetics And Isotherm ................................ ................................ .... 73 Statistical Analyses ................................ ................................ .......................... 74 Results and Discussion ................................ ................................ ........................... 75 Sorbent Properties ................................ ................................ ........................... 75 As(V) And Pb(II) Sorption Kinetics And Isotherms ................................ ........... 77 A(V) Sorption Mechanisms ................................ ................................ ............... 78 Pb(II) Sorption Mechanisms ................................ ................................ ............. 79 Su mmary and Conclusion ................................ ................................ ....................... 80 5 SUMMARY AND CONCLUSION ................................ ................................ ............ 94 LIST OF REFERENCES ................................ ................................ ............................... 97 BIOGRAPHIC AL SKETCH ................................ ................................ .......................... 113
7 LIST OF TABLES Table page 2 1 Elemental composition of biochar made from loblolly pine, c itrus , a lfalfa , and s witchgrass pyrolysized at 300, 450 and 600 o C ................................ ................ 42 2 2 Basic physical and chemical properties of biochars ................................ ........... 43 2 3 Bulk density and water holding ability of the biochars ................................ ........ 43 2 4 biochar ................................ ................................ ................................ ................ 44 2 5 Pearson correlation coefficients (r) and significance of linear relationship between As(V) and Pb(II) sorption and physiochemical properties of biochars .. 44 3 1 Elemental composition and BET surface area of loblolly wood biochar and hematite modified biochar ................................ ................................ .................. 64 3 2 Kinetics and isotherm models and best fit parameters of As(V) sorp tion onto pine wood biochar and hematite modified biochar ................................ .............. 64 4 1 Elemental composition, surface area and pore volume of PB, MPB and BPB .. 81 4 2 Best fit parameters for kinetics and isotherm models of As(V) sorption onto PB, MPB and BPB ................................ ................................ ............................. 82 4 3 Best fit parameters for kinetics and isotherm models of Pb(II) sorption onto PB, MPB and BPB ................................ ................................ ............................. 83
8 LIST OF FIGURES Figure page 2 1 Linear regression analysis between pH value of biochar and content s of cationic elements in biochar ................................ ................................ ............... 45 2 2 T hermogravimetric analysis between 25 and 200 o C. ................................ ........ 46 2 3 Thermogravimetric curves of biochars derived fr om pinewood, citrus wood , alfalfa, and switchgrass . ................................ ................................ ..................... 47 2 4 Sorption of As(V) and Pb(II) onto the biochars. Error bars represent standard deviations. ................................ ................................ ................................ .......... 48 2 5 Linear regression analysis between Pb(II) sorbed and elemental content and pH of biochar ................................ ................................ ................................ ...... 49 2 6 Linear regression analysis between As(V) sorbed and elemental content and pH of biochar ................................ ................................ ................................ ...... 50 3 1 Thermogravimetric analysis of pine biochar and hematite modified biochar ...... 65 3 2 SEM EDS elemental mapping a nalysis for HPB ................................ ................. 66 3 3 XPS diffraction patterns of Fe2p1/2 and Fe2p3/2 spectra of HPB. ..................... 67 3 4 Magnetic pr operties of hematite mineral and hem atite modified biochar ............ 67 3 5 XRD diffraction patterns of hematite particles and HPB ................................ ..... 68 3 6 Kinetics and isotherm data and fitted models for As(V) sorption by loblolly pine wood biochar and hematite modified biochar .. ................................ ........... 69 4 1 XPS spectra of MnCl 2 2 O modified biochar and biochar modified with s ynthesized birnessite ................................ ................................ ........................ 84 4 2 Thermogravimetric (TG) c urves of pine wood biochar , MnCl 2 2 O modified biochar and biochar modified w ith synthesized birnessite ................................ . 85 4 3 XRD diffraction patterns of MnCl 2 2 O mo dified biochar before after As(V) and after Pb(II) sorption. ................................ ................................ ..................... 86 4 4 XRD diffraction patterns of biocha r modified with birnessite before , after As(V), and after Pb(II) sorption. ................................ ................................ .......... 87 4 5 As(V) and Pb(II) sorption kinetics data and fitted models ................................ ... 88 4 6 As(V) and Pb(II) sorption isotherm data and fitted models ................................ . 89
9 4 7 SEM image and correspondin g EDS spectra of modified biochars after As(V) sorption ................................ ................................ ................................ ............... 90 4 8 SEM image and corresponding EDS spectra of modified biochars after Pb(II) sorption. ................................ ................................ ................................ .............. 91 4 9 SEM/EDS elem ental mapping analysis of MnCl 2 2 O modified biochar after Pb(II) sorption. ................................ ................................ ................................ .... 92 4 10 SEM/EDS ele mental mapping analysis of biochar modified w ith synthesized birnessite after Pb(II) sorption. ................................ ................................ ............ 93
10 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 IRON (FE) AND MANGANESE (MN) OXIDE MINERAL MODIFIED BIOCHARS: CHARACTERIZATION AND REMOVAL OF ARSENATE AND LEAD By Shengsen Wang December 201 4 Chair: Yuncong Li Cochair: Bin Gao Major: Soil and Water Science Arsenic (As) and lead (Pb) in water are detrimental to human and environmental health. Adsorption is an effective technique for heavy metal removal from contaminated water. Biochar is an emerging sorbent for heavy metals because of its special surface properties, which vary with feedst ock types and pyrolysis temparture. To improve the heavy metal sorption abilit y of biochars, various methods have been developed in this study to modify pristine biochars. T he objectives of this study were to characterize iron (Fe) and manganese (Mn) oxid es mineral modified biochars and to determine their ability to remove arsenate (As(V)) and lead (Pb(II)) from aq u eous solutions . Pristine biochars were first made from four types of feedstock at three pyrolysis temperature (300, 450 and 600 o C). Natural he matite and synthetic birnessite were then used to modify biochars derived from pinewood to improve their sorption ability of As(V) and Pb(II). Experimental results indicated that sorption ability of the pri stine biochars to both As(V) and Pb(II) was affec ted by feedstock types significantly (p<0.001), but less affected by pyrolysis temperature. Electrostatic interaction played an important role in
11 controlling the sorption of both As(V) and Pb(II) onto the pristine biochars. In the modified biochars, Fe 2 O 3 ) as confirmed by X ray diffraction (XRD) analysis. Hematite modification nearly doubled As sorption of Fe 2 O 3 particles on the carbon surface served as sorption sites through s MnO 2 ) are the dominant crystals for the MnCl 4 Â·H 2 O modified biochar (MPB) and biochars modified with birnessite (BPB). Both MPB and BPB demonstrated much improved sorption ability over that of the unm odified biochars. While manganosite did not help the sorption of As(V) and Pb(II) onto the MPB, birnessite in BPB showed strong affinity to the heavy metals. Findings from this study can be used to develop strategies to apply biochar technology for remedi ation of heavy metals in the environment.
12 CHAPTER 1 INTRODUCTION AND LITERATURE REVIEW Introduction Environmental pollution from inorganic and organic contaminants pose s great threat to human health ( Greer, Goodman, et al., 2002 , Van Oostdam, Gilman, et al., 1999 ) and causes environmental damage and ecological deterioration ( Dong, Ma, et al., 2011 , Dong, Ma, et al., 2013 , Lee, 1973 , Liu, Huang, et al., 2013 , Zhang and Gao, 2013 ) . Besides its common use for carbon sequestration and soil amendment ( Gaskin, Steiner, et al., 2008 , Lehmann, Gaunt, et al., 2006 , Park, Hung, et al., 2013 ) , biochar is attracting more interest as good sorbent of contaminants ( Dong, Ma, et al., 2011 , Dong, Ma, et al., 2013 , Park, Hung, et al., 2013 , Yao, Gao, et al., 2011a , Yao, Gao, et al., 2011b , Yao, Gao, et al., 2012 ) . The removal efficiency of contaminants for biochar is affected by biochar characteristics associated with feedstock types and pyrolysis temperature ( Dong, Ma, et al., 2011 , Sun, Gao, et al., 2014 , Yao, Gao, et al., 2012 ) . Recent studies showed increased sorption for organic and inorganic contaminants of bioachars modified with iron (Fe) and aluminum (Al) salts ( Zhang and Gao, 2013 , Zhang, Gao, et al., 2013 ) , and montmorillonite and kaolinite ( Yao, Gao, et al., 2014 ) . These studies have launched a new research area for biochar research and Fe 2 O 3 ) was known for its sorption of heavy metals and organic contaminants, e.g., lead ( Pb ) ( O'Reilly and Hochella Jr, 2003 ) , molybdenum ( Mo ) ( Goldberg, Forster, et al., 1996 ) , copper ( Cu ) ( Grossl, Sparks, et al., 1994 ) , arsenic ( As ) ( Dixit and Hering, 2003 ) , and cipr ( Chen, Ma, et al., 2013 ) . Mn oxides were studied for their extraordinary sorption for Pb ( O'Reilly and Hochella Jr, 2003 ) an d arsenite (As(III)) ( Lafferty, Ginder Vogel, et al., 2010 , Lafferty, Ginder Vogel, et al.,
13 2011 ) . Once biochar is applied to soil, it becomes coated with some soil minerals as indicated by high er Fe , Si and Al content ( Xu, Wei, et al., 2013 ) .Thus, we hypothesize the modification of biochar with commonly occuring soil minerals that have high affinity for metals would improve efficacy of biochar for metal contaminant remediati on. Therefore, this dissertation was proposed to investigate biochar properties as affected by feedstock types and charring temperature , produce Fe and Mn oxide mineral modified biochar, and study the characteristics and mechanisms contributing to the so rption for contaminants. Biochar Biochar is pyrogenic charcoal as a product of thermal conversion of lignocellulose materials under oxygen free or limited condition as well as low value byproducts of thermal pyrolysis of biomass during biofuel productions ( Lehmann, Gaunt, et al., 2006 , Park, Hung, et al., 2013 , Rutigliano, Romano, et al., 2014 , Sun, Gao, et al., 2014 , Yao, Gao, et al., 2011a ) . Biochars are characterized by high carbon (C) content, large surface area, and abundant functional groups. The mass based carbon content of biochar is enriched relative to original feedstock and may range from 69 84 %, 70 7 9 %, 66 81%, 24 87%, 51 85%, and 74 86 % for hickory, bagasse, bamboo, loblolly p ine wood (Pinus toeda), Brazilian pepper, and peanut hull respectively as temperature increase from 300 700 o C ( Dong, Ma, et al., 2013 , Park, Hung , et al., 2013 , Sun, Gao, et al., 2014 , Yao, Gao, et al., 2011a ) . As C content increases, the oxygen ( O ) and hydrogen ( H ) decrease, which is an indication of condensation of biochar structure ( Harvey, Herbert, et al., 2011 ) . Surface area is crucial for chemical reactions and retention of chemicals. Surface area of biochars increased two orders of magnitude when production temperature increased from 300 to 600 o C and the surface area at
14 600 o C may range from 200 to 400 m 2 /g ( Sun, Gao, et al., 2014 , Yao, Gao, et al., 2011a ) . Biochar has become of increased interest to soil professionals as well as farmers because of their beneficial aspects to soil, including improvement of soil quali t y (water retention, soil aeration, soil p H, organic matter, soil microorganisms) ( Bell and Worrall, 2011 ) and soil fertility ( Zhang, Bian, et al., 2012 ) , alleviation of global warming (carbon seques tra tion and retention of greenhouse gases, e.g., N 2 O, CO 2 ) ( Case, McNamara, et al., 2012 , Cayuela, van Zwieten , et al., 2013 , Lehmann, Gaunt, et al., 2006 ) , and remediation of soil contaminated by heavy metal and organic contamin ants ( Dong, Ma, et al., 2013 , Mohan, Pittman, et al., 2007 ) . Biochars have been used in many settings . Lehmann et al ( Lehmann, Gaunt, et al., 2006 ) first proposed incorporation of biochar into soil as a C sequestration technique to mitigate global warming. Biochar production is the process to convert the plant biomass into recalcitrant carbon so that it can reduce production of carbon dioxide (CO 2 ) and its subsequent release into atmosphere ( Baldock and Smernik, 2002 , Rutigliano, Romano, et al., 2014 , Sohi, Krull, et al., 2010 ) . B iochar could also reduce N 2 O gas emission from soil, which account s for two thirds of nitrous oxide ( N 2 O ) emission int o atmosphere ( Case, McNamara, et al., 2012 , Cayuela, van Zwieten, et al., 2013 , Zhang, Cui, et al., 2010 ) , but has v aried potential for methane reduction ( Karhu, Mattila, et al., 2011 , Zhang, Bian, et al., 2012 ) . In addition, biochar is used as soil amendmen t to increase soil fertility and crop yield ( Lehmann, Gaunt, et al., 2006 , Marris, 2006 , Zhang, Bian, et al., 2012 , Zhang, Cui, et al., 2010 ) , and to favorably impact soil physical and chemical properties ( Githinji, 2014 ) , e.g., increasing soil organic
15 carbon and soil pH , decreas ing soil bulk density ( Bell and Worrall, 2011 , Xu, Wei, et al., 2013 , Zhang, Bian, et al., 2012 ) , and increas ing water holding capacity and reducing nutrient leaching ( Bell and Worrall, 2011 , Karhu, Mattila, et al., 2011 ) . Although biochar can release some nutr ients especially calcium ( Ca ) and magnesium ( Mg ) into soil solution as a slow release material to supplement plant nutrition ( Deenik, McClellan, et al., 2010 , Mukherjee and Zimmerman, 2013 ) mainly related to high retention capacity of soil nutrients such as P and nitrogen ( N ) ( Knowles, Robinson, et al., 2011 , Xu, Wei, et al., 2013 , Yao, Gao, et al., 2012 ) which result in higher soil nutrients content. Recent studies showed that biochars have good sorption capacity for organic and ino rganic contaminants from soil and water. For example, biochar is gaining wide recognition for its capacity on sorption of inorganic contaminants such as chromium (Cr) ( Dong, Ma, et al., 2011 ) , mercury (Hg) ( Dong, Ma, et al., 2013 ) an d Pb ( Uchimiya and Bannon, 2013 ) , Cu ( Trakal, KomÃ¡rek, et al., 2011 , Uchimiya and Bannon, 2013 ) and zinc (Zn) ( Trakal, KomÃ¡rek, et al., 2011 ) . Biochar also inter act s wi th organic molecules e.g., biochar decrease simazine biodegradation and leaching ( Jones, Edwards Jones, et al., 2011 ) , methylene blue ( Zhang and Gao, 2013 ) . The functionality of biochars in environment is related to several factors and/or mechanisms. In general, large surface areas are favorable for sorption capacity of biochar for these organic and inorganic chemicals. There are mainly three mechanisms for biochar sorption of organic and inorganic chemicals, including electrostatic attraction, ion exc hange and surface complexion ( Dong, Ma, et al., 2011 , Harvey, Herbert, et al., 2011 , Mohan, Pittman, et al., 2007 ) . The sorption mechanism also varie s
16 for different target chemicals. The mechanisms were discussed by some author s. Dong ( Dong, Ma, et al., 2011 ) examined sorption of biochar d erived from sugar beet tailing for Cr(VI), and found sorption may induce change of functional groups, and thus hypothesized the sorption was related to electrostatic attraction of Cr(VI) to protonated carboxylic, alcohol and hydroxyl functional groups, and surface complexation of Cr(III) with functional groups followed by reduction of Cr(VI). Dong ( Dong, Ma, et al., 2013 ) also characterized the sorption of Hg i n Brazilian pepper biochar, and found surface complexation was the dominating sorption mec hanisms which was supported with Fourier transform infrared spectroscopy (FTIR) and X ray photoelectron spectroscopy (XPS) analysis. Harvey et al ( Harvey, Herbert, et al., 2011 ) use flow adsorption microcalorimetry method to determine the energy change associated with sorption proce ss and states that sorption of c adium by biochar is thr ough complexation process. Mohan et al ( Mohan, Pittman, et al., 2007 ) conclude sorption of arsenic, cadmium and lead by biochar happens with ion exchange process based on their observation of displacement of H and cations whe n sorption occurs. However, Yao et al ( Yao, Gao, et al., 2011b ) did not find change of three significant functional groups on the biocha r (O H, C O and C H binding) when phosphorus (P) was sorbed on the biochar. Instead, Yao et al ( Yao, Gao, et al., 2011a ) detected appear ance of nano sized periclase (MgO) on the biochar surface and attributed enhanced P sorption to the formation of this metal oxide, which was confirmed with Scanning electron microscope (SEM) and Energy dispersive X ray fluorescence spectroscopy (EDS) analy sis showing the elevated P peak on MgO crystal on the biochar post sorption. Their hypotheses was also confirmed with Mg enriched biochar derived from tomato biomass with higher Mg concentrations,
17 which showed greater sorption of P associated with formatio n of MgO and Mg hydroxides ( Yao, Gao, et al., 2013 ) . Electrostatic force was also supported with comparison of two biochars which showed less sorption on sugar beet tailing biochar with more negative zeta potential ( Yao, Gao, et al., 2011a ) . Feedstock Feedstock type is an important factor to determine biochar properties. Biochar properties including cation exchange capacity (CEC), surface area, pH, and functional groups determine its practic al application as soil amendment and potential sorbent. Wood and grass are common lignocellulose biomass for biofuel production. Woody and herbaceous plant materials are representative feedstock biomass, having different biomass composition. Pine wood has higher cellulose, hemicellulose and lignin than herbaceous species, such as switch grass and alfalfa ( Lupoi and Smith, 2012 ) . Lupoi and Smith ( Lupoi and Smith, 2012 ) characterized the properties and structures of several grass and wood species in cluding switchgrass, alfalfa, and pine wood, and found pine wood has greater lignin and cellulose content on dry biomass basis. Harvey et al ( Harvey, Herbert, et al., 2011 ) found cinnamic acids are dominant functional groups for grass derived biochars, and benzoic acid derivatives are dominant groups for wood derived biochars, and thus attributed the gr eater sorption for K on grass biochar to ethylenic group in cinnamic acids grass biochar. The higher lignin content in plant biomass w as reported to yield higher bio char production rate which decreased with increasing temperatures, and cellulose decomposition is more favorable at faster temperature increase ( Demirbas, 2004 ) .Thus, it would be he lp ful to understand the sorption mechanisms between grass and wood derived biochar which will dete rmine its possible use.
18 Florida citrus production contributed to 65 % of total U.S. citrus production in the 2011 2012 seasons ( Summary, 2012 ) . The Florida citrus industry is suffering from a severe infectious disease Huanglongbing (HLB), caused by phloem limited, gram ( Bove, 2006 , Coletta, Takita, et al., 2005 ) . Citrus HLB was firstly reported and described as a yellowing and leaf mottle in Southern China in the late twentieth century (Zhao, 1981). In United States, citrus HLB disease was first confirmed in Florida in 2005, and big acreage and yield losses were reported eig ht years later. HLB disease, also known as greening disease, was considered as one of the oldest and most destructive bacterial citrus disease s due to its infectious nature and absence of effective control measures (BovÃ©, 2006). The HLB disease was reporte d to cause an extensive loss across all the infected regions in Asia and Africa (Aubert, 1992). HLB infection disrupts transport of carbohydrate from source leaves to roots ( Rosales and Burns, 2011 ) , which contributes to starch accumulation in symptomatic leaves ( Kim, Sagaram, et al., 2009 ) , and low level in roots ( Etxeberria, Gonzalez, et al., 2009 ) . A recent study showed HLB infection decreased fibrous root mass density by 27 40% ( Graham, Joh nson, et al., 2013 ) . Although intensive biological and chemical management practices have been implemented in Florida ( Qureshi and Stansly, 2009 , Zhang, Powell, et al., 2011 ) , the tree decline is still unavoidable. The dead trees are usually remov ed from the grove to eliminate the source of infectious bacteria ( Gottwald, 2010 ) . To our knowledge, no attempts have been made to make biochar from HLB affected citrus trees for removal of contaminants. The use of citrus derived biochar could not only eliminate the bacteria but also have potential to mitigate environmental contaminati on.
