Characterization of Biochar and Coal Combustion Residues and Mechanisms of Biochar-Induced Immobilization and Transform...

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Characterization of Biochar and Coal Combustion Residues and Mechanisms of Biochar-Induced Immobilization and Transformation of Soluble Heavy Metals
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Dong, Xiaoling
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Degree:
Doctorate ( Ph.D.)
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University of Florida
Degree Disciplines:
Soil and Water Science
Committee Chair:
Ma, Lena Q
Committee Co-Chair:
Li, Yuncong
Committee Members:
Gao, Bin
Harris, Willie G, Jr
Rathinasabapathi, Balasubramani

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Subjects / Keywords:
arsenic -- biochar -- chromium -- coal -- combustion -- mercury -- residues
Soil and Water Science -- Dissertations, Academic -- UF
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Soil and Water Science thesis, Ph.D.
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theses   ( marcgt )
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Abstract:
Heavy metals Cr(VI), Hg(II) and As(III) are highly toxic to humans and animals. Biochar has high affinity for heavy metals and can be used to reduce their bioavailability. The objectives of this study were to investigate the effectiveness and mechanisms of biochar in removing Cr(VI), Hg(II) and As(III) from aqueous solution and to characterize Florida coal combustion residues (CCR). SBT (sugar beet tailing biochar) effectively removed Cr(VI) with maximum sorption capacity of 123 mg/g. Desorption and X-ray photoelectron spectroscopy (XPS) analysis showed Cr was primarily bound to SBT as Cr(III), indicating reduction mechanisms governed Cr(VI) immobilization. BP (Brazilian pepper) biochars produced at 300, 450, and 600°C had an Hg sorption capacity of 24.2, 18.8 and 15.1 mg/g based on Langmuir isotherm. XPS analysis showed 23-31% and 77-69% of sorbed Hg was associated with carboxylic and phenolic hydroxyl groups in BP300 and 450 whereas 91% of sorbed Hg was associated with graphite like domain on aromatic structure in BP600, which were consistent with flow calorimetry data. Biochars release dissolved organic matter (DOM), which is known to serve as both electron donor and acceptor during redox reactions. Cr(VI) and As(III) transformations were enhanced with increasing DOM concentrations; however, as pH increased, Cr(VI) reduction decreased while As(III) oxidation increased. Electron spin resonance studies suggested semiquinoic radicals were responsible for As(III) oxidation. During freezing processes of aqueous solutions, conversion of Cr(VI) and As(III) increased with elevated concentrations of solutes in the grain boundary. Though DOM enhanced both Cr(VI) reduction and As(III) oxidation, Cr(VI) reduction coupled with As(III) oxidation occurred in absence of DOM. Total and SPLP (Synthetic precipitation leaching procedure) concentrations of heavy metals in 27 Florida CCR samples were examined. Total concentrations of V, Mn, Pb, Ni, Cr, Cu, As, and Se and SPLP concentrations of Al, Fe, V, Ni, As, Mo and Sb concentrations from most samples were higher than their maximum contaminant levels. Biochar has potential for use in immobilizing heavy metals. Due to potential leaching of heavy metals from CCR, caution needs to be excised in their utilization and disposal.
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by Xiaoling Dong.
Thesis:
Thesis (Ph.D.)--University of Florida, 2013.
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Adviser: Ma, Lena Q.
Local:
Co-adviser: Li, Yuncong.
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RESTRICTED TO UF STUDENTS, STAFF, FACULTY, AND ON-CAMPUS USE UNTIL 2014-08-31

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1 CHARACTERIZATION OF BIOCHAR AND COAL COMBUSTION RESIDUES AND MECHANISMS OF BIOCHAR INDUCED IMMOBILIZATION AND TRANSFORMATION OF SOLUBLE HEAVY METALS By XIAOLING DONG A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2013

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2 2013 Xiaoling Dong

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3 To my family

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4 ACKNOWLEDGMENTS I would like to ex press my sincere gratitude and appreciation to Dr. Lena Q Ma, my advisor, for her guidance, trust and patience and for giving me the opportunity to learn and develop my interests as her graduate student I also would like to thank my committee members, Dr. Yuncong Li, Dr. Bin Gao, Dr. Willie Harris and Dr. Bala Rathinasabapathi for their support and assistance. I want to thank Dr. Gao and Dr. Rathinasabapathi for providing me biochar samples. I would like to thank Dr. Harris for assisting me with the SEM and XRD analysis. Additionally, I want to acknowledge Dr. Li for helping me to get coal ash samples. Also I want to thank Dr. Dean Rhue for his help in flow calorimetry experiment and Dr. Tim Townsend for helping me get the ash samples from Florida coal power plants. Many thanks go to my lab mates: Shiny Mathews, Jay Lessl, Hao Chen, Piyasa Gosh, Rujira Tisarum, Yingjia Zhu and Ky Gress for their supports in my research and personal life, especially for Ky Gress for helping me with the ash project. Last but not least I would like to recognize my parents for providing me emotional support and giving me all the opportunities and education necessary to pursuit many of my dreams and professional goals.

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5 TABLE OF CONTENTS page ACKNOWLEDGMENTS ................................ ................................ ................................ .. 4 LIST O F TABLES ................................ ................................ ................................ ............ 8 LIST OF FIGURES ................................ ................................ ................................ ........ 10 LIST OF ABBREVIATIONS ................................ ................................ ........................... 13 ABSTRACT ................................ ................................ ................................ ................... 14 CHAPTER 1 INTRODUCTION ................................ ................................ ................................ .... 16 2 LITERATURE REVIEW ................................ ................................ .......................... 18 Chromium, Me rcury and Arsenic in the Environment ................................ .............. 18 Chromium ................................ ................................ ................................ ......... 18 Mercury ................................ ................................ ................................ ............ 21 A rsenic ................................ ................................ ................................ ............. 23 Biochar ................................ ................................ ................................ .................... 25 Physical and Chemical Properties ................................ ................................ .... 26 Sugar B eet Tailing and Brazilian Pepper Biochar ................................ ............. 28 Coal Combustion Residues ................................ ................................ .................... 29 Production of Coal Combustion Residues ................................ ........................ 29 Characteristics of Coal Combustion Residues ................................ ................. 31 Potential Risk of Coal Combustion Residues to the Environment .................... 34 Coal Combustion Residues Regulatory Rules ................................ .................. 34 Research Objectives ................................ ................................ ............................... 35 3 CHARACTERISTICS AND MECH ANISMS OF HEXAVALENT CHROMIUM REMOVAL BY BIOCHAR FROM SUGAR BEET TAILING ................................ ..... 36 Introduction ................................ ................................ ................................ ............. 36 Materials and Methods ................................ ................................ ............................ 37 Biochar Preparation and Characterization ................................ ........................ 37 Cr(VI) Sorption and Desorption Experiments ................................ ................... 39 Fourier Transform Infrared (FTIR) and X ray Photoelectron Spectroscopy (XPS) Analysis ................................ ................................ .............................. 40 Cr(VI) Analysis ................................ ................................ ................................ 40 Resul ts and Discussion ................................ ................................ ........................... 41 Characteristics of SBT Biochar Before and After Reaction with Cr(VI) ............. 41 Characteristics of Cr(VI) Sorption by SBT Biochar ................................ ........... 44 Reduction of Cr(VI) to Cr(III) by SBT Biochar ................................ ................... 48

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6 Desorption of Cr from SBT Biochar ................................ ................................ .. 51 Mechanisms of Cr(VI) Removal by SBT Biochar ................................ .............. 53 Research Findings ................................ ................................ ................................ .. 54 4 MECHANISTIC INVEST IGATION OF MERCURY SORPTION BY BRAZILIAN PEPPER BIOCHARS OF DIFFERENT PYROLYTIC TEMPERATURES BASED ON X RAY PHOTOELECTRON SPECTROSCOPY AND FLOW CALORIMETRY ................................ ................................ ................................ ...... 55 Introduction ................................ ................................ ................................ ............. 55 Materials and Methods ................................ ................................ ............................ 57 Biochar Preparation and Characterization ................................ ........................ 57 Hg Sorption Isotherm by Biochars ................................ ................................ .... 59 Fourier Transform Infrared (FTIR) and X ray Photoelectron Spectroscopy (XPS) Analysis ................................ ................................ .............................. 59 Flow Calorimetr y Experiments ................................ ................................ .......... 60 Mercury and Statistical Analysis ................................ ................................ ....... 61 Results and Discussion ................................ ................................ ........................... 61 Biochar Produced at Low Temperature Contained More Functional Groups ... 61 Hg Speciation Impacted Its Sorption by Biochars at Different pHs ................... 67 Carboxylic and Phenolic Hydroxyl Groups Were Responsible for Hg Sorption ................................ ................................ ................................ ......... 69 Hg Sorption onto Biochars Was Probably Via Complexation with Functional Groups ................................ ................................ ................................ .......... 74 Research Findings ................................ ................................ ................................ .. 82 5 ENHANCED CR(VI) REDUCTION AND AS(III) OXIDATION IN ICE: IMPORTANT ROLE OF DISSOLVED ORGANIC MATTER FROM BIOCHAR ...... 84 Introduction ................................ ................................ ................................ ............. 84 Materials and Methods ................................ ................................ ............................ 86 Preparation and Char acterization of DOM from Biochars ................................ 86 Cr(VI) Reduction and As(III) Oxidation by DOM ................................ ............... 87 Simultaneous Cr(VI) Reduction and As(III) Oxidation ................................ ...... 88 Fourier Transform Infrared (FTIR) Analysis of DOM Loaded with Cr and As ... 88 Electron Spin Resonance (ESR) Analysis o f DOM ................................ ........... 89 Chemical and Statistical Analysis ................................ ................................ ..... 89 Results and Discussion ................................ ................................ ........................... 9 0 Similar Structure and Functional Groups Existed in DOM from Biochar and Soil ................................ ................................ ................................ ................ 90 Cr(VI) Reduction Was Enhanced by DOM and in the Ice Phase ...................... 93 As(III) Oxidation Was Enhanced by DOM and In Ice Phase ............................. 97 Simultaneous Cr(VI) Reduction and As(III) Oxidation With and Without DOM ................................ ................................ ................................ ............ 102 Research Findings ................................ ................................ ................................ 104 6 CHARACTERIZATION OF FLORIDA COAL COMBUSTION RESIDUES ............ 107

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7 Introduction ................................ ................................ ................................ ........... 107 Material s and Method s ................................ ................................ .......................... 108 CCR Samples ................................ ................................ ................................ 108 Sampling Methods ................................ ................................ .......................... 110 Analytical Methods and Quality Assurance ................................ .................... 112 Results and Discussion ................................ ................................ ......................... 113 pH ................................ ................................ ................................ ................... 113 Total Concentrations ................................ ................................ ...................... 114 SPLP Concentrations ................................ ................................ ..................... 125 Res earch Findings ................................ ................................ ................................ 138 7 CONCLUSIONS ................................ ................................ ................................ ... 140 LIST OF REFERENCES ................................ ................................ ............................. 145 BIOG RAPHICAL SKETCH ................................ ................................ .......................... 159

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8 LIST OF TABLES Table page 2 1 Concentrations of various elements in fly ash samples (presen ted as ranges of means: mg/kg) ................................ ................................ ............................... 31 2 2 Concentrations of various elements in bottom ash samples (presented as ranges of means: mg/kg). ................................ ................................ ................... 32 2 3 Concentrations of vario us elements in FGD samples (presented as ranges of means: mg/kg). ................................ ................................ ................................ ... 33 3 1 Elemental concentrations of SBT biochar (mg/g) and in solution (mg/L) after SBT biochar reaction with 100 mg/L Cr(VI) a nd DI water at pH 2 for 24h. ......... 41 4 1 C, N, H composition and atomic ratios of biochar samples derived from Brazilian pepper (BP) under different pyrolysis temperatures (300, 450 and 600 C). ................................ ................................ ................................ ............... 62 4 2 Properties of biochar samples derived from Brazilian pepper (BP) under different pyrolysis temperatures (300, 450 and 600 C). ................................ ..... 63 4 3 Parameters of Langmuir model for Hg sorption onto biochars derived from Brazilian pepper (BP) under different pyrolysis temperatures (300, 450 and 600 C). ................................ ................................ ................................ ............... 67 4 4 Surface composition f or Hg loaded BP biochars based on XPS analysis ........... 70 4 5 Heats of sorption and quantity of Hg sorbed after reaction with 50 mg/L Hg(NO 3 ) 2 solution at pH 6.0 during flow calorimetry and batch sorp tion experiments. ................................ ................................ ................................ ....... 78 5 1 Characteristics of DOM SBT DOM BP and DOM S ................................ ................. 91 5 2 Reduction of Cr(VI) to Cr(III) (mg/L) by DOM SBT DOM B P and DOM S at various pH in both aqueous and ice phases (10 mg/L Cr(VI) and 10 mgC/L DOM). ................................ ................................ ................................ ................. 95 5 3 ESR data of DOM samples. ................................ ................................ .............. 100 5 4 Oxidation of As(III) by DOM SBT DOM BP and DOM S at various pH in both aqueous and ice phases (10 mgL As(III) and 10 mgC/L DOM). ....................... 101 5 5 Simultaneous reduction of Cr(VI) and oxidation of As(I II) at various pH in the ice phase (g/L) (As(III)= 1130 g/L, Cr(VI) = 52 g/L at pH 2.0; As(III)= 113 g/L, Cr(VI) = 520 g/L at pH 6.0 and 10.0; DOM= 10 mgC/L). ....................... 103

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9 6 1 Summary of facility configuration ................................ ................................ ...... 109 6 2 CCR sampling methods for target CCR samples at each facility (Townsend, 2012) ................................ ................................ ................................ ................ 111 6 3 pH of CCR samples ................................ ................................ .......................... 117 6 4 Total concentrations of Ca, K, Na, Mg, Al, Fe, and Zn in CCR samples (mg/kg) ................................ ................................ ................................ ............. 118 6 5 Total concentrations of V, Mn, Pb, Ni Cr, Cu, As and Se in CCR samples (mg/kg) ................................ ................................ ................................ ............. 120 6 6 Total concentrations of Be, Co, Mo, Sb, Tl, Hg and Cd in CCR samples (mg/kg) ................................ ................................ ................................ ............. 123 6 7 SPLP concentrations of elements Ca, K, Na, Mg, Al, Fe, and Zn (g/L) .......... 129 6 8 SPLP concentrations of elements V, Mn, Pb, Ni, Cr, Cu, As and Se (g/L) ..... 131 6 9 SPLP concentrations of elements Be, Co, Mo, Sb, Tl, Hg and Cd (g/L) ......... 134 6 10 Solution pH after SPLP test ................................ ................................ .............. 136 6 11 Florida soil cleanup target level (mg/kg) and maximum contaminant level in drinking water (g/L) ................................ ................................ ......................... 137

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10 LIST OF FIGURES Figure page 2 1 Cr(III) speciation as a function of pH (ionic stre ngth=0.01 M and Cr(III)=1 mg/L) ................................ ................................ ................................ ................. 19 2 2 Calculated abundance of chromium(VI) species in aqueous solution at total Cr(VI) concentrati on 10 M and at pH 1 14 ................................ ....................... 20 2 3 Hg(II) aqueous diagram with initial concentration of 5 M as a function of pH. .. 22 2 4 pe pH diag ram for aqueous As species in the system As O 2 H 2 O at 25 C and 1 bar pressure. ................................ ................................ ................................ ... 24 3 1 Scanning electron micrographs and EDS spectra of SBT biochar before (a,c) and after (b,d) reaction with 10 0 mg/L Cr(VI) for 24 h at pH 2.0. ........................ 42 3 2 Fourier transform infrared spectra of SBT biochar before and after reaction with 100 mg/L Cr(VI) for 24 h at pH 2.0. ................................ ............................. 43 3 3 Effect of reaction time ( a ), pH (b ) and biochar mass ( c ) on Cr removal by SBT biochar after reacting with 100 mg/L Cr(VI). ................................ ............... 46 3 4 Langmuir and Freundlich p lot for Cr(VI) sorption on SBT biochar. ..................... 47 3 5 XPS spectra of the SBT biochar after reacting 2 g/L SBT biochar with 100 mg/L Cr(VI) for 24 h at pH 2. ................................ ................................ .............. 50 3 6 Concentrations of Cr(VI) and Cr(III) in solution during Cr desorption from SBT biochar by 0.1 M NaOH (a) and 0.1 M H 2 SO 4 (b). ................................ .............. 52 4 1 Scanning electron micrographs before (a) and after (b) reaction with 50 mg/L Hg for 24h at pH 6.0. .................. 64 4 2 Fourier transform infrared spectra of BP biochars (a, b, c) before and after re action with 50 mg/L Hg for 24 h at pH 6.0. ................................ ...................... 65 4 3 XPS spectra of survey scan of BP biochars. ................................ ...................... 66 4 4 Influence of pH on Hg remova l by BP biochars after reacting with 20 mg/LHg for 24 h. ................................ ................................ ................................ .............. 68 4 5 XPS spectra of survey scan of BP biochars after reaction with 50 mg/L Hg for 24 h at pH 6.0. ................................ ................................ ................................ .... 72 4 6 XPS spectra of high resolution scan of Hg4f for BP biochars after reaction with 50 mg/L Hg for 24 h at pH 6.0. ................................ ................................ .... 73

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11 4 7 Heat changes for K/Ca sorption at pH 6 before and after Hg sorption onto Brazilian pepper biochars produced at 300 (a and d), 450 (b and e) and ................................ ................................ ................................ ... 76 4 8 Heat changes of Hg sorption at pH 6 by BP biochars produced at 300 (a), ................................ ................................ ........................ 77 4 9 Heat changes of Hg sorption at pH 6 by BP 450 biochar during the first cycle (a) and second Hg cycle (b) ................................ ................................ ............... 80 4 10 Mechanisms of Hg removal by BP biochars produced at different temperatures. ................................ ................................ ................................ ..... 82 5 1 Fourier transformation infrared spectra of control, Cr loaded, and As loaded DOM SBT (a), control, Cr loaded, and As loaded DOM BP (b) and control, Cr loaded, and As loaded DOM S (c). ................................ ................................ ........ 92 5 2 Cr(III) appearance and DOM disappearance in DOM SBT (a,d), DOM BP (b,e) and DOM S (c,f) tre atment in both aqueous and ice phase s at pH 2.0 after 24 h reaction (10 mgC/L DOM and 10 mg/L Cr(VI)). ................................ ............... 94 5 3 Effect of DOM concentration on reduction of 10 mg/L Cr(VI) (a) and oxidation of 10 mg/L As(III) (b) by DOM SBT after 24 h. ................................ ....................... 97 5 4 As(III) oxidation by DOM SBT (a), DOM BP (b) and DOM S (c) in both aqueous and ice phase s at pH 10.0 after 24 h reaction (10 mgC/L DOM and 10 mg/L A s(III)). ................................ ................................ ................................ ................ 99 5 5 Mechani s ms of Cr(VI) reduction and As(III) oxidation by DOM in the ice phase. ................................ ................................ ................................ ............... 105 6 1 Total concentration of Al in CCR samples. ................................ ....................... 119 6 2 Total concentration of Fe in CCR samples. ................................ ...................... 119 6 3 Total concentration of V in CCR samples. ................................ ........................ 121 6 4 Total concentration of Pb in CCR samples. ................................ ...................... 121 6 5 Total concentration of Ni in CCR samples. ................................ ....................... 122 6 6 Total concentration of As in CCR samples. ................................ ...................... 122 6 7 Total concentration of Mo in CCR samples. ................................ ..................... 124 6 8 Total concentration of Sb in CCR samples. ................................ ...................... 124 6 9 SPLP concentration of Al in CCR samples. ................................ ...................... 130

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12 6 10 SPLP concentration of Fe in CCR samples. ................................ ..................... 130 6 11 SPLP concentration of V in CCR samples. ................................ ....................... 132 6 12 SPLP concentration of Pb in CCR samples. ................................ ..................... 132 6 13 SPLP concentration of Ni in CCR samples. ................................ ...................... 133 6 14 SPLP concentration of As in CCR samples. ................................ ..................... 133 6 15 SPLP concentration of Mo in CCR samples. ................................ .................... 135 6 16 SPLP concentration of Sb in CCR samples. ................................ ..................... 135

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13 LIST OF ABBREVIATIONS BP Brazili an pepper biochar CEC Cation exchange capacity DOC Dissolved organic carbon DOM Dissolved organic matter EDX Energy dispersive X ray spectroscopy ESR Electron spin resonance spectroscopy FGD Flue gas desulfurization FTIR F ourier transform infrared spect roscopy HG AFS Hydride generation atomic fluorescence ICP AES Inductively coupled plasma atomic emission spectroscopy ICP MS Inductively coupled plasma mass spectrometry SBT Sugar beet tailing biochar SEM EDX Scanning electron microscope SPLP Synthetic precipitation leaching procedure XPS X ray photoelectron spectroscopy ZPC Zero point of charge

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14 Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degre e of Doctor of Philosophy CHARACTERIZATION OF BIOCHAR AND COAL COMBUSTION RESIDUES AND MECHANISMS OF BIOCHAR INDUCED IMMOBILIZATION AND TRANSFORMATION OF SOLUBLE HEAVY METALS By Xiaoling Dong August 2013 Chair: Lena Q. Ma Co C hair: Yuncong Li Major: So il and Water Science H eavy metals Cr(VI), Hg(II) and As(III) are highly toxic to humans and animals Biochar has high affinity for heavy metals and can be used to reduce their bioavailability T he objective s of this study w ere to investigate the effective ness and mechanisms of biochar in removing Cr(VI), Hg(II) and As(III) from aqueous solution and to characterize Florida coal combustion residues (CCR). SBT (sugar beet tailing biochar) effectively removed Cr(VI) with maximum sorption capacity of 123 mg/g. Desorption and X ray photoelectron spectroscopy (XPS) analysis showed Cr was primarily bound to SBT as Cr(III), indicating r eduction mechanisms governed Cr(VI) immobilization. BP (Brazilian pepper) biochars produced had an Hg sorptio n capacit y of 24.2, 18.8 and 15.1 mg/g based on Langmuir isotherm. XPS analysis show ed 23 31% and 77 69% of sorbed Hg was associated with carboxylic and phenolic hydroxyl groups in BP300 and 450 whereas 91% of sorbed Hg was associated with graphite like do mai n on aromatic structure in BP60 0 ; these results were consist ent with flow calorimetry data. Biochar s release d issolved organic matter (DOM), which is known to serve as both

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15 electron donor and acceptor during redox reaction s. Cr(VI) and As(III) transform ation s were enhanced with increas ing DOM concentration s ; however, as pH increased, Cr ( VI ) reduction decreased while As ( III ) oxidation increased Electron spin resonance stud ies suggested semiquinoic radicals were responsible for As ( III ) oxidation. During f reezing processes of aqueous solutions, conversion of Cr ( VI ) and As ( III ) increased with elevated concentrations of solutes in the grain boundary. Though DOM enhanced both Cr ( VI ) reduction and As ( III ) oxidation, Cr ( VI ) reduction coupled with As ( III ) oxidati on occurred in absence of DOM. Total and SPLP ( Synthetic precipitation leaching procedure ) concentrations of heavy metals in 27 Florida CCR samples were examined. Total concentrations of V, Mn, Pb, Ni, Cr, Cu, As, and Se and SPLP concentrations of Al, Fe, V, Ni, As, Mo and Sb for most samples were higher than t heir maximum contaminant levels. B iochar has potent ial for use in immobilizing heavy metals Due to potential leaching of heavy metals from CCR caution needs to be ex er cised in the ir utilization and disposal.

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16 C HAPTER 1 INTRODUCTION This research covers two aspects: biochar and coal combustion residues. Biochar is a carbon rich b yproduct derived from the pyrolysis of organic ma terials such as plant, plant residues, manures, bones and so on ( Lehmann, 2007 ) I n rec ent years biochar has gained important due to the global issues of climate change and recognition for a more sustainable soil management approach ( Namgay et al. 2010 ) Biochar has also been investigated for its efficiency as an adsorbent to remove toxic metals from water ( Mohan and Pittman Jr, 2007 ) Coal combustion residues are the materials remained after coal burning for electricity generation Coal is formed from the decomposition of organic materials th at have been subjected to geological heat and pressure over million years, the conversion is called carbonization ( Finkelman, 1999 ) Coal contains detectable concentrations of virtually every element in the p eriodic table. Many of these elements, including many toxic metals, are enriched in coal ( Finkelman, 1999 ) Through the coal burning process, substantial amount of trace metals are enriched in the coal combus tion residues. Coal combustion residues have been widely used in soil and mine waste treatments, admixtures in cement and concrete, making bricks and other ceramic products, fill materials in civil engineering projects, and extraction of valuable materials ( Querol et al. 2002 ) EPA has reported at least 137 cases of surface water or groundwater contaminat ion from coal combustion residues in at least 34 states Hence, concerns need to be expressed over coal combustion residues. Although both biochar and coal combustion residues are originated from organic materials, biochar is widely used to solve environme ntal problems, whereas coal

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17 combustion residues pose risks to human health and ecological systems. The major objectives in this research were (1) to study the effectiveness and mechanisms of biochar in immobilizing and transforming Cr, Hg and As from water ; (2) to characterize Florida coal combustion residues, especially their total trace metal contents and their leachabilities.