19 Loblolly pine ( Pinus taeda ) is a perennial pine species, native in Southeastern U.S., and a very common wood species in the U.S. (Brender et al., 1981). The pine is also cultivated for its important commercial use in lumber and the pu lp /paper industry as pu lpwood ood (Langdon, 1979). The pine is also a feedstock for biofuel production. Recently, pine wood was thermally pyrolyzed under N 2 condition to make biochar and was investigated with respect to its sorption capacity for cadmium (Cd) ( Harvey, Herbert, et al., 2011 ) . The loblolly pine wood biochar made under pure N 2 and 93% N 2 plus 7% O 2 were comp ared for their surface and elemental characteristics and sorption for endocrine disrupting compounds, indicating N 2 biochar has higher surface area and pore volume resulting in higher sorption capacity for organic compounds ( Jung, Park, et al., 2013 ) . Pine wood biochar also showed strong sorption capacity for ammonium and nitrate N ( Sika and Hardie, 2014 ) . Switchgrass ( Panicum virgatum L .) is a native perennial grass which can be used as hay and pasture. Switchgrass has intensive and robust root systems extending as much as 3 m below ground, which may be related to its low requirement of soil fertility, e.g., optimum yie ld was observed at N rate 120 kg N ha( 1), and good adaption to unfavorable soil conditions such as drought and flooding ( Jensen, Clark, et al., 2007 , Vogel, Brejda, et al., 2002 ) . Due to low cost to ma intain productivity and high yielding potential, e.g., 1 L ethanol may cost averaged $0.13 input based on observation from Perrin et al., Switchgrass are now widely used as feedstock for biofuel production ( Jensen, Clark, et al., 2007 , Perrin, Vogel, et al., 2008 ) . As a byproduct biofuel conversion process, switchgrass biochar has been used for soil amendment. Switchgrass biochar increased moisture storage of Ultisols and Aridisols which was
20 affected by pyrolysis temperature ( Novak, Busscher, et al., 2012 ) . Nutrient retention capacity for major soil macro and micro nutrients of switchgrass ( Panicum virgatum L.) derived biochar was assessed in two Arid sols, showing higher retention capacity for switchgrass biochar made at 250 o C compared to 500 oC ( Ippolito, Novak, et al., 2012 ) . Alfalfa ( Medicago sativa L .) is a perennial legume ( Fougere, Lerudulier, et al., 1991 ) which is widely planted in m an y countries. U nited States contains 41% o f world alfalfa hectares and has largest productions (USDA National Agricultural Statistics Service, available online: http://www.nass.usda.gov). Alfalfa is mainly planted for livestock fodder. Like the perennial gras s switchgrass, alfalfa can also be used for biofuel production due to its good potential for biomass conversion to bioethanol ( Dien, Miller, et al., 2011 , Fiasconaro, Gogorcena, et al., 2012 , Hendrickson, Schmer, et al., 2013 ) . Alfalfa is usually symbiotically associated with rhizobia and thus increases nitrogen fixing capacity, fostering high fermentable carbohydrate concentrations ( Fiasconaro, Gogorcena, et al., 2012 ) . Due to its use in biofuel production, the byproducts will be produced during the process of converting biomass to ethanol. Arsenic (As) A rsenic is a toxic heavy metal loid with atom ic number 33 and is the 20 th most abundant element in natural environment ( Mohan and Pittman Jr, 2007 ) . As is det rimental to ecological safety and human health and is listed as a hazardous substance ( Graeme Md and Pollack Jr, 1998 , Manning and Goldberg, 1996 , Payne and Abel Fattah, 2005 , Singh, Ghosh, et al., 2014 ) . Exposure to As may cause acute poisoning causing abdominal pain and death, and chronic disease such as kidney, lung, bladder cancer ( Mohan and Pittman Jr, 2007 ) . Arsenic contamination in ground water and soil has been documented in man y countries ( Gulz, Gupta, et al., 2005 , Tuutijarvi,
21 Lu, et al., 2009 ) , including some Asian counties and the United States (U.S.) . Its acute poisoning causes 20,000 deaths each year ( Payne and Abel Fattah, 2005 ) . Due to its low lethal dose and severe toxicity , a very low level (0.01 m g L 1 ) of allowance concentration is prescribed for drinking water by world health organization (WHO) and U.S. Environmental Prote ction Agency (U.S.EPA) ( M ohan, Pittman, et al., 2007 , Singh, Ghosh, et al., 2014 , U.S.EPA, 2002 ) . In the U.S., as per U.S.EPA report in 2002, western states tend to have higher groundwater As levels than states in Midwestern a nd North Central regions ( U.S.EPA, 2002 ) . The background As concentration in uncontaminated soil is 0.1 40 mg kg 1 with the average content 5 to 6 mg kg 1 ( Gulz, Gupta, et al., 2005 , Gulz, 2002 ) . In Florida, the concentration ranges and ranges 0.03 50.6 and 0.01 38.2 mg kg 1 with upper baseline concentration as 6.21 and 7.63 mg kg 1 respectively in undisturbed and disturbed soil , respectively ( Chen, Ma, et al., 2001 ) . The maximum As allowance concentration in agricultural use is 20 mg kg 1 by European Union (EU) ( Bhattacharya, Samal, et al., 2010 ) and 10 mg kg 1 by Swedish environmental protection agency, 25 mg kg 1 by Canada, 20 mg kg 1 by New Jersey US, 0.8 m g kg 1 for residential soil by Florida state ( Chen, Ma, e t al., 2001 ) . A rsenic exists in both organic and inorganic forms in aqueous systems depending on medium pH and oxidation conditions ( Chowdhury, Yanful, et al., 2011 , Graeme Md and Pollack Jr, 1998 ) . It mainly exist s in different oxidization states such as trivalent (As( I II ) ) and heptavalent ( As(V) ), although neutral (As( 0 ) ) and negative trivalent forms can also be found ( Chowdhury, Yanful, et al., 2011 , Cullen and Reimer, 1989 ) . Inorganic forms such as arsenate (V) and arsenite (III) are predo minant forms in natur e and are the main sources of As related environmental concerns ( Chowdhury,
22 Yanful, et al., 2011 , Cullen and Reimer, 1989 , Graeme Md and Pollack Jr, 1998 , Mohan, Pittman, et al., 2007 ) . A s (III) is the most toxic form among all inorganic As species ( Lafferty, Ginder Vogel, et al., 2011 , Lafferty, Ginder Vogel, et al., 2010 , Winship, 1984 ) . pH and redox potential dominate inorganic As speciation ( Mohan, Pittman, et al., 2007 , Wang and Mulligan, 2006 , Wang, Lee, et al., 2012 ) . Wang and Mulligan ( Wang and Mulligan, 2006 ) diagrammed the As ion speciation at different pH and redox potential (Eh). Mohan et al ( Mohan, Pittman, et al., 2007 ) summ arized that at pH below about 9.2, neutral H 3 AsO 4 is the dominant species under oxidizing condition while HAsO 2 is dominant under reducing conditions . However, HAsO 4 2 dominate s at pH above 9.2 , independent of oxidation reduction potential . At natural pH of aqueous environment, H 3 AsO 3 and H 2 AsO 2 are the main forms for arsenite (pK a1 =9.2 and pK a2 =12.7 for H 3 AsO 3 ), and H 2 AsO 4 and HAsO 4 2 the main forms for arsenate (pK a1 =2.3 and pK a2 =6.8 and pK a3 =11.6 for H 3 AsO 4 ) ( Goldberg, 2002 , Masue, Loeppert, et al., 2007 ) . A rsenic accumulation may be from both natural and anthroponic activities. It naturally occurs in soil formation derived from parent materials and can be mobilized by weathering reaction, biological activity and geochemical reactions ( Mohan and Pit tman Jr, 2007 ) . Anthropogenic As from discharge and disposal of As containing compounds is another source for As contamination ( Mandal and Suzuki, 2002 ) . Anthropogenic sources , which are three fold greater in extent than naturally occurring As, include 80% from agricultural use in 1970s, insectisid es, herbicides, desiccants, wood preservatives, feed additives, drugs, and mining of As bearing ore ( Ferguson and Gavis, 1972 , Mandal and Suzuki, 2002 ) .
23 Considering the severe detriment to environmental and human health, enormous efforts have been made for As reclamation. According to U . S . EPA, the common remediation techniques for As removal include immobilization of the As using solidification/stabilization (S/S), vitrification, soil washing, pyrometallurgical recovery and in situ soil flushing for soil and waste, precipitation/coprecipitation, membrane filtration, adsorption, ion exchange, permeable reactive barriers for water tr eatment, eletrokinetic treatment, phytoremediation ( Ma, Komar, et al., 2001 ) , and biological treatment for soil, waste and water treatment ( U.S.EPA, 2000 , U.S.EPA, 2002 ) . Biorem ediation by the plant hyperaccumulation is also a good alternative for As contaminated soils. For examples, fern Pteris vittata can accumulate As in fronds up to 20 g kg 1 dry mass ( Ma, Komar, et al., 2001 ) . Adsorption is currently a prevailing technique for As removal. Mohan et al . ( Mohan and Pittman Jr, 2007 ) rev iewed the low cost sorbent s for a rsenate and arsenite removal, including clay minerals, metal (hydr)oxides, and activated carbon. As an emerging technology, thermal pyrolysis derived material also shows significant sorption capac ity for As. For example, As concentration was increased in pore water but bio available As was reduced based on their observation of reduced As concentration in biomass of tomato (Solanum lycopersicum L.) ( Beesley, Marmiroli , et al., 2013 ) . Water soluble As was retained by biochar but As concentration in leachate was not reduced significantly with a column leaching experiment ( Beesley and Marmiroli, 2011 ) . Modified biochar with AlCl 3 and FeCl 3 showed good affinity for arsenate based on batch experiment ( Zhang and Gao, 2013 , Zhang, Gao, et al., 2013 ) . Lead (Pb) Lead is one of the most common toxic heavy metal with atom ic number 82 and is the 37 th It mainly exist in P b 2+ (PbO) form such
24 as in the minerals massicot and litharge but can also occur in Pb 4+ form as in the mineral minium ( lead tetroxide ; Pb 3 O 4 ) where it exists in both oxidation states . Although lead may occur in nature, the majority of lead in soils is anthropogenic. The history of lead use date s back to thousands of years. Lead may originate from fuel combustion, paint, printing, battery , and m ining ( Lee, Kim, et al., 2013 , Mohan, Kumar, et al., 2014 , Mohan and Pittman Jr, 2007 ) . The maximum allowance concentr ation recommended by USEPA is 0.015 mg L 1 . In a U.S. Geological Survey report in 1999, 19% and 1.3% of total and dissolved Pb concentrations in ground water samples collected statewide exceeds 0.015 mg L 1 , respectively ( Katz, Berndt, et al., 1999 ) . According to a USGA lead background survey, the highest concentrations are found in the states of Nevada, Pennsylvania , and Washington states with mean soil lead level above 40 mg kg 1 , while in Florida, soil lead concentration is minimal with a mean level of 6.4 mg kg 1 ( Smith, Cannon, et al., 2013 ) . Lead contamination causes damage to nerve s ystems and is more serious to children ( Needleman, 2004 ) their exposure to lead such as lead contaminated house dust ( Lanphear, Matte, et al., 1998 ) and lead contaminated soil ( Weitzman, Aschengrau, et al., 1993 ) . A number of studies have been conducted to test Pb remediation approaches . Pb concentration in soil correlates well with Mn oxide concentrations ( McKenzie, 1980 ) . Futher studies showed birnessite, a common Mn oxide with layer ed structure , has capacity to captu re Pb in the interlayers ( Lee, Kim, et al., 2013 , McKenzie, 1980 , O'Reilly and Hochella Jr, 2003 ) . Iron oxide s such as ferrihyite removes Pb from aqueous solution by forming dentated complex es ( Trivedi, Dyer, et al., 2003 ) . Phosphate rock can remove Pb by precipitation mechanism ( Cao, Ma, et al., 2004 ) . Like
25 As bioremediation, P b contaminated soil can also be reclaimed by p hytoremediation such as with Brassica junce, Vetiveria zizanioides, Cardaminopsis halleri , and microorganisms to reduce Pb to no n toxic forms ( Akhtar, Chali, et al., 2013 ) . Apart from the se common sorbents, biochar also shows affinity for Pb retention ( Mohan, Kumar, et al., 2014 , Mohan, Pittman, et al., 2007 ) . Hydrous Iron and Manganese Oxides Hydrous and anhydrous oxides, hydroxides, and oxyhydroxides of metals are weathering products of primary minerals under intense weathering condition s such as high temperature and heavy rainfall . Iron and Mn rank 4 th and 12 th Both oxides play important roles in retaining soil nutrient such as P , especially in sandy and highly weather ed soils in which CEC tends to be low . They are also good sorbent s for organic and inorganic contaminants. (Hydro) oxide s of Fe are the most abundant metal oxide minerals in soil ( Essington, 2004 ) . Fe 2 O 3 ), Fe 2 O 3 ), ferrihyrite (~Fe 5 HO 8 Â·4H 2 O), and magnetite (Fe (+2) Fe (3+) 2 O 4 ) ( Kim, Suh, et al., 2012 ) . Goethite is the most common secondary Fe mineral in soil, and hematite is the second most common which is more easily found under warm climate ( Majzlan, Grevel, et al., 2003 ) . Hematite is the main ore of Fe mining. Hematite has a dioctahedral structure of the trigonal crystal system and rhombohedral lattice system . Hematite is positively c harged in soil environments with point of zero charge (pHpzc) at around 8.5 ( Essington, 2004 ) . Maghemite is a polymorph of hematite , having the same composition but a different crystal structure (cubic crystal system) . Of these iron oxide minerals, maghemite and magnetite are sufficiently magnetic to be attracted by an external magnet.
26 Hydrous Mn oxides are the second most abundant oxide minerals in soil ( Post, 1999 ) . The Mn oxides are very reactive minerals in soil , participating in important redox reactions . M anganese ores are used for pigments, catalyst and mainly for steel production ( Post, 1999 ) . Reducing conditions are more favorable to formation of these oxide minerals. Unlike Fe oxides, Mn oxi des usually have poor crystallin ity( small crystal size and poorly ordered structure ) ( Essington, 2004 ) . The common Mn oxides in surface environment s include phyllomanganate and tectomanganate. Birnessite, one of the most abundant Mn oxides, is a representative phyllomanganate which has layers of edge sharing MnO 6 octahedra ( Gaillot, Drits, et al., 2007 ) . T he dominant Mn form in octahedral is Mn 4+ , and the vacancy can be compensated with Mn 3+ which produces permanent negative charge. Exchangeable hydrated cations such as potassium (K), sodium (Na), and Ca apprear in the interlayer to balance the negative charges induced by isomorphoric substitution. However, these ions are not essential, and synthetic birnessite may not contain these ions. The incorporation of these cations do es affect the crystalline, e.g., d spacing in XRD analysis ( Cheney, Bhowmik, et al., 2008 , Gaillot, Drits, et al., 2004 , Lee, Kim, et al., 2013 ) . The birnessite has low pHpzc which is around 3. I n contrast, t ectomanganate, e.g., todorokite, has tunnel structure, where the edge sharing chains of Mn octahedral are connected with corners from other chains ( Essington, 2004 ) . Iron oxide s can be synthesized in the lab oratory . Hematite was prepared by hydrolysis of Fe(III) solution, and goethite by heating Fe(III) and alkaline solution ( O'Reilly and Hochella Jr, 2003 ) . M a ghemite was synthesized by coprecipitation of Fe 2+ /Fe 3+ (1:2) and oxidation with HClO 4 ( Tronc, Ezzir, et al., 2000 ) . Magnetite was
27 produced by reaction of Fe(II ) and 25% NH 4 OH solution, maghemite was obtained by heating magnetite at 200 o C, goethite was synthesized with Fe(III) and NaHCO 3 , and hematite was synthesized by heating goethite between 600 and 900 o C ( Legodi and de Waal, 2008 ) . As seen from synthesis process, the iron bearing oxides undergo transformations with change of conditions. Iron oxide minerals may undergo transformation to other forms of iron oxides with changes of composition or crystalline structure ( Essington, 2004 ) . This transformation is mainly affected by temperature a nd availibity of oxygen and carbon. For examples, goethite and siderite ( FeCO 3 ) can be thermally converted to hematite at 350 and 500 o C ( Ramirez Muniz, Jia, et al., 2012 ) . Hematite may be transformed to maghemite at above 530 o C when carbon is available, and to magn etite when N and S are present ( Minyuk, Subbotnikova, et al., 2011 ) . Similarly, Mn oxide can also be synthesized in laboratory. There are several methods to synthesize birnessite. Bascially, the synthesis may include oxida tion of manganous hydroxide or the reduction of potassium permanganate ( McKenzie, 1971 ) . Apart from the traditional synthetic methods, such as reduction of potassium permanganate (KMnO 4 ) with concentrated HCl, it can also be synthesized by thermal decomposition of KMnO 4 at high temperature such as from 170 Â°C to 1000 o C ( Gaillot, Drits, et al., 2007 , Gaillot, Drits, et al., 2004 ) or slow cooling process of brown birnessite to nanozised black birnessite with thermal stability up to 400 o C ( Cheney, Bhowmik, et al., 2008 ) . Hydrous metal oxide minerals are usually characterized by notably reactive surface ( Peacock and Sherman, 2004 ) and amphoteric changes ( ( Essington, 2004 ) . Arsenate tends to accumulate on the surface s of Al , Fe, and Mn oxides and clay
28 minerals ( Ferguson and Gavis, 1972 , Mandal and Suzuki, 2002 ) such that these components are usually associated with increased sorption capacity f or As ( Goldberg, 2002 ) ( Ferguson and Gavis, 1972 , Gulz, 2002 , Mandal a nd Suzuki, 2002 ) . For example, As concentration in lake sediments of Florida was found to be well correlated with Al and Si content ( Whitmore, Riedinger Whitmore, et al., 2008 ) . A rsenic is more bioavailable to plants and has higher phytotoxicity to plants in sandy soils than other soils and Gulz et al ( Gulz, Gupta, et al., 2005 ) ascribed the greater plant uptake of As to higher As concentration in soil solution resulting from low Fe and Al oxides content. Thus, minerals such as Fe oxide s ( Gimenez, Martinez, et al., 2007 ) , Al oxide s ( Anderson, Ferguson, et al., 1976 , Goldberg, 2002 , Masue, Loeppert, et al., 2007 ) , Mn oxide s ( Lafferty, Ginder Vogel, et al., 2010 , Lafferty, Ginder Vogel, et al., 2011 , Lafferty, Ginder Vogel, et al., 2010 , Ying, Kocar, et al., 2012 ) and to a less extent naturally occurring kaolinite, montmorillonite and illite ( Goldberg, 2002 , Manning and Goldberg, 1996 , Mohapatra, Mishra, et al., 2007 ) showaffinity for As and good poten tial as sorbent for As removal. Hematite is mostly studied for its removal of As contamination in water and soil systems ( Catalano, Park, et al., 2008 , GimÃ©nez, MartÃnez, et al., 2007 , Guo, Stuben, et al., 2007 , Mamindy Pajany, Hurel, et al., 2009 , Ramirez Muniz, Jia, et al., 2012 ) . Thus, in this study we proposed to modify biochars with representative Fe and Mn oxide minerals such as hematite and birnessite , respectively . The sorption of metal and organic compounds is mainly associated with inner sphere surface complexation ( Chen, Ma, et al., 2013 , Dixit and Hering, 2003 , Goldberg, Forster, et al., 1996 , O'Reilly and Hochella Jr, 2003 ) . The mechanism of As(V) sorption onto either natural or synthetic
29 hematite was investigated with both macroscopic and microscopic approaches. The macroscopic approach involved the stud y of As sorption behavior at different ionic strength or change of point of zero change (pHzpc) ( Goldberg and Johnston, 2001 , Mamindy Pajany, Hurel, et al., 2009 ) . For examples, As(V) sorption on natural hematite was shown to form both inner sphere and outer sphere complexes (including electrostatic attraction and binding to surface oxygen co ntaining functional group using X ray scattering analysis ( Catalano, Park, et al., 2008 ) . As (V) sorption on commercial hematite does not change with ionic strength and thus involve s formation of inner sphere surface c omplex ( Mamindy Pajany, Hurel, et al., 2009 ) . As(V) sorption on amorphous Fe oxides as inner sphere surface complexes was inferred based on the observation that point of zero charge changed with As (V) sorption ( Goldberg and Johnston, 2001 ) . Surface complex models which consider inner sphere sorption are also successfully implemented in agreement with macroscopic data ( Goldberg, 2002 ) . Microscopic approaches, which agree well with macroscopic approach es , are carried out with the aid of modern techniques or instruments such as X ray scattering analysis ( Catalano, Pa rk, et al., 2008 ) . Due to impurity of naturally occurring hematite, their sorption for As has been compromised. Many authors instead synthesize the iron oxides to enhance their sorption for As. Therma l l y converted hematite showed high sorption for As (V ) (5.69 8.94 mg g 1 ) ( Ramirez Muniz, Jia, et al., 2012 ) . S ynthetic iron oxide nanoparticles ha ve enhanced As sorption by virtue of their greater reactivity because of elevated specific surface area ( Auffan, Rose, et al., 2008 ) . Birnessite is a phyllomanganate where layer MnO 6 octahedron share edges with adjacent octahedral unit, and isomorphic substitution of cations such as Mn 3+ for Mn 4+
30 generates structural negative changes ( Essington, 2004 , Gaillot, Drits, et al., 2004 , Gaillot, Flot, et al., 2003 ) . Birnessite is a very important scavengers of cationic heavy metals ( O'Reilly and Hochella Jr, 2003 ) , especially Pb which is captured in the between layers. For example, previous work showed Mn oxide c oating in soil has higher contributioin for Pb sorption than Fe oxide coating ( Dong, Nelson, et al., 2000 ) . Because Mn exists mainly as Mn 4+ form in birnessite, this oxide shows oxi di zing potential to react with reducing reagents. For examples, birnessite was found to oxidize As (III) to As(V) in accordance with Mn 4+ reduction to Mn 3+ , while the As(V) form ed an As(V) MnO 2 complex ( Manning, Fendorf, et al., 2002 , Post, 1999 ) . Hypotheses and Research Objectives Feedstock types and pyrolysis temperature impact biochar properties, which may affect the ir sorption capacity for As (V) and Pb (II) . The Fe and Mn oxide minerals are good sorbent s for As(V) and Pb(II) , and thus modification of biochar with these minerals may increase biochar sorption capacity. Th us , the formulated hypotheses for this study were: Hypotheses Feedstock types and pyrolysis temperature impact biochar properties and thus sorption of As(V) and Pb(II) Hematite modified biochar has increas ed sorption capacity for As (V), possibly due to the effect of maghemite nano particles Synthetic b irnessite precipitates onto biochar which increase s sorption capacity for As(V) and Pb(II) . The sorption is mainly associated with birnessite on biochar modified by birnessite.