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18 CHAPTER 2 LITERATURE REVIEW C hromium, Mercury and Arsenic i n the Environment Chromium Chromium is released into the environment by various industrial operations such as electro plating, chromate manufacturing, leather tanning and wood preservation ( Papp, 2001 ) Though chromium exists in many Cr(III) and Cr(VI) are of major environmental significance because of their stability in the environment ( Emsely, 1989 ) In natural waters, chromium exists as Cr(VI) and Cr(III). Cr(III) species are much less soluble and relatively stable. However, Cr(VI) species, including chromate (CrO 4 and HCrO 4 ) and dichromate (Cr 2 O 7 ) are highly soluble and m obile in aqueous solutions. The latter is of significant environment concern because of its carcinogenic, mutagenic, and teratogenic behavior in biological systems ( Fendorf et al. 2000 ; Costa, 2003 ) The Cr (III) species in the environment depends on pH. In the absence of complexing agents, Cr(III) exists as hexa aquachromium and i ts hydrolysis products (Figure 2 1 ) ( Kumral, 2007 ) Cr(H 2 O) 6 3+ is a moderately strong acid (pK~4) and its deprotonated forms are present as CrOH 2+ Cr(OH) 2 + and Cr(OH) 3 which dominat e in pH 4 10. Trihydroxochro mium is sparing ly soluble in pH range of 5.5 12, and overlaps considerably with the pH range of natural waters ( Ball and Nordstrom, 1998 ) Dominant forms of Cr(III) are Cr(OH) 2 + and Cr(OH) 3 in the environment with amphoteric behavior. At higher pH, it is transformed into soluble tetrahydroxo complex, Cr( OH) 4 (pK=15.4 or 18.3). At more concentrated Cr(III) solutions (>10 6 M) the polynuclear hydrolytic

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19 products, Cr 2 (OH) 2 4+ Cr 3 (OH) 4 5+ and Cr 4 (OH) 6 6+ are aslo expected ( 2000 ) Figure 2 1 Cr(III) speciation as a function of pH (ionic streng th = 0.01 M and Cr(III )= 1 mg/L) (Adapted from Kumral, Elif. 2007. Master Thesis (Page 6, Figure 1 3). Chemistry Department, ) Cr(VI) forms several species, the relative proportions of which depend on both pH and total chromium concentration. Its dependence on pH is illustrated in Figure 2 2 ( and Stasicka, 2000 ) H 2 CrO 4 is a strong acid ( Sperling et al. 1992 ) and at pH>1 its deprotonated forms are prevailing: above pH 7 only Cr O 4 2 ions exist; at 1
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20 strong acidic medium as described in Equation 2 1 In more basic solution the reduction of CrO 4 2 generates OH against a gradient as describe d by Equation 2 2 The reduction potential of Cr(V) to Cr(III) is 0.13 in basic medium ( Kumral, 2007 ) HCrO 4 + 7H + +3e = Cr 3+ +4H 2 O (2 1 ) CrO 4 2 + 4H 2 O +3e = Cr(OH) 3 + 5OH ( 2 2) Figur e 2 2 Calculated abundance of chromium(VI) species in aqueous solution at total Cr(VI) concentration 10 M and at pH 1 14 (The dashed, vertical lines represent the pH range of natural waters) (Adapted from and Stasicka 2000)

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21 C hromium pollution in the environmental occurs naturally from erosion and weathering of chromium bearing minerals. The wide use of chromium in leather tanning, ferrochrome, pigments, electroplating and photography has led to its increase in the environment The Cr(VI) concentration in wastewater produced by industries are at 0.1 to 200 mg/L, but much higher concentration was also observed (3950 mg/L in tannery plant wastewater) ( Owlad et a l. 2009 ) Usually, Al(III) exists together with Cr(VI) in wastewaters from acidic mining drainage ( Aksu et al. 2002 ) w hich may influence Cr(VI) adsorption by bioso rbents. However, the limit for ch r omium in water and discharge into inland surface water recommended by Environmental Protection Agency are 0.05 mg/L and 0.1 mg/L, respectively ( EPA, 1990 ) Mercury Mercury is one of the most toxi c trace elements ( Gosar et al. 1997 ) It exists in three different oxidation states: 0, +1, and +2. However, Hg + is not stable, is easily oxidized to Hg 2+ Hg 0 (with high vapor pressure and relatively low solubility in water) and Hg 2+ (with high affinity for ligands such as chlorine, sulphur, hydroxyl ions and dissolved organic carbon) are the major forms in the environment In addition, it also exists in organ omercurial forms such as methylmercury (CH 3 Hg + ), which shows a great ability of bioaccumulation, and is recognized as the most toxic Hg form to humans ( Gosar et al. 1997 ) In aquatic systems, mercury exists in elemental, inorganic, and organic forms. Hg 0 is the only metal which exists in liquid form at room temperature. It has high volatility and relatively low water solubility ( Lindgvist and Rodlhe, 1985 ) Hg 2+ is common in the environment ( Loux, 1998 ) which is why it has been the subject of many stu dies (Figure 2 3) ( Park, 2011 ) Speciation analyses show that the hydrolysis

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22 reactions are dominant for aqueous Hg(II) in the absence of strong complexing ligands ( Kim et al. 2004a ) At low pH, the hexaqua ion Hg(H 2 O) 6 2+ is coordinated octahedrally by water molecules. HgOH + and Hg(OH) 2 aqueous specie s are formed at higher pH. Organic mercury may be divided into two categories: (1) c ovalently bonded organomercurials, such as m ethylHg and dimethyl mercury (the latter is less important than MethylHg in Hg transport and transformation), and (2) mercuric c omplexes with organic matter, such as humic substances ( Gill and Bruland, 1990 ) It is believed that suspended organic matter pla ys an important role in d etermin ing whether Hg is dissolved or remains as partic ulates ( Meili, 1997 ) Figure 2 3. Hg(II) aqueous diagram with initial concentration of 5 M as a function of pH (Adapted from Park, Changmin. 2011. Dissertation (Page 21, Figure2 3). The University of Texas at Austin ) T he major sources of mercury pollution in the aquatic environment are from various industries such as chloralkali, paint, pulp and paper, oil refining, electrical, rubber processing and fertilizer ( Namasivayam and Kadirvelu, 1999 ) The typical

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23 concentration of Hg(II) in chloralkali industry wastewater was 2 30 mg/L ( Manohar et al. 2002 ) The wastewater obtained from ore mining contain ed mercury at levels ranging from 70 to 90 m g/L ( Monteagudo and Ortiz, 2000 ) In the wastewater, not only mercury exists, but other hea vy metal like Pb coexists with mercury which may affect mercury sorption ( Manohar et al. 2002 ) The World Health Organization (WH O) recommends a maximum human Hg uptake of 0.3 concentration of 1 ( Wase and Forster, 1997 ) The USEPA permitted discharge limit of wastewater for total Hg is 10 water is 2 ( EPA, 2001 ) Japan Ministry of the Environment establishes even more stringent limits at 5 an d 0.5 ( Takahashi et al. 2001 ) respectively. It is therefore of great importance to remediate Hg contaminated water. Arsenic Arsenic is a carcinogenic metalloid of major environmental concern. It is the 20 th th in seawater and 12 th in human body. The elevated As in the environment is mainly attributed to anthropogenic sources, including industrial products and wastes, agri cultural pesticides and wood preservatives ( Ong et al. 1997 ; Wilkie and Hering, 1998 ) Arsenic occur s in the environment in several oxidation stat es ( 3, 0, +3 and +5) ( Smedley and Kinniburgh, 2002 ) But in natural waters, t he predominant forms of As are inorganic As(V) and As(III). As(III) specie is more toxic and mobile than As(V) ( Manning and Goldberg, 1997 ) Therefore, Cr(VI) reduction and As(III) oxidation are desirable to reduce their toxicity Organic As such as monomethylarsonic acid (MMA), demethylarsinic acid (DMA), and gaseous derivatives of arsine may be produced by

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24 biological activity, mostly in surface waters, but are rarely quantitatively important in the environment ( Smedley and Kinniburgh, 2002 ) The species of As in water depends on the pH and redox potential as indicated in the p e /pH diagram (Figure 2 4 ) ( Smedley and Kinniburgh, 2002 ) As(V) is the dominant specie under aerobic conditions. At pH less than 6.9, H 2 AsO 4 is dominant, while at higher pH, HAsO 4 2 becomes dominant. H 3 AsO 4 0 and AsO 4 3 may be present in extremely acidic and alkaline conditions respectively Under reducing conditions at pH less than 9.2, neutral As(III) specie s H 2 AsO 3 0 will predominate ( Korte and Fernando, 1991 ; Masscheleyn et al. 1991 ) Figure 2 4 pe pH diagram for aqueous As species in the system As O 2 H 2 O at 25 C and 1 bar pressure ( Adapted from Smedley and Kinniburgh 2002 ).

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25 Arsenic is mobilized by natural weathering reactions, biological activity, geochemical reactions, volcanic emissions and o ther anthropogenic activities. Soil erosion and leaching contribute to 61 2 00 and 238 0 00 ton /year of arsenic, respectively, in dissolved and suspended forms in the oceans ( Mackenzie et al. 1979 ) Some As problems in the environment result from mobilization under natural conditions. However, mining activities, combustion of foss il fuels, use of arsenic pesticides, crop desiccants and use of arsenic additives to livestock feed add additional arsenic to the environment Arsenic pollution has been reported worldwide in cluding USA, China, Chile, Bangladesh, Taiwan, Mexico, Argetina, Poland, Cannada Japan Indian and New Zealand ( Mohan and Pittman Jr, 2007 ) The largest population at risk with known groundwater a rsenic contamination is in Bangladesh, followed by west Bengal in India ( Mohan and Pittman Jr, 2007 ) The maximum permissible conc entration of As in drinking water is 50 g /L and recommended value is 10 g /L by EPA and WHO ( Smedley and Kinniburgh, 2002 ) Biochar Due to the toxic effect of Cr(VI), Hg(II) and As(III) on the environ ment it is important to develop effective and economical methods to reduce their impact A number of approaches have been applied for remediation of sites contaminated with Cr, Hg and As. Among the various methods like oxidation reduction, precipitation, co precipitation, and biosorption, biosorption technique is the most common and cost effective ( Demirbas, 2008 ) In addition, biosorbents are readily available in large quantities and they are environmental friendly. Biosorbents like agricultural waste ( Demirbas, 2008 ) marine seaweed ( Yu et al. 2003 ) and sludge biomass ( KilI et al.

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26 2008 ) have been used as promising and low cost sorbents to immobilize heavy metals in contaminated water. Due to these advantages, there has been ex tensive research exploring appropriate biosorbents capable of effectively immobilizing Cr(VI) Hg(II) and As(III) from water, such as sawdust ( Dakiky et al. 2002 ) walnut shell ( Wang et al. 2009b ) activated carbon from biomaterial ( Kadirvelu et al. 2004 ) and agricultural byproduct such as sunflower stem ( Jain et al. 2009 ) and rice husk ( Amin et al. 2006 ) H owever, little research has been done on the mechanisms of Cr(VI), Hg(II) and As(III) removal by biochar. Mohan et al. (2011) studied adsorption of Cr(VI) on bio chars from oak wood and oak bark versus pH, temperature and solid to liquid ratio. Maximum sorp tion capacity was o b tained at pH 2.0. Cr(VI) removal increased with an increase in temperature They is th e dominant mechanism governing the Cr(VI) removal by coconut coir derived biochar ( Shen et al. 2012 ) Kong et al. (2011) Investigated Hg(II) sorption by s oybean stalk based biochar. The maximum sorption capacity is 675 g/g at pH7.0 according to Langmuir model. Ion exchange and physicochemical adsorption are proposed as the mechanisms ; however, there is no direct evidence provided Physical a nd Chemical Properties Biochar is a product of thermal degradation of organic materials in the absence of air (pyrolysis). It has the ability to store a massive amount of greenhouse gas, increase crop yield, improve soil and water quality, lower soil emiss ions of other greenhouse gases (e.g. nitrous oxide and methane), decrease leaching of nutrients and

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27 reduce irrigation and fertilizer requirements ( Lehmann, 2007 ; Bird et al. 2008 ; Cui et al. 2008 ; Kimetu et al. 2008 ; Nguyen et al. 2009 ) Even more important is that it shows a great affinity for heavy metal ions, like As, Cd, and Pb ( Mohan et al. 2007 ; Cao et al. 2009 ; Liu and Zhang, 2009 ) The physical and chemical properties of biochar significantly influence its adsorption ability. I n order to explore the mechan ism governing metal ions removal by biochar, its major physical and chemical properties need to be investigated first. Surface area, porosity and pore size are the major physical properties that significantly impact the affinities of biochar for organic a nd inorganic pollutant. The surface area of biochar strongly affects its reactivity and combustion behavior. Chen et al. ( 2009 ) found that large surface area led to higher sorption of naphthalene on orange peel derived biochar. Both the porosity and pore size determine the adsorption properties of bio chars. For instance, small pore size w ill not trap large adsorbate and large pores may not be able to retain small adsorb ate regardless of their charges and polarity ( Ahmedna et al. 2004 ) Biomass rich in lignin (bamboo and coconut shell) develop biochar with macroporous structure, while biomass r ich in cellulose (husks) yield a predominantly microporous structured biochar ( Joseph S.D., 2007 ) Compared with the physical properties, the influence of chemical properties is even more significant. Many function groups such as carboxylic, amino, and hydroxyl groups are present on the surface of biochar which vary with original biomass stock It is well known that some function groups play an important role in the sorption process. For example, carboxyl groups played a major role in the comlexation of Cd and Pb by

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28 brown seaweed Sargassum fluitans ( Pan et al. 2000 ) Amine, carboxyl and hydroxy l groups contribute to Hg(II) sorption by Aspergillus versicolor biomass ( Das et al. 2007 ) Metal concentration on the biochar is another important factor. The rel ative proportion of the mineral content will influence the final chemical composition and structure of the bio char. For instance, after treat ing with acid to remove the hydrogen consumption ions ( Ca 2+ and Mg 2+ ), sludge biomass provided more hydrogen ions for Cr(VI) reduction process, and its Cr(VI) reducing capacity was increased by 20% ( Wu et al. 2010 ) Metal concentration is even more important for ion exchange governed adsorption process. For instance, Ca 2+ Mg 2+ Na + K + and H + i ons are released from the biomass into the solution as a result heavy metals (such as Cu, Mn, Zn, Cd, and Sr) are sorbed onto the biomass ( Treen Sears et al. 1984 ; Avery and Tobin, 1993 ; Muraleedharan et al. 1994 ) Sugar Beet Tailing and Brazilian Pepper B iochar Many biomass feedstock can be u sed to make biochar, such as wood, grasses, paper sludge, manure and green waste ( Wool f, 2008 ) Sugar beet tailing and invasive plant Brazilian pepper derived biochar s may also be used to remediate heavy metal s in contaminated environment. Sugar beet, a commercial crop for sugar production, is mainly grown in the north central and north w estern U.S. In 2007 08, Florida contributed an estimated 24.3 % of the total sugar (from sugarcane and sugar beet) produced in the U.S. Sugar beet tailing is a by product from sugar industry. It consists of small beets, broken or damaged beets, soil and oth er material not suitable for sugar production. It is high in moisture ( ~ 80%) and variable in nutrient content. Usually, tailings are stockpiled outside the factory and then disposed into landfills or land applied on nearby farmland at a

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29 significant cost. F or example, American Crystal Company spends 1 million dollars per year for disposing of 400 tons of tailings that are generated daily at its East Grand Forks, Minnesota ( Liu et al. 2008 ) Brazilian pepper is the most aggressive, evergreen shrub like tree in Florida. It invades many habitats by forming large dense forests ( FDEP., 2008 ) It poses a serious threat to species diversity in Florida cause several health and safety problems because it is a relative of poison ivy and can cause dermatitis. Now it covers more than 700,000 acres in south and central Florida, as well as many of the islands on the east and west coas ts of the state. In an effort to completely remove Brazilian pepper from 6,600 acre wetlands in the Everglades, the National Park Service spent 90 120 million by scraping the soil down to the bedrock ( Service, 2008 ) However, the question remains as to what to do with the seven million tons of removed soil. Although 46 herbivorous insects have been collected from Brazilian pepper in south Florida, they do not appear to retard plant growth or vigor. Coal Combustion Residues Production of Coal Combustion Residues Coal combustion residues (CCR) are the materials that remain after burning coal for electricity, which include fly ash, bottom ash, coal slag and flue gas desulfurization (FGD) r esidues ( EPA, 2009 ) Besides mining wastes, CCR are the second largest waste streams generated in t he US ( Fitzgerald, 2010 ) Almost 108 mi llion metric tons of CCR are generated on an annual basis within the United States ( ACAA, 2003 ) Florida ranks the 7 th in the country in term of CCR production with annual production between 160,000 and 1,185,000 tons ( EPA, 2009 )

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30 Fly ash is produced in the boiler during the combustion of coal. Due to its fine size (0.5 to 100 m) fly ash is carried out of th e boiler in the flue gas and is separated from the air stream by the air emission control devices to minimize its release to the atmosphere. The air emission control devices include cyclone, baghouse and electrostatic precipitators, and electrostatic precipitators are used in over 95% of US coal power plants ( Schnell and Brown., 2002 ) Ga sification slag is produced from conversion of coal to gaseous and liquid fuels, and is similar to fly ash but contains a higher proportion of coarse particulate materials ( Rowe et al. 2002 ) Bottom ash and coal slag are the residues remained in the boiler after removing fly ash Bottom ash is created in dry bottom boilers and stokers. Coal slag is produced from wet bottom boiler and cyclone boiler, in which the combu stion temperature is higher than the ash fusion temperature; therefore, ash materials are present in a molten state. After removed from boiler, the molten ash materials undergo rapid cooling fracture and crystallization to form the hard black, angular an d glassy pellets ( EPA, 2012 ) FGD residues are produced by air emission control devices to remove the SO 2 from the flue gas, which is a pollutant that must be removed before released to the atmosphere FGD systems are usually classified as wet or dry. Wet FGD system is most commonly used. It contains lime or limestone based materials which could trap SO 2 after contacting with the exhau st gas to form calcium sulfite or calcium sulfate ( EPA, 2012 ) However, large volumes of wastewater and sludge are produced. Dry FGD system is primarily used in coal power plant with low sulfur coal ( Gainer, 1996 ) There are three major categories: spray dryers, circulating spray dryers and dry injection

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31 systems. They all absorb SO 2 after contacting with flue gas and then the dry product is removed with the fly ash in electrostatic precipitators or baghouse. Characteristics of C oal Combustion Residues Fly ash consists of fine particle s (<0.625 m) with spherical shape in amorphous state. Fly ash has a specific g ravity between 2.1 to 3.0 and specific surface are a of 170 to 1000 m 2 /kg ( ASTMC618, 1994 ) Fly ash color varies from tan to black, as a result of carbon content remaining in the ash. Chemical pr operties of fly ash are less constant than physical properties. The considerable variation is due to the difference in coal source, coal preparation method, combustion conditions, furnace type and collection, handling and storage conditions. But it all con tains substantial amounts of SiO 2 Al 2 O 3 Fe 2 O 3 and CaO ( McCarthy et al. 1990 ) Table 2 1 Concentrations of various elements in fly ash samples (presented as ranges of means: mg/kg) (Adapted from Rowe, Hopkins and Congdon. 2002. (Page 213, Table 2). Journal of Environmental Monitoring and Assessment and EPRI. 2010. Gove rnment Report (Page 2 3, Table 2 1). Palo Alto, CA. ) Fly ash As Cd Cr Pb Hg Se Sb Be Mo #sites 59 28 59 58 19 59 38 229 57 Max 1,385 17 651 2,120 1.3 47 131 826 236 Min 8.1 <0.11 11 13 <0.0025 <1.4 <7.2 <0.4 4 Co Cu Mn Ni Tl V Zn Fe Al #sites 3 57 49 57 59 39 59 66 42 Max 124 1,452 1,332 353 85 652 2,880 175,550 152,000 Min 7.3 45 44 23 <0.17 <43.5 25 17,000 46,000 NA: data not available CaO content is an important indicator for fly ash classification. Fly ash is usually classified as cl ass C or c lass F: c lass C fly ash contains CaO over 15%, while c lass F only contains CaO less than 6% ( Joshi and Lohtia., 1997 ) CCR also contains a b r oad

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32 range of trace metals. Compared to coal, CCR is relatively rich in trace metals such as Cu, Zn, As, Se and Hg. In addition, CCR is also enriche d in other trace metals including Ba, B, Co, Cd, Cr, Co, Pb, Sr, Mo, and Ti. However, trace metals concentrat ions are quite variable (Table 2 1) ( Rowe et al. 2002 ; EPRI, 2010 ) Table 2 2. Concentrations of various elements in bottom ash samples (presented as ranges of means: mg/kg) (Adapted from Rowe, Hopkins and Congdon. 2002. (Page 213, Table 2). Journal of Env ironmental Monitoring and Assessment and EPRI. 2010. Government Report (Page 2 3, Table 2 1). Palo Alto, CA. ) Bottom Ash As Cd Cr Pb Hg Se Sb Be Mo #sites 37 37 37 37 160 37 37 152 37 Max 56 <5.5 4,710 843 1.3 8.2 8.4 568 46 Min <1.3 <5.5 <24 <2.1 <0.002 <1.25 <7 <0.064 <1.4 Co Cu Mn Ni Tl V Zn Fe Al #sites NA 37 37 37 21 37 37 37 42 Max NA 146 1940 1,267 59 275 717 199,500 145,000 Min NA 20 73 <12 <0.5 <50 3.8 21,600 30,500 NA: data not available Bottom ash and slag are the heavy, coarse, granular, incombustible particles remaining in the bottom of the boiler. Bottom ash has a specific gravity of 2.1 to 2.7 and the typical par ticle size varies between 75 m and 2 mm The specific gravity of slag is similar to bottom ash, 2.2 to 2.6 ( Joshi and Lohtia., 1997 ) Compared to bottom ash, slag has smaller and more constant particle size 0.5 to 5 mm. Chemical properties of bottom ash and slag are similar ( TFHRC, 2002 ) Both are composed primarily of silica, alumina, i r on a s well as large amounts of calcium, magnesium and other trace metals ( TFHRC, 2002 ) The chemical properties are also varied considerably according to the coal source and operating parameters (Table 2 2 ).

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33 Tabl e 2 3. Concentrations of various elem ents in FGD samples (presented as ranges of means: mg/kg) (Adapted from Rowe, Hopkins and Congdon. 2002. (Page 213, Table 2). Journal of Environmental Monitoring and Assessment and EPRI. 2010. Government Report (Page 2 3, Table 2 1). Palo Alto, CA. ) FGD As Cd Cr Pb Hg Se Sb Be Mo #sites 27 26 27 27 27 27 26 27 26 Max 11 0.37 24 2.0 1.5 32 2.0 <0.1 3.1 Min <1.9 <0.02 0.60 <1 0.0075 <2.5 <0.4 <0.1 <0.02 Co Cu Mn Ni Tl V Zn Fe Al #sites 26 26 27 26 26 26 26 26 NA Max <1 3.2 129 2.4 8.6 8.6 23 1, 823 NA Min <1 <0.4 <1 <0.2 <1 <1 <1.25 130 NA NA: data not available FGD also consists of fine particle s usually between 0.4 m and 0.5 m with specific gravity of 2. 2 to 2.6 ( TFHRC, 2002 ) The yellow color mainly comes from the sorption of SO 2 Chemical properties of FGD vary significantly with different methods of SO 2 removal (Table 2 3) Fly ash and spen t sorbent are the dominant component s if the FGD scrubbers are used before the air emission control devices ( ICAC, 1995 ) The fly ash content decreases significantly if the FGD scrubbers are used after removing the fly ash. Besides the fly ash and spend sorbent, calcium sulfate and calcium sulfite are also present ( EPRI, 1999 ) The relative contribution of these two products depends on the FGD material. For example, i n the wet syste m with lime based reagent, calcium sulfite is the main product while calcium sulfate dominant the limestone based wet FGD scrubber ( EPA 2012 )

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34 Potential R isk of C oal Combustion Residues to t he Environment CCR are usually disposed in two ways: dry disposal in landfill or wet disposal at on site storage pond. Both ways may cause a variety of environmental problems due to the high concen tration of soluble salt, trace metals and other pollutants that may leach out from CCR ( Fitzgerald, 2010 ) EPA has reported at least 137 cases of surface water or groundwater contamination from CCR in at least 34 states. The state with the most coal ash cases is Wisconsin with 13, followed by I llinois with 12, North Carolina with 10, Indiana and Pennsylvania with 9 and Florida with 8 ( EPA, 2007a ) The largest one ns of CCR storage in the dike flooded the surrounding area. The leachable contaminants As, Se, B, Sr and Ba showed effects on the quality of the impacted area by an 18 month survey ( Ruhl et al. 2010 ) T he As concentration was extremely high, up to 2000 g/ L, ( Ruhl et al. 2010 ) Coal Combustion Residues Regulatory Rules CCR wa s regulated as nonhazardous materials in 1993 by EPA. The need for national regulations of CCR disposal is heightened by th e Kingston CCR spill from a surface wet pond. EPA is currently proposing to regulate CCR to address the potential risks from CCR disposal. EPA is considering two possible options: 1) EPA will list CCR as special wastes subject to regulation under subtitle C of RCRA (Resource Conservation and Recovery Act) which will be disposed in landfills or surface impoundments; 2) EPA will regulate CCR under subtitle D o f RCRA as a non hazardous wastes. EPA considers each option to have its advantages and disadvantage s. It is still in the development stage.

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35 Research Objectives Th is study wa s divided into two sections : biochar and CCR. The overall objective of the first section wa s to investigate the effectiveness and mechanisms of Cr(VI), Hg(II) and As(III) immobilizat ion and transformation by biochar in aqueous solution In order to meet this objective, biochar characteristics, impact factors and mechanisms were investigated. The overall objective of the second section wa s to characterize Florida CCR samples. In order to meet this objective, 27 CCR samples were collected from 7 Florida coal power plants and the total concentration and SPLP concentration were analyzed T his dissertation wa s structured as follows: Chapter 2 focus ed on the effectiveness and mechanisms of C r(VI) removal by SBT biochar C hapter 3 emphasize d the mechanism of Hg(II) removal by BP biochar produced at different temperature Chapter 4 evaluate d Cr(VI) reduction and As(III) oxidation by biochar derived dissolved organic matter ( DOM ) Chapter 5 inve stigate d the total and SPLP concentration of metals in 27 CCR samples collected from 7 Florida coal power plants. Chapter 6 provide d a general conclusion and emphasize d the major observations in this study.

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36 C HAPTER 3 CHARACTERISTICS AND MECHANISMS OF HEX AVALENT CHROMIUM REMOVAL BY BIOCHAR FROM SUGAR BEET TAILING Introduction Because of the global awareness of its underlying detrimental effects, chromium contamination in the environment has received great attention in recent decades. As a major pollutant i n surface water and groundwater, Cr is released to the environment by various industrial activities including electro plating, chromate manufacturing, leather tanning and wood preservation ( Papp, 2001 ; Dnmez and Aksu, 2002 ; Kim et al. 2002 ) In natural waters, Cr exists primarily as Cr(VI) and Cr(III). Cr(III) is much less soluble and relatively stable. In comparison, Cr(VI) is highly soluble and mobile in aqueous solutions. The latter is of significant environment concern due to its carcinogenic, mutagenic and tertogenic effects on biological system ( Fendorf et al. 2000 ; Costa, 2003 ) The Cr(VI) concentration in wastewater produced by various industries range from 0.1 an d 200 mg/L, but as high as 3950 mg/L has been observed in tannery plant wastewater ( Owlad et al. 2009 ) The USEPA limit for Cr in potable water and discharge into inland surface water is 0.05 mg/L and 0.1 mg/L respectively ( EPA, 1 990 ) Due to its toxic effect, it is important to remove Cr(VI) from water system. produced from carbon rich biomass. Recent developments of renewable bioenergy technologies make it possible to convert waste biomass into value added biochar and at the same time produce bioenergy. Sugar beet tailing (SBT), a by product from sugar industry, can be used to produce biochar. SBT consists of small beets, broken beets, soil and other mate rial unsuitable for sugar production. It is high in moisture (~80%) and

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37 variable in nutrient content. Usually, they are stockpiled outside the factory, which are disposed into landfills or land applied on nearby farmland at a significant cost ( Liu et al. 2008 ) Biochar shows a great affinity for heavy metals and their sorption capacity is comparable with other biosorbents. For example, th e Pb sorption capacity of various biochars ranges from 2.4 (rice husk) to 20.5 mg/g (sugarcane bagasse) ( Mohan et al. 2007 ; Cao et al. 2009 ; Liu and Zhang, 2009 ; Ding et al. 2010 ) Though little information is available on Cr removal by biochar, the effectiven ess of Cr removal by various biomaterials has been investigated ( Park et al. 2008 ; Abdullah and Prasad, 2009 ) Removal of Cr(VI) is mainly attributed to Cr(VI) reduction in an acid medium. Park et al. ( 2006 ) proposed that adsorption coupled with reduction is the main mechanism of Cr(VI) removal by biomaterials. Du e to its high redox potential value (+1.3 V), Cr(VI) is easily reduced to Cr(III) under acid conditions in the presence of organic matter. However, little work has been done to examine the mechanisms and effectiveness of Cr(VI) removal by biochar. The obj ectives of this investigation were to (1) investigate the effectiveness of SBT biochar in removing Cr(VI) from water via sorption/desorption studies; (2) evaluate the effects of various parameters including pH, initial Cr(VI) concentration, contact time an d biomass on its effectiveness, and (3) determine the mechanisms governing Cr(VI) removal by SBT biochar. Materials and Methods Biochar Preparation and Characterization Fresh sugar beet tailing collected from a sugar industry in Florida, U.S. was oven dr scale tubular reactor within a muffle furnace

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38 (1500M, Barnstead Intl. Corp, IA). Nitrogen gas at 10 psi was used to maintain an oxygen free environment. The furnace temperature was programmed to increase at a rate of 10 was allowed to cool at room temperature under a flow of N 2 gas. It was then washed The drie d biochar was stored in air tight plastic containers prior to use and will be referred as SBT biochar. The surface morphology of SBT biochar was examined using JSM 6400 scanning electron microscope (SEM, JEOL USA, Peabody, MA) at 15 keV equipped with energ y dispersive X ray spectroscopy (EDS, Oxford Instruments USA, Concord, MA). Evaluation of its zero point of charge (ZPC) was based on the mass titration procedure of Valds et al. ( 2002 ) Three aqueous solutions of different pH (3, 6 and 11) were pr epared. Different amounts of the adsorbent (0.05, 0.1, 0.5, 1, 3, 7 and 10 %, w/w) were added to 20 mL of each solution. The aqueous suspensions were then allowed to equilibrate for 24 h under agitation (150 rpm) at room temperature ach solution was measured using a digital pH meter and the ZPC was determined as the converging pH value from the pH vs. sorbent mass curve (Table 3 1). Surface area was analyzed using a NOVA 1200 surface area analyzer (Quantachrome Instrument, Boynton Bea ch, FL) with the Bruanuer Emmett Teller (BET) nitrogen adsorption method at 77 K. Surface area was 137 m 2 /g (CO 2 ) and 0.2 m 2 /g (N 2 ) (Table 3 1). The elemental composition of the SBT biochar was determined by inductively coupled plasma atomic emission spect roscopy (ICP AES, Plasma 3200,

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39 Perkin Elmer Crop, MA) after digesting with HNO 3 /H 2 O 2 hot block digestion procedure ( EPA, 1986 ) Cr(VI) Sorption and Desorption Experiments The stock solution of 1000 mg/L Cr(VI) was prepared by dissolving analytical grade K 2 Cr 2 O 7 (analytical reagent gra de) in deionized water. All working solutions were freshly prepared before use by diluting the stock solution with deionized water, and the pH was adjusted to the desired values with 1 M HNO 3 or 1 M NaOH solution. Batch kinetic experiments were carried ou t at constant pH of 2.0 with 100 mg/L Cr(VI) and sorbent dose of 2 g/L at room temperature. After shaking, aliquots of solution samples were withdrawn at different time intervals. Batch equilibrium experiments were conducted using 60 mL plastic bottles wit h sample volume of 50 mL, and Cr(VI) concentrations of 50, 100, 200, 400, and 800 mg/L. The concentrations were within the range found in industry wastewater samples ( Owlad et al. 2009 ) The samples were agitate reaction time, a known volume of the solution was removed and centrifuged at 14,400 rpm for 10 min for Cr(VI) analysis. The effect of pH on Cr(VI) sorption onto SBT biochar was inve stigated by varying solution pH from 1.0 to 6.0 while keeping Cr(VI) at 100 mg/L and biochar at 2 g/L. The effect of biomass concentration was studied in the range of 0.2 8 g/L while keeping Cr(VI) at 100 mg/L and pH at 2. Desorption experiment was carrie d out as follows: after sorption experiment with 100 mg/L Cr(VI) and 2 g/L biochar at pH 2.0 reached an equilibrium and washed with DI water several times, solutions containing 0.1 M H 2 SO 4 or 0.1 M NaOH was used to desorb from the biochar. The solutions we re agitated on a shaker at 200 rpm and room temperature for 24 h. Samples for analysis of Cr(VI) and total Cr concentrations were

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40 removed periodically. All experiments were conducted in tri plicate and the results were calculated as the mean values of the s amples. Fourier Transform Infrared (FTIR) and X ray Photoelectron Spectroscopy (XPS) A nalysis FTIR analyses (Bruker Optics Inc., Billerica, MA) were used to identify the functional groups on the biochar before and after Cr(VI) sorption. Prior to FTIR anal range from 500 to 4000 cm 1 with a resolution of 8 cm 1 The resulting spectra were the average of 32 scans, which were used to identify the functional groups based on their characteristic absorbance peaks A ll experiments were conducted in triplicate The valence state of the Cr bound on the SBT biochar was determined by an X ray Photoelectron Spectrometer (Perkin Elmer Model 5100, USA). The Cr laden biochar was obtained by shaking 2 g/L biochar with 100 mg/L of Cr(VI) at pH 2.0 for 24 h. The sample was analyzed at an Al X pass energy of 0.1 eV, with 10 high resolution scans. The system was operated at a base pressure of 210 8 mbar The calibration of the binding energy of the spectra was performed with the C1s peak of the aliphatic carbons at 284.6 eV. Cr(VI) Analysis The concentrations of Cr(VI) were analyzed by measuring the absorbance of the purple complex of Cr(VI) wit h 1,5 diphenylcarbohydrazide at 540 nm using a UV spectrophotometer (Shimadzu Corporation, Japan) ( Eaton et al. 1995 ) To determine the total Cr, Cr(III) was first converted to Cr(VI) at high temperature (13 0 adding excess potassium permanganate prior to reacting with 1,5 diphenylcarbohydrazide ( Park et al. 2008 ) The concentrations of Cr(III) were

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41 calculated as the difference b etween the total Cr and Cr(VI) concentrations. The detection limit of this method was 0.03 mg/L. Results and Discussion Characteristics of SBT Bio char Before and After Reaction w ith Cr(VI) Based on the chemical analysis, SBT biochar contained substantial amounts of cations, including Ca, K, Mg, Fe, Al and Na, with Ca being the highest at 2.8% (Table 3 1). Assuming all the Ca is soluble, the calculated concentration is 56.4 mg/L. The actual Ca concentrations after reacting with 100 mg/L Cr(VI) or DI water w ere 28.3 29.0 mg/L (Table 3 1), indicating 50 51% of the Ca was retained on the SBT biochar. The surface characteristics of SBT biochar before and after reaction with Cr(VI) is shown in a SEM micrograph (Fig ure 3 1a, b). Consistent with the chemical analys is, the EDS spectra also indicated that the SBT biochar contained substantial amounts of K, Mg and Ca (Fig ure 3 1c, d). The presence of Cr on the biochar was obvious after reaction with 100 mg/L Cr(VI) (Fig ure 3 1d). Table 3 1. Elemental concentrations of SBT biochar (mg/g) and in solution (mg/L) after SBT biochar reaction with 100 mg/L Cr(VI) and DI water at pH 2 for 24h. Metals 100 mg/L Cr(VI) (Mg/L) DI Water ( mg/L ) SBT biochar(mg/g) Na 1.170.49 1.710.39 2.650.22 K 9.510.05 8.700.68 18.960.85 Ca 29.00.36 28.30.14 28.210.54 Mg 9.090.10 9.030.06 9.860.21 Al 0.500.03 0.700.01 2.950.06 Fe 0.320.004 0.790.04 4.410.09 Initial pH 2.0 2.0 Final pH 2.410.005 2.220.005 Point zero charge 7.32 Surface area (m 2 /g CO 2 ) 137 Surface area ( m 2 /g N 2 ) 0.2 Particle size

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42 Figure 3 1. Scanning electron micrographs and EDS spectra of SBT biochar before (a,c) and after (b,d) reaction with 100 mg/L Cr(VI) for 24 h at pH 2.0. The FTIR spectra of SBT biochar before and after Cr(VI) sorption were used to determine the vibration frequency changes in its functional groups. The spectra displayed a number of sorption peaks indicating the complex nature of SBT biochar (Fig ure 3 2). Before sorption, the broad sorption peak at 3317 cm 1 was indicative of the exi stence of bonded hydroxyl group. The sorption peak around 2927 cm 1 was assigned to CH stretching ( Kapoor and Viraraghavan, 1997 ) The peak at 1620 cm 1 b a c d

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43 represented a chelated form of the carbonyl from the carboxyl group ( Yun et al. 2001 ) In some studies this peak was described as the region of both ionized non coordinated ( Drake et al. 1995 ) When an amide group is present, it could represent the CO stretching mode conjugated with the NH 2 (amide 1 band) group ( Kapoor and Viraraghavan, 1997 ) The peak at 1420 cm 1 corresponded to aromatic C=C ring stretchin g ( zimen an d Ersoy Meriboyu, 2010 ) The peaks observed at 1375, 1317, 1049 and 780 cm 1 could be assigned to C=O stretch of carboxylate ions, hydroxyl bending vibration, CO stretching vibration of the alcoholic groups and aromatic compounds, respectively ( Elangovan et al. 2008 ; Jain et al. 2009 ; Kousalya et al. 2010 ; Wang et al. 2010 ) In short, the functional groups present on the biochar included CH, OH, C=O, and CO, which may be responsible for Cr sorption. Figure 3 2. Fourier transform infrared spect ra of SBT biochar before and after reaction with 100 mg/L Cr(VI) for 24 h at pH 2.0.

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44 Changes in the function groups of SBT biochar were visible after Cr sorption (Fig ure 3 2 ). The small and sharp peak at 1317 cm 1 ( OH) in the native SBT biochar disappeare d after Cr(VI) sorption and meanwhile one peak at 1700 cm 1 appeared, which represented the C=O stretch of carboxylate ions. These changes showed that hydroxyl and carboxylate groups may be involved in Cr removal. These results are in good agreement with o ther studies ( Komy et al. 2006 ; Gabr et al. 2008 ; Joo et al. 2 010 ; Wang et al. 2010 ) which concluded that carboxyl and hydroxyl groups are the main functional groups for sorption of heavy metals. Characteristics of Cr(VI) Sorption b y SBT Biochar To better under stand the Cr(VI) sorption characteristics by SBT biochar, both kinetic and equilibrium models were used to describe the data. Changes in Cr concentrations over time are shown in Fig ure 3 3a. Cr removal by SBT biochar was rapid within the first 2 h and afte r that it decreased gradually, reaching equilibrium after 16 h. The data indicated that Cr sorption can be divided into two stages: at early stage the sorption rate was very rapid (78% Cr was sorbed in 2 h), followed by a stage where the sorption rate was slow. Based on this study, 24 h was used in all following experiments. The kinetics of Cr(VI) sorption on SBT biochar were analyzed with the pseudo second order model, which is a better fit than pseudo first order (R 2 >0.99 vs. R 2 =0.95; data not shown). Th e pseudo second order equation is based on the assumption that the sorption rate is controlled by both sorbent capacity and sorbate concentration ( Ar slan and Pehlivan, 2007 ) and is expressed as: ( 3 1)

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45 where K 2 is the pseudo second order rate constant (g/mg h), q e and q t represent the amount of Cr(VI) sorbed (mg/g) at equilibrium time and at time t, respectively. The equilibrium sorption capacity (q e = 47.9 mg/g), and the second order constants (K 2 = 0.027) were determined from the slope and intercept of the plot t/q versus t (data not shown). The good fit of the data to the model implied that the removal of Cr (VI) by SBT biochar was closer to chemi sorptions, i.e., a new chemical species were generated at the sorbent surface. Furthermore, the sorption capacity obtained from pseudo second order equation fitting result was consistent with the experimental value of 47.8 mg/g (data not shown). In additi on to kinetic model, the sorption data were analyzed with Langmuir equation, which provided a slightly better fit th an Freundlich isotherm models ( Fig ure 3 4; R 2 =0.990 VS. R 2 =0.985). The Freundlich isotherm model assumes that the adsorption occurs on a het erogeneous surface by multilayer sorption and that the amount of sorbate sorbed increases infinitely with increasing concentration ( Jain et al. 2009 ) The Freudlich isotherm model was applied for Cr(VI) and is expressed as ( Jain et al. 2009 ) : (3 2 ) where C e is the equilibrium concentration (mg/L), q e is the amount of Cr(VI) adsorbed at equilibrium time (mg/g), K f and n are the adsorption capacity and intensity incorporating all factors affecting the adsorption process. Both K f and n affect the adsorption isotherm, the larger the K f and n values, the higher the adsorption capacity.

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46 Figure 3 3. Effect of reaction time ( a ), pH (b ) and biochar mass ( c ) on Cr removal by SBT biochar after reactin g with 100 mg/L Cr(VI). a b c

PAGE 47

47 Figure 3 4. Langmuir and Freundlich plot for Cr(VI) sorption on SBT biochar. On the other hand, the Langmuir model assumes that the uptake of metal ions occurs on a homogeneous surface by monolayer sorpt ion without interaction between sorbed ions. The model assumes uniform energies of sorption onto the surface and no transmigration of sorbate in the plane of the surface. The Langmuir isotherms are represented by the following equation ( Jain et al. 2009 ) : ( 3 3) where Q 0 (mg/g) is the maximum quantity of metal ions per unit biomass to form a complete monolayer on the surface and b is a constant related to the affinity of binding sites with the metal ion. The values of Q m (137 m g/g) and b (0.052 L/g) were calculated from Fig ure 3 4. The sorption capacity of SBT biochar was better than ot h er

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48 biosorbents reported in the literature, which range from 21.3 mg/g by wheat biochar ( Wang et al. 2010 ) to 123 mg/g by Rhizopus nigricans ( Loukidou et al. 2004 ) In addition to reaction time, the effects of pH and biochar mass on Cr(VI) sorption were also investigated. Fig ure 3 3b shows that removal of Cr(VI) decreased significantly with increased solution pH. All Cr(VI) was removed at pH 2 compared to 16% at pH 3. The low removal of Cr(VI) at pH 3 may imply the importance of Cr(VI) reduction to Cr(III) for Cr sorption onto the biochar. At pH 3, the SBT biochar was still positively charged (PZC=7.32; Table 3 1) and should be able to sorb Cr(VI) if this process was dominant. The decrease in Cr(VI) removal as pH increased from 3 to 6 could be attributed to the increase of OH competition for Cr(VI) species for the sorption sites on the biochar. The optimum solution pH for removal of Cr by SBT biochar was 2, where 37% of the initial Cr(VI) were present in solution as Cr(III) (Fig ure 3 3b). Compared to Cr(VI), Cr(III) is much less mobile and less toxic in the environment ( Fendorf et al. 2000 ) Reduction of Cr(VI) to Cr( III ) b y SBT Biochar As expected, the removal of Cr(VI) increased with increasing biochar mass (Fig ure 3 3c). At >4 g/L biochar, n o Cr(VI) was detected, i.e, all Cr in the solution was present as Cr(III) (Fig ure 3 3c). Hence, the increase in Cr removal of 11.2 mg/L as the biochar increased from 4 to 8 g/L could be attributed to Cr(III) sorption onto the biochar. In this case, then at least 11.2% of the Cr on the biochar was present as Cr(III). Concentrations of Cr(VI) and Cr(III) in the solution during Cr(VI) removal by SBT b iochar were monitored (Figure 3 3a). The Cr(VI) concentration decreased dramatically with time ; after 24 h, ~98 % of Cr(VI) was removed from the aqueous solution. Meanwhile, Cr(III), which did not exist initially, increased in proportion to the Cr(VI) depletion, with final Cr(III) concentr ation being 18.3 mg/L (Figure 3 3a). These results

PAGE 49

49 indicated that some of the Cr(VI) was reduced to Cr(III) when in contact with SBT biochar, and part of the converted Cr(III) was retained by the biochar and part was released into the solution. In aqueous solution at pH 2, Cr(III) is predominantly present as Cr 3+ whereas Cr(VI) is u nstable. Since the PZC of SBT biochar was 7.32 (Table 3 1), it was positively charged at pH 2. This means that SBT biochar was able to sorb Cr(VI), which was consistent with the continued decrease in Cr(VI) in solution (Figure 3 3a). However, the biochar w as ineffective in sorbing Cr(III), which was reflected by 18.3 mg/L Cr(III) in the solution after 24 h. Still at the end of reaction, ~ 78% Cr was sorbed onto the biochar. As a result of SBT biochar sorption of Cr(VI) and Cr(VI) reduction to Cr(III) during the 24 h reaction, Cr(VI) concentrations decreased whereas Cr(III) concentrations increased (Figure 3 3a). However, the appearance rate of Cr(III) was much less than the disappearance rate of Cr(VI). For example, from 2 to 4 h, Cr(III) concentration incre ased by 1.4 mg/L, at the same time, Cr(VI) concentration decreased by 9.7 mg/L (Figure 3 3a). The decrease in aqueous Cr(VI) can be attributed to two processes: 1) sorption by the positively charged SBT biochar, and 2) reduction to Cr(III). Based on mass b alance, part of Cr(III) was released into solution (18.3 mg/L) and part was present on the biochar. To investigate the valence state of the Cr bound on SBT biochar, X ray photoelectron spectroscopy was used. Similar to aqueous solution, both Cr(III) and Cr (VI) were de tected on the biochar (Figure 3 5). Significant bands appeared at binding energies of 577.0 580.0 (Cr 2p 3/2 orbitals, which were used to identify the valence state of Cr) and 587.0 588.0 eV (Cr 2p 1/2 orbitals). The Cr 2p 3/2 orbitals are assigne d at 577.2

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50 eV (CrCl 3 ) and 576.2 576.5 eV (Cr 2 O 3 ) for Cr(III) compounds, while Cr(VI) compounds are characterized by higher binding energies such as 580.1 eV (CrO 3 ) or 580.6 eV (K 2 Cr 2 O 7 ) ( Merri tt et al. 1983 ; Stypula and Stoch, 1994 ) The binding energy of Cr 2p3/2 orbitals at 577.50 eV was attributed to Cr(III) (Figure 3 5). Thus, the spectra of Cr loaded SBT biochar clearly indicated t he presence of Cr(III) on its surface. The small peak at 580.5 might indicated the presence of Cr(VI). However, due to the noise, it was impossible to quantify the ratio of Cr(VI) to Cr(III). Figure 3 5. XPS spectra of the SBT biochar after reacting 2 g /L SBT biochar with 100 mg/L Cr(VI) for 24 h at pH 2.

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51 Even at pH < PZC (7.32; Table 3 1), the number of negatively charged sites on SBT biochar may be sufficient to bind all of Cr(III) ions, which is the possible mechanism for Cr(III) sorption. Arslan et al. ( Arslan and Pehlivan, 2007 ) reported that at pH 2.0 to 3.2, the carboxylic acid sites could be appreciably deprotonated (COO ), which might bin d some Cr(III) ions. Gardea Torresdey et al. ( Gardea Torresdey et al. 2000 ) demonstrated that the reduced Cr(III) is possibly bound to the oxygen containing ligands of carboxyl groups of oat biom ass. Carboxyl groups iden tified on SBT biochar (Figure 3 2) may have participated in Cr(III) sorption onto biochar. By adding SBT biochar to a solution, a proportion of it readily dissolved. At pH 2.0, the dissolved organic carbon content remained constan t at 80.75 mg C/L after 24 h. In the treatment containing only 80.75 mg C/L DOC, at pH 2.0, 39.6 mg/L Cr(VI) was reduced The observation s indicated that Cr(VI) removal by SBT biochar included two combined processes: (1) adsorption and reduction of Cr(VI) on the biochar surface and (2) reduction of Cr(VI) by biochar derived dissolved organic carbon. Desorption of Cr f rom SBT Biochar Based on XPS, both Cr(VI) and Cr(III) peaks were observed. To better understand the Cr species distribution on the biochar, concentration changes in Cr(VI) and Cr(III) during desorption of Cr on SBT biochar by 0.1 M H 2 SO 4 (pH=1) or 0.1 M NaOH ( pH=13) was determined (Figure 3 6). It is well known that solution pH greatly affects the stability and speciation of Cr(VI) and Cr(II I). Cr(III) is predominantly present as Cr 3+ at pH=1 whereas ~ 25% as Cr(OH) 3 and ~ 75% as Cr(OH) 4 ( Richard and Bourg, 1991 ) At pH range of 6 12, Cr(OH) 3 is also stable. On the other hand, Cr(VI) is predominantly p resent as

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52 CrO 4 2 positively charged during desorption using H 2 SO 4 and negatively charged during desorption using NaOH. Hence, desorption of Cr from SBT biochar was complicated, which was affected b oth by the biochar charges and Cr species. Figure 3 6. Concentrations of Cr(VI) and Cr(III) in solution during Cr desorption from SBT biochar by 0.1 M NaOH (a) and 0.1 M H 2 SO 4 (b). During NaOH de sorption, both Cr(VI) and Cr(III) were detected (Figure 3 6a). The initial Cr(VI) and Cr(III) concentration were 0.87 and 1.9 mg/L. The initial appearance of Cr(VI) and Cr(III) indicated the presence of both Cr species on the a b

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53 biochar. Park et al. ( 2005 ) also found that NaOH was more effective than H 2 SO 4 in desorbing Cr bound on the dead fungal biomass of Aspergillus niger. After 5 h, Cr(VI) in the solution is higher than Cr(III), and after 10 d, Cr bound on the biomass is all eluted as Cr(VI). In our case, the concentrations of both Cr(VI) and Cr(III) increased with time, reaching 21.7 and 14.2 mg/L after 24 h (Figure 3 6a). The increase of Cr(VI) in the solution could be explained by the oxidation of Cr(III) on the biochar under alk aline condition, which was then released into solution as Cr(VI). Similarly, during H 2 SO 4 desorption, both Cr(VI) and Cr(III) appeared in the solution, but at much higher concentrations (7.1 and 3.5 mg/L) than those from NaOH desorption. However, Cr(VI) d isappeared completely af ter 4 h (Figure 3 6b). The initial appearance of Cr(VI) and Cr(III) again suggested the presence of both species on the biochar. The Cr(VI) was subsequently reduced to Cr(III) under the strong acidic environment ( Kratochvil et al. 1998 ) Based on these results, it could be concluded that both Cr(VI) and Cr(III) existed on the SBT biochar. Mechanisms of Cr(VI) Removal b y SBT Biochar We hypothesized that SBT biochar effectively removed Cr(VI) via electrostatic attraction of Cr(VI) coupled with Cr(VI) reduction to Cr(III) and Cr(III) complexation mechanisms ( Park et al. 2005 ; Kousalya et al. 2010 ) Firstly, under strongly acidic condition, the negatively charged Cr(VI) species were migrated to the positively charged surfaces o f SBT biochar (protonated carboxylic, alcohol and hydroxyl groups; Fig ure 3 2; Table 3 1) with the help of electrostatic driving forces; Secondly, with the participation of hydrogen ions and the electron donors from SBT biochar, Cr(VI) was reduced to Cr(II I) (Fig ure 3 5); Finally, part of the converted Cr(III) was released to the medium (Fig ure 3 3), with the rest being bonded with function groups through complexation

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54 (Fig ure s 3 2, 3 5). Furthermore, besides biochar particles, the dissolved organic carbon r eleased from biochar was also responsible for Cr(VI) reduction. Research Findings SBT biochar was effective in remov ing Cr(VI) from aqueous solution. The Cr(VI) removal mechanisms included electrostatic attraction of Cr(VI) coupled with Cr(VI) reduction a nd Cr(III) complexation. FTIR analysis suggested that carboxylate and hydroxyl groups may be involved in Cr removing process. The optimum pH for Cr sorption was 2 with 98% removal of Cr(VI). The Cr(VI) sorption increased from 19.8 to 88.5% as the biochar c ontent increased from 0.2 to 8.0 g/L. The Cr(VI) sorption process can be described by the pseudo second order model (R 2 >0.99) and reached equilibrium after 16 h. The sorption data can be fitted to the Langmuir model (R 2 =0.99) and the maximum sorption capac ity was 137 mg/g. Desorption of Cr from SBT biochar was 47 and 32% by 0.1 M NaOH and 0.1 M H 2 SO 4 Based on our data, SBT biochar may serve as a potential alternative for removal of toxic Cr(VI) from aqueous solution.