31 Objectives The overall objective of this study was to compare properties of biochar produced at different pyrolysis temperature and feedstock, and modify pris tine biochar s with hematite and birnessite as well as minerals to increase sorption for As (V) and Pb (II) . The specific objectives addressed in this dissertation were to: Compare biochar physical and chemical properties as affected by pyrolysis t emperature and feedstock types and the sorption potential for As(V) and Pb(II) . Evaluate the effect of hematite modification on biochar properties and As (V) sorption and i nvestigate the possible mechanism associated with the As(V) sorption. Evaluate two synthetic methods to produce Mn oxides modified biochars and i nvestigate the possible mechanism associated with As(V) and Pb(II) sorption.
3 2 CHAPTER 2 PYROLYSIS TEMPERATURES AFFECTED CHARACTERISTICS OF BIOCHAR PRODUCED FROM WOODY AND GRASS FEEDSTOCKS AND THEIR SORPTION OF ARSENATE AND LEAD Introduction Biochar is pyrogenic organic matter derived from the pyrolysis of biomass, such as wood and grass, under N 2 or oxygen limited conditions ( Lehmann, Gaunt, et al., 2006 , Rutigl iano, Romano, et al., 2014 ) . Recent studies have suggested that biochar land application can mitigate global warming by converting plant biomass into recalcitrant carbon ( Lehmann, Gaunt, et al., 2006 ) and reducing N 2 O gas emission from soil ( Spokas, Koskinen, et al., 2009 ) . In addition, biochar amendment can, in some instances, increase soil fertility and crop yield ( Major, Rondon, et al., 2010 ) , and improve soil physical and chemical properties such as water holding and nutrient r etention capacity ( Bell and Worrall, 2011 , Karhu, Mattila, et al., 2011 ) . Biochar properties are strongly affected by their peak pyrolysis temperature ( Pa rk, Hung, et al., 2013 ) . As pyrolysis temperature increases, the degree of carbonization of the feedstock increase, as indicated by increased carbon (C) content as well as decreased hydrogen (H) and oxygen (O) contents in the resulting biochar ( Harvey, Herbert, et al., 2011 , Uchimiya, Wartelle, et al., 2011 ) . Biochar made at low temperature generally has lower pH, higher water holding capacity, lower specific surface area, more carboxylic group and phenolic hydroxyl functional groups and higher cation exchang e capacity (CEC) ( Gaskin, Steiner, et al., 2008 , Ippolito, Novak, et al., 2012 , Novak, Busscher, et al., 2012 ) . These properties deter mine its potential environmental applications. For instance, biochar surface properties such as BET
33 surface area and pore volume were correlated with its ability to sorb phenanthrene ( Par k, Hung, et al., 2013 ) . Feedstock types may also affect the physicochemical properties of the biochar. For instance, woody biomass often has higher cellulose, hemicellulose, and lignin contents than herbaceous or grass species ( Lupoi and Smith, 2012 ) . The higher lignin content in plant biomass was reported to promote carbonization and to increase biochar production rate ( Demirbas, 2004 ) . Furthermore, previous studies have shown that surface area, pH, and functional groups, to affect their potential environmental applications ( Kloss, Zehetner, et al., 2012 , Sun, Gao, et al., 2014 ) . Because of its special surface properties, biochar has been suggested to be a low cost adsorbent for soil remediation and water treatment ( Cao, Ma, et al., 2009 , Inyang, Gao, et al., 2011 , Yao, Gao, et al., 2011 ) . Soil contamination by heavy metals is detrimental to crop production and human health. Arsenic (As) and lead (Pb) are carcinogenic trace metals and are of great concern for human and animal health. Of the many well develo ped approaches for As and Pb removal from soil and groundwater systems, adsorption is the one of the most discussed techniques ( Mohan and Pittman Jr, 2007 ) . Several studies have reported the strong sorption of heavy metals, particularly lead, by biochars produced from various feedstocks and under different conditions ( Inyang, Gao, et al., 2012 , Uchimiya, Lima, et al., 2010 , Xue, Gao, et al., 2012 ) . However, it is still unclear which types of biochar remove heavy metals most efficiently and what biochar char acteristics are most important in this process. To our knowledge, no previous works have systematically investigated the separate and
34 combined effect of pyrolysis temperature and feedstock type on biochar sorption of heavy metals in aqueous solutions. The main objective of this work was to understand how the physicochemical properties of biochar, arising from variations in pyrolysis temperature and feedstock were selected as the feedstocks to produce biochars at three pyrolysis temperatures. A range of laboratory tools were used to characterize and compare the properties of the biochars. In addition, batch sorption experiments were conducted to measure their sorption of As and Pb in aqueous solutions. Materials and Methods Reagents All chemicals of used in this study were of analytical grade and were purchased (Nanopure water, Barnstead) to produce stock solutions of sodium arsenate dibasic heptahydrate (Na 2 HAsO 4 2 O) and lead nitrate (Pb(NO 3 ) 2 ) which were stored in a refrigerator and later diluted with the same water to make experimental solutions. Biochar Production Loblolly pine ( Pinus taed ) and Hamlin citrus ( C. sinensis L. Osb ) trees, which are distributed widely in the southeastern U.S., were selected as the woody feedstock. The Hamlin citrus tree s were infected with Huanglongbing (HLB) disease and are usually burned on site. Switchgrass ( Panicum virgatum L.) and alfalfa ( Medicago sativa ), which are common bioenergy crops, were selected as the herbaceous feedstocks. The loblolly pine wood (PB) and HLB affected Hamlin citrus wood (CT) were collected from the Citrus Research and Education Center of the University of Florida, Gainesville, FL. The
35 switchgrass (SW) and alfalfa (AF) were obtained from USDA ARS, Prosser, WA. The feedstock materials were ri nsed with DI water several times except citrus wood. Citrus wood were washed with soap and cleaned with a brush to remove pesticide deposits and nutrient because the citrus wood was obtained from an intensively managed grove. The feedstocks were oven dried at 60 o C for 48 h and then crushed to a grain size of less than 2 mm with a mechanical mill. About 15 20 grams of crushed dry samples were loaded into a quartz tubes and put inside a bench top furnace ( Barnstead 1500 M ). The whole pyrolysis process was op erated with constant N 2 gas flow inside the tubes at peak temperatures of 300, 450, or 600 o C. The resulting biochars were designated by their biomass type and peak temperature such as PB300, CT300, SW300, AF300, etc. The biochars were passed through a sie ve to isolate the 0.5 1 mm particle size intervals, and were then rinsed with tap water for one hour and DI water for 10 min and oven dried at 80 o C for 12 h. These were saved in sealed containers for subsequent analyses. Biochar C haracterization Total C, N and H content in the biochar samples were analyzed with a CHN Elemental analyzer (Carlo Erba NA 1500). The inorganic element contents (Ca, Mg, K, P, Al and Fe) of the biochar sampler were determined using the AOAC method by an inductively couple d plasma atomic emission spectrometry (ICP AES, Perkin Elmer Plasma 3200). Oxygen content was calculated as the weight difference between the raw biochar and sum of C, H, N and non volatile elements ( Peterson, Appell, et al., 2013 ) . Total surface area was measured with NOVA 1200 analyzer using Brunauer Emmett Teller (BET) method using N 2 sorption adsorption isotherms. The pH value of biochar samples was determined using published method ( Sun, Gao, et al., 2014 ) .
36 Briefly, 1 g of sample was added to 20 ml DI water. The suspension was shaken with a mechanic shaker at 40 rpm for 1 h, and equilibra ted for 5 min before measuring pH with a pH meter (Fisher Scientific Accumet basic AB15). Thermal stability analysis of biochar samples was done under an air atmosphere using a Mettler Toledo thermogravimetric analyzer (TGA) between 25 and 700 o C. Ash con tent was calculated as the weight left after thermal combustion at 700 o C relative to the starting biochar weight. As and Pb S orption For sorption experiment, solutions of As(V) (10 mg L 1 ) and Pb(II) (20 mg L 1 ) were made from the stock solutions (1000 mg L 1 ). About 0.05 g of each biochar was added to 20 ml solutions in 68 ml digestion vessels (Environmental Express) The vessels were placed onto a shaker and agitated at 40 rpm until sampling at room temperature (22 Â± 0.5 o C). After 24 h equilibrium (pr edetermined), the suspensions were membrane) and filtrate was analyzed for As and Pb using the ICP AES. Amount sorbed was calculated as the difference between initial and final sol ution metal concentrations. Each treatment was repeated three times and average values with error bars were reported. Statistical A nalysis Data for As(V) and Pb (II) sorption experiment were analyzed using a factorial analysis of variance (ANOVA; S ASÂ® 9.3; SAS Institute, Cary, NC) and DuncanÂ´s multiple range tests. TGA curves were made with Excel Â® 2010 and Sigmaplot Â® 12.0 software (Systat Software, Inc., San Jose, California, USA). Paterson correlation
37 analysis will be conducted to investigate the correlat ion between As and Pb sorption and elemental composition as well as surface area. Results and Discussion Elemental C omposition All the tested biochars were rich in carbon (C, 64% 86%), oxygen (O, 11% 32%), and hydrogen (H, 2% 5%) ( Table 2 1 ), which is normal for biochars prepared from raw woody and grass feedstocks ( Uchimiya, Wartelle, et al., 2011 ) . When the pyrolysis temperature increased, the C contents of the biochar also increased, but the O and H contents decreased. These trends are consistent with the findings of previous studies that pyrolysis can concentrate C of the biomass feed stocks ( Kloss, Zehetner, et al., 2012 , Sun, Gao, et al., 2014 ) . The ratios of H/C and O/C were also used as carbonization indicators. Decreased H/C and O/C ratios of all the biochars made at higher temperatures indicated they lost more water and O containing functional groups, and formed more aromatic and graphic str ucture to condense carbon ( Keiluweit, Nico, et al., 2010 , Uchimiya, Wartelle, et al., 2011 ) . Table 2 1 also shows that, under the same pyrolysis conditions, biochars derived from the wood feedstock had higher C contents than the grass biochars. This is probably because the woody biomass has higher cellulose and lignin content but lower hemicell ulose and ash content than herbaceous biomass. The nitrogen (N) contents of the biochars were between 0.3% and 3%, which are also within the commonly observed value of previous studies ( Sun, Gao, et al., 2014 ) . There was no obvious trend on the effect of pyrolysis temperature on N contents; however, the grass biochars contained more N than the wood biochars made at the same conditions. All of the tested biochars were low in non volatile elements, except the CT and AF series had slightly high Ca contents ( Table 2 1 ). Because high
38 temperature can accelerate biomass decomp osition, the higher temperature biochars contained more non volatile elements than the low temperature ones. Basic P roperties Biochar production rate (28% 50%) was dependent on both pyrolysis temperature and feedstock type ( Table 2 2 ), which agrees with the findings of other studies ( Keiluweit, Nico, et al., 2010 , Sun, Gao, et al., 2014 ) . All the biochars were alkaline ( Table 2 2 , pH 7.1 10), which is common for biochars prepared with dry pyrolysis technology ( Lehmann, Gaunt, et al., 2006 , Sun, Gao, et al., 2014 ) . In general, pH of the biochars increased with pyrolysis temperature ( Table 2 2 ). This is because high pyrolysis temperature can increase the amount of alkaline ca tions (e.g., Ca, Mg, K). Correlation analysis indicated that the pH values of the biochars were strongly correlated with the (Ca+Mg+K) contents ( Figure 2 1 ). The surface area of the biochars also strongly depended on both pyrolysis temperature and feedsto ck type ( Table 2 2 ). Wood biochars made at the two lower temperatures (i.e., 300 and 450 o C) and all the grass biochars had small surface area (0.1 15 m 2 g 1 ), but PB600 and CT600 had surface area of 209 and 183 m 2 g 1 , respectively. Ash contents of the te sted biochars increased with pyrolysis temperature with an exception for PB300 ( Table 2 2 ). The Ash contents of the grass biochars were higher than that of the corresponding wood biochars, which can be attributed to the fact that herbaceous biomass has hig her ash than the woody biomass. The PB and SW series had the lowest and the highest ash contents, respectively, compared to other biochars made at the same pyrolysis temperatures. All the biochars had good water holding capacity and maintained 72% 86% of saturation under free drainage conditions ( Table 2 3 ). Overall, the grass biochars
39 showed slightly better water holding ability than the wood biochars; and SW was the most capable to retain water and CT was the least capable. The water holding capacity of the biochars was further evaluated with TGA analyses of the water saturated biochars ( Figure 2 2 ), which showed the similar trends. TGA analysis was carried out to study thermal stability of biochars in air ( Figure 2 3 ). When heating temperature was imple mented between 25 and 100 o C, the mass of all the biochars only decreased slightly (due to water loss) and then became stable until at least 200 o C. No phase change was found on all the biochars at 100 200 o C, indicating decomposition did not occur at this temperature range. Therefore, 100 o C was used as the starting points for plotting the TGA curves in Figure 2 3 . As pyrolysis temperature increased, the biochars became more resistant to decompose or more stable, which agrees well with the results of previ ous studies ( Sun, Gao, et al., 2014 ) . As(V) and Pb (II) S orption Biochars were evaluated for their ability to sorb anionic (As(V)) and cationic (Pb(II)) heavy metals. Under the tested experimental conditions, all the biochars showed limited sorption ability (110 280 mg kg 1 ) to As(V) in aqueous solution ( Figure 2 4 ). Am ong all the biochars, the AF450 sorbed the greatest amount of As(V). For the rest of the biochars, pine derived biochars sorbed more As(V) than the others made under the same conditions. The sorption of Pb(II) on the biochars was higher than the As(V) ( Fig ure 2 4 ). However, pine derived biochars showed significantly lower Pb sorption than other biochars derived from other feedstocks. The SW450 sorbed about 950 mg/kg of Pb, which is much higher than all the other biochars. Statistical analyses indicated th at feedstock type had a significant effect on both Pb(II) and As(V) sorption (p<0.0001). Pyrolysis temperature only had statistically
40 significant effect on As(V) sorption (p=0.0036), but not on Pb(II) sorption (p=0.122). .05) showed that PB and AF sorbed As(V) to the greatest extent (without significant difference between these), followed by SW, and then CT ( Table 2 4 ). For Pb(II) sorption, no statistically significant difference was found among CT, AF, and SW samples, but PB samples were lower. All the biochars made at 300 and 450 o C showed no statistically significant difference in As sorption, but all of them showed statistically larger As sorption than the 600 o C biochars. Sorption of heavy metals onto biochars has bee n ascribed to several potential mechanisms, including electrostatic attraction, surface complexation with function groups, and ion exchange ( Ahmad, Rajapaksha, et al., 2014 , Mohana, Sarswata, et al., 2014 ) . To better understand interaction mechanisms controlling the sorption of As and Pb on the tested biochars, regression analyses were conducted to identify the key properties of the biochars affecting the sorption. The results showed that surface area had ver y weak correlation with the sorption of the two heavy metals under the tested experimental conditions, suggesting physisorption may not be the governing mechanism. Solution pH was strongly correlated with both Pb(II) and As(V) sorption 0.808 and 0.661, respectively ( Table 2 5 and Figure 2 5 and 2 6 ). This indicates that electrostatic interactions may control the sorption of Pb and As onto the biochars because solution pH strongly affects the sorbate species and surface charge of the sor bent. Strong statistic correlations were also identified for the alkaline cations, particularly (Ca+Mg+K), and the sorption of the two heavy metals ( Table 2 5 and Figure 2 5 and 2 6 ), further reflecting the importance of the electrostatic interaction to th e sorption of Pb and As on the biochars. While no correlations between the sorption
41 of As and other elemental compositions was identified, the sorption of Pb was correlated with the contents of P, N, and, C. This result is consistent with the findings in t he literature that sorption of Pb onto the biochar is controlled by multiple mechanisms, including precipitation (with phosphate ions) and surface complexation (with functional groups) ( Ahmad, Rajapaksha, et al., 2014 , Cao, Ma, et al., 2009 , Mohana, Sarswata, et al., 2014 ) . For example, strong correlati on (r=0.803) between Pb(II) sorption and the N content suggests N containing function groups such as amine may involve Pb sorption process ( Yantasee, Lin, et al., 2004 ) . Summary and C onclusion Twelve biochars, derived from four feedstocks at three pyrolysis temperatures, were tested and compared for their properties and sorption of two heavy metals (Pb and As). Among them, the pine biochar had the highest production rate, followed by two grass species, and then the citrus. Compared to low pyrolysis temperature, feedstock pyrolyzed at higher temperature showed lower production rate, higher C content, lower N, H and O content, generally higher non volatile elemental contents, higher ash content, higher surface area, and higher thermal stability. Feedstock type affected As(V) and Pb(II) sorption on the biochar significantly with p<0.001, while pyrolysis temperature had no statistically significant effect on the biochar sorption of Pb(II). Higher As(V) sorption was found on PB than on CT, and AF450 was the most efficient for As (V) sorption. The lowest Pb(II) sorption was observed for PB, which were several times lower than CT, SW and AF. Statistically analyses showed that electrostatic interaction played an important role in controlling the sorption of both Pb(II) and As(V) onto the biochar. Other mechanisms, such as precipitation and surface complexation, could also control the sorption of Pb(II) onto the biochars.