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55 CHAPTER 4 MECHANISTIC INVESTIGATION O F MERCURY SORPTION BY BRAZILIAN PEPPER BIOCHARS OF DIFFERENT PYROLYTIC TEMPERATURES BASED ON X RAY PHOTOELECTRON SPECTROSCOPY AND FLOW CALORIMETRY Introduction Mercury (Hg) is one of the most toxic trace elements in the environment ( Gosar et al. 1997 ) Due to its toxicity to living organism at low concentrations, Hg contamination has received great attention in recent years. Exposure to Hg can damage the nervou s system in humans, especially for the developing nervous system of young children ( Yuan et al. 2011 ) Mercury is released to the environment by various industries including chlorakali, paint, pulp and paper, and oil refining ( Namasivayam and Kadirvelu, 1999 ) Due to lack of knowledge and regulation, factory effluents containing Hg were commonly released into the environment in the past. To protect public health, ( Wase and Forster, 1997 ) for drinking water ( EPA, 2001 ) It is therefore of great importance to reduce Hg contamination in the environment. Brazilian pepper (Schinus terebinthifolius) is the most aggressive evergreen shrub like tree in Florida. It invades many habitats by forming large dense forests ( FDEP., 2008 ) ecosystems, and several health and safety problems because it is a relative of poison ivy and can ca use dermatitis. Now it covers more than 700,000 acres in south and central Florida, as well as many of the islands on the east and west coasts of the state ( FDEP., 2008 ) Recent developments of renewable bioenergy technologies make it

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56 possible to convert this invasive plant into value added biochar and at the same time produce bioenergy. Biochar has affinity for heavy metals and their sorption capacity is comparable with other biosorbents. For example, the sorption capacity of sugarcane bagasse biochar for Pb i s 20.5 mg/g ( Ding et al. 2010 ) sugar beet tailing biochar for Cr is 123 mg/g ( Dong et al. 2011 ) and pine bark biochar for As is 12.2 mg/g ( Mohan et al. 2007 ) Though little information is available on Hg sorption by biochar, its sorption by various biomaterials has been investigated. Das et al. ( Das et al. 2007 ) proposed that complexation through functional groups such as carboxyl and hydroxyl groups is the main mechanism for Hg removal by Aspergillus versicolor biomass. Cation exchange is involved in Hg removal by activated sludge biomass ( KilI et al. 2008 ) which is supported by higher release of alkali metal ions (Ca 2+ Na + and K + ) during Hg removal. Flow calorimetry is a powerful too l to study mechanism of metal sorption on different solids ( Appel et al. 2002 ; Harvey et al. 2011 ) It provid es a direct measure of the heat of reaction on a liquid/solid interface ( Appel et al. 2002 ) It has several advantages over conventional batch experiments, including separating reactions occurring simultaneously but at different rates, applying multiple sorption/ desorption cycles to the same sample, and distinguishing between reversible and irreversible processes ( Kabengi et al. 2006 ) For example, Cao et al. ( Cao et al. 2004 ) identified that Cu a nd Zn sorption onto phosphate rock was through cation exchange since the heat release of K/Ca exchange before and after Cu and Zn sorption were unchanged. However, by comparison, the heat release for K/Ca exchange after Pb sorption was half of that before Pb sorption, consistent with an irreversible precipitation of

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57 uoropyromorphite ( Cao et al. 2004 ) To our knowledge, flow calorimetry has not been applied to study Hg sorption on biochar. In addition to flow calorimetry, FTIR and XPS are also powerful tools in identifying functional groups and the surface speci ation involved in metal sorption. For example, carboxyl and hydroxyl groups were identified to be involved in Cr(VI) removal by sugar beet tailing biochar by FTIR analysis and the XPS analysis confirmed both Cr(VI) and Cr(III) were present ( Dong et al 2011 ) In this study, biochars were produced from Brazilian pepper (BP) via slow pyrolysis using a laboratory investigate the characteristics and mechanisms of Hg sorption by BP biochars using various techniques. The specific objectives were to (1) investigate the effecti veness of BP biochars in Hg sorption under different conditions, and (2) determine the mechanisms governing Hg sorption by BP biochars using flow calorimetry, SEM EDX, FITR and XPS. Materials and Methods Biochar Preparation and Characterization Fresh Braz ilian pepper plant collected from Florida, U.S. was oven dried at scale tubular reactor within a muffle furnace (1500M, Barnstead Intl. Corp, IA). Nitrogen gas at 10 psi was used to maintain an oxygen free environment. The fur nace temperature was programmed to increase at a rate of procedure took ~ 2 h. The biochars were allowed to cool at room temperature under a flow of N 2 gas. They were the

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58 tight plastic containers prior to use and will be referred as BP300, BP450 and BP600. The surface morphology of BP biochars were examined using JSM 6400 scanning electron microscope (SEM, JEOL USA, Peabody, MA) at 15 keV equipped with energy dispersive X ray spectroscopy (EDS, Oxford Instruments USA, Concord, MA). Evaluation of its zero point of charge (pHZPC) was based on the mass titration procedure of Valds et al ( 2002 ) The elemental composition of the BP biochars were determined by inductively coupled plasma atomic emission spectroscopy (ICP AES, Plasma 3200, Perkin Elmer Crop, MA) after digesting with HNO 3 /H 2 O 2 ho t block digestion procedure ( EPA 1986 ) Elemental C, N, and H concentrations were determined using a CHN Elemental Analyzer (Carlo Erba NA 1500) via high temperature catalyzed combustion followed by infrared detection of the resulting CO 2 H 2 and NO 2 gases. Cation exchange capacity (C EC) was determined as described by Gaskin et al. ( Gaskin et al. 2008 ) 0.2 g of biochar was leached with 20 mL of deionized water five times, and the leachates were collected together. The K + Na + Ca 2+ and Mg 2+ in the leachate were determined as the soluble base cations of the biochar. After the fifth run of the leaching with deionized water, the biochar was leac hed with 20 mL of 1 M Na acetate (pH 6 ) five times, and the leachates collected together. The K + Na + Ca 2+ and Mg 2+ in the leachates were determined as the exchangeable base cations. The biochar samples were then washed with 20 mL of ethanol five times to remove the excessive Na + Afterwards, the Na + on the exchangeable sites of the biochar was displaced by 20 mL of 1 M NH 4 acetate (pH 6 ) five times, and the CEC of the biochar was calculated from the Na + displaced by NH 4 The Ca 2+ Mg 2+ K + and Na + in the

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59 leachates were determined using the same methods described above ( Gaskin et al. 2008 ) The content of carboxyli c and phenolic hydroxyl groups was determined by Bohem titration method ( Boehm, 1966 ) Hg Sorption Isotherm b y Biochars The stock solution of 1,000 mg/L Hg was prepared by dissolving analytical grade Hg(NO 3 ) 2 in deionized wate r. All working solutions were freshly prepared before use by diluting the stock solution with deionized water, and the pH was adjusted to 6 with 1 M HNO 3 or KOH solution. Batch experiments were conducted using 60 mL polyethylene bottles with sample volume of 50 mL, and Hg concentrations of 1, 5, 10, 20, and 50 mg/L. The concentrations were within the range found in industry wastewater samples ( Monteagudo and Or tiz, 2000 ; Manohar et al. 2002 ) The samples were agitated on a volume of solution was removed and filtered with The effect of pH on Hg sorption by BP biochars was investigated at pH from 2.0 to 8.0 while keeping Hg at 20 mg/L and biochar at 2 mg/L. Fourier Transform Infrared (FTIR) and X ray Photoelectron Spectroscopy (XPS) Anal ysis BP300, BP450 and BP600 biochar samples before and after reacting with 50 mg/L Hg for 24 h at pH 6.0 and 2 g/L biochar were analyzed by FTIR and XPS. FTIR spectroscopy (Bruker Optics Inc., Billerica, MA) was used to determine changes in functional grou ps on the biochar after Hg sorption. The spectra were obtained from 500 to 4000 cm 1 with a resolution of 8 cm 1 The resulting spectra were the average of 32 scans. The valence state of the Hg sorbed onto BP biochars was determined by XPS

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60 (Perkin Elmer Model 5100, USA). The sample was analyzed at an Al X excitation at 14 86.6 eV) at 100 W and pass energy of 0.1 eV, with 10 high resolution scans. The system was operated at a base pressure of 210 8 mbar. The calibration of the spectra binding energy was performed with the C 1s peak of the aliphatic carbons at 284.6 eV. Flow Calorimetry Experiments Flow calorimetry was used to study Hg sorption characteristics onto biochars. The flow calorimeter was made at University of Florida and its detailed operation was described by Rhue et al ( Rhue et al. 2002 ) All experiments were carried out at a solution flow rate of 0.30 0.35 mL/min, and ionic strength of 30 mM at pH 6.0 with 25 40 mg of biochar. Prior to the sorption experiments, biochar samples were hydrated in 30 mM KNO 3 packed into the c alorimeter, and flushed with 30 mM KNO 3 until a stable baseline (indicating complete K saturation and equilibrium) was obtained. After a satisfactory baseline was obtained with 30 mM KNO 3 equilibration, solution of 15 mM Ca(NO 3 ) 2 was then introduced to the K saturated BP biochar to determine the heat exchange of Ca for K. The K and Ca solutions were cycled through the BP biochar samples several times. This was done to check for heterogeneity between BP biochar samples, helping to determine the correction fa ctor for differences in BP samples, which was based on exchange thermodynamics instead of BP weight. Furthermore, this ensured that the system to was function ing normally when K and Ca peak areas were of equal and opposite magnitude, indicating an exchange process. Once the K/Ca cycle reached equilibrium, 29.5 mM KNO 3 solution containing 50 mg/L Hg as Hg(NO 3 ) 2 was introduced to the K saturated BP biochar samples to obtain

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61 reaction heat of Hg for K. The system was then flushed with 30 mM KNO 3 to remove the s orbed Hg on the biochar. All three BP biochar samples underwent several cycles of K/Ca solution before and after reaction with Hg. This was done to compare the energetics of K/Ca exchange before and after Hg sorption to ascertain the changes in the energet ics was due to Hg introduction. The biochar residues were then removed from the flow calorimeter column, washed with deionized H 2 O, air dried, and digested in HNO 3 /H 2 O 2 with USEPA Method 3050. The amount of Hg remained on the biochar was via chemisorption as they were not exchangeable with 30 mM KNO 3 Mercury and Statistical Analysis The concentrations of Hg were analyzed by hydride generation atomic fluorescence spectrometry (HG AFS, AI3300, Aurora Instruments Ltd, Vancouver, BC). The detection limit of wa s 0.3 ng/L 1 All experiments were conducted in triplicate and the results were calculated as the mean values of the samples. Data were analyzed using JMP, version 9. R esults a nd Dis cu ssion Biochar P roduced at Low Temperature Contained More Functional Grou ps The N content was relatively stable in all three biochars whereas H, C and O conte nt varied significantly (Table 4 1). With increasing temperature, C content increased whereas H and O decreased. The progressive decrease in atomic ratio of H/C (4 times) and O/C (7 times) with temperature indicated condensation of the biochar structure ( Keiluweit et al. 2010 ; Harvey et al. 2011 ) The carboxylic group and phenolic hydroxyl group content were 3.15 and 14.3 cmol c /kg for BP300 biochar, 1.82 and 10.1 cmol c /kg for BP450 biochar, 1.79 and 4.67 cmol c /kg for BP600 biochar (Table 4 2). Their

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62 contents decreased with i ncreasing pyrolysis temperature, indicating loss of oxygen containing groups. Disappearance of function groups in biochars may coincide with dehydration and rearrangement of molecules at high temperature ( Harvey, Herbert et al. 2011 ) Elemental analysis showed that BP biochars contained substantial amounts of cations including Ca, K, Na, Mg, Al and Fe and they were mostly present as metal hydroxides, with higher temperature resulting i n higher concentrati ons (Table 4 2). Among all cations, Ca concentration was the highest, ranging from 2.56 to 19.3 mg/g (Table 4 2). The surface characteristics of BP biochars based on SEM before and after reacti on with Hg are shown in Figure 4 1. Consistent with the chemica l analysis, the EDS spectra also indicated substantial amount s of Ca in BP biochars (Figure 4 1). Table 4 1. C, N, H composition and atomic ratios of biochar samples derived from Brazilian pepper (BP) under different pyrolysis temperatures (300, 450 and 6 00 C). Biochar Component a % Atomic ratio a C N H O b H/C O/C N/C BP300 50.83.32 c 1.880.01 4.330.11 42.53.21 1.02 0.63 0.03 BP450 75.65.13 1.520.02 3.540.09 17.20.85 0.56 0.18 0.02 BP600 84.84.38 1.670.01 2.030.07 8.460.09 0.29 0.09 0.0 2 a : Values are on ash free basis b : Determined by mass difference assuming sample mass was made up of the tested elements in both Table 4 1 and Table 4 2. c : Data were expressed as mean standard error of mean. BP biochar samples were also characterize d by FTIR spectroscopy (Figure 4 2). Pyrolysis temperature impacted the functional groups on the biochar, with biochars produced at lower temperature showing more functional groups. The BP300 biochar spectra displayed a number of peaks indicat ing its compl ex nature (Figure 4 2a). The

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63 broad peak at 3428 cm 1 represented the bonded hydroxyl of phenol group. The peak around 2912 cm 1 was indicative of aliphatic CH stretching ( Kapoor and Viraraghavan, 1997 ) The peaks at 1702 and 1509 cm 1 were related to C=O stretching of carboxyli c group ( Noguchi and Sugiura, 2003 ; Elangovan et al. 2008 ; Gupta and Rastogi, 2008 ) The peak at 1615 and 1111 cm 1 corresponded to C=C stretch and C O C bonds associated with cellulose, hemicellulose, and lignin in wood ( Blzquez et al. 2010 ) Table 4 2. Properties of biochar samples derived from Brazilian pepper (BP) under different pyrolysis temperatures (300, 450 and 600 C). Metals BP300 BP450 BP600 Ca (mg / g) 2.560.14 a 12.30.4 5 19.30.47 K (mg / g) 0.660.03 1.800.06 2.210.06 Na (mg / g) 1.280.08 5.220.16 7.280.16 Mg (mg / g) 0.420.03 1.380.05 1.370.03 Al (mg / g) 0.090.001 0.130.003 0.100.007 Fe (mg / g) 0.090.005 0.240.003 0.390.008 Point zero charge 6.650.08 8.35 0.11 9.910.09 Cation exchange capacity (CEC:cmol c /kg) 2.830.04 1.510.01 1.570.01 Carboxylic group ( cmol c /kg ) 3.15 0.0 2 1.82 0.0 1 1.79 0.0 1 Phenolic hydroxyl group 14.3 0.0 6 10.1 0.0 3 4.67 0.0 7 ( cmol c /kg ) Particle size ( m) a : Data were expressed as mean standard error of mean.

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64 Figure 4 1. before (a) and after (b) reaction with 50 mg/L Hg for 24h at pH 6.0. Compared to BP300 biochar, BP450 biochar had fewer peaks, with the peaks at 1615 and 1111 cm 1 (C=C stretch and C O C bond) being reduced while those at 2912, 1702, and 1509 cm 1 ( CH and carboxyli c groups) disappearing (Figure 4 2b). A new peak between 885 and 750 cm 1 representing aromatic C H groups was consistent with condensation of biochar structure. BP600 biochar had the fewest peaks, with all peaks being reduced. Appearance of three aromatic C H bands between 885 and 750 cm 1 confirmed further co ndensation of biochar structure. Disappearance of O containing a b

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65 functional groups and increase of aromatic structure with temperature were clear evidence of condensation of biochar. These results confirmed that most function groups of lignocellulosic materi als were lost during pyrolysis at high temperature ( Bilba and Ouensanga, 1996 ) Figure 4 2. Fourier transform infrared spectra of BP biochars (a, b, c) before and after reaction with 5 0 mg/L Hg for 24 h at pH 6.0. a b c

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66 XPS was also used to analyze the functional groups present on the outer surface of BP biochars (Figure 4 3). For all BP biochars, the C 1s spectra at 284.6 eV consisted of three peaks: graphitic and aromatic carbon at 284.6 eV, phenolic hydroxyl or ether groups at 286 eV, and carboxylic or ester groups at 288.6 eV; the O 1s peak included oxygen carbonyl or quinone at 530.7 eV, oxygen in hydroxyl or ester at 532.1 eV and oxygen in anhydride, lactone or carboxylic acids at 533.3 eV ( Valds et al. 2002 ) A Ca 2p peak was observed on BP 600 biochar, which was consisted with its relative high Ca content. It was interesti ng to note the absence of N signal, which indicated the presence of N in the inner part of biochar since XPS is sensitive to surface 5 10 nm ( Renme et al. 2010 ) Figure 4 3. XPS spectra of survey scan of BP biochars.

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67 Hg S peciation Impacted Its S orption by Biochars at D ifferent pHs The Hg sorption capacity of biochars was deter mined by Langmuir model (Table 4 3) It decreased from 24 to 15 mg/ g as pyrolysis temperature increased, implying that low temperature was beneficial to form biochars with higher Hg sorption capaci ty, which was related to more functional groups on the biochars as supported by FTIR data (Figure 4 2). The sorption capacity of BP300 and BP450 biochars was within the range calculated based on carboxylic and phenolic hydroxyl content assuming monodentate or bidentate Hg complex ( Lv et al. 2012 ) Hence it was possible that both monodentate and bidentate Hg complexes were present on biochars. However, the sorption capacity for BP600 biochar w as higher than the calculated value based on monodentate Hg complex, indicating different mechanism from low temperature biochars. Table 4 3. Parameters of Langmuir model for Hg sorption onto biochars derived from Brazilian pepper (BP) under different pyr olysis temperatures (300, 450 and 600 C). Sorbents Q 0 a (mg g 1 ) K L b (L mg 1 ) R 2 BP300 24.2 0.015 0.962 BP450 1 8 8 0.070 0.978 BP600 15.1 0.053 0.961 a : Langmuir constant related to the maximum quantity of metal ions per unit weight of biomass b : Lang muir constant related to the affinity of binding sites with the metal ions. Mercury sorption by biochars depended on the interactions between Hg ions and the functional groups on biochar, which was impacted by solution pH. The starting Hg concentration w as 20 mg/L, which was high enough to provide sufficient Hg to saturate biochars and low enough to maintain Hg soluble in pH range of 2 8 ( Walcarius and

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68 Delacte, 2005 ) Hg sorption by biochars increased with increasing pH, reaching a pl ateau at pH of 4.0 8.0 (Figure 4 4). At pH 2.0, the amount of Hg sorbed was 3.2 7.0, 2.6 4.6, 2.1 3.0 mg/g and at pH 4 .0 8 .0 it was 8.7 9.0, 6.1 7.8 and 6.3 7.0 mg/g for BP300, BP450 and BP 600, respectively. Based on ana lysis using the speciation model MINEQL+, at pH of 2 3, Hg(OH) + and Hg 2+ were the dominant species in the solution, accounting for >70%. At this pH, biochars were positively charged as their pH PZC was high (6.65, 8.35 and 9.91; Table 4 2 ). The low Hg sor ption at this pH was attributed to the positive charge of the sorbent surface (< pH ZPC ) as well as the competition from H + in the solution. This was consistent with the pK a values of 3.5 5.0 for carboxylic group ( COOH) and around 10.0 for phenolic group ( C OH) ( Lv et al. 2012 ) Hence, at acidic solution, less Hg was sorbed with protonated carboxylic and phenolic hydroxyl groups on biochars. Figure 4 4. Influence o f pH on Hg removal by BP biochars after reacting with 20 mg/LHg for 24 h.

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69 2 was the dominant species, accounting for 95% of aqueous Hg. At pH 6 at which most experiments were conducted, it increased to 99.9% At this pH range, carboxylic and phenolic hydroxyl group on biochars became more deprotonated ( COO an d C O ), which may complex with Hg(OH) 2 forming COOHg + and C OHg + With less electrostatic repulsion from biochars at pH range 4 8, complexation of Hg(OH) 2 with surface functional groups like carboxylic and phenol groups was likely more favored which was consistent with the higher sorption capacity at this pH range Mercury sorption onto 2 mercaptobenzimidazole impregnated clay and sewage sludge derived activated carbon were also increased with increasing pH reaching a plateau at the pH range of 4.0 8. 0 ( Manohar et al. 2002 ; Zhang et al. 2005 ) Carboxylic and Phenolic Hydroxyl Groups Were Responsible for Hg Sorption Due to the similar Hg sorption capacity at pH 4 8, pH 6.0 was selected for all experiments. The FTIR spectra of BP biochars before and after Hg sorption were used to determine the vibration frequency changes in function groups. Hg sorption cha nged the peaks of functional groups in all biochars. The disappearance of carboxyl stretching band at 1702 cm 1 and C O C band at 1111 cm 1 was noted in Hg loaded BP 300 biochar (Figure 4 2a), typical of interaction of carboxyl group with metal ions ( Noguchi and Sugiura, 2003 ; Elangovan et al. 2008 ) Further, the peak intensity at 3428 cm 1 due to bonded hy droxyl group decreased, indicating the involvement of hydroxyl groups in Hg sorption (Figure 4 2a) ( Blzquez et al. 2010 ) Compare d to BP300 biochar, the impact of Hg sorption on FTIR spectra of BP450 and BP600 biochars were limited (Figure 4 1b & c), consistent with their low Hg sorption (13.2 16.4 mg/g) compared with BP300 (21.3 mg/g). For BP450 biochars, the decrease of both carbo xyl and hydroxyl

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70 from phenol groups peaks was observed, typical of their interactions with metal ions. These results are also confirmed by other studies, which concluded that carboxyl and hydroxyl groups are the main functional groups for metal sorption ( Jianlong, 2002 ; Das et al. 2007 ) However, for BP600 biochar, instead of phenolic hydroxyl, the decreased of C =C was observed, indicating a different sorption mechanism from BP300 and BP450 biochar. Table 4 4. Surface composition for Hg loaded BP biochars based on XPS analysis Element Functional Groups Binding Energy (eV) Atomic Content (%) BP300 BP450 BP600 C 1s Total 7 9.9 82.5 88.6 graphitic aromatic(C C) 284.6 71.4 69.7 62.5 C in hydroxyl, ethers 286.0 23 25.8 33.7 C in carboxyl 288.6 5.60 4.5 3.80 O 1s Total 18.9 16.4 11.0 carbonyl ,quinone 13.9 6.5 3.0 h ydroxyl or ethers 532.1 52.7 51.2 53 .5 anhydride, lactone, carboxylic acids 533.3 33.4 42.3 43.5 Hg 4f 1.1 0 .7 0.4 phenolic Hg 101.1 76.9 69.5 0 carboxylic Hg 101.3 23.1 30.5 9.0 graphite Hg 102.0 0 0 91

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71 The XPS analysis of BP biochars after Hg sorption was used to gain a better understanding of Hg sorption mechanisms. The presence of Hg 4f clearly confirmed the presence of Hg on biochar, which corresponded to the decrease of O 1s peak area (Figure 4 5), suggesting the involvement of oxygen containing functional groups in Hg sorptio n. After Hg sorption, total oxygen content on biochar surface decreased by 2.2%, 1.7% and 1.1% on BP300, BP450, and BP600, respectively (Table 4 4) The high resolution spectra of Hg 4f on BP biochars could be fitted with two doublet peaks, both separated by 4.10.1eV (the distance between Hg4f 7/2 and 4f 5/2 peaks) (Figure 4 6). The absence of Hg4f 7/2 peaks at bending energy < 100.6 eV indicated that no elemental Hg was formed, which was at 99.9 eV ( Hutson et al. 2007 ) Hg4f 7/2 for the two peaks observed on BP300 and BP450 biochars were situated at 101.1 and 101.3 eV. Based on peak area r atios, surface atomic ratios of Hg 101.1 to Hg 101.3 were 3.33 and 2.28 for BP300 and BP450 biochar, which indicated the molar ratio of these two Hg species. Bond et al ( Bond et al. 2000 ) attributed two Hg4f 7/2 peaks to two different Hg sorption sites on H 2 SO 4 /KMnO 4 treated carbon nanotube: the peak at 101.1 eV was assigned to phenol group while the one at 101.3 eV to carboxylic group, and the ratio of these two Hg sorption sites was 1.86. However, the two Hg4f 7/2 doublet peaks appeared at 101.3 and 102.0 eV for BP600 biochar, and the atomic ratio was 0.1. The domain of typical bending energy for Hg4f 7/2 was 99.9 101.6 eV, the peak at 102.0 eV was outside this range ( Wang et al. 2009a ) This could be caused by the formation of Hg C like structure (C=C), which was a stack of flat aromatic (graphene) sheet crosslinked in a random manner ( Harvey et al. 2011 ) Hg(OH) 2 is a softer acid than metal cations and as a rule, the interaction of

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72 Hg(OH) 2 with graphite like structure (soft bases) is likely favore d. In addition, an abundance of delocalized lone graphite like domains of plant derived biochars ( Harvey et al. 2011 ) It is the lone electrons likely formed Hg lik e structure had been reported to be responsible for Hg sorption on hybrid ligand modified activated carbon ( Zhu et al. 2009 ) Figure 4 5. XPS spectra of survey scan of BP biochars after reaction with 50 mg/L Hg for 24 h at pH 6.0.