42 Table 2 1. Elemental composition of biochar made from loblolly pine ( LP ), Citrus (CT), Alfalfa (AL), and Switchgrass (SW) pyrolysized at 300, 450 and 600 o C under N 2 Biochar C N H O Al Ca Fe K P Mg O/C atomic ratio H/C atomic ratio %, mass PB300 69.24 0.34 4.95 25.17 0.028 0.14 0.019 0.02 0.03 0.078 0.27 0.86 PB450 80.18 0.31 3.31 15.76 0.036 0.20 0.022 0.02 0.04 0.12 0.15 0.50 PB600 85.68 0.33 2.13 11.40 0.041 0.19 0.025 0.05 0.04 0.12 0.10 0.30 CT300 64.48 1.53 4.73 27.07 0.012 1.51 0.015 0.34 0.10 0.21 0.31 0.88 CT450 67.63 1.23 3.61 25.02 0.011 1.71 0.017 0.41 0.11 0.25 0.28 0.64 CT600 78.28 1.28 2.08 14.90 0.013 2.28 0.019 0.66 0.15 0.35 0.14 0.32 AF300 64.72 3.10 5.26 24.24 0.023 1.19 0.039 0.51 0.15 0.56 0.28 0.98 AF450 69.66 2.42 3.01 22.13 0.024 1.33 0.03 0.70 0.17 0.53 0.24 0.52 AF600 73.25 2.22 1.91 19.43 0.025 1.53 0.038 0.76 0.19 0.65 0.20 0.31 SW300 59.32 2.34 4.64 31.68 0.039 0.43 0.066 0.84 0.21 0.44 0.40 0.94 SW450 64.02 2.23 2.87 28.27 0.039 0.56 0.072 1.05 0.30 0.59 0.33 0.54 SW600 68.15 1.90 2.21 24.99 0.042 0.52 0.067 1.21 0.32 0.61 0.28 0.39
43 Table 2 2 . Basic physical and chemical properties of biochars Biochar pH Ash Production (%) BET Surface area BJH pore vol (des. leg) (%) rate (%) (m 2 g 1 ) (cc g 1 ) PB300 7.10 5.55 49.8 0.2 0.002 PB450 7.61 3.65 31.3 0.1 0 PB600 7.05 4.01 28.3 209 0.003 CT300 7.76 9.00 41.6 0.8 0.005 CT450 10.00 10.31 29.7 2.8 0.009 CT600 9.48 13.57 27.6 182 0.013 AF300 8.38 10.51 42.1 0.6 0.006 AF450 9.17 11.59 32. 4 0.7 0.005 AF600 10.35 12.52 28. 2 0.2 0.006 SW300 8.21 18.23 43.7 1.2 0.01 SW450 9.74 19.17 31.3 10 0.02 SW600 9.84 21.25 30.1 15 0.024 Table 2 3 . Bulk density and water holding ability of the biochars Biochar Bulk Density (g cm 3 ) Water retained (%) PB300 0.19 73.24 PB450 0.19 79.22 PB600 0.21 77.87 CT300 0.21 74.59 CT450 0.26 75.34 CT600 0.23 72.38 AL300 0.20 77.59 AL450 0.15 80.57 AL600 0.16 75.54 SW300 0.13 86.05 SW450 0.12 85.52 SW600 0.12 83.33
44 Table 2 4 . biochar As(V) (mg kg 1 ) Pb(II) (mg kg 1 ) Feedstock Pine 201.89 A* 2466.4 B Alfalfa 191.72 A 7486.2 A Switchgr 132.64 B 7781.9 A citrus 118.36 C 7337 .0 A Temperature 300 o C 165.41 A 5976.5 A 450 o C 187.78 A 6187.8 A 600 o C 130.28 B 6639.4 A Table 2 5 . Pearson correlation coefficients (r) and significance of linear relationship between As(V) and Pb(II) sorption and physiochemical properties of biochars Factors Pb(II) As(V) r Significance level r Significance Surface a rea 0.321 NS 0.218 NS pH 0.808 ** 0.661 ** Ca 0.613 * 0.615 * Mg 0.801 ** 0.522 NS K 0.826 *** 0.653 * Ca+Mg+K 0.931 *** 0.790 *** P 0.790 ** 0.601 NS Al 0.318 NS 0.292 NS O 0.500 NS 0.157 NS N 0.803 ** 0.484 NS C 0.609 * 0.233 NS
45 Figure 2 1 . Linear regression analysis between pH value of biochar and contents of cationic elements (Ca+Mg+K) in biochar
46 Figure 2 2. Thermogravimetric analysis (TGA) between 25 and 200 o C for pinewood biochar (A), citrus wood biochar (B), alfalfa biochar (C), and switchgrass biochar (D) pyrolyzed at 300, 450 and 600 o C.
47 Figure 2 3 . Thermogravimetric curves of biochars derived from pinewood (A), citrus wood (B), alfalfa (C), and switchgrass (D).
48 Figure 2 4 . Sorption of As(V) (A) and Pb(II) (B) onto the biochars. Error bars represent standard deviations.
49 Figure 2 5 . Linear regression analysis between Pb(II) sorbed and elemental content and pH of biochar
50 Figure 2 6 . Linear regression analysis between As(V) sorbed and elemental content and pH of biochar
51 CHAPTER 3 REMOVAL OF ARSENIC BY MAGNETIC BIOCHAR PREPARED FROM PINE WOOD AND NATURAL HEMATITE 1 Introduction Arsenic (As) is a carcinogenic trace metal that is toxic to human and animals. It occurs naturally in soils and can be mobilized by weathering reactions and biological activity and may lead to con tamination of surface or groundwater aquifers ( Mohan and Pittman Jr, 2007 ) . In addition, anthropogenic As contamination by discharge and disposal of As containing compounds may lead to even greater As concentrations ( Mandal and Suzuki, 2002 ) . Because of the severe toxicity of arsenic to humans, a very strict drinking water allowance limit of 10 Âµg L 1 was prescribed by the US Environmental Protection Agency. Various removal techniques such as precipitation, adsorption, membrane separation, ion exchange and permeable reactive b arriers, have been developed to treat As contaminated water and soils ( Tuutijarvi, Lu, et al., 2009 ) . Among them, adsorption is the one of the most commonly used method for the removal of As from aqueous solutions ( Mohan and Pittman Jr, 2007 ) . Many low cost sorbents such as carbonace ous materials, clay minerals and metal oxyhydroxides, have been successfully applied for As sorption ( Goldberg, 2002 , Zhou, Gao, et al., 2014 ) . Several studies have demonstrated that As tends to accumulate on the surface of metal oxyhydroxides and clay minerals containing Fe, Mn, Al, Cu, and Co ( Ferguson and Gavis, 1972 , Mandal and Suzuki, 2002 ) . Strong sorption of As from aqueous solution by 1 Reprinted with permission from Wang, S., B. Gao, A.R. Zimmerman, Y. Li, L. Ma, W.G. Harris, et al. 2015. Removal of arsenic by magnetic biochar prepared from pinewood and natural hematite. Bioresource Technology 175: 391 395. doi:http://dx.doi.org/10.1016/j.biortech.2014.10.104.
52 Fe, Al, and Mn oxides and naturally occurring clay minerals, such as kaolinite, m ontmorillonite, and illite, has been reported frequently in the literature ( Gimenez, Martinez, et al., 2007 , Goldberg, 2002 ) . Hematite is one of t he most abundant natural iron oxide mineral and shows good As sorption ability ( Gimenez, Martinez, et al., 2007 ) . The sorption behaviors and mechanisms of As on hematite have been investigated for various solution chemistry conditions ( Goldberg and Johnston, 2001 ) . Using X ray scattering analysis, Catalano et al. ( 2008 ) found that the sorption of As on hematite was mainly controlled by electrostatic interactions to form inner and outer sphere surface complexes. Surface complex models have also successfully simulated experimental sorption data ( Goldberg, 2002 ) . Several studies have synthesized iron oxides to mimic the natural hematite and found that the synthesized hematite showed better As sorption ability because of higher purity and larger surface a rea ( Auffan, Rose, et al., 2008 ) . In addition, ability to As in aqueous solution ( Ramirez Muniz, Jia, et al., 2012 ) . Biochar is a pyrogenic carbon material produce d by combustion of biomass under oxygen limited conditions. Because of its unique properties such as high surface area and cation exchange capacity, biochar can be used for such applications as soil improvement, fertility enhancement and carbon sequestrati on ( Mohana, Sarswata, et al., 2014 , Zimmerman, Gao, et al., 2011 ) . It also has shown great potential to remove heavy metals from aqueous solution and to reduce their mobility and bioavailability in soils ( Ahmad, Rajapaksha, et al., 2014 , Zhou, Gao, et al., 2013 ) . Because the surfaces of most of the biochars are predominantly net negatively charged ( Mukherjee,
53 Zimmerman, et al., 2011 , Yao, Gao, et al., 2012 ) ; however, their sorption of aqueous As, which is in anionic forms of either ars enate (As(V)) or arsenite (As(III)), is relatively low ( Beesley and Marmiroli, 2011 ) . Several methods have thus been developed to modify biochar to enhance its sorption of As. In particular, biochar modified with colloidal and nano sized oxyhydroxides showed strong ability to remove As from aqueous solution ( Chen, Chen, et al., 2011 , Zhang and Gao, 2013 ) . F or example, iron oxide biochar composites were prepared by either pyrolyzing iron chloride (FeCl 3 ) modified biomass or precipitating Fe 3+ /Fe 2+ on biochar surfaces greatly enhanced the As sorption ability of the biochars ( Chen, Chen, et al., 2011 , Zhang, Gao, et al., 2013 ) . However, the methods used to create these biochar nanocomposites are relatively complex and costly. Thus, additional investigations thus are needed to develop simple and cost effective methods to modify biochars with iron oxide particles, particularly with natural iron oxide minerals. The objective of this work was to develop and evaluate a new method to prepare iron oxide biochar composites from biomass and natural hematite. Hematite treated pine wood was used as the feedstock to produce the biochar through pyrolysis. Physicochemical p roperties of the resulting biochar were measured in laboratory and sorption ability of the biochar to As was assessed through batch sorption experiments. Materials and Methods Reagents All chemicals used in this work were analytical grade and were dissolv ed in heptahydrate (Na 2 HAsO 4 Â·7H 2
54 mineral was obtained from Wards Natural Science Establishment, Inc. (Minnesota, USA ) and X ray diffraction (XRD) analysis was conducted to confirm its purity. The mineral was crushed with hammer and ground with a pestle and mortar and sieved to particle sizes of between 38 and 75Âµm for use in subsequent experiments. Commercial loblolly p ine ( Pinus taeda ) wood was oven dried overnight at 80 o C and then chopped with a mechanic mill. The crushed feedstock was then sieved to between 0.425 and 1 mm. Biochar P reparation The method of preparing hematite modified biochar was similar to that used by Yao et al. ( 2014 ) . A hematite suspension was prepared by mixing 2 g of crushed mineral particles in 40 ml of DI water. The suspension was s tirred and sonicated for 30 min with an ultrasonicator (3510R DTH, Bransonic Ultrasonics Corporation) to form a stable suspension. About 10 g of the feedstock was then well mixed with the suspension for 2 h and was then oven dried at 80 o C. The hematite tr eated biomass was then pyrolyzed in a tube furnace (MTI, Richmond, CA) under N 2 at a peak as control and went through the same pyrolysis process. The biochars produced wer e sieved to obtain particles between 75 Âµm and 0.425 mm and rinsed with DI water several times. Biochar was then oven dried overnight at 80 o C and saved in a sealed container for later use. The biochars with and without hematite modification were denoted a s HPB and PB, respectively. C haracterization Total carbon (C), nitrogen (N), and hydrogen (H) content in the biochar samples were analyzed with a CHN Elemental analyzer (Carlo Erba NA 1500). The inorganic
55 element content of the biochar samples were determ ined using the AOAC method and analyzed by inductively coupled plasma atomic emission spectrometry (ICP AES, Perkin Elmer Plasma 3200). Oxygen was determined as the weight difference between the raw dried biochar and sum of C, H, N, and other non volatile elements. Total surface area was measured using N 2 sorption on a NOVA 1200 analyzer and calculated using Brunauer Emmett Teller (BET) method. Scanning electron microscope (SEM) images were obtained with a JEOL JSM 6400 Scanning Microscope. Energy dispersiv e X ray spectroscopy (EDS, Oxford Instruments Link ISIS) was coupled with SEM to examine surface elemental composition, surface elemental distribution maps. Surface elemental composition was also analyzed by X ray photoelectron spectroscopy (XPS) with a P HI 5100 series ESCA spectrometer (Perkin Elmer). An Al X ray source was used with a 93.90 eV passing energy between 0 and 1400 eV binding energy. High energy resolution scans of Fe2p and C1s peaks were obtained with pass energy of 23.50 eV in the 705 734 e V and 281 293 eV binding energy ranges, respectively. Surface crystallinity was analyzed to identify Fe bearing minerals using an X ray diffractometer (XRD) (Philips Electronic Instruments) equipped with a stepping motor, a graphite crystal monochromator, Thermal stability of PB and HPB was examined using thermogravimetric analysis Temperature was increased by 10 o C per minute between 25 and 700 o C under air atmosphere.
56 Adsorpti on Kinetics and I sotherm Sorption kinetics of As at a constant concentration (20 mg L 1 ) onto biochar were examined using the method of Zhang and Gao ( 2013 ) . Briefly, about 0.05 g of biochar (2.5 g L 1 ) was added to 20 ml As solutions (pH~7) in 68 ml digestion vessels (En vironmental Express) at room temperature (22 Â± 0.5 o C).The vessels with sorption mixtures were placed on a shaker and agitated at 50 rpm until sampling. At each sampling time (0.5 1, 2, 4, 8, 12, 24 and 48 h), the suspensions were immediately membrane). Concentrations of As in the filtrates were determined with an inductively coupled plasma atomic emission spectrometry (ICP AES, Perkin Elmer Plasma 3200) and the residues were examin ed spectroscopically as described above. Sorbed As(V) was calculated as the difference in As concentration between initial and final solution. Kinetics data were fitte d using various kinetic models . Adsorption isotherms were constructed using 20 ml solutions As(V) concentrations ranging 1 50 mg L 1 with biochar (2.5 g L 1 ). The suspension was agitated on a shaker for 24 h, a period of time determine to be sufficient to have reached apparent sorption equilibrium, and then treated as described abov e. Isotherm data were simulated with various isotherm models . Sorption M odels Kinetics data were fitted by pseudo first order, pseudo second order and the Elovich models. Governing eq uations of the models can be written as follows ( Yao, Gao, et al., 2011 ) : (3 1 )
57 ( 3 2) (3 3 ) where equations 3 1, 3 2 and 3 3 represent first order, second order and Elovich model respectively. q t and q e are the amount of phosphate adsorbed at time t and at equilibrium, respectively (mg kg 1 ), and k 1 and k 2 are the first order and second order apparent adsorption rate constants (h 1 ), respectively. Also, is the initial adsorption rate (mg kg 1 ) and is the desorption constant (kg mg 1 ). The first order and second order models describe the kinetics of the solid solution system based on mononuclear and binuclear adsorption, respectively, with respect to the sorbent capacity ( Gerente, Lee, et al., 2007 ) , while the Elovich model is an empirical equation considering the contribution of desorption . Isotherm data were simulated with Langmuir and Freundlich isotherm models, and the governing equations can be written as follows ( Yao, Gao, et al., 2011 ) : ( 3 4) ( 3 5) where equations 3 4 and 3 5 represent Langmuir and Freundlich model respectively. K and K f represents the Langmuir bonding term related to interaction energies (L mg 1 ) and the Freundlich affinity coefficient (mg (1 n) L n kg 1 ) , respectively, Q denotes the Langmuir maximum capacity (mg kg 1 ), C e is the equilibrium solution concentration (mg L 1 ) of the sorbate, and n is the Freundlich linearity constant. The Langmuir model assumes monolayer adsorption onto a homogeneous surface with no
58 interactions between the adsorbed molecules. The Freundlich model, however, is an empirical equation, often used to describe chemisorptions onto heterogeneous surface Statistical A nalysis TGA curves were made with Excel Â® 2010 and Sigmaplot Â® 12.0 s oftware (Systat Software, Inc., San Jose, California, USA). Sorption kinetics and isotherms analys is was conducted with SigmaPlot Â® 12.0 using user defined function to edit the simulation equations. Results and Discussion B iochars P roperty Elemental analysis showed that the Fe content of the HPB was 118 times greater than that of the PB ( Table 3 1 ), likely due to the addition of Fe bearing oxides in natural hematite minerals. This is also supported by the higher ash content of HPB after combustion at 700 o C under air ( Table 3 1 ). The amounts of volatile elements such as C, H, and N, and non volatile elements such as Ca, K, and P in the HPB was lower, possibly due to the dilution effects of the hematite addition. Because hematite also has Al and Mg, the modification also slightly increased their contents in the HPB ( Table 3 1 ). The BET N 2 surface area decreased only slightly after hematite addition suggesting that 1) hematite had a specific surface area similar to that of the pristine biochar, and 2) the added hematite did not bloc k pore openings of the biochar. TGA analysis showed the HPB to be slightly more thermally stable than the pristine biochar, both beginning to decompose at a higher temperature and losing less weight overall (Figure 3 1). This is just as expected given that hematite is far less volatile than biochar.
59 The SEM/EDS analyses of the HPB suggest that pyrolysis soldered and stabilized the hematite particles on biochar surfaces (Figure 3 2 ). SEM image (Figure 3 2 A and 3 2 E) clearly show m any small aggregates/particles stabilized on carbon surface. The EDS mapping carbon (Figure 3 2 B and 3 2 F), oxygen (Figure 3 2 C and 3 2 G), and iron (Figure 3 2 D and 3 2 H) were used to identify the surface elemental compositions, where the dark particles in the carbon map (absence of C) matched well with the light areas on iron and oxygen (presence) maps. XPS analysis of the HPB showed that the binding energies of Fe2p1/2 and Fe2p3/2 were 710.85 and 724.48 eV, respectively (Figure 3 3 ). These values are close to reported values for Fe 3+ ( Zhang, Gao, et al., 2013 , Zhong, Hu, et al., 2006 ) , indicating that the iron particle s on the biochar surface are Fe 2 O 3 . Although the original hematite is weakly magnetic, the HPB demonstrated strong magnetic property and could be easily attracted by a permanent magnet ( Figure 3 4 ) . This result indicated that the pyrolysis process not only produced the hematite biochar magnetic property of the hematite. Previous studies have shown that thermal treatment can enhance the magnetic properties of hematite ( Mansilla, Zysler, et al., 2002 , Zhang, Paterson, et a l., 2012 ) . Natural hematite is mainly Fe 2 O 3 with rhombohedral st ructure, which has weak magnetic property, whereas thermally Fe 2 O 3 ), cubic structures with strong magnetic properties ( Zhang, Paterson, et al., 2012 ) . Comparisons of the XRD patterns of the orig inal hematite with that of the HPB sample suggests that the pyrolysis process did alter the crystal structure of the hematite of the HPB (Figure 3 5 ). The d spacing of the original hematite used in this work matched well with the values
60 from published data base for natural hematite minerals ( Schimanke and Martin, 2000 ) . The diffraction pattern of the HPB (Figure 3 5 B), however, was different from the Fe 2 O 3 ) (Figure 3 5 A). The Bragg peaks of XRD pattern of the HPB were at 30.2, 35.5, 43.2, 57.3, and 62.9 o , which can be assigned to Fe 2 O 3 ( Schimanke and Martin, 2000 , Zhang, Gao, et al., 2013 ) . Adsorption K inetics The sorption of As (V) to both the PB and the HPB samples showed two phases: a rapid initial sorption during the first two hours followed by a much slower phase till reaching sorption equilibria after roughly 20 hr (Figure 3 6 A). The first phase could be ascribed to rapid occupa tion of easily accessible external surface sorption sites such as outer sphere complexation ( Essington, 2004 ) . The slow phase could be related to the formation of inner layer complexes ( Essington, 2004 ) . Alternatively, the slower phase may be related to kinetic inhibition of As movement through narrow pore channels. In comparison with the PB, the HPB showed faster and higher sorption of As, suggesting Fe 2 O 3 particles may serve as adsorption sites for As in aqueous solution. Several sorption kinetic models were applied to simulate the experimental data ( Table 3 2 ). Among all the tested models, Elovich model, an empirical model, described the As sorption kinetics the best (Table 3 2). This suggests that the sorption of As (V) onto the biochar samples may be controlled by multiple interaction mechanisms or processes ( Sparks, 1999 ) . Sorption I sotherm The (Figure 3B), which is commonly observed for As sorption onto many carbonaceous sorbents ( Chang, Lin, et al., 2010 , Zhang, Gao, et al., 2013 ) . Sorption of As onto both
61 biochars increased with aqueous As concentration but seem to plateau at concentrations above 30 40 mg L 1 . Sorption of As onto the hematite modified biochar was rou ghly double that of the pristine biochar at all concentrations, This further suggesting that the iron oxide particles serve as adsorption sites with higher capacity for As in aqueous solution than the unmodified biochar. Both Langmuir and Freundlich model were applied to simulate the sorption isotherms ( Table 3 2 ). Although the two models both describe the experimental data well, the simulations of the Langmuir model fit the isotherms better (Table 3 2). Although Langmuir model is the most common model use d to describe monolayer adsorption processes, it has also be used to describe the sorption of chemical compounds on composited materials ( Yao, Gao, et al., 2011 , Zhang, Gao, et al., 2013 ) including As (V) onto iron oxide nanoparticles or onto iron oxide biochar nanocomposites ( Chowdhury, Yanful, et al., 2011 , Zhang, Gao, et al., 2013 ) . Values of the Langmuir maximum sorption capacity of As for PB and HPB were 265 and 429 mg kg 1 , respectively, also indicating that the hematite modification roughly doubled the As sorption ability of the biochar. The Langmuir maximum sorption capacity of As(V) of HPB is lower than that of an iron oxide biochar nanocomposites developed by Zhang et al ( 2013 ) , but it is similar to that of sorption capacity of magnetic biochars derived from chemical treated agricultural residuals ( Chen, Chen, et al., 2011 ) . Sorption M echanisms Because both PB and HPB removed As from aqueous solution, the sorption of As(V) onto HPB cou ld be controlled by multiple processes associated with both carbonaceous and iron oxide surfaces. Previous studies have suggested that biochar can encourage the formation of precipitates which can then remove anions and cations,
62 such as lead and phosphate, from aqueous solutions ( Inyang, Gao, et al., 2012 , 2011 , Yao, Gao, et al., 2013 ) . The XRD analysis of the post sorption biochar samples did not show any new peaks ( Figure 3 5 C), suggesting that no new minerals formed and the precipitation mechanism may not be important for As removal by HBP. The sorption of As onto a solid surface is most likely controlled by two factors, namely speciation of the As and the charge of the sorbent surface ( Tuutijarvi, Lu, et al., 2009 ) . Under the tested experimental conditions (pH~7), As(V) mainly exists as HAsO 4 2 . Some of the functional groups of the biochar are protonated and thus positively charged as the point of zero charg e for biochar pyrolyzed at 600 o C is usually above 7 ( Mukherjee, Zimmerman, et al., 2011 ) . As a result, the HAsO 4 2 may interact with the positively charged functional groups on both PB and HP B surfaces through electrostatic attractions. In addition, iron oxide also has a pH dependent charge derived from protonation and deprotonation of surface OH groups . The pH zpc of the iron oxide particles on the HPB surfaces is around 7.5 ( Chowdhury, Ya nful, et al., 2011 , Tuutijarvi, Lu, et al., 2009 ) , therefore, they are also predominantly positively charged under the tested experimental conditions. As a result, both the iron particles and surface functional groups of the HPB can serve as the sorption site for As in aqueous solut ion, which Summary and Conclusion A magnetic biochar was synthesized by direct pyrolysis of hematite treated biomass. The thermal treatment not only produced a biochar based composite Fe 2 O 3 particles with strong magnetic properties. Fe 2 O 3 particles can serve as sorption sites for As(V) in aqueous solutions, and thus greatly improved the As(V) removal ability of the biochar. Because both biomass
63 and hematite mineral are low cost natural materials that are abundant and inexpensive, the magnetic biochar developed in this work can be used as an alternative remediation agent in many environmental applications to mitigate the risks of As contamination.
64 Table 3 1. Eleme ntal composition and BET surface area of loblolly wood biochar (P B) and Hematite modified biochar (H PB) Biochar C N H O Fe Al Ca Mg K P BET surface area m 2 g 1 %, mass based PB 85.7 0.3 2.1 11.4 0.02 0.04 0.19 0.12 0.05 0.04 209.6 HPB 51.7 0.2 1.4 43.1 2.95 0.24 0.10 0.14 0.04 0.03 193.1 Table 3 2. Kinetics and isotherm models and best fit parameters of As(V) sorption onto pine wood biochar (PB) and hematite modified biochar (HPB) Model/Equations Biochar Parameter1 Parameter2 R 2 Kinetics First order PB q e =128.5 mg kg 1 k 1 =1.66 h 1 0.814 HPB q e =231.5 mg kg 1 k 1 =0.828 h 1 0.829 Second order PB q e =138.6 mg kg 1 k 2 =0.0162 kg mg 1 h 1 0.900 HPB q e =248.4 mg kg 1 k 2 =0.0051 kg mg 1 h 1 0.911 Elovich PB 1 h 1 1 0.961 HPB 1 h 1 1 0.960 Isotherm Langmuir PB S max =265.2 mg kg 1 K=0.0724 L mg 1 0.951 HPB S max =428.7 mg kg 1 K=0.159 L mg 1 0.968 Freundlich PB n=0.448 K f =37.55 mg (1 n) L n kg 1 0.870 HPB n=0.305 K f =121.8 mg (1 n) L n kg 1 0.893
65 Figure 3 1 . Thermogravimetric analysis (TGA) of pine biochar (PB) and hematite modified biochar (HPB)
66 1000x 10,000x Figure 3 2 . SEM EDS elemental mapping analysis for HPB (1000x & 10000x), including surface topography (A), compositional map for (B) C, (C) O and (D) Fe. The red rectangles highlight the zoom in location that shows an iron oxide particle on carbon surface with surface topography (E), compositional map for (F) C, (G) O and (H) Fe.
67 Figure 3 3. XPS diffraction patterns of Fe2p1/2 and Fe2p3/2 spectra of HPB. Figure 3 4 . Magnetic properties of hematite mineral (A) an d hematite modified biochar (HPB) (B)
68 Figure 3 5 . XRD diffraction patterns of (A) hematite particles (B) HPB (C) post sorption HPB
69 Figure 3 6 . Kinetics (A) and isotherm (B) data and fitted models for As(V) sorption by loblolly pine wood biochar (PB) and hematite modified biochar (HPB). Error bars represent standard deviation.
70 CHAPTER 4 MANGANESE OXIDE MODIFIED BIOCHARS: PREPARATION, CHARACTERIZATION, AND SORPTION OF LEAD AND ARSENATE Introduction Arsenic (As) is a carcinogenic metalloid and long term exposure may cause acute poisoning (abdominal pain and death) and chronic disease (kidney, lung, bladder cancer) ( Mohan and Pittman Jr, 2007 ) . As contamination may be naturally occurring but more often it is from anthropogenic sources including mine wastes, fossil fuel combustion, paints , semi conductors , alloys , wood preservatives , pesticides and fertilizers ( Goldberg, 2002 ) . Lead (Pb) is a toxic metal that can be inhaled or ingested, damaging nerve systems and kidneys ( Cheng and Hu, 2010 , Needleman, 2004 ) . Pb pollution may result from natural lead containing o re or anthropogenic sources such as waste incineration, coal burning, and leaded gasoline ( Cheng and Hu, 2010 ) . Reclamation of heavy metals via adsorption is a convenient and prevailing approach to remove As and Pb from wastewater ( Zhang, Niu, et al., 2010 ) . Recently, res earch has been directed at producing and optimizing adsorbent materials that have characteristics similar to activated carbon (C) materials but are more cost efficient and environmentally friendly. For example, biochar is pyrogenic C produced by thermal c onversion of lignocellulose biomass under oxygen free or limited conditions ( Lehmann, Gaunt, et al., 2006 ) . Biochar has been recognized as a candidate sorbent for As ( Mohan, Pittman, et al., 2007 ) and Pb ( Mohan, Pittman, et al., 2007 , Uchimiya and Bannon, 2013 , Uchimiya, Chang, et al., 2011 ) . However, the As and Pb sorption capacities of most of the pristine biochars are lower than some commercial adsorbents. Thus, the goal of this research was to enhance sorption capacity of biochar by producing biochar based composites.
71 Manganese (Mn) oxide minerals are commonly found in soil, especially those that have undergone cycles of wetting and drying ( Essington, 2004 , Post, 1999 ) . Of more than thirty Mn oxides found in terrestrial environments, birnessite is one of the most abundant ( O'Reilly and Hochella Jr, 2003 , Post, 1999 ) . It has been noted for their high adsorption capacity of b oth a rsenite (As(III)) and arsenate (As(V)) ( Lenoble, Laclautre, et al., 2004 , Manning, Fendorf , et al., 2002 ) . In addition, birnessite also has high oxidization potential and can convert As (III) to less toxic As (V) ( Lafferty, Ginder Vogel, et al., 2010 , Post, 1999 ) . The sorption of As(V) by MnO 2 is pH dependent, but there is only a gradual decrease of sorption between pH 2 10 ( Zhang and Sun, 2013 ) . This is because surface complexation interactions between A s(V) and Mn oxides may occur at appropriate ambient pH and the oxygen moiety of As(V) displaces a surface hydroxyl group on Mn oxide to generate an inner sphere complex ( Manning, Fendorf, et al., 2002 , Zhang and Sun, 2013 ) . Mn oxides also have a high potential for Pb(II) sorption ( Matocha, Elzinga, et al., 2001 , McKenzie, 1980 , O'Reilly and Hochella Jr, 2003 ) . Lead sorption was observed to be associated with internal reactive sites. For example, Pb(II) may enter the space ers and tunnel into the internal cryptomelane structure ( Lee, Kim, et al., 2013 , O'Reilly and Hochella Jr, 2003 ) . The specific objectives of this work were to: (1) prepare and characterize two Mn oxide biochar composites, (2) compare the As(V) and Pb(II) sorption capacities of the composit es with that of pristine biochar, and (3) explore the possible mechanisms involved in the As(V) and Pb(II) sorption.
72 Materials and Methods Reagents Sodium arsenate dibasic heptahydrate (Na 2 HAsO 4 Â·7H 2 O), l ead nitrate (Pb (NO 3 ) 2 ), potassium permanganate (KMnO 4 ), manganese chloride tetrahydrate (MnCl 2 Â·4H 2 O) and concentrated hydrochloric acid (HCl) of analytical grade were (Nanopure water, Barnstead). Sorbent P reparation Mn oxide modified pine biochar (MPB) Loblolly pine ( Pinus taeda ) wood was oven dried overnight at 80 o C, then crushed and sieved to obtain the 0.425 1 mm size fraction. After sieving, 14 g of the feedstock was immersed in 100 mL of 5 g L 1 MnCl 2 Â·4H 2 O solutio n for 2 h. The mixture was oven dried at 80 o C overnight, and then pyrolyzed in a tube furnace (MTI, Richmond, CA) under flowing N 2 by ramping the temperature at a rate of 10 o C per min pine biochar (PB), used as the control treatment, was made similarly but without pre soaking the feedstock in Mn solution. Birnessite modified pine biochar (BPB) Birnessite was synthesized using a modified KMnO 4 precipitation method as described by McKenzi e ( McKenzie, 1980 , O'Reilly and Hochella Jr, 2003 ) . Briefly, 3.15 g KMnO 4 was dissolved in 50 mL DI water, and a 5 g of PB was added to the solution and agitated for 2 h with a mag netic stirrer. The suspension was then boiled for 20 min, followed by drop wise addition of 3.3 mL concentrated HCl. The reaction was continued for additional 10 min under vigorous stirring. The solution was then allowed to cool to
73 room temperature (22 Â± 0 .5 o birnessite composite was rinsed thoroughly with DI water, oven dried overnight at 80 o C and stored in a closed jar until analysis. Sorbent C haracterization Total carbon (C), nitrogen (N), and hydrogen (H) content in the biochar samples was analyzed with a CHN elemental analyzer (Carlo Erba NA 1500). The inorganic element composition of biochar samples was determined according to the AOAC method (AOAC, 1990) and ana lyzed by inductively coupled plasma atomic emission spectrometry (ICP AES, Perkin Elmer Plasma 3200). Oxygen was determined as the weight difference between the raw dried biochar and sum of C, H, N, and other non volatile elements. Total surface area was m easured on a NOVA 1200 analyzer using N 2 sorption Brunauer Emmett Teller (BET) method ( Hammes, Smernik, et al., 2008 , Peterson, Appell, et al., 2013 ) . Scanning electron microscope (SEM) images were obtained on a JEOL JSM 6400 Scanning Microscope. Energy dispersive X ray fluorescence spectroscopy (EDS, Oxford Instrumen ts Link ISIS) was coupled with SEM in order to map surface elemental distributions and associations in relation to particles embedded in the biochar matrix ( Newbury and Ritchie, 2013 ) . Surface elemental composition was also analyzed by X ray photoelectron spectroscopy (XPS) with a PHI 5100 series ESCA spectrometer (Perkin Elmer). An X ray diffractometer (XRD) (Ultima IV X Ray Diffractometer, Rigaku Corporation, Japan) equipped with a stepping motor and radiation source was used, and scan s were made between 0 and 80 o Thermogravimetric analyses (TGA) of original and modified biochar samples were done
74 using a Mettler Toledo TGA/DSC1 analyzer. The temperature was increased by 10 o C per minute between 25 and 700 o C under air atmosphere. Adsorption Kinetics And Isotherm Investigation of As(V) (10 mg L 1 ) and Pb(II) (50 mg L 1 ) sorption kinetics by biochar followed the methods of Zhang et al ( Zhang, Gao, et al., 2013 ) . Briefly, about 0.05 g of biochar was added to 20 mL of the sorbate solution in 68 mL digestion vessels (Environmental Express) at room temperature (22 Â± 0.5 o C). Thus, sorbent concentrations were about 2.5 g L 1 for all tr eatments. The vessels were placed onto a rotary shaker, and shaken at 40 rpm until sampling. At each sampling time (0, 0.5, 1, 2, nylon membrane). As(V) and Pb(II) were measured in the filtrate and sorption was calculated as the difference in initial and final solution concentration of the sorbate. Adsorption isotherms were determined for these metals using the same procedure as de scribed above but using a range of As(V) (20 ml, 1 20 mg L 1 ) and Pb(II) (20 ml, 1 300 mg L 1 ) sorbate solution concentrations and 24 h contact period. The kinetic and isotherm experiments were run in triplicate. The pH for As(V) and Pb(II) were adjusted t o around 7 and 5.5, respectively. Statistical Analyses Modeling of sorption kinetics and isotherm data was conducted with SigmaPlot 12.0 using user defined functions to minimize the residuals between model calculated and measured values.
75 Results and Disc ussion Sorbent Properties Compared to pristine biochar, bulk carbon content of MPB and BPB was 6.7 and 24.1% less, respectively (Table 4 1). Similar trends were also found for Mg, Al, N, and H contents because of the dilution effect. However, the concentration of K in BPB increased several times over the pristine biochar as K was introduced in the synthesis process. Chemical modifications a lso increased Mn concentrations in MPB and BPB by about 182 and 354 times, respectively, compared to PB (Table 4 1). These results were confirmed with XPS analysis, which showed that Mn content was 3.7 % (atomic) in MPB (Figure 4 1 a ) compared to 9.4 % (ato mic) in BPB (Figure 4 1b ). The ash content of PB, MPB and BPB were 4.02, 14.0 and 33.4%, respectively (Figure 4 2 ). The higher ash contents of MPB and BPB likely resulted from the addition of Mn oxides. Conversion of PB to MPB resulted in more than a doub ling of BET surface area and an increase in pore volume of more than seven times (Table 4 1). This is likely due to the formation of new Mn bearing minerals. However, while the pore volume of BPB was 22 times that of PB, the BET surface area decreased by t wo thirds, suggesting that the birnessite modification may change the pore size distribution of the biochar. The MPB showed ex cellent crystallinity (Figure 4 3 a ). Its XRD p eaks at d=2.562, 2.220, 1.570, 1.339, and 1.281 Ã… are corresponding to manganosite ( MnO) ( Chen, Xing, et al., 2009 ) . The presence of other subdominant XRD peaks may indicate formation of additional phases on MPB. The peaks at d spacing of 4.935 Ã… may represent formation of other Mn containing compounds based on referenced peak values. Direct precipitation with KMnO 4 an d HCl on PB (i.e., BPB) showed relatively poorly crystallinity
76 (Figure 4 4 a ) with an XRD pattern matching birenessite ( O'Reilly and Hochella Jr, 2003 ) . The peak at around d=7.303 Ã… (001) , 2.450 Ã… (100) and 1.420 Ã… (110) are MnO 2 ) ( McKenzie, 1980 ) . TGA analysis indicates thermal stability of MPB is lower than BPB (Figure 4 2 ). TGA curve s include a relative stable phase (first phase) and a rapid mass loss phase (second phase) after combustion in the air. DTG curves indicate inflammability of biochars (data not shown). The main peak in response to the turning point for rapid weight loss, appeared at 380 and 335 o C for PB and MPB respectively. This trend is contrary to the magnesium ( Yao, Gao, et al., 2013 ) and hematite ( Wang, Gao, et al., 2014 ) modified biochars, where modified biochars have higher thermal stability than pristine biochar. This may be related to the transition of Mn oxides during heating process in air atmosphere. Thermal transformation and TG characteristics of different Mn oxides under inert and air atmosphere had been studied ( Gonzalez, Gutierrez, et al., 1996 ) . The authors ( Gonzalez, Gutierrez, et al., 1996 ) found that manganite (MnOOH) and pyrolusite (MnO 2 ) exhibited faster weight loss at air, inert and reducing a tmosphere, e.g., weight loss start ing from around 60 o C, rapid weight loss start ing from below 300 o C and 600 o C under inert atmosphere, respectively. It is possible that some Mn oxides in MPB underwent transformation resulting in lower thermal stability. In the second phase of TGA curve (Figure 4 2 ), PB dropped more rapidly than MPB, indicating after the in flection point MPB became more stable. However, the weight of BPB did not maintain stable as PB and MPB, instead it kept dropping. The relationship between weight and temperature showed a negative slope ( R 2 =0.92) until around 400 o C. This gradual weight loss in the first period of the BPB may be due largely to the water loss
77 entrapped in the interlayers of birnessite ( Gonzalez, Gutierrez, et al., 1996 , Post, 1999 ) . The higher thermal stability of BPB may be related to the prop erties of birnessite, which is relatively stable until 620 o C in air atmosphere ( Gonzalez, Gutierrez, et al., 1996 ) . As ( V ) And Pb ( II ) Sorption Kinetics And Isotherms As (V) and Pb (II) sorption by both composites as well as the pristine biochar was biphasic, with a rapid initial phase over the first few hours followed by a much slow sorption phase (Figure 4 5 ). The rapid sorption phase may be ascribed to the rapid occu pation of easily accessible external surface sorption sites ( Chowdhury, Yanful, et al., 2011 ) , likely via physical sorption. For example, rapid sorption kinetics were reported for As(V) sorption by Fe modified activated carbon ( Payne and Abel Fattah, 2005 ) and for Pb(II) sorption between adjacent layers (not interlayers) of birnessite ( O'Reilly and Hochella Jr, 2003 ) . The slow sorption phase may be attributed to specific (chemo ) and irreversible sorption. The sorption kinetics data was fitted with pseudo first order, pseudo second order, and Elovich kinetic models to provide insight into the potential sorption mechanisms (Figure 4 5 ). As(V) sorption by PB, MPB and BPB was best fit by the Elovich kinetic model with R 2 >0.98 (Table 4 2). This was also true for Pb(II) sorption kinetics of the PB and MPB. For the BPB, however, the Pb(II) sorption kinetics were better described by the first and second order kinetic equations (Table 4 3). These results suggested that sorption of heavy meta ls on the biochar samples used in this work could be controlled by multiple mechanisms. of MPB and BPB were more than 72.3 and 2.5 times greater than that of PB, respectiv
78 also much lower than that of PB. These results indicated that, in comparison to the pristine bi ochar, the modified biochars had greater affinity to the two metals. The sorption isotherms were generated by varying the ratio of sorbate to sorbent. At low sorbate concentrations (below about 5 mg kg 1 As(V) or Pb(II)), sorption isotherms increased rapi dly with increasing equilibrium sorbate concentrations (Figure 4 6 ). Sorption increased much more slowly at higher sorbate concentrations. Both Langmuir and Freundlich models reproduced the As(V) sorption isotherm data well (R 2 > 0.98, Table 4 2). The Lang muir maximum sorption capacities (S max ) of MPB and BPB to As(V) were about 3 and 4.7 times greater than that of PB, respectively. Similarly, both the Langmuir and Freundlich models described the sorption of Pb(II) to the sorbents well (Table 4 3). The S max of MPB and BPB to Pb(II) were about 2.1 and 20.0 times greater than that of PB, respectively. These results confirmed that the modification enhanced the As(V) and Pb(II) sorption ability of the biochar. A ( V ) Sorption Mechanisms For the MPB, the EDS spectr a of the post sorption samples did not show the interaction between sorbed As(V) and Mn oxide (mainly MnO) particles (Figure 4 7 a and b ), which might be because the content of As adsorbed on the MPB was below the determination limit of EDS ( Hu, Ding, e t al., 2015 ) . The XRD pattern of the post sorption MPB (Figure 4 3b ) were similar to the original sample (Figure 4 3a ), suggesting precipitation might not play an important role in As sorption. Additional investigations are still needed to better unders tand the effects of MnO modification on the sorption of As(V) onto MPB.