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73 Figure 4 6. XPS spectra of high resolution scan of Hg4f for BP biochars after reaction with 50 mg/L Hg for 24 h at pH 6.0.

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74 Hg Sorption onto Biochars Was Probably Via Complexation w ith Functio nal Groups The FTIR and XPS data indicated that functional groups such as carboxyl and hydroxyl groups were responsible for Hg sorption by biochars. In addition to complexation by functional groups, other mechanisms such as cation exchange might have parti cipated in Hg sorption onto biochars. To further determine the role of cation exchange and complexation in Hg sorption, flow calorimetry experiment was conducted. Heat change based on flow calorimetry indicated that Ca sorption onto K saturated biochars w as endothermic (Figure 4 7a c). An exothermic signal with a slightly different shape but similar area corresponding to Ca displacement by K was observed when K re saturated the biochars. The two equal peaks indicated K/Ca exchange was an exchangeable react ion. The heat associated with the peaks was 87.4, 76.4, and 48.3 mJ/g for BP 300, BP 450 and BP 600 biochar. Since K and Ca sorption was exchangeable, the amount of K and Ca sorbed on the biochar could be used to represent the CEC of biochars. The molar he at of sorption ( or ) was calculated as the ratio of amount of heat produced to the amount of K or Ca sorbed, which corresponded to 3.09, 3.07, and 3.08 kJ/mol c for BP 300, BP 450 and BP 600 biochar. The constant molar heat of sor ption of K/Ca onto biochar suggested similar mechanism of K/Ca sorption onto the three biochars. Furthermore, the relative low value of and also confirmed K and Ca sorption on biochar was a physical process ( Gueu et al. 2007 ) The molar heat of K sorption on plant derived biochar was 3.2 7 .9 kJ/mol c independent of biochar pyrolysis temperature ( Harvey et al. 2011 ) The primary sources of CEC on biochar included carboxylic group, hydroxyl in phenol group and metal hydroxides ( Harvey et al. 2011 ) At pH 6.0, since phenol grou p was a we a k acid

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75 with pKa of 10, Ca or K would not react with phenol to produce acid ( Ahmady Asbchin et al. 2008 ) The contribution o f metal hydroxide was also likely small since they typically have pH ZPC > 7. Hence, deprotonated carboxylic group was likely the primary source of CEC on biochar. Furthermore, the carboxyl group content in BP biochars was close to the CEC value (Table 4 2) indicating that carboxylic group was probably the major cation exchange sites for K/Ca exchange on biochars. Similar conclusion was made by Harvey et al ( Harvey et al. 2011 ) on plant derived biochars. The heat of exchange for K/Ca after Hg sorpt ion decreased by 41%, 70% and 100% for BP300, BP450 and BP600 biochar, consistent with a non exchangeable Hg sorption onto biochars (Figure 4 7d f). The data also indicated that similar amount of charge at 0.12, 0.11, and 0.16 mmol c /kg for BP300, BP450 and BP600 biochar was consumed during Hg sorption ( ). The smaller heat signals suggested that Hg sorption occurred on the same cation exchange sites and this process was not exchangeable with K and Ca. Stability constants of Ca, K, and Hg with ca rboxylic group on biochar are not available, but the stability constant of carboxyl group in citric acid are available, which were log k=2.3, 1.4 and 11 ( Martell et al. 1998 ) The much greater value for Hg than K/Ca (~5 8 fold) indicated that Ca or K could not replace Hg. Cao et al ( Cao et al. 2004 ) also observed the heat signal for K/Ca exchange after Pb sorption on phosphate rock decreased by 50%, which attributed to an irreversible reaction Bond et al ( Bond et al. 2000 ) attributed Hg sorption onto carbon nanotubes (CNTs) to formation of (CNT COO) 2 Hg. Our resea rch together with literature were consistent with that carboxylic group was responsible for Hg sorption. Furthermore, the decreased but not complete disappear ance of the K/Ca heat signal also indicated that one Hg cycle did not occupy

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76 all cation exchange s ites on BP300 and BP450 biochar. However, one cycle consumed all sites on BP600 biochar, which might indicate different sorption mechanism. Figure 4 7. Heat changes for K/Ca sorption at pH 6 before and after Hg sorption onto Brazilian peppe r biochars produced at 300 (a and d), 450 (b and e) and a b c d e f

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77 Figure 4 8. Heat changes of Hg sorption at pH 6 by BP biochars produced at 300 (a), 450 (b) a b c

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78 After exposing to one Hg cycle and the signal returned to baseline, biochar samples were returned to a K cycle (Figure 4 8). Comparison of Hg sorption response to those for K/Ca exchange and coupled with the essentially undetectable heat si gnal of K replacing Hg suggested that Hg sorption occurred via a different mechanisms than simple reversible cation exchange. Mercury sorption was also kinetically different from cation exchange. Typically, K/Ca exchange reached completion within 10 min, w hile Hg sorption lasted o ver 20 min (Figure 4 7 & 4 8). Kinetically different phosphate sorption versus K/Ca exchange o n a Ultisol was also observed. K/Ca exchange on a Ultisol was rapid and took ~10 min for the heat signal to return to baseline, but phosp hate reaction, which was explained by rapid ligand exchange and slower diffusion limited secondary reactions continued after 10 min ( Rhue et al. 2002 ) Table 4 5. Heats of sorpti on and quantity of Hg sorbed after reaction with 50 mg/L Hg(NO 3 ) 2 solution at pH 6.0 during flow calorimetry and batch sorption experiments. Sample m g /g Heats ( kJ/mol) Flow a Batch b BP300 2.61 21.3 19.67 BP450 2.81 16.4 18.28 BP600 14.8 13.2 25.36 a Flow rate = 0.30 0.35 mL/min until reaching baseline b Biochar : solution ratio= 2 g : 1 L shaken for 24 h The molar heat of for Hg sorption onto biochars was calculated based on the slope of the cumulative en ergy released and the amount of Hg adsorbed, which were 19.7, 18.3 and 25.4 kJ/mol for BP300, BP450 and BP600 (Table 4 5). These values were consistent with reported values for Hg sorption onto activated carbon and

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79 carbonized sugi wood, which were 23.6 and 19.4 kJ/mol. ( Mohan et al. 2000 ; Pulido Novicio et al. 2001 ) The was 3 kJ/mol f or BP biochars, which was significantly lower than value. This indicated that Hg bonded more tightly to the biochars surface via a non exchangeable process ( Harvey et al. 2011 ) The amounts of Hg sorbed by BP300, BP450 and BP600 bioch ars were 2.61, 2.81, and 14.8 mg/g Hg ( ), accounting for 13%, 17% and 112% of the Hg sorbed by batch experime nt with 50 mg/L at pH 6 (Table 4 5). Figure 4 9 showed the heat signals of reaction obtained in the first and second Hg treatment cycl e for BP450 biochar they were comparable. The amount of sorbed Hg after two Hg cycles was 5.32 mg/g, which was much almost as twice as injecting one cycle of Hg. The data implied Hg did not react with all sorption sites after one Hg cycle for BP300 and BP 450 biochar. Rhue et al ( Rhue et al. 2002 ) observed that phosphate saturation in a Ultisol soil was completed in five 40 min cycles. However, for BP600 biochar, all sorption site s were occupied after one cycle, coupled with the longer time required for the heat signal to return to base line (~50 min, twice as BP300 and BP450), which indicated that Hg sorption onto these sites was different from that in BP300 and BP450 biochars. Th is result was confirmed by the appearance of a new Hg 101.9 (graphite like structure) peak during XPS analysis on BP600 biochar. That Hg sorption occurred via complexation onto different functional groups on biochar produced at different temperature was con sistent with the conclusions of Harvey et al ( Harvey et al. 2011 ) They concluded that sorption of Cd 2+ on plant derived biochars produced at <550C occurred via complexation to carboxylic group, while Cd 2+ sorption on biochars produced at >550C o ccurred via interaction between Cd 2+ and graphite like domains on biochar.

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80 Figure 4 9 Heat changes of Hg sorption at pH 6 by BP 450 biochar during the first cycle (a) and second Hg cycle (b) b a

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81 Th e amount of Hg sorbed ( ) onto biochars suggested the involvement of other functional groups in addition to carboxylic groups. If Hg sorption existed solely on the carboxylic groups, the major cation exchange sites on biochar, the ratio of the amount of charges consumed during Hg sorption ( ) to the amount of Hg sorbed ( ) on the biochars can be useful. Lv et al ( Lv et al. 2012 ) used surface complexation mo del and synchrotron X ray spectroscopy to study Hg sorption on lignin, a major component of plant biomass, monodentate and bidentate Hg complexes are formed, with bidentate complex being predominant species at pH>5.5. If these were the case, the mole of charge consumed per mol of Hg sorbed should be 2 for bidentate complexation. However, the ratio of was 0.90, 0.75 and 0.21 for BP300, BP450 and BP600, pointing the sorption also occurred on undeprotonated functional groups, possibly phenol group and graphite like structure These results were consistent with XPS analysis, where Hg 101.3 (phenolic hydroxyl group) and Hg 101.9 (graphite like struc ture) peaks were identified. On this consideration, Hg complexation with phenolic hydroxyl group and graphite like structure were also involved. During Hg sorption, solution pH increased by 0.32 to 0.52 unit, which could be explained by the complexation be tween deprotonated carboxylic group and Hg(OH) 2 the dominant Hg species at pH 6.0. Moreover, base d on the Hg speciation of 50 mg/ L Hg at pH 6, Hg precipitation was unlikely. Thus, Hg sorption on BP biochars might be explained by the following equations: carboxylic and phenolic hydroxyl were the predominantly binding sites for BP300 and BP450 biochars; while carboxylic and graphite like structure were the major binding sites for BP600 biochar (Figure 4 10 ) (4 1)

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82 (4 2) (4 3) Figure 4 10 Mechanisms of Hg removal by BP biochars produced at different temperatures Research Findings This study investigated the mechanisms of Hg sorption by biochars from Brazilian pepper. The abundance of functional gr oups and Hg sorption capacity decreased with the increase of biochar pyrolysis temperature. Effect of pH observed that Hg sorption reached a plateau at pH 4.0 8.0. FTIR analysis suggested that carboxylic and phenol groups were involved in Hg sorption proce ss. XPS analysis

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83 confirmed the Hg species formed on BP biocars. The two Hg4f 7/2 peaks on BP300 and BP450 biochars were situated at 101.1 and 101.3 eV with the atomic ratio of 3.33 and 2.28 respectively, which were related with carboxylic and phenolic hydro xyl groups. However, the two Hg4f 7/2 doublet peaks appeared at 101.3 and 102 eV for BP600 biochar with atomic ratio of 0.10, instead of phenolic hydroxyl group, graphite like structure was form ed. Flow calorimetry experiment suggested that K/Ca sorption on to biochars was through reversible cation exchange on deprotonated carboxylic group with molar heat ( or ) of 3.1 kJ/ mol. The decrease of heat for K/Ca exchange after Hg sorption, and the essentially undetectable heat signal of K displacing Hg indicated that Hg sorption was via a different and irreversible process. The ratio of the amount of charges consumed during Hg sorption ( ) to the amount of Hg sorbed ( ) on the biochars further confirmed that Hg sorpt ion did not solely occur on deprotonated carboxylic group s but other groups also participated. The decrease heat for post Hg K/Ca exchange coupled with less sorbed Hg comparing with the batch experiment (accounting for 13%, 17% and 112% for BP300, BP450 a nd BP600, respectively) implied that after one Hg cycle part of the sorption site s w ere still available for BP300 and BP450 biochar, however, all the sorption sites on BP600 biochar were occupied coupled with twice Hg sorption time, which implied that a d ifferent sorption site graphite like structure was involved. The net cumulative Hg sorption was exothermic and the molar heat was 19.7, 18.3 and 25.4 kJ/ mol for BP300, BP450 and BP600, respectively. Our research suggested that Hg was irreversibly sorbed via complexation with phenolic hydroxyl and carboxylic groups in low temperature biochars (BP300 and BP450) and graphite like structure in high temperature biochar (BP600).

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84 CHAPTER 5 ENHANCED CR(VI) REDUCTION AND AS(III) OXIDATION IN ICE: IMPORTANT ROLE OF DISSOLVED ORGANIC MATTER FROM BIOCHAR Introduction Arsenic and chromium are toxic to humans and animals and are of global environmental concerns. Chromium is released into the environment by various industries including electro plating, chromate manuf acturing, leather tanning and wood preservation ( Papp, 2001 ; Dnmez and Aksu, 2002 ; Kim et al. 2002 ) In the environment, Cr exists primarily as Cr ( VI ) and Cr ( III ) and Cr ( VI ) is of significant concern due to its carcinogenic ity ( Fendorf et al. 2000 ; Costa, 2003 ) Arsenic is ubiquitous in the sources, including industrial waste produ cts, agricultural pesticides and wood preservatives ( Ong et al. 1997 ; Wilkie and Hering, 1998 ) The predominant forms of As in the environment incl ude As ( V ) and As ( III ) As ( V ) is mainly present in aerobic soils as H 2 AsO 4 and HAsO 4 2 while As ( III ) is present in sediments as H 3 AsO 3 0 at pH below 9.2 ( Korte and Fernando, 1991 ; Masscheleyn et al. 1991 ) As ( III ) is more toxic and mobile in the environment than As ( V ) ( Manning and Goldberg, 1997 ) Therefore, Cr ( VI ) reduction and As ( III ) oxidation are desirable to reduce their adverse environment impact. Dissolved organic matter (DOM) in soils is known to serve as both an electron donor and acceptor during redox reaction, which is a n important factor influencing Cr and As biogeochemistry in the environment ( Wittbrodt and Palmer, 1997 ; Ko et al. 2004 ) Redman et al. s howed that DOM samples (10 mg C / L pH 6.0) from Brazil and the US were able to oxidize 25 40 g / L As ( III ) to As ( V ) within 90 h ( Redman et al.

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85 2002 ) However, the mechanis ms of these oxidation reactions were not investigated. Jiang et al ( Jiang et al. 2009 ) s uggested that semiquinone radicals were the main electron accepting moieties in a model DOM AQDS (9,10 anthraquinone 2,6 disulfonic acid) for As ( III ) oxidation. The As ( III ) oxidation reaction strongly depend s on semiquinone radical content with AsIII oxida tion rate being increased from 13% to 67% when the semiquinone radicals increased from 2.5 10 20 to 5.6 10 18 spins / L. In addition, reduction of Cr ( VI ) to Cr ( III ) is more likely to occur in DOM rich environment ( Richard and Bourg, 1 991 ) Bolan et al observed there is a significant positive relationship between CrVI reduction and the amount of easily oxidizable DOM in a soil, suggesting that DOM may act as an electron donor during Cr ( VI ) reduction to Cr ( III ) with itself being oxidi zed to CO 2 ( Bolan et al. 2003 ) Biochar a low density carbon rich material produced by heating biomass under low temperature and minimum oxygen, is widely used as a soil conditioner and fertilizer ( Steiner et al. 2007 ) Over time, biochar partic les release DOM to soil due to natural weathering or oxidative depolymerization of biochar ( Kim et al. 2004b ) In addition a large portion of the organic carbon in soils and sediments might be of biochar origi nation (5 40%). T he first molecular evidences of the existence of biochar structure in DOM was observed by Kim et al in samples collected from Rio Nergo in Brazil and then again in samples from a blackwater stream in the New Jersey Pine Barrens ( Kim et al. 2004b ) Hence, biochar derived DOM may play an important role in CrVI and AsIII transformation and in controlling their speciation in soils. I t is well known that chemical reaction s generally decrease with d ecreasing temperature s H owever, some reactions are accelerated in the ice phase F or example,

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86 Kim et al reported that Cr ( VI ) reduction by model organic acids such as citric, oxalic acid, and humic acid was accelerated 2 3 fold in ice phase compared to aq ueous phase ( Kim and Choi, 2011 ) T he solutes were concentrated as they are excluded from crystalline ice into the liquid like grain boundary. However, redox reactions of Cr ( VI ) and As ( III ) with DOM from biochar have not been studied despite their importance in the e nvironment. According to the thermodynamic calculation, As ( III ) and Cr ( VI ) can serve as a redox couple, which has been confirmed by Kim et al. In the system containing DOM, a s well as the Cr ( VI ) and As ( III ) redox couple whether or not DOM dominates the Cr ( VI ) and /or As ( III ) redox conversions or the Cr ( VI ) /As ( III ) redox couple plays the primary role has not been elucidat ed ( Kim and Choi, 2011 ) Two plant derived biochar s were selected, including sugar beet tailing (SBT; byproduct from sugar industry) and Brazilian pe pper (Schinus terebinthifolius, BP; invasive plant in Florida) and one DOM sample from Florida soil was used as a control The objective s of this study were to (1) investigate Cr ( VI ) reduction and As ( III ) oxidation by DOM from SBT and BP biocha r (DOMSBT, a nd DOMBP) in the ice and aqueous phase ; ( 2 ) quantify parameters such as pH and DOM concentration affecting the process; and (3) determine the mechanism s governing Cr ( VI ) and As ( III ) conversion in the system containing DOM, Cr ( VI ) and As ( III ) Materials an d Methods Preparation and Characterization of DOM from Biochars T he sugar beet tailing biochar (SBT) and Brazilian pepper (BP) biochar were produced at 300 in a muffle furnace (1500M, Barnstead Intl. Corp, IA) under N 2 at 10 psi to maintain an oxygen free environment. The detail of biochar preparation method

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87 and basic characteristics of the two biochar samples were previously described. ( Dong et al. 2011 ; Dong et al. 2013 ) A typical sandy loam soil was collected in Gainesville, Florida, as a comparison A sample o f 8 g of dry SBT BP biochar and soil was suspended in a reaction vessel with 1 L of deionized water, and w as agitated on a shaker at 200 rpm at room for 1 day (reached an apparent equilibrium) The suspension was then passed through a 0.2 membrane filter T he concentration of DOM was analyzed by a TOC analyzer (TOC 5050A; Shimadzu, Kyoto, J apan) The DOM samples were analyzed for metal content by ICP AES ( Plasma 3200, Perkin Elmer Crop, MA) Carbon, nitrogen and oxygen content were determined by a CHN elemental analyzer (Carlo Erba NA 1500). Adsorption of 254 nm UV radiation was measured by a UV spectrophotometer (Shimadzu Corporation, Japan) Fourier t ransform i nfrared (Bruker Optics Inc., Billerica, MA) spectra w as also collected of all DOM samples. Cr ( VI ) Reduction and As ( III ) Oxidation by DOM S tock solutions of 1000 mg/L Cr ( VI ) and As ( II I ) were prepared by dissolving analytical grade K 2 Cr 2 O 7 and NaAsO 2 in deionized water. All working solutions were freshly prepared before use by diluting the stock solution with deionized water, and the pH was adjusted to the desired values with 1 M HNO 3 o r 1 M NaOH solution. Solution s containing 15 m L of 10 mg / L Cr ( VI ) or As ( III ) with or without 10 mg C/ L DOM ( DOM SBT DOM BP or DOM S ) w ere put in to two 20m L scintillation vials with one set being kept in the aqueous phase ( ) and the other set in the ice phase ( 20 ) The solution samples were frozen within 1 h. Aliquots of samples were withdrawn at different time intervals for Cr ( VI ) As ( III ) and DOM concentration analysis. T he effect of

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88 pH on Cr ( VI ) reduction and As ( II I ) oxidation by DOM was investigated by varying solution pH from 2.0 10.0 while keeping Cr ( VI ) or As ( III ) at 10 mg/ L. The impact of DOM concentration was tested using DOM SBT sample by varying DOM c oncentration s from 3 to 300 mgC/ L, which covered the DOM co ncentrations used in this experiment For the ice phase treatment, the samples were put in waterbath (ISOTEMP 210, Fisher Scientific) at 40 for 10 min to thaw the samples. Simultaneous Cr ( VI ) Reduction and As ( III ) Oxidation T o determine the role of the DOM and Cr ( VI ) /As ( III ) redox couple in Cr ( VI ) and /or AsIII redox conversion, an excess amount of Cr ( VI ) or As ( III ) was added to solution co ntaining 10 mgC / L DOM The DOM was sufficient to reduce 52 g / L Cr ( VI ) or oxidize 113 g / L As ( III ) To determine the role of AsIII and DOM in CrVI reduction, solution s containing 1130 g / L As ( III ) and 52 g / L Cr ( VI ) with and without 10 mgC / L DOM were mixed in a vial, which was adjusted to pH 2.0 ( Cr ( VI ) reduction was effective at low pH ) and placed at 20 To determine the role of Cr ( VI ) and DOM in As ( III ) oxidation, solution s containing 113 g / L As ( III ) and 520 g / L Cr ( VI ) with and without 10 mg C / L DO M were mixed in a vial at pH 6.0 or 10 .0 (As ( III ) oxidation was effective at high pH) which we re placed in ice phase for 24 h. Fourier Transform Infrared (FTIR) Analysis of DOM Loaded with Cr and As T o determine the changes in chemical structure and functional groups involved in As and Cr redox conversion, 100 ml DOM samples containing 10 mgC / L DO M before and after reacting with 10 mg / L Cr ( VI ) at pH 2.0 or 15 mg / L As ( III ) at pH 10.0 for 24 h were freeze dried and analyzed using FTIR spectrometers (Bruker Optics Inc., Billerica, MA) The spectra were obtained in the range from 500 to 4 000 cm 1 with a resolution of

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89 8 cm 1 The resulting spectra w ere the average of 32 scans, which were used to identify the functional groups based on their characteristic absorbance peaks. Electron Spin Resonance (ESR) Analysis of DOM To determine the changes in functional groups involved in As redox transformation 100 mL DOM samples containing 10 mgC / L DOM before and after reacting with 1 0 mg / L As ( III ) at pH 7.0 and 10.0 for 24 h were freeze dried and then placed in glass capillaries. ESR spectra at the X band o f 9.8 GHz were recorded by a Bruker EMS spectrometer equipped with a TE102 cavity (Bruker Instruments, MA). Sample temperature was controlled at 5 K during data acquisition. Effective microwave and modulation fields of the sample were calibrated with perox ylamine disulfonate. The ESR was operated at a peak to peak modulation of 2.0 G, a microwave attenuation of 40 db, a modulation frequency of 100 kHz, a conversion time of 10 ms, and a time constant of 1.28 ms. Records of each spectrum were obtained by digi tal signal averaging and were electronically integrated by standard Bruker data acquisition software. The standard compound 3 carbamoyl 2,2,5,5 tetramethyl 3 pyrrolin 1 yloxy was used as a reference for spin count calibration. Chemical and Statistical Ana lysis T he concentrations of Cr ( VI ) were analyzed by measuring the absorbance of the purple complex of Cr ( VI ) with 1,5 diphenylcarbohydrazide at 540 nm using a UV spectrophotometer (Shimadzu Corporation, Japan) ( Eaton et al. 1995 ) To determine the total Cr, Cr ( III ) was first converted to Cr ( VI ) at high temperature (130 adding excess potassium permanganate prior to reacting with 1,5 diphenylcarbohydrazide ( Park et al. 2008 ) The concentrations of Cr ( III ) were calculated as the difference between the total Cr and Cr ( VI ) concentrations. The