79 Association of As(V) to Mn oxide (mainly birnessite) particles of post sorption BPB samples was detected by the EDS analysis (Figure 4 7c and d). This result indicated that interactio n between As(V) and birnessite particles could play an important role in As(V) sorption onto the BPB. The XRD pattern of the post sorption BPB (Figure 4 4c ) did not show any new peaks, suggesting precipitation might not be an important mechanism for As(V) sorption onto the BPB. Previous studies have demonstrated the strong interaction between As(V) and birnessite particles and the sorption is found to be related to edge site of birnessite ( Manning, Fendorf, et al., 2002 , To urnassat, Charlet, et al., 2002 ) . Pb ( II ) Sorption Mechanisms T he SEM/EDS analyses of the post sorption MPB samples could not verify that the sorbe d Pb(II) was related to the MnO particles in the MPB (Figure 4 8 and 4 9 ). However, the SEM/EDS results of the post sorption BPB showed the Pb(II) might be associated with the birnessite particles in the BPB (Figure 4 10 ). This result suggested that birnes site particles in the modified biochars served as Pb(II) sorption site. The EDS mapping of the post sorption MPB showed no obvious association of Pb(II) and MnO particles, but did show a large Pb particle (Figure 4 9 e and 4 8b ), suggesting the precipitatio n mechanism ( Inyang, Gao, et al., 2012 ) . This mechanism was further confirmed by comparing the XRD spectra of the MPB before and after Pb(II) sorption (Figure 4 3 ). XRD spectra of the post sorption MPB showed new peaks at d=3.597 and 3.409 Ã… (Figure 4 3c ), corresponding to Pb bearing particles . The XRD spectra of the post sorption BPB did not show new peaks, suggesting precipitation might not be an important mechanism for Pb(II) sorption onto BPB. The
80 EDS mapping of the post sorption BPB (Figure 4 10 ), however, showed some association between sorbed Pb(II) and the birnessite particles in the BPB. Previous studies have suggested that Pb(II) cations are sorbed either through diffusion into birnessite interlayers ( O'Reilly and Hochella Jr, 2003 ) , by replacing cations located above or below the vacancy sites of birnessite ( Lee, Kim, et al., 2013 ) , or by coordinatio n to vacancy sites of hexagonal birnessite surface structure ( Lafferty, Ginder Vogel, et al., 2010 , Matocha, Elzinga, et al., 2001 ) . Summary and Conclusion Biochar/manganosite (i.e., MPB) and biochar/birnessite (i.e., BPB) composites were synthesized. In comparison to the pristine biochar, although both MPB and BPB showed enhanced sorption of As(V) and Pb(II), the BPB had much better sorption ability. The enhanced As(V) and Pb(II) sorption by BPB was mainly due to the presences of birnessite particles within the BPB, which showed strong interactions with the two heavy metal s and thus played an important role in the sorption. Findings from this work suggested that birnessite modification of biochar provides an effective way to prepare low cost carbon adsorbents for heavy metal treatment.
81 T able 4 1. modified biochar (MPB) and biochar modified with synthesized birnessite (BPB) Biochar C N H O Mn K Ca Mg Al P BET Surface area m 2 g 1 BJH pore vo l (des. leg) cc g 1 %, mass based PB 85.68 0.33 2.13 11.19 0.023 0.052 0.186 0.120 0.041 0.039 209.6 0.003 MPB 78.95 0.26 1.86 14.58 4.19 0.018 0.056 0.062 0.008 0.020 463.1 0.022 BPB 61.54 0.25 1.85 27.65 8.14 0.163 0.246 0.143 0.008 0.012 67.4 0.066
82 Table 4 2 . Best fit parameters for kinetics and isotherm models of As(V) sorption onto pine wood biochar (PB), MnCl 2 2 O modified biochar (MPB) and biochar modified with synthesized birnessite (BPB) Model/Equations Biochar Parameter1 Parameter2 R 2 Kinetics First order PB q e =0.1285 g kg 1 k 1 =1.66 h 1 0.814 MPB q e =0.5092 g kg 1 k 1 =2.25 h 1 0.938 BPB q e =0.7539 g kg 1 k 1 =1.18 h 1 0.907 Second order PB q e =0.1386 g kg 1 k 2 =16.2 kg g 1 h 1 0.900 MPB q e =0.5344 g kg 1 k 2 =7.1 kg g 1 h 1 0.967 BPB q e =0.8098 g kg 1 k 2 =2.1 kg g 1 h 1 0.968 Elovich PB 1 h 1 1 0.961 MPB 1 h 1 1 0.991 BPB 1 h 1 1 0.991 Isotherm Langmuir PB S max =0.201 g kg 1 K=0.3155 L g 1 0.991 MPB S max =0.594 g kg 1 K=2.913 L g 1 0.997 BPB S max =0.932 g kg 1 K=1.592 L g 1 0.980 Freundlich PB n=0.3322 K f =0.0681 g (1 n) L n kg 1 0.995 MPB n=0.0958 K f =0.454 g (1 n) L n kg 1 0.999 BPB n=0.1817 K f =0.566 g (1 n) L n kg 1 0.997
83 Table 4 3 . Best fit parameters for kinetics and isotherm models of Pb(II) sorption onto pine wood biochar (PB), MnCl 2 2 O modified biochar (MPB) and biochar modified with synthesized birnessite (BPB) Model/Equations Biochar Parameter1 Parameter2 R 2 Kinetics First order PB q e =1.944 g kg 1 k 1 =0.273 h 1 0.923 MPB q e =3.792 g kg 1 k 1 =0.823 h 1 0.890 BPB q e = 16.26 g kg 1 k 1 = 0.663 h 1 0.984 Second order PB q e =2.151 g kg 1 k 2 =0.180 kg g 1 h 1 0.949 MPB q e =4.071 g kg 1 k 2 =0.303 kg g 1 h 1 0.956 BPB q e =17.50 g kg 1 k 2 = 0.0551kg g 1 h 1 0.989 Elovich PB 1 h 1 1 0.954 MPB 1 h 1 1 0.980 BPB 1 h 1 1 0.939 Isotherm Langmuir PB S max = 2.352 g kg 1 K=0.0841 L g 1 0.895 MPB S max =4.913 g kg 1 K=0.193 L g 1 0.916 BPB S max =47.05 g kg 1 K= 0.709 L g 1 0.923 Freundlich PB n=0.292 K f =0.574 g (1 n) L n kg 1 0.940 MPB n=0.278 K f =1.485 g (1 n) L n kg 1 0.990 BPB n=0.223 K f = 16.80 g (1 n) L n kg 1 0.808
84 Figure 4 1. XPS spectra of MnCl 2 2 O modified biochar (MPB) (a) and biochar modified with synthesized birnessite (BPB) (b).
85 Figure 4 2 . Thermogravimetric (TG) curves of pine wood biochar (PB), MnCl 2 2 O modified biochar (MPB) and biochar modified with synthesized birnessite (BPB)
86 Figure 4 3. XRD diffraction patterns of MnCl 2 2 O modified biochar (MPB) before (a), after As(V) (b), and after Pb(II) (c) sorption.
87 Figure 4 4. XRD diffraction patterns of biochar modified with birnessite (BPB) before (a), after As(V) (b), and after Pb(II) (c) sorption.
88 Figure 4 5. As(V) (a) and Pb(II) (b) sorption kinetics data and fitted models for pine wood biochar (PB), MnCl 2 2 O modified biochar (MPB) and biochar modified with synthesized birnessite (BPB)
89 Figure 4 6. As(V) (a) and Pb(II) (b) sorption isotherm data and fitted models for pine wood biochar (PB), MnCl 2 2 O modified biochar (MPB) and biochar modified with synthesized birnessite (BPB)
90 Figure 4 7. SEM image and corresponding EDS spectra of modified biochars after As(V ) sorption: (a b) MnCl 2 2 O modified biochar (MPB) and (c d) biochar modified with synthesized birnessite (BPB).
91 Figure 4 8 . SEM image and corresponding EDS spectra of modified biochars after Pb(II) sorption: (a b) MnCl 2 2 O modified biochar (MPB) and (c d) biochar modified with synthesized birnessite (BPB).
92 Figure 4 9 . SEM/EDS elemental mapping analysis (1000 x) of MnCl 2 2 O modified biochar (MPB) after Pb(II) sorption, including surface topography (a), and individual elemental distribution map of C (b), Ca (c), Mn (d), Pb(II) (e), and O (f).
93 Figure 4 10 . SEM/EDS elemental mapping analysis (1000 x) of biochar modified with synthesized birnessite (BPB) after Pb(II) sorption, including surface topography (a) and individual elemental distribution map of C (b), O (c), Pb(II) (d), Mn (e), and K (f).
94 CHAPTER 5 SUMMARY AND CONCLUSION As per literature review and findings of this work, carbon enriched bichars can serve as an inexpensive and efficient sorbent for environmental remediation. Th is work investigated the effects of two important factors, feedstock types and char r ing temperature, on biochar properties and sorption of As(V) and Pb(II) . To improve sorption capacity, Fe and Mn oxide minerals have been infused into pristine biochars aiming to increase the sorption capacity for these he avy metals. To gain a n overall understanding of the effects of temperature and feedstock types on biochar propert ies and performance in heavy metal sorption, in this work, four feedstocks were used to make biochars at three charring temperatures. The resul ts showed feedstock types affected physicochemical properties of biochar , including production rate, elemental composition, pH, surface area, and thermal stability. Results of c orrelation analysis between biochar propert ies and arsenate (As(V)) and lead (P b(II)) sorption capacit ies suggested that feedstock type is the main factor dominating sorption capacity while temperature effect is minimal. Over all , the sorption for Pb (II) is positively related to alkaline metal content such as Ca, Mg and K while As (V) sorption relates to acidifying metals such as Al. Thus, I concluded that electrostatic interaction played an important role in controlling the sorption of both Pb(II) and As(V) onto the biochar . (Hydro) oxides of metals are reactive minerals in soil and reported repeatedly for good As (V) sorption. Hence, natural hematite as one of the most abundant Fe oxides was used to modify biochars through a single step modification method. The XRD and SEM analyses proved that a bio Fe 2 O 3 composite was produced with nearly
95 doubled maximal As sorption capacity as compared to un modified pine bio char. The Fe 2 O 3 paticles as confirmed by EDS mapping analysis, and thus increase its specific surface area. Besides, the Fe 2 O 3 co mposite is more magnetic which facilitates removal from aqueous s o lutions. The sorption mainly occurs on Fe 2 O 3 particles on the carbon surface through electrostatic interactions . Compared to other magnetic biochars, this new composite was more easily fab ricated with inexpensive materials and thus ha s good potential for environmental application . Apart from Fe oxides, Mn oxides are very reactive minerals for both As (V) and especially Pb (II) sorption . Birnessite , abundant in environment where Mn oxide resides, was used to modify pine biochar. The modification changed elemental composition, greatly decreased specific surface area and thermal stability and increased pore volume . Sorption isotherms study showed that the biochar/birnessite composite had extraordinary sorption for Pb (II) and greatly increased sorption for As (V) . The sorption of biochar/birnessite composite for Pb (II) is mainly attributed to birnessite. However, another modification method which produc es biochar/MnO composite is not as effective, although the sorption capacity was also increased . Instead, Pb bearing particles were formed on the biochar/MnO surface indicating precipitation is the major mechanismThe sorbed As concentration is below the det ection limit of EDS , making it difficult to infer possible mechanisms. The results of this work enlighten the use of biochars for As (V) and especially Pb (II) remo v al from aq u eous solutions . The modification with oxides of Fe and Mn is an effective way to increase sorption capacity of As and Pb and thus have good environmental implications. The sorbents can be used for industrial wastewater
96 remediation. The particle size of biochar/metal oxides composites is larger than the commonly used synthetic metal ox ide minerals. This make it easier to be separated from aqueous solution after treatment process.
97 LIST OF REFERENCES Ahmad, M., A.U. Rajapaksha, J.E. Lim, M. Zhang, N. Bolan, D. Mohan, et al. 2014. Biochar as a sorbent for contaminant management in soil and water: A review. Chemosphere 99: 19 33. Akhtar, M.S., B. Chali and T. Azam. 2013. Bioremediation of arsenic and lead by plants and microbes from contaminated s oil. Research in Plant Sciences 1: 68 73. doi:10.12691/plant 1 3 4. Anderson, M.A., J.F. Ferguson and J. Gavis. 1976. Arsenate adsorption on amorphous aluminum hydroxide. Journal of Colloid and Interface Science 54: 391 399. doi: http:/ /dx.doi.org/10.1016/0021 9797(76)90318 0 . Auffan, M., J. Rose, O. Proux, D. Borschneck, A. Masion, P. Chaurand, et al. 2008. Enhanced adsorption of arsenic onto maghemites nanoparticles: As(III) as a probe of the surface structure and heterogeneity. Langmu ir 24: 3215 3222. doi:10.1021/1a702998x. Baldock, J.A. and R.J. Smernik. 2002. Chemical composition and bioavailability of thermally altered Pinus resinosa (Red pine) wood. Organic Geochemistry 33: 1093 1109. doi: http://dx.doi.org/10.1016/S0146 6380(02)00062 1 . Beesley, L. and M. Marmiroli. 2011. The immobilisation and retention of soluble arsenic, cadmium and zinc by biochar. Environmental Pollution 159: 474 480. doi: http://dx.doi.org/10.1016/j.envpol.2010.10.016 . Beesley, L., M. Marmiroli, L. Pagano, V. Pigoni, G. Fellet, T. Fresno, et al. 2013. Biochar addition to an arsenic contaminated soil increases arsenic concentrations in th e pore water but reduces uptake to tomato plants (Solanum lycopersicum L.). Science of the Total Environment 454: 598 603. doi:10.1016/j.scitotenv2013.02.047. Bell, M.J. and F. Worrall. 2011. Charcoal addition to soils in NE England: A carbon sink with env ironmental co benefits? Science of The Total Environment 409: 1704 1714. doi: http://dx.doi.org/10.1016/j.scitotenv.2011.01.031 . Bhattacharya, P., A.C. Samal, J. Majumdar and S.C. Santra. 201 0. Arsenic contamination in rice, wheat, pulses, and vegetables: a study in an arsenic affected area of West Bengal, India. Water Air Soil Pollut 213: 3 13. doi:10.1007/s11270 010 0361 9. Bove, J.M. 2006. Huanglongbing: A destructive, newly emerging, centu ry old disease of citrus. Journal of Plant Pathology 88: 7 37. Cao, X., L.Q. Ma, D.R. Rhue and C.S. Appeal. 2004. Mechanisms of lead, copper, and zinc retention by phosphate rock. 131: 435 444.
98 Cao, X.D., L. Ma, B. Gao and W. Harris. 2009. Dairy manure der ived biochar effectively sorbs lead and atrazine. Environ Sci Technol 43: 3285 3291. doi:Doi 10.1021/Es803092k. Case, S.D.C., N.P. McNamara, D.S. Reay and J. Whitaker. 2012. The effect of biochar addition on N2O and CO2 emissions from a sandy loam soil T he role of soil aeration. Soil Biology and Biochemistry 51: 125 134. doi: http://dx.doi.org/10.1016/j.soilbio.2012.03.017 . Catalano, J.G., C. Park, P. Fenter and Z. Zhang. 2008. Simultaneous in ner and outer sphere arsenate adsorption on corundum and hematite. Geochimica Et Cosmochimica Acta 72: 1986 2004. doi:10.1016/j.gca.2008.02.013. Cayuela, M.L., L. van Zwieten, B.P. Singh, S. Jeffery, A. Roig and M.A. SÃ¡nchez Monedero. 2013. Biochar's role in mitigating soil nitrous oxide emissions: A review and meta analysis. Agriculture, Ecosystems & Environment. doi: http://dx.doi.org/10.1016/j.agee.2013.10.009 . Chang, Q.G., W. Lin and W.C. Ying . 2010. Preparation of iron impregnated granular activated carbon for arsenic removal from drinking water. J Hazard Mater 184: 515 522. doi:10.1016/j.jhazmat.2010.08.066. Chen, B.L., Z.M. Chen and S.F. Lv. 2011. A novel magnetic biochar efficiently sorbs o rganic pollutants and phosphate. Bioresource Technol 102: 716 723. doi:DOI 10.1016/j.biortech.2010.08.067. Chen, H., L.Q. Ma, B. Gao and C. Gu. 2013. Effects of Cu and Ca cations and Fe/Al coating on ciprofloxacin sorption onto sand media. Journal of Hazar dous Materials 252 253: 375 381. doi: http://dx.doi.org/10.1016/j.jhazmat.2013.03.014 . Chen, L., H. Xing, Y. Shen, J. Bai and G. Jiang. 2009. Solid state thermolysis of [MnO]12 containing molec ular clusters into novel MnO nano and microparticles. Journal of Solid State Chemistry 182: 1387 1395. doi: http://dx.doi.org/10.1016/j.jssc.2009.03.002 . Chen, M., L.Q. Ma, C.G. Hoogeweg and W.G. Harris. 2001. Arsenic Background Concentrations in Florida, U.S.A. Surface Soils: Determination and Interpretation . Environmental Forensics 2( 2 ) : 117 126. Cheney, M.A., P.K. Bhowmik, S.Z. Qian, S.W. Joo, W.S. Hou and J.M. Okoh. 2008. A new method of synth esizing black Birnessite nanoparticles: From brown to black birnessite with nanostructures. Journal of Nanomaterials. Journal of Nanomaterials. doi:10.1155/2008/763706. Cheng, H. and Y. Hu. 2010. Lead (Pb) isotopic fingerprinting and its applications in le ad pollution studies in China: A review. Environmental Pollution 158: 1134 1146. doi: http://dx.doi.org/10.1016/j.envpol.2009.12.028 .