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90 detection limit of this method wa s 0.03 mg / L. As ( V ) and As ( III ) were separated using an arseni c speciation cartridge (Waters Corporation, MC), which retains As ( V ) ( Mathews et al. 2010 ) Total arsenic and As ( III ) were determined by graphite furnace atomic absorption spectrophotometry ( GFAAS; AA240Z, Varian Inc., CA). As(V) concentration was determined by the difference between total and As(III). To check the reliability of this method, As speciation was also analyzed using HPLC ICP MS, which is comparable. ( Singh and Ma, 2006 ) Total DOM concentrations were analyzed on a TOC analyzer (Shimadzu Corporation, Japan) All reactions were conducted in triplicate and the results were calcu lated as the mean values of the samples. Data were statistically analyzed using JMP, version 9. Results and Discussion Similar Structure and Fu nctional Groups Existed in DOM f rom Biochar and Soil In order to und its basic ch aracteristics were determined. Three DOM samples were used, two from biochars (DOM B including DOM SBT and DOM BP ) and one from soil (DOM s ). DOM SBT sample contained the highest DOM at 323 mg C/ L whereas DOM S had the lowest (Table 5 1). The DOM in herbal biomas s biochar is much easier to leach than wood biomass biochar since the breakdown of grass begins at lower temperature than that of wood ( Keiluweit et al. 2010 ) Carbon normalized ultraviolet radiation absorbance at 254 nm (SUVA) is commonly used as an indicator of aromatic character in DOM samples which results from ( Traina et al. 1990 ) The absorption intensity varied, with DOM SBT >DOM BP >DOM S (T able 5 1). This observation suggested that DOM from biochar contained a relatively high er amount of aromatic or polyphenolic organic compounds than present in DOM s Concentrations of

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91 Fe, Mn, Al and Cu in all 3 samples were negligible, indicating that Fe an d Mn are not responsible for As oxidation (Table 5 1). Table 5 1. Characteristics of DOM SBT DOM BP and DOM S Fe (g/L) b Mn (g/L) b SUVA254 b (L/(mgXcm)) DOM (dissolved organic matter) (mgC/L) DOM SBT a 1.60.15 5.80.18 2.46 3230.89 DOM BP a ND* ND* 1.39 37.90.67 DOM S a ND* ND* 0.63 8.910.11 ND*: Not detected. The detection limits of ICP AES for Fe and Mn are 6.2 and 30 g/L. a : DOM SBT DOM BP and DOM S stands for dissolved organic matter derived from sugar beet tailing biochar, Brazilian pepper bio char and sandy soil. b : Dissolved organic matter concentration was 10 mgC/L, and pH was adjusted to 7.0 prior to analysis. The FTIR spectra displayed a number of sorption peaks indicating the complex nature of DOM (Figure 5 1). All DOM samples exhib ited similar peaks at 3340 3420, 2925 2927, 1612 1622, 1380 1382 and 1042 1390 cm 1 which were a ssigned to hydroxyl, CH 2 /CH 3 stretching carboxylic, C=O stretch of carboxylate ions, and CO stretching vibration of the alcoholic groups respectively ( Noguchi and Sugiura, 2003 ; Elangovan et al. 2008 ; Gupta a nd Rastogi, 2008 ) The DOM BP exhibited carboxylic and N H bending at 1706 cm 1 and 1457 cm 1 ( Xuan et al. 2006 ; Yahaya et al. 2009 ) T wo peaks at 674cm 1 and 765 cm 1 were assigned to the bending modes of the aromatic compounds for DOM S ( Bilba and Ouensanga, 1996 ) The similarity of these absorbance bands indicated that many similar structur al and functional groups exist in DOM from biochar and soil.

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92 Figure 5 1. Fourier transformation infrared spectra of control, Cr loaded, and As loaded DOM SBT (a), control, Cr loaded, and As loaded DOM BP (b) and control, Cr loaded, and As loaded DOM S (c). b a c

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9 3 Cr ( VI ) Reduction Was Enhanced by DOM and in t he Ice Phase Reduction of Cr ( VI ) was investigated at pH 2.0 with and without 10 mg C/ L DOM serving as an electron donor. Cr ( VI ) reduction is favorable at low pH, which is common in acid mine drainage and ind ustrial wastewater ( Regenspurg and Peiffer, 2005 ; Dong et al. 2011 ) In the aqueous phase, there was no C r ( VI ) reduction in the absence of DOM (data not shown). With the addition of DOM, 34 49% of Cr ( VI ) was reduced (Figure 5 2a c). Cr ( VI ) reduction to Cr ( III ) was 1.5 1.8 fold faster in the ice p hase ( 2 0 C) co mpared to the aqueous phase (25 C) (Figure 5 2a c ) After 24 h, DOM facilitated Cr ( VI ) reduction was more rapid with DOM B (80 86%) than DOM S (52%) in the ice phase (Figure 5 2a c). Cr reduction was slow in ice phase during the first 4 h. After placing in ice phase, it took time for the solutes to exclude from crystalline ice lattice into the liquid like grain boundary. It seemed that DOM B was more effective than DOM S which might be explained by their relatively higher content of polyphenolic organic compounds (Table 5 1). Phenolic compound s are expected to be easily oxidized to carboxylic acid by Cr ( VI ) in acidic media ( Wang et al. 2008 ) During Cr ( VI ) reduction, DOM concentrations in all three samples decreased (F igure 5 2d f). After 24 h, the ratio of CO 2 (based on decreased DOM concentration) to reduced Cr ( VI ) in the ice phase was 0.50, 0.48 and 0.57 mol / mol for DOM SBT DOM BP and DOM S respectively. In the aqueous phase the ratio increased to 0.55, 0.54 and 0.66 However, the ratio i s 0.75 mol/ mol during the reaction between organic matter and Cr ( VI ) re duction during application of the Walkley Black method ( Bolan et al. 2003 ) The relatively low ratio indicated that some of the DOM was completely oxidized to CO 2 whereas some was only partially oxidized, with the ice phase promoting more partial

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94 DOM oxidation Hsu et al. also observed partia l DOM oxidation during Cr ( VI ) reduction by DOM from fungal biomass of Neurospora crassa ( Hsu et al. 2010 ) Figure 5 2. Cr ( III ) appearance and DOM disappearance in DOM SBT (a,d), DOM BP (b,e) and DOM S (c,f) treatment in both aqueous and ice phase s at pH 2.0 after 24 h reaction ( 10 mg C / L DO M and 10 mg/ L Cr ( VI ) ). a b c d e f

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95 Besides temperature, solution pH also affected Cr reduction. W hen solution pH increased from 2.0 to 4.0, Cr ( VI ) reduction in the ice phase decreased from 86% to 52% for DOM SBT and 80% to 48% for DOM BP (Table 5 2). A s imilar trend was also observed in the aqueous phase (Table 5 2). The reduction of Cr ( VI ) in both ic e and aqueous phase s became insignificant above pH 6.0 (data not shown). Besides pH, electron donor concentration (DOM) was another key factor influencing Cr ( VI ) reduction. W hen DOM concentration increased from 3 to 30 mg C/ L, Cr ( VI ) reduction increased fro m 20% to 100% i n the ice phase (Figure 5 3a). At the same concentrations at room temperature, Cr ( VI ) reduction increased from 2.5% to 83% in the aqueous phase Table 5 2. Reduction of Cr ( VI ) to Cr ( III ) ( mg/ L) by DOM SBT DOM BP and DOM S at various pH in both aqueous and ice phase s (10 mg/ L Cr ( VI ) and 10 mg C / L DO M) DOM SBT DOM BP DOM S pH Aqueous Ice Aqueous Ice Aqueous Ice 3.0 3.670.21 7.880.28 3.390.12 6.820.21 2.870.23 6.540.16 4.0 2.560.18 5.210.03 2.010.08 4.810.01 1.760.45 4.040.02 5.0 2.2 30.14 3.790.11 2.170.06 3.060.05 1.170.28 2.850.38 6.0 1.680.02 2.140.07 1.150.01 1.980.02 0.680.08 1.620.29 The increased conversion of Cr ( VI ) in the ice phase compared to the aqueous phase was likely due to the fre eze concentration effect In ice phase, the degradation rate of p nitroanisole (PNA) was enhanced up to 40 times, probably resulting from concentrated content of solutes as they are excluded from crystalline ice lattice into the liquid like grain boundary ( Grannas et al. 2007 ) The concentrations of DOM, protons and Cr ( VI ) are highly concentrated in the ice grain boundary regi on, which accelerate d Cr ( VI ) reduction. For example, at DOM=10 mg C/ L and Cr ( VI ) =10 mg / L, Cr ( VI ) reduction

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96 in aqueous phase a t pH 3 was comparable to the ice phase at pH 5 (Table 5 2). In addition, at pH 2.0, 86.4 % of Cr ( VI ) was reduced in the ice phase at 1 0 mg C/ L DOM, which was comparable to 82.5 % of Cr ( VI ) in the aqueous phase a t 30 0 mg C/ L DOM (Figure 5 3a). Hence, concentrating either protons or DOM in aqueous phase had similar effect as in ice phase. Therefore, the accelerated reduction of Cr ( VI ) in ice phase was most likely due to the freeze concentration of both H + and DOM in the ice grain boundary. Heger et al. also reported that the transition point (pH T ) of cresol red solution, was due to an increase in the concentration of protons in the ice grain boundary ( Heger et al. 2006 ) Kim et al. estimated that the enhanced Cr ( VI ) reduction by citric and oxalic acids in the ice phase was due to the increase of solutes in the grain bo undary ( Kim a nd Choi, 2011 ) The coupled redox reactions of Cr ( VI ) reduction and DOM oxidation was expressed in Equation 5 1 The reduction of Cr ( VI ) was more favorable with at lower pH and higher DOM, which was consistent with other studies ( Kim and Choi, 2011 ) ( 5 1 ) The FTIR spectra of freeze dried DOM sample s before and after Cr ( VI ) reduction showed vibration frequency changes in their functional groups (Figure 5 1). After Cr ( VI ) reduction, peaks for carboxylic groups (1700 1720 and 1380 cm 1 ) showed a visible increase in all DOM samples ; while peaks for hydroxyl groups (3340 3420 cm 1 ) exhibited a remarkable decrease in all DOM samples. The increase of peaks for carboxylic groups probably result ed from partial oxidation of DOM. More oxidized functional groups s uch as carbonyl and carboxylic groups were generated on the

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97 activated carbon after oxidizing by ozone ( Valds et al. 2002 ) The increase in carboxylic peaks was also observed in Cr ( VI ) reduction by DOM derived from funga l biomass of N. crassa ( Hsu et al. 2010 ) Figure 5 3. Effect of DO M concentration on reduction of 10 mg/ L Cr ( VI ) (a) and oxida tion of 10 mg/ L As ( III ) (b) by DOM SBT after 24 h As ( III ) Oxidation Was Enhanced by DOM and In Ice Phase DOM is also known to serve as an electron acceptor Redman et al. reported that DOM samples collected from Brazil and US rivers were able to oxidize 25 40 g/ L As ( III ) within 90 h at 10 mg C/ L and pH 6.0 ( Redman et al. 2002 ) However, the effect of biochar derived DOM in As ( III ) oxidation is unclear ( Zimmerman, 2010 ) Figure 5 4 showed As ( III ) oxidation by 10 mg C/ L DOM at pH 10.0 in both ice and aqueous phases. Such pH is common during bleach cleaning of decks and stairs made of CCA t reated wood (copper, chromium and arsenic) where both As and C r are leached into soils underneath ( Cooper et al. 2001 ) In addition, semiquinone radical a b

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98 was stable at high pH. Hence, pH 10.0 was selected ( Jiang et al. 2009 ) In the aqueous phase, there was no As ( III ) oxidation in the absence of DOM (data not shown); however, all three DOM samples were effective in oxidizing As ( III ) with DOM S being more effective than DOM B After 24 h, 21% As ( III ) was converted to As ( V ) by DOM S compared to 16 18% by DOM B (Figure 5 4 ). As(III) oxidation in the ice phase was 1. 1 fold higher than in the aqueous phase. H owever, DOM was less effective in As ( III ) oxidation compared to Cr ( VI ) reduction, indicat ing th at DOM was a better reductant than oxidant. Fe and Mn in DOM and O 2 in the air might contribute to As ( III ) oxi dation. Since Fe and Mn content s in all thre e DOM samples w ere almost negligible, they were not important. The fact that the control treatment wi thout DOM showed no As ( III ) oxidation in 24 h (data not shown) indicated that O 2 in As ( III ) oxidation. Hence, DOM was the most likely electron acceptor during As ( III ) oxidation. ESR spectroscopy is used to measure free radic a l content, a property that can ( Chen et al. 2002 ) Table 5 3 provides a summary of relevant ESR parameters for the th ree DOM samples. Their splitting factors (g value) ranged from 2.0043 to 2.0048, which were consistent with the presence of free semiquinone radical with a typical g value of ~2.0044 ( Paul et al. 2006 ) Semiquinone radical was i dentified as the primary organic radical in DOM samples, functioning as the main electron accepting moiet y ( Cory a nd McKnight, 2005 ) After reacting with As ( III ) for 24 h, the electro spin content in DOM samples was undetectable, indicating semiquinone radical was responsible for As ( III ) oxidation (data not shown). Jiang et al. ( Jiang et al. 2009 ) also confirmed that semiquinone ra dicals derived from a model

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99 Fig ure 5 4. As ( III ) oxidation by DOM SBT (a), DOM BP (b) and DOM S (c) in both aqueous and ice phase s at pH 10 .0 after 24 h reaction ( 10 mg C / L DO M and 10 mg/ L As ( III ) ). a b c

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100 quinone (AQDS, 9,10 anth raquinone 2,6 disulfonic acid) were strong oxidants of As(III) to As(V). Comparison of spin concentrations revealed soil DOM was more effective with DOM S > DOM BP > DOM SBT (Table 5 3). This order seemed consistent with the effectiveness of AsIII oxidation by DOM samples, with higher spin concentration correlated with higher oxidation efficiency. With regard to spin concentration, the same pattern was observed for all 3 DOM samples: high signal intensity for samples at pH 10.0 but no signal at pH 7.0 (data not shown). High pH was known to stabilize semiquinone radic a l s, an effect confirmed for DOM with the highest spin concentration at alkaline pH ( Paul et al. 2006 ) Table 5 3. ESR data of DOM samples. Spin concentration (10 17 /g c arbon) Line width (G) g value DOM SBT 2.52 3.1 2.0048 DOM BP 4.81 3.5 2.0046 DOM S 44.5 5.0 2.0043 It is known that Cr ( VI ) reduction is favored at low pH whereas As ( III ) oxidation is favored at high pH (Tables 5 2, 5 4 ). Same phenomena wa s observed by Jiang et al. where oxidation of As ( III ) to As ( V ) by semiquinone radical depend ed strongly on pH with more As ( III ) (67.3%) being oxidized at pH 11.0 compared to pH 7.0 (12.6%) and pH 3.0 (0.5%) ( Jiang et al. 2009 ) A possible explanation is that radicals at lower pH was n ot as stable as at higher pH ( Jiang et al. 2009 ) The effect of DOM concentration on As ( III ) oxidation was similar to Cr ( VI ) reduction, with higher DOM concentration resulting in more As ( III ) oxidation (Figure 5 3b), probably due to increased concentration of semiquinone radical. The relative ly higher oxidation rate in the ice phase could also be

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101 explained by the freeze concentration effect. For example, at DOM=10 mg C/ L and As ( III ) =10 mg / L, As ( III ) oxidation at pH 9 in the aqueous phase was comparable to that in ice phase a t pH 8 (Table 5 4). In addition, a similar amount of As ( III ) (around 19%) was oxidized by 1 00 mg C/ L DOM in the aqueous phase and by 10 mg C/ L DOM in the ice phase a t pH 10.0 (Figure 5 3b). The reaction can be expressed by Equation 5 2 which indicated tha t As ( III ) oxidation was more favor ed with more hydroxyls and DOM. ( 5 2 ) Table 5 4. Oxidation of As ( II I ) by DOM SBT DOM BP and DOM S at various pH in both aqueous and ice phase s (10 mgL As ( III ) and 10 mg C / L DO M) DOM SBT DOM BP DOM S pH Aqueous Ice Aqueous Ice Aqueous Ice 6.0 0.040.01 0.110.05 0.100.07 0.150.02 0.070.05 0.120.02 7.0 0.150.08 0.380.04 0.170.01 0.310.05 0.200.06 0.390.04 8.0 0.390.11 0.710.01 0.320.02 0.750.01 0.420.01 0.790.07 9.0 0.680.05 0.950.08 0.710. 22 1.200.10 0.870.13 1.560.14 The changes in functional groups of DOM after As ( III ) oxidation were evaluated using FTIR (Figure. 5 1 ). Compared to DOM SBT the As ( III ) loaded DOM SBT was different: a peak at 1612 cm 1 decreased and a peak at 1380 incre ased and shifted to 1401 cm 1 They were both signed to carboxylic group, which contained semiquinone radicals ( Yun et al. 2001 ; Elangovan et al. 2008 ) A new peak appeared at 847 cm 1 which was signed as As O band of As ( V ) The As O band for AsIII at around 780 cm 1 was not found, further supporting As ( III ) oxidation ( Pena et al. 2006 ) The As ( III ) loaded DOM BP and DOM S samples showed similar changes. In short, carboxylic grou p probably

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102 participated in As ( III ) oxidation, and the presence of As ( V ) was confirmed by the appearance of the As ( V ) O band in DOM Simulta neous Cr ( VI ) Reduction and As ( III ) Oxidation With and Without DOM Besides DOM, Cr ( VI ) and As ( III ) also could serve as a redox couple (Eq. 5 4 ). The positive standard redox potential indicates that redox reaction could occur spontaneously with a theoretical molar ratio of 1.5 (As ( III ) to Cr ( VI ) ). To confirm the redox reaction between Cr ( VI ) and As ( III ) the ir stoichiome tric relationship was investigated at various pH (2.0, 6.0 and 10.0) in the ice phase. Although the rates of Cr ( VI ) reduction and As ( III ) oxidation by DOM were significant only at acidic pH for Cr reduction and alkali ne pH for As oxidation, the Cr ( VI ) and As ( III ) redox reaction was efficient over a wide pH range (Table 5 5). The molar ratio of oxidized As ( III ) to reduced Cr ( VI ) in ice ranged from 1.39 to 1.49, which was close to the theoretical value 1.5. Kim et al. also observed the simultaneous Cr/As conv ersion with molar ratio of 1.2 1.5 in ice after 1 h (Equation 4 3 ) ( Kim and Choi, 2011 ) ( 5 3 ) Both DOM and Cr ( VI ) /As ( III ) could serve as electron acceptor and donor in ice phase, however, which played a more dominant role during the reaction was still unclear. Hence, two separate experiment s with/without 10 mgC / L DOM (enough to reduce 52 g/L Cr(VI) or oxidize 113 g/L As(III)) were conducted: an ex cess amount of As ( III ) at 1 1 3 0 g/L with Cr ( VI ) = 52 g/L with and without was conducted at pH 2.0 or excess amount of Cr ( VI ) at 5 2 0 g/L with As ( III ) = 113 g/L with and without 10 mgC / L DOM was conducted at pH 6.0 and 10.0. At pH 2.0 both Cr ( VI ) reduction and As ( III ) oxidation was observed in the absence of DOM. While with DOM, all Cr ( VI ) was reduced to Cr ( III ) with no As ( III )

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103 oxidation (Table 5 5). HCrO 4 and HAsO 2 were the dominant Cr and As species at this pH. At this low pH, As ( III ) oxidation by DOM was unfavorable (Table 5 5). DOM and As ( III ) probably competed for Cr ( VI ) ions as electron donor due to insuff icient Cr ( VI ) present at 52 g/L (Eqs. 5 1 and 5 3 ). Since no As ( III ) was oxidized with or without DOM, DOM was probably the preferred electron donor for Cr ( VI ) reduction at pH 2 ( Eq uation 5 1) was the dominant reaction. Table 5 5. Simultaneous reduction of Cr ( VI ) and oxidation of As ( III ) at various pH in the ice phase (g/L) (As ( III ) = 1130 g/L, Cr ( VI ) = 52 g/L at pH 2.0; As ( III ) = 113 g/L, Cr ( VI ) = 520 g/L at pH 6.0 and 10.0; DOM= 10 mgC/L). pH 2.0 pH 6.0 pH 10.0 As ( V ) Cr ( III ) As ( V ) Cr ( III ) As ( V ) Cr ( III ) Control a 1125.41 52.1 3 70 1118.02 54.33.17 1132.63 56.25.56 DOM SBT 1.132.8 8 53.0 5 46 1 12 5.18 51.55.15 1 11 5.78 0.162.72 DOM BP 1.891.3 5 51.9 4.84 1 13 9.60 55.06.65 1 12 3.45 0.102.84 DOM S 1.600.4 3 51.8 1.09 1 13 7.58 46.31.4 0 1 12 6.12 0.101.35 a : Control was treatment without DOM. It contained 1130 g/L As ( III ) 52 g/L Cr ( VI ) at pH 2.0; and 113 g/L As ( III ) 520 g/L Cr ( VI ) at pH 6.0 and 10.0; At pH 6.0, both As ( III ) oxidation and Cr ( VI ) reduction were observed and the re were no significant differences with and without DOM (Table 5 5). HCrO 4 and HAsO 2 were still the dominant Cr and As species, and DOM and Cr ( VI ) probably competed for As ( III ) ions as electron acceptor (Eqs. 5 2 and 5 3 ). With insufficient amount of As ( I II ) at 1 13 g/L if Eq. 5 2 was favored, all As ( III ) would have been oxidized by DOM and there

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104 would be no As ( III ) left to reduce Cr ( VI ) The presence of Cr ( III ) in DOM treatments implied Cr ( VI ) /As ( III ) redox reaction or Eq. 5 3 probably dominated at this pH. At pH 10.0, As ( III ) oxidation was observed with and without DO M but Cr ( VI ) reduction occurred without DOM Cr ( VI ) reduction by DOM was not favorable at this pH (Table 5 5), implying the impact of DOM on Cr ( VI ) reduction (Eq. 5 1 ) was negligible. The difference with and without DOM treatments confirmed that DOM was preferred for As ( III ) oxidation at this pH ( Eq uation 5 2) Furthermore, the stoichiometric ratio of As ( V ) to Cr ( III ) without DOM during the pH range 2.0 to 10.0 was 1.36 1.49, which was cons istent with the result of Kim et al. after 1 h reaction ( Kim and Choi, 2011 ) Hence, both DOM and the Cr ( VI ) /As ( III ) redox couple were responsible for the Cr and As redox conversion, but their respective roles varied with pH. DOM dominated the redox conversion under acidic and alkaline conditions; whereas the Cr ( VI ) and As ( III ) redox reaction dominated at neutral pH. Research Findings DOM from sugar beet tailing, Brazilian pepper and soil were more effective in Cr ( VI ) reduction and As ( III ) oxidation in ice phase than aqueous phase (Table 5 2 and 5 4 ). Cr ( VI ) reduction by DOM decreased while As ( III ) oxidation increased with increasing pH (Eqs. 5 1 and 5 2). FTIR data suggested that carboxylic group was probably responsible for Cr ( VI ) reduction and As ( III ) oxidation. The increased peak of carboxylic group during Cr ( VI ) reduction was probably due to the oxidation of polyphenolic organic compounds of DOM samples; whereas the decrease of carboxylic group peak during As ( III ) oxidation might be due to the reduction of DOM samp les, especially consumption of the semiquinone radicals. The accelerated conversion of Cr ( VI ) and As ( III ) with DOM serving as electron donor and accepter was consistent with

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105 the increased concentration of DOM, protons, hydroxyls and semiquinoic radicals in the grain boundary. The conversion reactions in the grain boundary was supported by our experiment data and other studies ( Heger et al. 2006 ; Kim and Choi, 2011 ) Besides DOM, As ( I II ) could also serve as electron donor for Cr ( VI ) reduction while it was oxidized to As ( V ) The simultaneous conversion of Cr ( VI ) and As ( III ) coupled with the ( VI ) and oxidize As ( III ) demonstrated that both the redox couple and DO M were responsible for their transformation, but their significance varied. DOM dominated the redox conversion at acidic and alkali condition whereas Cr ( VI ) and As ( III ) redox reaction dominated at neutral pH (Figure 5 5) Figure 5 5. Mechani s ms of Cr ( VI ) reduction and As ( III ) oxidation by DOM in the ice phase Chromium and arsenic can be present as co contaminants in the environment. The presence of DOM in a frozen landscape may affect their speciation, fate and

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106 transport in soil and water. For example, during cold periods, Cr ( VI ) in contaminated acidic sites and As ( III ) in contaminated alkaline sites may undergo accelerated conversion in the presence of DOM. The coupled reaction of Cr reduction with As oxidation may occur when Cr ( VI ) and As ( III ) coexist in presence of low DOM. In winter time in cold regions, Cr ( VI ) and As ( III ) in frozen soils can be converted to the less toxic form at an accelerated rate.

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107 CHAPTER 6 CHARACTERIZATION OF FLORIDA COAL COMBUSTION RESIDUES Introduction C oal combustion residu es (CCR) consist of fly ash, bottom ash sl a g and flue gas desulfurization residues (FGD) Besides mining waste, CCR are the second largest waste streams generated in the US ( Fitzgerald, 2010 ) As reported by American Coal Ash Association, approximately six hundred power plants generate 130 mil lion tons of CCR per year, of which 56% are stored in landfill and surface impoundments, while the remainders are reused in agriculture, commercial and engineering applications. The remainder i s managed in either landfills or disposed in surface impoundmen ts ( Ruhl et al. 2012 ) Whether fly ash is landfilled or disposed in surface impoundments, there is a concern about potential leaching of heavy metals from CCR partially due to the extr eme high pH and high concentration s of trace metals. For example, the CCR spill in Kingston in 2008 not only flooded more than 300 acres of land, damaging homes and property, but also released considerable amount of trace metals, particular Se, B, Sr and B a The As concentration was high up to 2 000 mg/kg, which showed adverse effects on the environments ( Ruhl et al. 2010 ) EPA also has also document ed case s of environmental damage by CCR in Florida. One site has been reclassified as a potential damage ca se based on documented exceedance of primary drinking water standards for Cd, Cr and F, and secondary drinking water standards for S, Cl, Mn and Fe in on site gr oundwater resulting from CCR. Another case has been categorized as a potential damage case for

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108 disposal facility, which does not impact drinking water wells offsite. It is for this reason characterization of CCR has attracted considerable attention. Due to the adverse impact of CCR spill in Kingston, EPA is currently proposing to regulate CCR destined for disposal in landfills or surface impoundments as a special waste. Hence, it is necessary to measure the concentrations and leachability of trace metals in CCR samples. This study a im ed to investigate the potential environmental impacts of CCR collected from Florida coal power plants The objectives were : (1) to measure the total elemental concentrations and pH of all CCR samples; (2) to asses s metal leach ing behavior from these CCR samples by leaching test Metal leach ing behavior was assessed by s ynthetic precipitation leaching procedure (SPLP) one of the most commonly used leaching test in the US. SPLP simulates an acid rain leachate and provides insight into whether the metals ha ve mobility potential in the unsaturated zone and the potential risk to groundwater conditions ( Jang et al. 2002 ) Material s and Method s CCR Sampl e s The 27 CCR samples tested in this study included 1 4 fly ashes, 9 bottom ashes, an d 4 FGD residues from 7 Florida coal power plants (Table 6 1). Coal power plants providing samples were identified by a code to allow specific facilities to remain anonymous. Table 6 1 summarizes the facilities provi di ng CCR samples, grouped by coal type, air pollution control configuration, and disposal method. In addition, description of CCR samples, sampling units and sampling methods was pr ovided in Table 6 2.

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109 Table 6 1. Summary of facility configuration Facility Type of coal Air pollution control co nfiguration Ash disposal method FGD disposal method A Eastern Bituminous Fabric filter a SCR b Wet Scrubber c Landfill, Sale Sale B Bituminous ESP d SCR, Wet Scrubber e Landfill, Sale Sale C Bituminous Fabric filter, SCR, Wet Scrubber Landfill, Sale S ale D Bituminous Fabric filter, SCR, Wet Scrubber Landfill, Sale Pond (not in use) Sale E Eastern Bituminous CGCU f HGCU g CT h Sale F Bituminous (Central Appalachian Coal) ESP; SCR, Wet Scrubber Landfill, Sale Landfill G Bituminous (Illinois basin coa l) ESP, SCR, Wet Scrubber Landfill, Sale Sale a : used for remove the fly ash particles from the flue gas b : Selective catalytic reduction, used for NO x Controls; c &e : used for SO 2 control; d : Electrostatic precipitator, used for attract and remove th e fly ash particles from the flue gas. f : C old gas clean up system, used for remove slag particles and H 2 S, COS ; g : Hot gas clean up system, used for remove slag particles and H 2 S, COS ; h : Combustion Turbine, used for low NO x Emissions

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110 Sampling Method s In order to get the representative CCR samples, composite sampling method was applied ( Townsend, 2012 ) Composite sampling is a technique that combines a number of discrete samples collected f rom a body of material into a single homogenized sample for purpose of analysis. Two methods developed by Townsend (2012) had been identified to collect CCR samples: in stream method and storage area method. The stream method entailed collect ing CCR sample s from temporary storage areas or conveyance systems (e.g., belts, pipes, and silos) ; w hile, the storage area method CCR samples were collected from storage areas such as storage piles. The in stream method allowed for the collection of CCR samples withou t the necessity of setting up field sampling points within a defined gird for representative sample collection. When the in stream method was applied, facility personal would collect daily CCR samples over a 5 day work week. The example below outlined the in stream sampling procedure. Example : O ne five gallon bucket was provided for each CCR sample On sampling day one (Monday), half gallon CCR sample was collected in a five gallon bucket. On sampling day two, another half gallon CCR sample was collected i n the sample five gallon bucket, etc. After samples wer e collected for 5 days, one composite CCR sample comprised of daily samples w as collected. The storage area method was used to collect CCR samples located in storage area throughout a facility. This s ampling method required field sampling points within a defined grid for representative sample collection. The number of samples collected from a storage area was dependent on the size of storage area and potential heterogeneities.

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111 The following procedure p rovided a framework for the storage area method. CCR samples would be collected from various depths of the CCR pile as described in EPA publication SW 846 chapter nine to demonstrate that samples were statistically representative. A grid system was used to locate and select random sampling locations. Compositing of a number of subsamples was used to reduce the effects of variation ( Townsend, 2012 ) Table 6 2 summarized the target CCR samples and their sampling methods at each facility. Table 6 2. CCR sampling me thods for target CCR samples at each facility ( Townsend, 2012 ) Facility CCR categories for characterization Sampling strategies In Stream Storage Area A Fly Ash 1 (Unit 1&3&4) F ly ash 2 (fly ash with low LOI) Fly ash 3 (fly ash with high LOI) Fly Ash 4 Bottom Ash (Unit 1&2&3&4) Slag FGD Residue X X X X X X X B Fly Ash 1 ( Unit 1&2) Fly Ash 2 (Unit 4) Fly Ash 3 (Unit 5) Bottom Ash 1 (Unit 1&2) Bottom Ash 2 (Unit 2) Bottom Ash 3 (Unit 4) Bottom Ash 4 (Unit 5) FGD Residue (Unit 4&5) X X X X X X X X C Fly ash Bottom ash X X D Fly ash X E Slag 1 ( used in grit blasting) Slag 2 ( used in cement) Slag 3 ( fine power, cement) X X X F Fly Ash Bottom Ash FGD Residue X X X G Fly Ash Bottom Ash FGD Residue X X X

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112 Analytical Method s and Quality Assurance The aqueous extractant for pH measurement were obtained by mechanically shaking the CCR samples with distilled water at a solid : liquid ratio of 1:10 (w/v) for 1 h and the n passing through 0.2 m membrane filter s CCR samples were digested by EPA method 305 0 to analysis the total elemental concentrations ( EPA, 1996 ) SPLP ( Synthetic precipitation leaching procedure ) was used to establish the metal leaching behavior of each CCR sample. The SPLP extractant consisted of an unbuffered solution made by adjusting 1 L of distilled water to pH 4.2 with a 60/40 ratio mix of concentrated H 2 SO 4 and HNO 3 The SPLP was done with 1:20 solid/extractant ratios (in this case, 5 g soil for 100 mL of extractant) and the CCR samples were mixed for 20 hours in a heads up, heads down rotary mixer. At the end of reaction time, resulting solutions were filtered (with GFF fiberglass filters), acidified (5% v/v with concentrated HNO 3 ) and pH was measured. SW 846 Method 6020 was u sed to analyze the metal contents in solutions by ICP MS ( Inductively coupled plasma mass spectrometry ) (NexION 300, Perkin Elmer Crop, MA ) ( EPA, 2007b ) All experiments were conducted in tri plicate and the results were calculated as the mean values of the samples. T he QA/QC for total and SPLP concentration ana lysis included blank, spiked samples and triplicates in every 30 samples T he method blank for total concentration was the concentrated nitric acid and for SPLP concentration was 60/40 ratio mix of concentrated H 2 SO 4 and HNO 3 Each method blank was carried through the entire digestion or leaching test protocol Triplicates were used for all 27 CCR samples A known quantity of elements of concern was added prior to digestion or leaching test to serve as spiked samples.

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113 Accuracy or recovery was determined by calculating the percent bias from a known standard (spiked samples). Precision was calculated as the percentage of relative standard deviation to the mean value. Recoveries for spiked samples were within the acceptable ranges (80 120%). P recision for tripl icate samples were acceptable ranging between 0.17% and 1 0 %. The performance of the instrument (ICP MS) was checked by running an intermediate calibration standard for every 15 samples. All the calibration standard checks were within the acceptable range (85 120%). Results and Discussion pH The 27 CCR samples collected from 7 Florida coal power plants consisted of 3 residues : fly ash, bottom ash and FGD residues. Proash and Ecotherm collected from facility A were fly ash samples after treatment with the tr iboelectric separation process ( Bittner et al. 2009 ) Proash was a consistent, low loss on ignition fly ash for use as a su bstitute for cement in concrete; while Ecotherm was a high loss on ignition fly ash and could be returned to the facility for the recovery of the fuel value or utilized at cement kiln s as both a fuel substitute for coal and as a source of mineral required for adjusting cement clinker chemistry ( Bittner et al. 2009 ) Economizer ash collected from company A was also similar to fly ash but was generally not suitable for use as a pozzolan in cement production. Gasification slag sample s coll ected from facility E belonged to the bottom ash category since they had similar property Due to the similarity of physical and chemical characteristics between slag and bottom ash samples, the slag sample collected from facility A was listed in the bottom ash category. Hence, there were 1 1 fl y ash samples, 12 bottom ash samples and 4 FGD residues.

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114 The pH of Florida CCR samples were grouped into three categories: low pH (pH < 4.00), medium pH (5.00 < pH <8.00), and high pH (pH > 8.00) (Table 6 3) Fl y ash samples were mainly with low or high pH (2 fly ash with pH<4.00, 7 fly ash samples with pH > 8.00); however, bottom ash and FGD samples were in the medium pH category. T he relatively low pH of fly ash samples from facility A and F was due to their NO X control technology via selective catalytic reduction process. A mmonia was sprayed into the flue gas to enhance the conversion of NO X to harmless nitrogen gas and water. Total Concentrations Total elemental concentrations of CCR samples were analyzed by ICP MS using E PA Method 6020 after acid digestion using EPA Method 3050. Major elements including Ca, K, Na, Mg, Al, Fe, and Zn showed the highest concentrations in the CCR samples (Table 6 4 Figure 6 1 6 2 ) followed by trace elements of V, Mn, Pb, Ni, Cr, Cu, As, and Se (Table 6 5 Figure 6 3 6 6 ), with trace elements of Be, Co, Mo, Sb, Tl, Hg and Cd showing the lowest concentration (Table 6 6 Fi gure 6 7 6 8 ). The highest concentrations of major elements such as Ca, K, Na, Mg, Al, Fe, and Zn were related to the abundance of these elements in bituminous coals ( Duliu et al. 2005 ) In CCR samples, trace metals can be categorized into three main groups accor ding to their partitioning behaviors during coal combustion ( Vejahati et al. 2010 ) Group 1 elements are mainly concentrated in the coarse residues (bottom ash or slag ), or are partitioned equally between coarse residues and fine grained particulates (fly ash). Examples of these elements are Mg and Mn ( Xu et al. 2004 ) Group 2 elements are concentrated more in the particula tes compared with coarse bottom ash/slag They are also enriched on the fine grained particles which may escape particulate control systems. (e.g. As, Cd, Cu, Pb, Sb, Se, Zn) ( Xu et al. 2004 ) Group 3 elemen ts which

PAGE 115

115 are readily volatilize during the combustion process and are mainly concentrated in the vapor or gas phase (e.g. Br, Hg, I) ( Xu et al. 2004 ) However, some of the elements, such as Cr, Ni, and V, ma y show intermediate partitioning behavior between Group 1 and 2 and volatile elements, such as Se, may display partitioning behavior intermediate between Group 2 and Group 3 ( Vejahati et al. 2010 ) Hence, t he abundance of trace metals of V, Mn, P b, Ni, Cr, Cu, As, Mo and Se might be due to the enrichment of these elements during combustion process. The lowest concentration of trace elements Be, Co, Sb, Tl, Hg and Cd probably caused by their relatively low content in coal source or their volatile behavior (such as Hg) in combustion systems remaining in the gas phase during passage through the plant. Among the 7 facilities, CCR samples from facility A showed the highest concentration s of Pb, As, Tl, Hg and Cd whereas highest concentrations of As, Mn Cr, Cu, Se, Sb, Co and Cd were observed in samples from facility C Facility E exhibit ed the highest concentration for V, Ni and Mo. In addition, higher As and Be concentration were detected in CCR samples from facility B and F respectively Th e variance of trace elements content from different facility might be due to the difference between the variable coal sources or different age of coals from the same source or even same seams Total elemental concentrations of Florida CCR samples were comp ared with other CCR samples collected from around the US (Table 2 1). E lement al concentrations of f our FGD samples in Florida were in agreement with concentrations of other US FGD samples. The Florida fly ash and bottom ash samples showed significant diffe rence s for

PAGE 116

116 elements Al, Fe, Ni and V. Concentrations of the rest elements were in agreement with concentrations from other US CCR samples. The Al concentration s of all Florida fly ash and bottom ash /slag s amples were in the range of 1,130 11,380 and 880 22 ,894 mg/kg which was significant ly lower than the minimum value s of the 42 US CCR samples (fly ash: 46,000 mg/kg ; bottom ash: 30,500 mg/kg ) (Figure 6 1) The Fe concentrations in fly ash samples from facilit y D E F and B ( fly ash from unit 1) and bottom ash samples from facilit y A ( slag sample ), B and F were also much lower than the minimum values of other US samples (fly ash: 17,000 mg/kg; bottom ash: 21,600 mg/kg) (Figure 6 2) For Ni and V, only facility C and E exceeded the maximum value of other US CCR samples (Ni: 1,267 mg/kg; V: 652 mg/kg) (Figure 6 3 6 4) The significant difference between Florida CCR and other 42 US CCR samples might be due to the difference coal sources. Furthermore, concentrations of elements differ significantly b etween coals from the same seams also has been reported ( Duliu et al. 2005 ) The cle anup target levels for residential and industrial soil s were used to assess the potential environmental risk of CCR samples (Table 6 1 1 ) Concentrations of all 19 elements in four FGD samples were below the residential soil cleanup level. The fly and botto m ash samples collected from facility A C and E exceeded the residential soil cleanup level for Pb, Cu, Sb, Tl and Hg In addition, As, Ni and V concentrations in fly ash and bottom ash samples collected from all 7 facilities were even higher than th eir i ndustrial soil cleanup levels These results indicated that major concern should be given to those 8 elements especially As, V and Ni for Florida fly ash and bottom ash samples as they may be used to improve soil properties and increase crop production.

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117 Numbers exceeding the residential soil cleanup level are labeled as R and exceeding industrial level are labeled as I. Table 6 3. pH of CCR samples Sample pH A Fly ash 1 9.57 A Fly ash 2 11.82 A Fly ash 3 8.91 A Fly ash 4 1.74 A Bottom ash 5.20 A FGD 6.82 A Slag 6.11 B Fly ash 1 6.89 B Fly ash 2 11.84 B Fly ash 3 11.89 B Bottom ash 1 10.75 B Bottom ash 2 9.66 B Bottom ash 3 8.66 B Bottom ash 4 8.88 B FGD 7.64 C Fly ash 5.84 C Bottom ash 8.44 D Fly ash 11.16 E Slag 1 6.91 E Slag 2 6.64 E Slag 3 5.23 F Fly ash 3.96 F Bottom ash 5.13 F FGD 8.33 G Fly ash 11.12 G Bottom ash 8.31 G FGD 7.96

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118 Table 6 4 Total concentrations of Ca, K, Na, Mg, Al, Fe, and Zn in CCR samples (mg/kg) Samples Ca K Na Mg Al Fe Zn A Fly ash 1 10,202 456 299 1,091 7,995 18,169 113 A Fly ash 2 18,576 489 322 1,339 7,868 17,491 104 A Fly ash 3 8,251 521 246 1,190 7,828 17,941 146 A Fly ash 4 10,483 347 44.4 107 1,133 199,709 R 10.7 A Bottom ash 21,333 995 353 2,909 22,894 82,534 R 160 A Slag 1,173 410 173 442 3,882 9,041 18.7 A FGD 65,418 27.3 18.3 375 178 302 10.5 B Fly ash 1 2,767 579 347 900 7,392 6,159 26.0 B Fly ash 2 24,505 982 1,137 1,594 9,460 22,118 147 B Fly ash 3 25,961 921 1,037 1,573 9,767 24,038 147 B Bottom ash 1 1,073 130 196 25 3 1,235 2,511 7.83 B Bottom ash 2 502 96 82 156 880 8,321 6.20 B Bottom ash 3 6,407 197 372 559 6,928 5,049 27.0 B Bottom ash 4 4,090 169 376 367 6,165 4,350 29.0 B FGD 51,731 36 110 115 209 786 4.69 C Fly ash 12,174 814 2,904 6,003 11,380 12,356 3,2 12 C Bottom ash 24,241 120 2,205 3,244 6,045 24,305 524 D Fly ash 5,561 931 1,400 1,411 8,533 6,546 67.0 E Slag 1 158 33.9 8.47 122 1,683 14,671 18.5 E Slag 2 285 53.9 108 321 1,183 13,368 1,388 E Slag 3 1,167 213 313 678 2,579 8,597 508 F Fly ash 3, 340 652 328 1,418 9,592 8,592 47.0 F Bottom ash 2,189 269 2,410 7,033 2,871 19,544 33.5 F FGD 1,490,609 100 266 2,212 291 398 66.4 G Fly ash 14,653 1,283 1,177 1,199 10,696 29,484 120 G Bottom ash 6,280 772 580 783 9,741 22,955 23.0 G FGD 58,429 66 25 5 232 299 1,104 6.74

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119 Figure 6 1 Total concentration of Al in CCR samples Figure 6 2 Total concentration of Fe in CCR samples

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120 Table 6 5 Total concentrations of V, Mn, Pb, Ni, Cr, C u, As and Se in CCR samples (mg/kg) Samples V Mn Pb Ni Cr Cu As Se A Fly ash 1 286 R 97.8 25.2 25.6 63.5 23.8 36.9 I 21.9 A Fly ash 2 340 R 97.1 24.2 29.6 74.2 25.9 36.8 I 24.3 A Fly ash 3 530 R 87.0 36.6 54.8 65.8 29.0 39.9 I 32.4 A Fly ash 4 8.04 269 1, 018 R 13.0 6.94 37.3 72.6 I 7.06 A Bottom ash 209 R 392 11.0 72.3 63.7 39.5 7.69 R 3.85 A Slag 93.1 R 22.5 5.06 62.0 14.5 10.9 0.738 A FGD 13.8 3.73 11.1 1.71 11.8 1.92 0.397 B Fly ash 1 76.0 R 44.0 21.0 24.0 16.0 61.0 73.0 I 21.0 B Fly ash 2 286 R 210 3 1.0 32.0 8.31 35.0 44.0 I 16.0 B Fly ash 3 297 R 223 30.0 33.0 8.44 34.0 41.0 I 16.0 B Bottom ash 1 6.00 20.0 0.61 5.97 2.03 11.0 3.49 R 0.90 B Bottom ash 2 3.86 103 0.54 7.93 2.40 15.0 1.55 0.80 B Bottom ash 3 33.0 42.0 4.06 7.70 2.45 6.98 3.90 R 1.0 5 B Bottom ash 4 26.0 38.0 3.35 7.54 2.15 7.75 2.59 R 0.83 B FGD 10.0 4.41 0.37 1.84 0.19 3.87 0.60 5.80 C Fly ash 30,796 I 629 427 R 9,768 R 398 R 692 R 73.0 I 58.0 C Bottom ash 18,536 I 602 369 5,992 R 249 R 299 R 53.0 I 39.0 D Fly ash 91.0 R 112 38.0 30.0 15.0 65.0 46.0 I 3.83 E Slag 1 322,642 I 14.9 4.85 121,182 I 391 R 72.9 61.2 I 1.99 E Slag 2 149,889 I 30.3 58.2 7,279 R 175 47.0 62.9 I 12.5 E Slag 3 4,297 R 47.9 114 1,655 R 53.0 17.3 46.6 I 26.5 F Fly ash 129 R 50.0 38.2 33.3 34.3 70.5 66.1 I 23.1 F Bottom ash 39. 2 49.1 11.9 23.9 10.7 25.8 10.6 R 12.6 F FGD 11.4 7.78 1.03 2.81 13.7 1.74 0.873 17.0 G Fly ash 207 108 32.0 38.0 10.0 55.0 49.6 I 14.0 G Bottom ash 61.0 73.0 3.68 17.0 5.56 14.0 7.61 R 2.37 G FGD 8.20 5.54 0.85 1.7 0.20 3.02 1.37 5.56 : Below the detec tion limit. The detection limit of ICP MS for Se is 0.52 g/L

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121 Figure 6 3 Total concentration of V in CCR samples. Figure 6 4 Total concentration of Pb in CCR samples.

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122 Figure 6 5 Total concentration of Ni in CCR samples. Figure 6 6 Total concentration of As in CCR samples.

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123 Table 6 6 Total concentrations of Be, Co, Mo, Sb, Tl, Hg and Cd in CCR samples (mg/kg) Samples Be Co Mo Sb Tl Hg Cd A Fly ash 1 2.62 5.87 71.7 22. 6 7.51 R 0.702 2.71 A Fly ash 2 2.73 5.59 75.2 1.22 7.37 R 0.593 3.03 A Fly ash 3 2.75 6.11 77.7 1.38 8.70 R 1.28 3.16 A Fly ash 4 0.233 2.62 63.6 0.732 11.3 R 8.73 R 0.299 A Bottom ash 4.85 13.3 21.6 0.099 1.58 4.78 R 1.16 A Slag 0.695 2.42 3.78 0.431 0.077 2.39 0.038 A FGD 0.099 0.197 1.45 0.169 0.165 4.20 R 0.393 B Fly ash 1 4.73 16.0 10.0 0.70 1.63 0.76 0.199 B Fly ash 2 3.72 8.31 81.0 1.71 5.57 1.12 2.51 B Fly ash 3 3.50 8.44 75.0 1.64 5.40 1.46 2.55 B Bottom ash 1 0.35 2.03 0.05 0.008 1.16 B Bottom ash 2 0.31 2.40 0.02 0.005 0.17 B Bottom ash 3 1.39 2.45 3.46 0.11 0.04 0.16 0.149 B Bottom ash 4 1.08 2.15 4.45 0.08 0.11 0.38 0.120 B FGD 0.03 0.19 1.30 0.02 0.01 0.66 0.013 C Fly ash 2.15 398 302 3 6.0 R 1.20 0.39 4.79 C Bottom ash 4.78 249 108 27.0 R 0.07 1.10 1.24 D Fly ash 4.24 15.0 15.0 0.85 0.80 0.15 0.587 E Slag 1 0.066 64.8 421 14.2 0.067 3.46 R 0.268 E Slag 2 0.100 37.7 227 15.5 0.570 2.82 1.71 E Slag 3 0.475 8.70 126 6.68 4.47 2.48 2.74 F Fly ash 6.08 13.7 33.9 2.01 2.60 0.793 F Bottom ash 1.57 5.72 4.71 0.131 0.262 0.294 F FGD 0.033 0.290 1.70 0.065 0.292 0.097 G Fly ash 3.90 10.0 68.0 0.60 6.15 0.12 2.78 G Bottom ash 2.20 5.56 7.74 0.1 0.70 1.19 0.472 G FGD 0.07 0.20 0.17 0.0 3 0.04 1.12 0.010 : B elow the detection limit The detection limit of ICP MS for Mo Sb, Hg, Cd is 0.76, 0.08, 1 and 0.17 g/L

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124 Figure 6 7 Total concentration of Mo in CCR samples. Figure 6 8 Total concentration of Sb in CCR samples.