99 Chowdhury, S.R., E.K. Yanful and A.R. Pratt. 2011. Arsenic r emoval from aqueous solutions by mixed magnetite maghemite nanoparticles. Environmental Earth Sciences 64: 411 423. doi:10.1007/s12665 010 0865 z. Coletta, H.D., M.A. Takita, M. Targon and M.A. Machado. 2005. Analysis of 16S rDNA sequences from citrus huan glongbing bacteria reveal a different "Ca. Liberibacter" strain associated with citrus disease in Sao Paulo. Plant Disease 89: 848 852. doi:10.1094/pd 89 0848. Cullen, W. and K. Reimer. 1989. Arsenic speciation in the environment. Chemical Reviews 89: 713 764. doi:10.1021/cr00094a002. Deenik, J.L., T. McClellan, G. Uehara, M.J. Antal and S. Campbell. 2010. Charcoal volatile matter content influences plant growth and soil nitrogen transformation s. Soil Science Society of America Journal 74: 1259 1270. doi:10 .2136/sssaj2009.0115. Demirbas, A. 2004. Effects of tempe rature and particle size on bio char yield from pyrolysis of agricultural residues. Journal of Analytical and Applied Pyrolysis 72: 243 248. doi:10.1016/j.jaap.2004.07.003. Dien, B.S., D.J. Miller, R. E. Hector, R.A. Dixon, F. Chen, M. McCaslin, et al. 2011. Enhancing alfalfa conversion efficiencies for sugar recovery and ethanol production by altering lignin composition. Bioresource Technology 102: 6479 6486. doi:10.1016/j.biortech.2011.03.022. Dixit, S. and J.G. Hering. 2003. Comparison of Arsenic(V) and Arsenic(III) Sorption onto Technology 37: 4182 4189. doi:10.1021/es030309t. Dong, D., Y.M. Nelson, L.W. Lion, M.L. Shuler and W.C. Ghiorse. 2000. Adsorption of Pb and Cd onto metal oxides and organic material in natural surface coatings as determined by selective extractions: new evidence for the importance of Mn and Fe oxides. Water Research 34: 427 436. doi: http://dx.doi.org/10.1016/S0043 1354(99)00185 2 . Dong, X., L.Q. Ma and Y. Li. 2011. Characteristics and mechanisms of hexavalent chromium removal by biochar from sugar beet tailing J. Hazard. Mater. 19 0: 909 915. Dong, X., L.Q. Ma, Y. Zhu, Y. L i and B. Gu. 2013. Mechanistic investigation of mercury sorption by brazilian pepper biochars of different pyrolytic temperatures based on x ray photoelectron spectroscopy and flow calorime try. Environmental Scien ce & Technology 47: 12156 12164. doi:10.1021/es4017816. Essington, M. 2004. Soil and Water Chemistry An Integrative Approach . CRC Press.
100 Etxeberria, E., P. Gonzalez, D. Achor and G. Albrigo. 2009. Anatomical distribution of abnormally high levels of starch in HLB affected Valencia orange trees. Physiological and Molecular Plant Pathology 74: 76 83. doi:10.1016/j.pmpp.2009.09.004. Ferguson, J.F. and J. Gavis. 1972. A review of the arsenic cycle in natural waters. Water Research 6: 1259 1274. doi: http://dx.doi.org/10.1016/0043 1354(72)90052 8 . Fiasconaro, M.L., Y. Gogorcena, F. Munoz, D. Andueza, M. Sanchez Diaz and M.C. Antolin. 2012. Effects of nitrogen source and water availability on stem carbo hydrates and cellulosic bioethanol traits of alfalfa plants. Plant Science 191: 16 23. doi:10.1016/j.plantsci.2012.04.007. Fougere, F., D. Lerudulier and J.G. Streeter. 1991. Effects of salt stress on amino acid, organic acid, and carbohydrate composition of roots, bacteroids, and cytosol of alfalfa (medicago satival). Plant Physiology 96: 1228 1236. doi:10.1104/pp.96.4.1228. Gaillot, A. C., V.A. Drits, A. Manceau and B. Lanson. 2007. Structure of the synthetic K rich phyllomanganate birnessite obtained by high temperature decomposition of KMnO4: Substructures of K rich birnessite from 1000 Â°C experiment. Microporous and Mesoporous Materials 98: 267 282. doi: http://dx.doi.org/10.1016/j.mi cromeso.2006.09.010 . Gaillot, A. C., V.A. Drits, A. PlanÃ§on and B. Lanson. 2004. Structure o f synthetic k rich birnessites obtained by high temperature decomposition of KMnO4. 2. Pha se and structural heterogeneit ies. Chemistry of Materials 16: 1890 1905. do i:10.1021/cm035236r. Gaillot, A. C., D. Flot, V.A. Drits, A. Manceau, M. Burghammer and B. Lanson. 2003. Structur e of synthetic k rich birnessite obtained by high temperature decomposition of KMnO4. I. Two layer polytype f rom 800 Â°C Experiment. Chemistry of Materials 15: 4666 4678. doi:10.1021/cm021733g. Gaskin, J.W., C. Steiner, K. Harris, K.C. Das and B. Bibens. 2008. Effect of low temperature pyrolysis conditions on biochar for agricultural use. Transactions of the Asabe 51: 2061 2069. Gerente, C., V.K.C. Lee, P. Le Cloirec and G. McKay. 2007. Application of chitosan for the removal of metals from wastewaters by adsorption Mechanisms and models review. Crit Rev Env Sci Tec 37: 41 127. Gimenez, J., M. Martinez, J. de Pablo, M. Rovira and L. Duro. 2007. Ar senic sorption onto natural hematite, magnetite, and goethite. Journal of Hazardous Materials 141: 575 580. doi:10.1016/j.jhazmat.2006.07.020.
101 GimÃ©nez, J., M. MartÃnez, J. de Pablo, M. Rovira and L. Duro. 2007. Arsenic sorption onto natural hematite, magne tite, and goethite. Journal of Hazardous Materials 141: 575 580. doi: http://dx.doi.org/10.1016/j.jhazmat.2006.07.020 . Githinji, L. 2014. Effect of biochar application rate on soil physical and hydraulic properties of a sandy loam. Archives of Agronomy and Soil Science 60: 457 470. doi:10.1080/03650340.2013.821698. Goldberg, S. 2002. Competitive adsorption of arsenate and arsenite on oxides and clay minerals. Soil Science Society of America Jour nal 66: 413 421. Goldberg, S., H.S. Forster and C.L. Godfrey. 1996. Molybdenum adsorption on oxides, clay minerals, and soils. Soil Science Society of America Journal 60: 425 432. Goldberg, S. and C.T. Johnston. 2001. Mechanism s of arsenic adsorption on am orphous oxides evaluated using macroscopic measurements, vibrational spectroscopy, and surface complexation modeli ng. Journal of Colloid and Interface Science 234: 204 216. doi: http://dx.doi.org/10.1 006/jcis.2000.7295 . Gonzalez, C., J.I. Gutierrez, J.R. GonzalezVelasco, A. Cid, A. Arranz and J.F. Arranz. 1996. Transformations of manganese oxides under different thermal conditions. Journal of Thermal Analysis 47: 93 102. doi:10.1007/bf01982689. Gottwal d, T.R. 2010. Current e pidemiological understanding of citrus huanglongbing. annual review of phytopatho logy 48: 119 139. doi:10.1146/annurev phyto 073009 114418. Graeme Md, K.A. and M.D.F.C.V. Pollack Jr. 1998. Heavy metal toxicity, Part I: Arsenic and Me rcury. The Journal of Emergency Medicine 16: 45 56. doi: http://dx.doi.org/10.1016/S0736 4679(97)00241 2 . Graham, J.H., E.G. Johnson, T.R. Gottwald and M.S. Irey. 2013. Presymptomatic fibrous r oot decline in citrus trees caused by huanglongbing and potential interaction with phytophtho ra spp. Plant Disease 97: 1195 1199. doi:10.1094/pdis 01 13 0024 re. Greer, M., G. Goodman, R. Pleus and S. Greer. 2002. Health effects assessment for environmenta l perchlorate contamination: The dose response for inhibition of thyroidal radioidine uptake in humans (vol 110, pg 927, 2002). Environmental Health Perspectives 110: A564 A564. Grossl, P.R., D.L. Sparks and C.C. Ainsworth. 1994. Rapid kinetics of Cu(II) a dsorption desorption on goethite. Environmental Science & Technology 28: 1422 1429. doi:10.1021/es00057a008. Gulz, P., S. K. Gupta and R. Schulin. 2005. Arsenic accumulation of common plants from contaminated soils. Plant Soil 272: 337 347. doi:10.1007/s11 104 004 5960 z.
102 Gulz, P.A. 2002. Arsenic uptake of common crop plants from contaminated soils and interaction with phosphate. Swiss Federal Institute of Technology Zurich. Guo, H., D. Stuben and Z. Berner. 2007. Removal of arsenic from aqueous solution by natural siderite and hematite. Applied Geochemistry 22: 1039 1051. doi:10.1016/j.apgeochem.2007.01.004. Hammes, K., R.J. Smernik, J.O. Skjemstad and M.W.I. Schmidt. 2008. Characterisation and evaluation of reference materials for black carbon analysis usin g elemental composition, colour, BET surface area and C 13 NMR spectroscopy. Applied Geochemistry 23: 2113 2122. doi:10.1016/j.apgeochem.2008.04.023. Harvey, O.R., B.E. Herbert, R.D. Rhue and L.J. Kuo. 2011. Metal I nteractions at the biochar water interfa ce: Ener getics and structure sorption relationships elucidated by flow adsorption microcalorim etry. Environmental Science & Technology 45: 5550 5556. doi:10.1021/es104401h. Hendrickson, J.R., M.R. Schmer and M.A. Sanderson. 2013. Water use efficiency by sw itchgrass compared to a native grass or a native grass alfalfa mixture. Bioenergy Research 6: 746 754. doi:10.1007/s12155 012 9290 3. Hu, X., Z. Ding, A.R. Zimmerman, S. Wang and B. Gao. 2015. Batch and column sorption of arsenic onto iron impregnated bioc har synthesized through hydrolysis. Water Res 68: 206 216. Inyang, M., B. Gao, Y. Yao, Y. Xue, A.R. Zimmerman, P. Pullammanappallil, et al. 2012. Removal of heavy metals from aqueous solution by biochars derived from anaerobically digested biomass. Bioreso urce Technology 110: 50 56. doi:10.1016/j.biortech.2012.01.072. Inyang, M.D., B. Gao, W.C. Ding, P. Pullammanappallil, A.R. Zimmerman and X.D. Cao. 2011. Enhanced lead sorption by biochar derived from anaerobically digested sugarcane bagasse. Separation Sc ience And Technology 46: 1950 1956. doi:10.1080/01496395.2011.584604. Ippolito, J.A., J.M. Novak, W.J. Busscher, B.M. Ahmedna, D. Rehrah and D.W. Watts. 2012. Switchgrass biochar affects two Aridisols. Journal of Environmental Quality 41: 1123 1130. doi:10 .2134/jeq2011.0100. Jensen, K., C.D. Clark, P. Ellis, B. English, J. Menard, M. Walsh, et al. 2007. Farmer willingness to grow switchgrass for energy production. Biomass and Bioenergy 31: 773 781. doi: http://dx.doi.org/10.1016/j.biombioe.2007.04.002 . Jones, D.L., G. Edwards Jones and D.V. Murphy. 2011. Biochar mediated alterations in herbicide breakdown and leaching in soil. Soil Biology and Biochemistry 43: 804 813. doi: http://dx.doi.org/10.1016/j.soilbio.2010.12.015 .
103 Jung, C., J. Park, K.H. Lim, S. Park, J. Heo, N. Her, et al. 2013. Adsorption of selected endocrine disrupting compounds and pharmaceuticals on activated biochars. J ournal of Hazardous Materials 263, Part 2: 702 710. doi: http://dx.doi.org/10.1016/j.jhazmat.2013.10.033 . Karhu, K., T. Mattila, I. BergstrÃ¶m and K. Regina. 2011. Biochar addition to agricultur al soil increased CH 4 uptake and water holding capacity Results from a short term pilot field study. Agriculture, Ecosystems & Environment 140: 309 313. doi: http://dx.doi.org/10.1016/j.agee.201 0.12.005 . Katz, B.G., M.P. Berndt, T.D. Bullen and P. Hansard. 1999. Factors controlling elevated lead concentrations in water samples from aquifer systems in Florida. U.S. Geological Survey, Tallahassee, Florida. p. 1 22. Keiluweit, M., P.S. Nico, M.G. J ohnson and M. Kleber. 2010. Dynamic molecular structure of plant biomass derived black carbon (biochar). Environmental Science & Technology 44: 1247 1253. doi:10.1021/es9031419. Kim, J.S., U.S. Sagaram, J.K. Burns, J.L. Li and N. Wang. 2009. Response of sw eet orang e (Citrus sinensis) to 'Candidatus Liberibacter asiaticus' i nfection: microscopy and micr oarray Analyses. Phytopathology 99: 50 57. doi:10.1094/phyto 99 1 0050. Kim, W., C. Suh, S. Cho, K. Roh, H. Kwon, K. Song, et al. 2012. A new method for the i dentification and quantification of magnetite maghemite mixture using conventional X ray diffraction technique. Talanta 94: 348 352. doi:10.1016/j.talanta.2012.03.001. Kloss, S., F. Zehetner, A. Dellantonio, R. Hamid, F. Ottner, V. Liedtke, et al. 2012. Ch aracterizatio n of slow pyrolysis biochars: effects of feedstocks and pyrolysis temperature on biochar properties . J Environ Qual 41: 990 1000. doi:Doi 10.2134/Jeq2011.0070. Knowles, O.A., B.H. Robinson, A. Contangelo and L. Clucas. 2011. Biochar for the mi tigation of nitrate leaching from soil amended with biosolids. Science of The Total Environment 409: 3206 3210. doi: http://dx.doi.org/10.1016/j.scitotenv.2011.05.011 . Lafferty, B., M. Ginder Vogel and D. Sparks. 2010. Ars enite oxidation by a poorly crystalline manganese oxide 1. Stirred flow experiments . Environmental Science & Technology 44: 8460 8466. doi:10.1021/es102013p. Lafferty, B., M. Ginder Vogel and D. Sparks. 2011. Arsenite oxidati on by a poorly crystalline manganese oxide. 3. Arsenic and manganese desorption . Environmental Science & Technology 45: 9218 9223. doi:10.1021/es201281u.
104 Lafferty, B.J., M. Ginder Vogel, M.Q. Zhu, K.J.T. Livi and D.L. Sparks. 2010. Arsenite oxidation by a poorly crystalline manganese oxide. 2. Results from X ray absorption spectroscopy and X ray diffraction. Environ Sci Technol 44: 8467 8472. doi:Doi 10.1021/Es102016c. Lanphear, B.P., T.D. Matte, J. Rogers, R.P. Clickner, B. Dietz, R.L. Bornschein, et al. 1998. The contribution of lead contaminated house dust and residential soil to children's blood lead levels: a pooled analysis of 12 epidemiologic studie s. Environmental Research 79: 51 68. doi: http://dx.doi.org/10.1006/enrs.1998.3859 . Lee, C. Y., T. Kim, S. Komarneni, S. K. Han and Y. Cho. 2013. Sorption characteristics of lead cations on microporous organo birnessite. Applied Clay Science 83 84: 263 269. doi: http://dx.doi.org/10.1016/j.clay.2013.08.035 . Lee, G.F. 1973. Role of phosphorus in eutrophication and diffuse source control. Water Research 7: 111 128. doi: http://dx. doi.org/10.1016/0043 1354(73)90156 5 . Legodi, M.A. and D. de Waal. 2008. The preparation of magnetite, goethite, hematite and maghemite of pigment quality from mill scale iron waste (vol 74, pg 161, 2007). Dyes and Pigments 78: 177 177. doi:10.1016/j.dyepi g.2008.01.011. Lehmann, J., J. Gaunt and M. Rondon. 2006. Bio char sequestration in terrestrial ecosystems a review Mitigation and Adaptation Strategies for Global Change 11: 403 427. Lenoble, V., C. Laclautre, B. Serpaud, V. Deluchat and J. C. Bollinger. 2004. As(V) retention and As(III) simultaneous oxidation and removal on a MnO2 loaded polystyrene resin. Science of The Total Environment 326: 197 207. doi: http://dx.doi.org/10.1016/j.scito tenv.2003.12.012 . Liu, M., B. Huang, X. Bi, Z. Ren, G. Sheng and J. Fu. 2013. Heavy metals and organic compounds contamination in soil from an e waste region in South China. Environmental Science Processes & Impacts 15: 919 929. doi:10.1039/c3em00043e. Lup oi, J.S. and E.A. Smith. 2012. Characterization of woody and herbaceous biomasses lignin composition with 1064 nm Dispersive Multichannel Raman Spectroscopy. Applied Spectroscopy 66: 903 910. doi:10.1366/12 06621. Ma, L., K. Komar, C. Tu, W. Zhang, Y. Cai and E. Kennelley. 2001. A fern that hyperaccumulates arsenic A hardy, versatile, fast growing plant helps to remove arsenic from contaminated soils. Nature 409: 579 579. doi:10.1038/35054664. Major, J., M. Rondon, D. Molina, S.J. Riha and J. Lehmann. 2010. Maize yield and nutrition during 4 years after biochar application to a Colombian savanna oxisol. Plant Soil 333: 117 128. doi:DOI 10.1007/s11104 010 0327 0.
105 Majzlan, J., K.D. Grevel and A. Navrotsky. 2003. Thermodynamics of Fe oxides: Part II. Entha lpies of formation and relative stability of goethite (alpha FeOOH), lepidocrocite (gamma FeOOH), and maghemite (gamma Fe2O3). American Mineralogist 88: 855 859. Mamindy Pajany, Y., C. Hurel, N. Marmier and M. RomÃ©o. 2009. Arsenic adsorption onto hematite and goethite. Comptes Rendus Chimie 12: 876 881. doi: http://dx.doi.org/10.1016/j.crci.2008.10.012 . Mandal, B.K. and K.T. Suzuki. 2002. Arsenic round the world: a review. Talanta 58: 201 235. doi: http://dx.doi.org/10.1016/S0039 9140(02)00268 0 . Manning, B. and S. Goldberg. 1996. Modeling arsenate competitive adsorption on kaolinite, montmorillonite and illite. Clays and Clay Minerals 4 4: 609 623. doi:10.1346/CCMN.1996.0440504. Manning, B.A., S.E. Fendorf, B. Bostick and D .L. Suarez. 2002. Arsenic(III) oxidation and a rsenic(V) adsorption reactions on synthetic birnessit e. Environmental Science & Technology 36: 976 981. doi:10.1021/es0110 170. Mansilla, M.V., R. Zysler, D. Fiorani and L. Suber. 2002. Annealing effects on magnetic properties of acicular hematite nanoparticles. Physica B Condensed Matter 320: 206 209. Marris, E. 2006. Putting the carbon back: Black is the new green. Nature 44 2: 624 626. Masue, Y., R. Loeppert and T. Kramer. 2007. Arsenate and arsenite adsorption and desorption behavior on coprecipitated aluminum : iron hydroxides. Environmental Science & Technology 41: 837 842. doi:10.1021/es061160z. Matocha, C.J., E.J. Elzing a and D.L. Sparks. 2001. Reactivity of Pb(II) at the Mn(III,IV) 2972. doi:10.1021/es0012164. McKenzie, R.M. 1971. The Synthesis of Birnessite, Crytomelane, and Some Other Oxides and Hydroxides of Manganese. Mineralogical Magazine 38: 493 502. McKenzie, R.M. 1980. The Adsorption of Lead and Other Heavy Metals on Oxides of Manganese and Iron Australian Journal of Soil Research 18: 61 73. Minyuk, P.S., T.V. Subbotnikova and A.A. Plyashkevich. 2011. Measurements of thermal magnetic susceptibility of hematite and goethite. Izvestiya Physics of the Solid Earth 47: 762 774. doi:10.1134/s1069351311080052. Mohan, D., H. Kumar, A. Sarswat, M. Alexandre Franco and C.U. Pittman. 2014. Cadmium and lead remediation using magnetic oak wood and oak bark fast pyrolysis bio chars. Chemical Engineering Journal 236: 513 528. doi:10.1016/j.cej.2013.09.057.
106 Mohan, D., C.U. Pittman, M. Bricka, F. Smith, B. Yancey, J. Mohammad, et al. 2007. Sorption of arsenic, cad mium, and lead by chars produced from fast pyrolysis of wood and bark during bio oil production. Journal of Colloid and Interface Science 310: 57 73. doi:10.1016/j.jcis.2007.01.020. Mohan, D. and C.U. Pittman Jr. 2007. Arsenic removal from water/wastewater using adsorbents A critical review. Journal of Hazardous Materials 142: 1 53. doi: http://dx.doi.org/10.1016/j.jhazmat.2007.01.006 . Mohana, D., A. Sarswata, Y.S. Okb and C.U.P. Jr. 2014. Organ ic and inorganic contaminants removal from water with biochar, a renewable, low cost and sustainable adsorbent A critical review. Bioresource Technol 160: 191 202. Mohapatra, D., D. Mishra, G. Chaudhury and R. Das. 2007. Arsenic(V) adsorption mechanism u sing kaolinite, montmorillonite and illite from aqueous medium. Journal of Environmental Science and Health Part a Toxic/hazardous Substances & Environmental Engineering 42: 463 469. doi:10.1080/10934520601187666. Mukherjee, A. and A.R. Zimmerman. 2013. Or ganic carbon and nutrient release from a range of laboratory produced biochars and biochar soil mixtures. Geoderma 193 194: 122 130. doi: http://dx.doi.org/10.1016/j.geoderma.2012.10.002 . Mukh erjee, A., A.R. Zimmerman and W. Harris. 2011. Surface chemistry variations among a series of laboratory produced biochars. Geoderma 163: 247 255. doi:DOI 10.1016/j.geoderma.2011.04.021. Needleman, H. 2004. Lead poisoning. Annual Review of Medicine 55: 209 222. Newbury, D.E. and N.W.M. Ritchie. 2013. Elemental mapping of microstructures by scanning electron microscopy energy dispersive X ray spectrometry (SEM EDS): extraordinary advances with the silicon drift detector (SDD). Journal of Analytical Atomic Sp ectrometry 28: 973 988. doi:10.1039/C3JA50026H. Novak, J.M., W.J. Busscher, D.W. Watts, J.E. Amonette, J.A. Ippolito, I.M. Lima, et al. 2012. Biochars impact on soil moisture storage in an ultisol and two aridisol s. Soil Science 177: 310 320. doi:10.1097/S S.0b013e31824e5593. O'Reilly, S.E. and M.F. Hochella Jr. 2003. Lead sorption efficiencies of natural and synthetic Mn and Fe oxides. Geochimica et Cosmochimica Acta 67: 4471 4487. doi: http://d x.doi.org/10.1016/S0016 7037(03)00413 7 . Park, J., I. Hung, Z. Gan, O.J. Rojas, K.H. Lim and S. Park. 2013. Activated carbon from biochar: Influence of its physicochemical properties on the sorption characteristics of phenanthrene. Bioresource Technology 1 49: 383 389. doi: http://dx.doi.org/10.1016/j.biortech.2013.09.085 .
107 Payne, K. and T. Abel Fattah. 2005. Adsorption of arsenate and arsenite by iron treated activated carbon and zeolites: Effec ts of pH, temperature, and ionic strength. Journal of Environmental Science and Health Part a Toxic/hazardous Substances & Environmental Engineering 40: 723 749. doi:10.1081/ESE 200048254|10.1081/ESE 20048254. Peacock, C.L. and D.M. Sherman. 2004. Copper(I I) sorption onto goethite, hematite and lepidocrocite: a surface complexation model based on ab initio molecular geometries and EXAFS spectroscopy. Geochimica et Cosmochimica Acta 68: 2623 2637. doi: http://dx.doi.org/10.1016/j.gca.2003.11.030 . Perrin, R., K. Vogel, M. Schmer and R. Mitchell. 2008. Farm scale production cost of switchgrass for biomas s. Bioenergy Research 1: 91 97. doi:10.1007/s12155 008 9005 y. Peterson, S.C., M. Appell, M.A. Jackso n and A.A. Boateng. 2013. Comparing corn stover and switchgrass biochar: characterization and sorption properties. Journal of Agricultural Science (Toronto) 5: 1 8. Post, J. 1999. Manganese oxide minerals: Crystal structures and economic and environmental significance. Proceedings of the National Academy of Sciences of the United States of America 96: 3447 3454. doi:10.1073/pnas.96.7.3447. Qureshi, J.A. and P.A. Stansly. 2009. Exclusion techniques reveal significant biotic mortality suffered by Asian citrus psyllid Diaphorina citri (Hemiptera: Psyllidae) populations in Florida citrus. Biological Control 50: 129 136. doi:10.1016/j.biocontrol.2009.04.001. Ramirez Muniz, K., F. Jia and S. Song. 2012. Adsorption of As V in aqueous solutions on porous hematite pr epared by thermal modification of a siderite goethite concentrate. Environmental Chemistry 9: 512 520. doi:10.1071/EN12120. Rosales, R. and J.K. Burns. 2011. Phytohormon e changes and carbohydrate status in sweet orange fruit from Huanglongbing infected t re es. Journal of Plant Growth Regulation 30: 312 321. doi:10.1007/s00344 011 9193 0. Rutigliano, F.A., M. Romano, R. Marzaioli, I. Baglivo, S. Baronti, F. Miglietta, et al. 2014. Effect of biochar addition on soil microbial community in a wheat crop. Europea n Journal of Soil Biology 60: 9 15. doi: http://dx.doi.org/10.1016/j.ejsobi.2013.10.007 . Schimanke, G. and M. Martin. 2000. In situ XRD study of the phase transition of nanocrystalline maghemite Fe 2 O 3 Fe 2 O 3 ). Solid State Ionics 136 137: 1235 1240. doi: http://dx.doi.org/10.1016/S0167 2738(00)00593 2 . Sika, M.P. and A.G. Hardie. 2014. Effect of pine wood biochar on ammonium nitrate leaching and availability in a South African sandy soil. European Journal of Soil Science 65: 113 119. doi:10.1111/ejss.12082.
108 Singh, S.K., A.K. Ghosh, A. Kumar, K. Kislay, C. Kumar, R.R. Tiwari, et al. 2014. Gro undwater arsenic contaminat ion and associated health ris ks in Bihar, India. International Journal of Environmental Research 8: 49 60. Smith, D.B., W.F. Cannon, L.G. Woodruff, F. Solano, J.E. Kilburn and D.L. Fey. 2013. Geochemical and mineralogical data for soils of the conterminous United States: U.S. Geological Survey Data In: http://pubs.usgs.gov/ds/801/ , editor. p. 19. Sohi, S.P., E. Krull, E. Lopez Capel and R. Bol. 2010. A review of biochar and its use and function in soil. Advances in Agronomy, Vol 105 105: 47 82. doi:10.1016/s0065 2113(10)05002 9. Sparks, D.L. 1999. Kinetics and mechanisms of chemical reactions at the soil mineral/water surface. Soil physical chemistry, Second edition. CRC Press LLC. p. 135 191. Spokas, K.A., W.C. Koskinen, J.M. Baker and D.C. Reicosky. 2009. Impacts of woodchip biochar additions on greenhouse gas production and sorption/degradation of two herbicides in a Minnesota soil. Chemosphere 77: 574 581. doi:DOI 10.1016/j.chemosphere.2009.06.053. Summary, C. F. 2012. USDA National Agricultural Statistics Service. Sun, Y., B. Gao, Y. Yao, J. Fang, M. Zhang, Y. Zhou, et al. 2014. Effects of feedstock type, production method, and pyrolysis temperature on biochar and hydrochar properties. Chemical Engineering Jour nal 240: 574 578. doi: http://dx.doi.org/10.1016/j.cej.2013.10.081 . Sun, Y.N., B. Gao, Y. Yao, J.N. Fang, M. Zhang, Y.M. Zhou, et al. 2014. Effects of feedstock type, production method, and pyrolys is temperature on biochar and hydrochar properties. Chem Eng J 240: 574 578. doi:DOI 10.1016/j.cej.2013.10.081. Tournassat, C., L. Charlet, D. Bosbach and A. Manceau. 2002. Arsenic(III) oxidation by birnessite and precipitation of manganese(II) arsenate. E nviron Sci Technol 36: 493 500. doi:Doi 10.1021/Es0109500. application to metal contaminated soil: Evaluating of Cd, Cu, Pb and Zn sorption behavior using single and multi elemen t sorption experiment. p. 372 380. Trivedi, P., J.A. Dyer and D.L. Sparks. 2003. Lead sorption onto ferrihydrite. 1. A macroscopic and spectroscopic assessment. Environmental Science & Technology 37: 908 914. doi:10.1021/es0257927.
109 Tronc, E., A. Ezzir, R. Cherkaoui, C. ChanÃ©ac, M. NoguÃ¨s, H. Kachkachi, et al. 2000. Surface Fe2O3 nanoparticles. Journal of Magnetism and Magnetic Materials 221: 63 79. doi: http://dx.doi.org /10.1016/S0304 8853(00)00369 3 . Tuutijarvi, T., J. Lu, M. Sillanpaa and G. Chen. 2009. As(V) adsorption on maghemite nanoparticles. Journal of Hazardous Materials 166: 1415 1420. doi:10.1016/j.jhazmat.2008.12.069. U.S.EPA. 2000. Technologies and Costs for Removal of Arsenic from Drinking Water. U.S.EPA. 2002. Arsenic Treatment Technologies for Soil, Waste, and Water. U.S. EPA/National Service Center for Environmental Publications (NSCEP). Uchimiya, M. and D. Bannon. 2013. Solubility of lead and copper in b iochar amended small arms range so ils: Influenc e of soil organic carbon a nd pH. J Agr Food Chem 61: 7679 7688. Uchimiya, M., S. Chang and K.T. Klasson. 2011. Screening biochars for heavy metal retention in soil: Role of oxygen functional groups. Journal of Hazardous Materials 190: 432 441. doi: http://dx.doi.org/10.1016/j.jhazmat.2011.03.063 . Uchimiya, M., I.M. Lima, K.T. Klasson, S.C. Chang, L.H. Wartelle and J.E. Rodgers. 2010. Immobilization of heavy metal ions (Cu II, Cd II, Ni II, and Pb II) by broiler litter derived biochars in water and soil. J Agr Food Chem 58: 5538 5544. doi:Doi 10.1021/Jf9044217. Uchimiya, M., L.H. Wartelle, K.T. Klasson, C.A. Fortier and I.M. Lima. 2011. Influence of p yrolysis temperature on biochar property and function as a heavy metal sorbent in Soil. J Agr Food Chem 59: 2501 2510. doi:Doi 10.1021/Jf104206c. Van Oostdam, J., A. Gilman, E. Dewailly, P. Usher, B. Wheatley, H. Kuhnlein, et al. 1999. Human health implica tions of environmental contaminants in Arctic Canada: a review. Science of The Total Environment 230: 1 82. doi: http://dx.doi.org/10.1016/S0048 9697(99)00036 4 . Vogel, K.P., J.J. Brejda, D.T. Walters and D.R. Buxton. 2002. Switchgrass biomass production in the Midwest USA: Harvest and nitrogen management. Agronomy Journal 94: 413 420. Wang, S., B. Gao, A. Zimmerman, Y. Li, L. Ma, W. Harris, et al. 2014. Removal of arsenic by magnetic biochar prepared from pinewood and natural hematite. Bioresources Technology doi: 10.1016/j.biortech.2014.10.104. Wang, S. and C. Mulligan. 2006. Occurrence of arsenic contamination in Canada: Sources, behavior and distribution. Science of the Total Environment 36 6: 701 721. doi:10.1016/j.scitotenv.2005.09.005.
110 Wang, Z., S. W. Lee, J.G. Catalano, J.S. Lezama Pacheco, J.R. Bargar, B.M. Tebo, et al. 2012. Adsorption of Uranium(VI) to manganese o xides: X ray absorption spectroscopy and surface complexation m odeling. E nvironmental Science & Technology 47: 850 858. doi:10.1021/es304454g. Weitzman, M., A. Aschengrau, D. Bellinger, R. Jones, J. Hamlin and A. Beiser. 1993. LEad contaminated soil abatement and urban children's blood lead levels. JAMA 269: 1647 1654. doi: 10.1001/jama.1993.03500130061033. Whitmore, T.J., M.A. Riedinger Whitmore, J.M. Smoak, K.V. Kolasa, E.A. Goddard and R. Bindler. 2008. Arsenic contamination of lake sediments in Florida: evidence of herbicide mobility from watershed soils. Journal of Paleo limnology 40: 869 884. doi:10.1007/s10933 008 9204 8. Winship, K. 1984. Toxicity of inorganic arsenic salts. Adverse Drug Reactions and Toxicological Reviews 3: 129 160. Xu, G., L.L. Wei, J.N. Sun, H.B. Shao and S.X. Chang. 2013. What is more important for enhancing nutrient bioavailability with biochar application into a sandy soil: Direct or indirect mechanism? Ecological Engineering 52: 119 124. doi: http://dx.doi.org/10.1016/j.ecoleng.2012.1 2.091 . Xue, Y., B. Gao, Y. Yao, M. Inyang, M. Zhang, A.R. Zimmerman, et al. 2012. Hydrogen peroxide modification enhances the ability of biochar (hydrochar) produced from hydrothermal carbonization of peanut hull to remove aqueous heavy metals: Batch and c olumn tests. Chem Eng J 200: 673 680. doi:10.1016/j.cej.2012.06.116. Yantasee, W., Y. Lin, G.E. Fryxell, K.L. Alford, B.J. Busche and C.D. Johnson. 2004. Selective Removal of Copper(II) from Aqueous Solutions Using Fine Grained Activated Carbon Functionali zed with Amine. Industrial & Engineering Chemistry Research 43: 2759 2764. doi:10.1021/ie030182g. Yao, Y., B. Gao, J.J. Chen, M. Zhang, M. Inyang, Y.C. Li, et al. 2013. Engineered carbon (biochar) prepared by direct pyrolysis of Mg accumulated tomato tissu es: Characterization and phosphate removal potential. Bioresource Technology 138: 8 13. doi:10.1016/j.biortech.2013.03.057. Yao, Y., B. Gao, J. Fang, M. Zhang, H. Chen, Y. Zhou, et al. 2014. Characterization and environmental applications of clay biochar c omposites. Chem Eng J 242: 136 143. doi:DOI 10.1016/j.cej.2013.12.062. Yao, Y., B. Gao, J. Fang, M. Zhang, H. Chen, Y. Zhou, et al. 2014. Characterization and environmental applications of clay biochar composites. Chemical Engineering Journal 242: 136 143. doi: http://dx.doi.org/10.1016/j.cej.2013.12.062 .
111 Yao, Y., B. Gao, M. Inyang, A.R. Zimmerman, X. Cao, P. Pullammanappallil, et al. 2011. Biochar derived from anaerobically digested sugar beet tai lings: Characterization and phosphate removal potential. Bioresource Technol 102: 6273 6278. Yao, Y., B. Gao, M. Inyang, A.R. Zimmerman, X. Cao, P. Pullammanappallil, et al. 2011a. Biochar derived from anaerobically digested sugar beet tailings: Characteri zation and phosphate removal potential. Bioresource Technology 102: 6273 6278. Yao, Y., B. Gao, M. Inyang, A.R. Zimmerman, X.D. Cao, P. Pullammanappallil, et al. 2011. Removal of phosphate from aqueous solution by biochar derived from anaerobically digeste d sugar beet tailings. J Hazard Mater 190: 501 507. doi:DOI 10.1016/j.jhazmat.2011.03.083. Yao, Y., B. Gao, M. Inyang, A.R. Zimmerman, X.D. Cao, P. Pullammanappallil, et al. 2011b. Removal of phosphate from aqueous solution by biochar derived from anaerobi cally digested sugar beet tailings. Journal of Hazard Materials 190: 501 507. Yao, Y., B. Gao, M. Zhang, M. Inyang and A.R. Zimmerman. 2012. Effect of biochar amendment on sorption and leaching of nitrate, ammonium, and phosphate in a sandy soil. Chemosphe re 89: 1467 1471. Ying, S.C., B.D. Kocar and S. Fendorf. 2012. Oxidation and competitive retention of arsenic between iron and manganese oxides. Geochimica et Cosmochimica Acta 96: 294 303. doi: h ttp://dx.doi.org/10.1016/j.gca.2012.07.013 . Zhang, A., R. Bian, G. Pan, L. Cui, Q. Hussain, L. Li, et al. 2012. Effects of biochar amendment on soil quality, crop yield and greenhouse gas emission in a Chinese rice paddy: A field study of 2 consecutive ric e growing cycles. Field Crops Research 127: 153 160. doi: http://dx.doi.org/10.1016/j.fcr.2011.11.020 . Zhang, A., L. Cui, G. Pan, L. Li, Q. Hussain, X. Zhang, et al. 2010. Effect of biochar amendme nt on yield and methane and nitrous oxide emissions from a rice paddy from Tai Lake plain, China. Agriculture, Ecosystems & Environment 139: 469 475. doi: http://dx.doi.org/10.1016/j.agee.2010.09. 003 . Zhang, C.X., G.A. Paterson and Q.S. Liu. 2012. A new mechanism for the magnetic enhancement of hematite during heating: the role of clay minerals. Studia Geophysica Et Geodaetica 56: 845 860. doi:10.1007/s11200 011 9018 4. Zhang, M. and B. Gao. 2013. Removal of arsenic, methylene blue, and phosphate by biochar/AlOOH nanocomposite. Chem Eng J 226: 286 292. doi:10.1016/j.cej.2013.04.077.
112 Zhang, M. and B. Gao. 2013. Removal of arsenic, methylene blue, and phosphate by biochar/AlOOH nanocomposite. Chemical Engineering Journal 226: 286 292. doi: http://dx.doi.org/10.1016/j.cej.2013.04.077 . Zhang, M., B. Gao, S. Varnoosfaderani, A. Hebard, Y. Yao and M. Inyang. 2013. Preparation and characterization o f a novel magnetic biochar for arsenic removal. Bioresource Technol 130: 457 462. doi:DOI 10.1016/j.biortech.2012.11.132. Zhang, M., B. Gao, S. Varnoosfaderani, A. Hebard, Y. Yao and M. Inyang. 2013. Preparation and characterization of a novel magnetic bio char for arsenic removal. Bioresource Technology 130: 457 462. doi: http://dx.doi.org/10.1016/j.biortech.2012.11.132 . Zhang, M.Q., C.A. Powell, L.J. Zhou, Z.L. He, E. Stover and Y.P. Duan. 201 1. Chemical Compounds Effective Against the Citrus Huanglongbing Bacterium 'Candidatus Liberibacter asiaticus' In Planta. Phytopathology 101: 1097 1103. doi:10.1094/phyto 09 10 0262. Zhang, S., H. Niu, Y. Cai, X. Zhao and Y. Shi. 2010. Arsenite and arsenat e adsorption on coprecipitated bimetal oxide magnetic nanomaterials: MnFe2O4 and CoFe2O4. Chemical Engineering Journal 158: 599 607. doi: http://dx.doi.org/10.1016/j.cej.2010.02.013 . Zhang, T. and D.D. Sun. 2013. Removal of arsenic from water using multifunctional micro /nano structured MnO 2 spheres and microfiltration. Chemical Engineering Journal 225: 271 279. doi: http://dx.doi.org/10.1016/j.cej.2013.04.001 . Zhong, L.S., J.S. Hu, H.P. Liang, A.M. Cao, W.G. Song and L.J. Wan. 2006. Self a ssembled 3D flowerlike iron oxide nanostructures and their application in water treatm ent. Advanced Materials 18: 2426 2431. doi:10.1002/adma.200600504. Zhou, Y., B. Gao, A.R. Zimmerman, H. Chen, M. Zhang and X.D. Cao. 2014. Biochar supported zerovalent iron for removal of various contaminants from aqueous solutions. Bioresource Technol 152: 538 542. doi:DOI 10.1016/j.biortech.2 013.11.021. Zhou, Y.M., B. Gao, A.R. Zimmerman, J. Fang, Y.N. Sun and X.D. Cao. 2013. Sorption of heavy metals on chitosan modified biochars and its biological effects. Chem Eng J 231: 512 518. doi:DOI 10.1016/j.cej.2013.07.036. Zimmerman, A.R., B. Gao and M.Y. Ahn. 2011. Positive and negative carbon mineralization priming effects among a variety of biochar amended soils. Soil Biol Biochem 43: 1169 1179. doi:DOI 10.1016/j.soilbio.2011.02.005.
113 BIOGRAPHI CAL SKETCH from Shandong Agricultural University, Taian, China. He was an exchange student at University of Florida/USDA ARS, Prosser, WA (2008 2009). He received his Ph.D degree in December 201 4 at Soil and Water Science Department, University of Florida, Gainesville, FL.