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125 SPLP Concentrations SPLP was conducted to test the leaching behavior of elements in CCR samples. The SPLP leaching fluid simulates acid rain, which was prepared using dilute sulfuric acid and nitric acid solution (60/40 mix) at pH 4.20 and solid to liquid ratio of 1:20 (in this case, 5 g soil for 100 mL extractant) Concentrations of major elements Ca, K, Na, Mg, Al, and Fe were the highest in CCR samples ; however, Ca, K, Na, and Mg are not regarded as elements o f concern. The groundwater maximum contaminant level for Al is 200 ug/L and almost all CCR samples exceeded this level except for 3 FGD samples, and 2 slag samples (Table 6 7 Figure 6 9 ). The relative ly low Al concentration from those 5 samples might b e due to their relative ly low total Al concentrations. For Fe, the maximum contaminant level is 300 ug/L and only 10 samples from facility A B D E and F met this standard (Figure 6 10) They were mainly FGD and bottom ash samples with lower Fe content. Fly ash from facility A leached a significant amount of Fe, 2, 20 5 m g /L The higher Fe concentration was related with the high er solubility of Fe at acidic pH ( Table 6 7 ) coupled with the relative high total Fe content. Besides the major elements, CCR sampl es also contained a broad range of trace metals. Their concentrations in CCR samples were generally low ; however, they were of primary concern regarding the potential for the m to leach into surface or groundwater, contaminating drinking water, surface wate r or living organisms. For V, 10 CCR samples which were mainly FGD and bottom ash samples met t he cleanup target level (Table 6 8 Figure 6 11 ). The higher concentration for samples from facility C and E might be due to the higher total V concentration a nd lower pH V

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126 had been reported to be less leachable with the increase of pH, especially at strongly alkaline pH values ( Izquierdo and Querol, 2012 ) For Mn, 5 fly ash, 2 bo ttom ash and 1 FGD samples from facility A C E and F exceeded the cleanup target level (Table 6 8 ) Fly ash sample collected from facility A showed a high concentration of Mn at 9 52 m g /L almost 190 times of the maximum contaminant level. The relativ e ly higher Mn concentration might be explained by the lower pH since Mn was insoluble under near neutral to alkaline conditions and dissolved Mn increased with decreasing pH ( Izqu ierdo and Querol, 2012 ) For Pb, 7 fly ash samples from facility A B C E F and 2 bottom ash samples from facility A and E were higher than the maximum contaminant level (Table 6 8, Figure 6 12) Fly ash from facility A w as much higher, which might b e related with the higher Pb content and lower pH since the acidic condition enhanced the Pb leaching ( Praharaj et al. 2002 ) For Ni, only 9 samples exceed ed the 100 ug/L maximum contaminant level They were fly ash and bottom ash samples from facility A C E and F (Table 6 8, Figure 6 13) The extremely high concentration in ash samples from facility C and E were related with the ir higher total Ni content. Besides the total Ni content, pH also played an impo rtant role for leaching of Ni. The lowest leachable levels of Ni were reported to be attained in the pH region of 8 10, which w ere consistent with our results ( Izquierdo and Qu erol, 2012 ) The higher Cr concentration in fly ash samples was related with the high Cr total concentration of these samples (Table 6 8) The higher Se concentration in fly samples

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127 from facility A B E F and G was related with the high solubility of S e at strongly acidic and alkali pH (Table 6 8 ) ( Izquierdo and Querol, 2012 ) In addition, a lmost all fly ash and bottom ash samples (except 1 slag from facility A 2 bottom ash from facility B and F and 1 slag from facility E ) exceeded t he maximum contaminant level for As (Figure 6 14) A rsenic is primarily associated with As bearing pyrite in coals which is decomposed during the combustion and gives rise to a dominant surf ace association of sparingly soluble arsenate species in the ash ( Goodarzi et al. 2008 ) As an oxyanionic species, As is characterized by a pH dependent leaching : As leach es from acidic fly ash increase with pH whereas in alkaline fly ash this trend is reversed displaying a plateau of maximum solubility in the pH 7 11 range ( Izquierdo and Querol, 2012 ) Such variability is ascrib ed to adsorption processes and the interaction b etween arsenic and other species in CCR such as calcium, Fe oxyhydroxides and Al and Mn oxides ( Izquierdo and Querol, 2012 ) Ca plays an important role in controlling As leaching behavior from CCR samples since t he formation of insoluble Ca arsenate is common in Ca rich CCR at alkaline pH, while low Ca CCR provides less chance for As to precipitate ( van der Hoek et al. 1994 ) Fe oxyhydroxides and Al and Mn oxides which exist in CCR samples play a major role in As scavenging from pH 3 to pH 8 ( Cornelis et al. 2008 ; Izquierdo and Querol, 2012 ) The affinity of arsenate for metal oxides decr eases above pH 8, which leads to the increase of As leachability Our results were consistent with this trend except two acidic fly ash samples from facility A and F. Highest As values were observed in these two samples, which might be explained by their r elatively higher total As concentration (72.6

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128 mg/kg for fly ash sample from facility A and 66.1mg/kg for fly ash sample from facility F) (Table 6 8). For Be except for 3 fly ash samples from facility A C and F all sample s met the cleanup level (Table 6 9 ). For Co, all samples met the maximum contaminant level except one fly ash sample from facility C (Table 6 9 ). For Cd, except for 1 bottom ash sample from facility S and 5 fly ash samples from facility A E and F all samples met the maximum contaminan t level (Table 6 9 ). For Mo, all fly ash samples and 1 bottom ash sample from facility A and 1 bottom ash sample from facility G exceeded the maximum contaminant level which was likely caused by the high leachability of Mo at alkali pH range (Table 6 9 Figure 6 15 ). For Sb, except facility G all the other 6 facility contained fly ash and/or bottom ash samples did not meet the maximum contaminant level ( 5 sample s from facility A 2 sample s from facility C and 1 sample from facility B D E and F ) (Figu re 6 16) The maximum contaminant level for Tl is 2 ug/L and 9 samples (8 fly ash samples and 1 bottom ash sample) were slightly higher than this value. For Mo, 1 Fly ash sample from facility A was almost 114 times higher than the maximum contaminant lev el which indicated that Mo leached concentration seemed to peak at extreme low pH (Table 6 9 Table 6 10 ). Only 1 Fly ash sample from facility A exceeded the maximum contaminant level for Hg, whic h w as related with the relatively high total Hg concentrat ion and low pH. Those numbers exceeding groundwater maximum contaminant level are labeled as M (Table 6 11 ).

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129 Table 6 7 SPLP concentrations of elements Ca, K, Na, Mg, Al, Fe, and Zn (g/L) Samples Ca K Na Mg Al Fe Zn A Fly ash 1 313,212 6,150 6,445 924 11,339 M 4,753 M 115 A Fly ash 2 520,270 6,811 7,051 2,018 6,897 M 11,46 9 M A Fly ash 3 271,887 8,920 8,802 8,534 9,507 M 4,584 M A Fly ash 4 329,370 5,307 2,086 5,081 50,673 M 2,204,797 M A Bottom ash 10,011 865 2,706 19,087 5,468 M 19,445 M A Slag 87. 3 331 825 99.3 A FGD 413,168 125 591 1,051 10.3 B Fly ash 1 30,905 3,741 3,709 2,061 678 M 270 B Fly ash 2 436,121 7,267 20,032 49.0 817 M 947 M 27.0 B Fly ash 3 446,825 7,985 18,914 792 730 M 575 M 34.0 B Bottom ash 1 13,776 317 183 167 1,079 M 3.35 B Bottom ash 2 4,283 271 31.0 3.65 521 M 69.0 6.20 B Bottom ash 3 13,142 388 1,427 808 664 M 745 M B Bottom ash 4 4,301 103 166 84.0 366 M 141 B FGD 270,772 47.0 833 297 119 140 C Fly ash 258,185 1,720 13,979 35,483 3,338 M 18,818 M 2,838 C B ottom ash 132,198 1,686 16,185 19,914 243 M 365 M 33.0 D Fly ash 97,560 6,293 35,608 40.0 8,103 M 244 4.50 E Slag 1 512 399 122 90.2 378 M 442 M E Slag 2 4,128 222 811 265 69.3 215 E Slag 3 5,342 87.3 2,063 888 1,882 M 1,381 M 3118 F Fly ash 74,353 10,69 8 5,318 13,257 16,439 M 3,523 M 479 F Bottom ash 70,890 10,454 159,911 431,781 911 M 944 M 393 F FGD 458,541 1,801 12,527 41,408 167 134 13.0 G Fly ash 249,201 14,759 23,410 347 2,282 M 3,178 M 25.0 G Bottom ash 29,088 1,538 4,330 265 880 M 885 M 4.75 G FGD 3 30,177 205 2,331 658 227 M 1,109 M : Below the detection limit. The detection limit of ICP MS for K, Fe Zn is 0. 1, 0.94 and 0.9 2 g/L

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130 Figure 6 9 SPLP concentration of Al in CCR samples. Figure 6 10 SPLP concentration of Fe in CCR samples.

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131 Table 6 8 SPLP concentrations of elements V, Mn, Pb, Ni, Cr, Cu, As and Se (g/L) Samples V Mn Pb Ni Cr Cu As Se A Fly ash 1 695 M 36. 7 2.67 135 M 31. 7 77.3 M 285 M A Fly ash 2 90 4 M 86.3 M 3.67 247 M 357 M 50.3 111 M 165 M A Fly ash 3 801 M 32.3 21.7 M 52.0 102 M 27.0 80. 7 M 506 M A Fly ash 4 282 M 9,520 M 11,403 M 348 M 123 M 748 587 M 101 M A Bottom ash 110 M 466 M 43.7 28.3 22. 7 25.0 M A Slag 2.67 196 M 161 4.17 A FGD 3.00 23.3 0.33 7.00 B Fly ash 1 66 .0 M 47.0 1.60 15.0 11.0 14.0 81.0 M 165 M B Fly ash 2 129 M 14.0 19.0 M 0.95 76.0 5.10 22.0 M 252 M B Fly ash 3 93.0 M 7.95 21.0 M 1.40 65.0 3.30 14.0 M 242 M B Bottom ash 1 55.0 M 0.30 0.95 1.60 27.0 M 1.75 B Bottom ash 2 22.0 0.65 2.20 0.95 18.0 15.0 M B Bottom ash 3 23.0 19.0 1.70 1.30 6.50 1.50 14.0 M 1.30 B Bottom ash 4 21.0 3.40 0.65 2.15 7.60 0.70 B FGD 2.15 8.25 0.15 1.70 1.55 0.65 11.0 C Fly ash 23,789 M 346 M 154 M 7,714 M 97.0 195 40.0 M 32.0 C Bottom ash 8,094 M 19.0 20.0 M 419 M 13.0 19.0 13.0 M 2.05 D Fly ash 252 M 6.75 3.00 1.60 189 M 12.0 52.0 M 7.05 E Slag 1 2,065 M 5,042 M 3.33 8.67 11.7 M 32.2 E Slag 2 454 M 10,083 M 0.33 1.67 8.67 106 M E Slag 3 6,344 M 125 M 43.0 M 15,974 M 11.3 7.67 87.2 M 82.2 M F Fly ash 112 M 307 M 35.3 M 310 M 54 .0 1,395 M 87.3 M 78.7 M F Bottom ash 27.0 372 M 10.0 318 M 3.00 4.33 F FGD 26.3 78.0 M 11.0 11.7 5.00 12.7 4.00 G Fly ash 192 M 15.0 11.0 5.05 125 M 21.0 46.0 M 58.0 M G Bottom ash 32.0 15.0 1.45 0.60 8.25 1.55 17.0 M 6.00 G FGD 8.65 16.0 1.75 0.05 16.0 4.20 3.95 19.0 : Below the detection limit. The detection limit of ICP MS for V, Mn Pb, Ni, Cr, Cu, As and Se is 0.31, 0.34, 0.23, 0.73, 0.5, 0.7, 0.64 and 0.52 g/L

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132 Figure 6 11 SPLP concentration of V in CCR samples. Figure 6 12 SPLP concentration of Pb in CCR samples.

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133 Figure 6 13 SPLP concentration of Ni in CCR samples. Figure 6 14 SPLP concentration of As in CCR samples.

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134 Table 6 9 SPLP concentrations of elements Be, Co, Mo, Sb, Tl, Hg and Cd (g/L) Samples Be Co Mo Sb Tl Hg Cd A Fly ash 1 2.00 1. 7 2,938 M 34.3 M 16.3 M 0.333 5.33 M A Fly ash 2 4.67 4.67 3,136 M 20.0 M 18.0 M 8.33 M A Fly ash 3 2.00 4.00 3,073 M 33.0 M 42. 7 M 5.67 M A Fly ash 4 3.00 70.3 395 M 13. 7 M 227 M 146 M 4.33 A Bottom ash 1.67 61.7 32.0 34. 7 M 3.00 1.00 0.67 A Slag A FGD 10.3 1.00 1.00 B Fly ash 1 0.40 7.25 298 M 29.0 M 1.30 0.10 B Fly ash 2 0.55 0.65 1,509 M 1.25 2.05 M 0.40 4.25 B Fly ash 3 0.30 0.35 1,375 M 0.55 3.25 M 3.80 B Bottom ash 1 13.0 2.15 1.40 B Bottom ash 2 0.05 0.15 3.60 0.30 B Bottom ash 3 0.10 0.25 15.0 0.25 0.10 B Bottom ash 4 0.10 16.0 0.35 0.10 B FGD 0.05 2.55 0.10 C Fly as h 9.00 M 282 M 1,038 M 98.0 M 7.10 M 4.1 C Bottom ash 0.40 15.0 387 M 91.0 M 0.10 0.05 D Fly ash 0.35 1.00 595 M 13.0 M 0.60 0.55 E Slag 1 12.0 88.0 M 2.00 E Slag 2 24.0 116 M 5.00 1.00 E Slag 3 0.50 67.5 620 M 73.5 M 4.00 M 23.2 M F Fly ash 3 6.0 M 83.7 64.0 M 19.0 M 43.0 M 1.67 24.0 M F Bottom ash 0.67 40.7 3.33 0.67 1.00 6.33 M F FGD 7.33 1.00 0.33 G Fly ash 1.15 1.50 2,269 M 2.20 5.50 M 1.80 4.50 G Bottom ash 0.15 0.30 81.0 M 0.90 0.75 G FGD 0.05 0.15 20.0 0.05 0.35 : Below the detection limit. The detection limit of ICP MS for Be, Co, Mo, Sb, Tl, Hg and Cd is 0.64, 0.41, 0.76, 0.08, 0.51, 1 and 0.17 g/L

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135 Figure 6 15 SPLP concentration of Mo in CCR samples. Figure 6 16 SPLP concentration of Sb in CCR samples.

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136 Table 6 10 Solution pH after SPLP test Sample pH A Fly ash 1 10.2 A Fly ash 2 11.6 A Fly ash 3 9.06 A Fly ash 4 1.83 A Bottom ash 6.41 A Slag 5.72 A FGD 7.06 B Fly ash 1 7.8 B Fly ash 2 12.1 B Fly ash 3 12.2 B Bottom ash 1 10.9 B Bottom ash 2 10.1 B Bottom ash 3 9.29 B Bottom ash 4 9.27 B FGD 8.23 C Fly ash 5.97 C Bottom ash 8.81 D Fly ash 11.5 E Slag 1 6.83 E Slag 2 6.56 E Slag 3 5.23 F Fly ash 3.97 F Bottom ash 4.99 F FGD 8.0 3 G Fly ash 11.6 G Bottom ash 9.21 G FGD 8.04

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137 Table 6 1 1 Florida soil cleanup target level (mg/kg) and maximum contaminant level in drinking water (g/L) Contaminants Residential soil Industrial soil Drinking Water Aluminum 80,000 NA 200 Antimony 2 7 370 6 Arsenic 2.1 12 10 Beryllium 120 1,400 4 Cadmium 82 1,700 5 Chromium IV 210 470 100 Chromium III 110,000 NA 100 Cobalt 1,700 42,000 140 Copper 150 89,000 1,000 Iron 53,000 NA 300 Lead 400 1,400 15 Manganese 3500 43,000 50 Mercury 3 17 2 Molybdenum 440 11,000 35 Nickel 340 35,000 100 Selenium 440 11,000 50 Thalium 6.1 150 2 Vanadium 67 10,000 49 Zinc 26,000 630,000 5,000 NA: not available

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138 Research Findings The 27 CCR samples collected from 7 Florida coal power facilities were groupe d into three categories: fly ash, bottom ash and FGD residues. There were 11 fly samples, 12 bottom ash samples (bottom ash and slag) and 4 FGD samples. The pH of Florida CCR samples were grouped into 3 categories: low medium and high pH (Table 6 3) Fly ash samples were mainly in low or high pH categories, while bottom ash and FGD samples were dominant ly in the medium category. Review of the data presented in Table 6 4, Table 6 5 and Table 6 6 showed that m ajor elements including Ca, K, Na, Mg, Al, Fe, and Zn were the highest in concentrations (>100 mg/kg) in CCR samples followed by trace elements of V, Mn, Pb, Ni, Cr, Cu, As, Mo and Se (10 100 mg/kg) The trace elements Be, Co, Sb, Tl, Hg and Cd show ed the lowest concentration (<10 mg/kg) The higher content of major elements could be explained by the relatively higher content of those elements in coal sources. The abundance of trace metals of V, Mn, Pb, Ni, Cr, Cu, As, Mo and Se might be due to the enrichment of these elements during combustion proces s. The lowest concentration of trace elements Be, Co, Sb, Tl, Hg and Cd probably caused by their relatively low content in coal source or their volatile behavior (such as Hg) in combustion systems remain ing in the gas phase during passage through the plant Fly ash and bottom ash samples s howed higher elemental content than FGD samples. Compared with other CCR samples collected from around the US, t otal Al and Fe concentrations were lower while Ni and V in facility C and E maximum val ue of other US CCR samples (Ni: 1,267 mg/kg; V: 652 mg/kg). The difference between Florida CCR sample and other US CCR samples might be due to the different coal source or even the diffe rent age of coals from the same source.

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139 Concentrations of the rest of the elements were in agreement with other US CCR samples. F ly and bottom ash samples collected from facility A C and E exceeded the residential soil cleanup level for Pb, Cu, Sb, Tl and Hg. The variance of these elements content might be due to the diffe rent coal source. In addition, As, Ni and V concentrations in fly ash and bottom ash samples collected from all 7 facilities were even higher than their industrial soil cleanup levels which could be explained by the enrichment behavior of As, Ni and V dur ing combustion process The leaching behavior of Florida CCR samples w as tested by modified SPLP method. The elemental concentrations in the leachate were compared with the maximum contaminant level in drinking water to show the potential risk for contamin ating drinking water. Fly ash and bottom ash samples showed up to 100 fold highe r concentration than FGD samples, which was explained by their higher total concentrations. Concentrations of major elements Ca, K, Na and Mg in the leachate were the highest ; however, Ca, K, Na, and Mg were not regarded as elements of concern. Al, Fe, V, Ni, As, Mo and Sb concentrations in the leachates from most samples were higher than the maximum contaminant level, which was related to the total concentrations and specific p H values. The leach ate results exceed the maximum contaminant level in drinking water for Zn, Mn, Pb, Cr, Cu, Se, Be, Co, Tl, Hg and Cd although these results were not observed in all CCR samples from Florida.

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140 CHAPTER 7 CONCLUSION S This study was divided in two sections : (1) study the mechanisms of Cr(VI), Hg(II) and As(III) removal by sugar beet tailing or Brazilian pepper derived biochars; and (2) characterize 27 coal combustion residues collected from 7 Florida coal power plants. For the first objectiv e, t o our knowledge, few data are available for Cr(VI), Hg(II) and As(III) removal by biochars as well as the associated underlying mechanisms. Hence, t his study was useful in determining the Cr(VI), Hg(II) and As(III) removal mechanisms and impact factors by sugar beet tailing or Brazilian pepper derived biochar S ugar beet tailing biochar was effective in remov ing Cr(VI) from aqueous solution. The Cr(VI) sorption process could be described by the pseudo second order model (R 2 >0.99) and reached equilibrium after 16 h. The sorption data fitted to the Langmuir model (R 2 =0.99) with the maximum sorption capacity being 123 mg/g. The optimum pH for Cr sorption was 2 with 98% removal of Cr(VI). The Cr(VI) sorption increased from 19.8 to 88.5% as the biochar conten t increased from 0.2 to 8.0 g/L at pH 2.0 FTIR analysis suggested that carboxylate and hydroxyl groups m ight be involved in Cr (VI) removing process. Desorption of Cr from SBT biochar was 47 and 32% by 0.1 M NaOH and 0.1 M H 2 SO 4 XPS analysis coupled with the desorption experiment showed that b oth Cr(III) and Cr(VI) were sorbed on biochar and most of the Cr bonded to SBT biochar was Cr(III). The electrostatic attraction of Cr(VI) by biochar surface, reduction of Cr(VI) to Cr(III) ion, and the bond formation groups through complexation were probably responsible for Cr(VI) removal by SBT biochar.

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141 T he mechanisms of Hg sorption onto bi ochars produced from Brazilian pepper have been studied by using different analytical techniques. The Hg sorption capacity of BP300, BP450 and BP600 was 24.2, 18.8 and 15.1 mg / g based on Langmuir isothe rm. FTIR data suggested the participation of phenolic hydroxyl and carboxylic groups in Hg sorption by biochars. XPS analysis showed that 23 31% and 77 69% of sorbed Hg was associated with carboxylic and phenolic hydroxyl groups in biochars BP300 450 where as 91% of sorbed Hg was associated with graphite like domain on aromatic structure in BP600 biochar, which were consistent with flow calorimetry data. Based on flow calorimetry, sorption of K and Ca onto biochar was exchangeable with the m olar heat of sorp tion of 3.1 kJ / mol By comparison, Hg sorption was probably via complexation with functional groups as it was not exchangeable by K or Ca with molar heat of sorption of 19.7, 18.3 and 25.4 kJ / mol for BP300, BP450 and BP600. Our research suggested that H g was irreversibly sorbed via complexation with phenolic hydroxyl and carboxylic groups in low temperature biochars (BP300 and BP450) and graphite like structure in high temperature biochar (BP600). Dissolved organic matter (DOM) can serve as both electron donor and acceptor during Cr reduction and As oxidation, reducing their adve rse impact on the environment. We evaluated the impact of DOM from two biochars (sugar beet tailing and Brazilian pepper) on Cr ( VI ) reduction and As ( III ) oxidation in both ice and aqueous phases with a soil DOM as control. We examined (1) the influence of pH, temperature and DOM concentration on Cr ( VI ) reduction and As ( III ) oxidation, and (2) the mechanisms governing Cr and As transformation in the system. Increasing DOM conc entrat ion from

PAGE 142

142 3 to 300 mg C/L enhanced Cr ( VI ) reduction from 20% to 100% and As ( III ) oxidation from 6.2% to 25%; however, Cr ( VI ) reduction decreased from 80 86% to negligible while As ( III ) oxidation increased from negligible to 18 19% with increasing pH from 2. 0 to 10.0. Electron spin resonance study suggested semiquinone radicals in DOM were involved in As ( III ) oxidation while fourier transform infrared analysis suggested that carboxylic groups in DOM participated in both Cr ( VI ) reduction and As ( III ) oxidation. During Cr ( VI ) reduction, part of DOM (~10%) was oxidized to CO 2 The enhanced conversion of Cr ( VI ) (1.5 1.8 fold) and As ( III ) (1.09 1.14 fold) in the ice phase compared to aqueous phase was due to the freeze concentration effect with elevated concentratio ns of electron donors (DOM and Cr ( VI ) ), electron acceptors (semiquinone radicals and As ( III ) ), protons and hydroxyls in the grain boundary. Though DOM enhanced both Cr ( VI ) reduction and As ( III ) oxidation, Cr ( VI ) reduction coupled with As ( III ) oxidation occ urred in absence of DOM. The role of DOM, Cr ( VI ) and/or As ( III ) in Cr and As transformation may provide new insights into their speciation and toxicity in cold regions. For the second objective, t he 27 CCR samples collected from 7 Florida coal power facili ties were grouped into three categories: fly ash, bottom ash and FGD residues. There were 11 fly samples, 12 bottom ash samples (bottom ash and slag) and 4 FGD samples. The pH of Florida CCR samples were grouped into 3 categories: low medium and high pH. Fly ash samples were mainly with low or high pH; while bottom ash and FGD samples dominant the medium category. The m ajor elements including Ca, K, Na, Mg, Al, Fe, and Zn showed the highest concentrations (>100 mg/kg) in CCR samples followed by trace el ements of V, Mn, Pb,

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143 Ni, Cr, Cu, As, Mo and Se (10 100 mg/kg) The trace elements Be, Co, Sb, Tl, Hg and Cd show ed the lowest concentration (<10 mg/kg) The higher content of major elements could be explained by the relatively higher content of those eleme nts in coal sources. The abundance of trace metals of V, Mn, Pb, Ni, Cr, Cu, As, Mo and Se might be due to the enrichment of these elements during combustion process. The lowest concentration of trace elements Be, Co, Sb, Tl, Hg and Cd probably caused by t heir relatively low content in coal source or their volatile behavior (such as Hg) in combustion systems remaining in the gas phase during passage through the plant. Fly ash and bottom ash samples showed higher elemental content than FGD samples. Compared with other CCR samples collected from around the US, total Al and Fe concentrations were up to 40 times lower, while Ni and V in facility C and E samples exceeded the maximum value of other US CCR samples (Ni: 1,267 mg/kg; V: 652 mg/kg). Concentrations o f the rest of the elements were in agreement with other US CCR samples. F ly and bottom ash samples collected from facility A C and E exceeded the residential soil cleanup level for Pb, Cu, Sb, Tl and Hg. In addition, As, Ni and V concentrations in fly ash and bottom ash samples collected from all 7 facilities were even higher than their industrial soil cleanup levels. The difference among Florida CCR samples and between Florida and other US CCR samples might be due to the different coal source or even the different age of coals from the same source. The leaching behavior of Florida CCR samples were tested by modified SPLP method. The elemental concentrations in the leachate were compared with the maximum contaminant level in drinking water to show the poten tial risk for contaminating drinking water. Fly ash and bottom ash samples showed up to 100 fold

PAGE 144

144 higher concentration than FGD samples, which was explained by their higher total concentrations. Al, Fe, V, Ni, As, Mo and Sb concentrations in the leachates f rom most samples were higher than the cleanup target level, which was related to the ir total concentrations and specific pH values. The leachate results exceed the maximum contaminant level in drinking water for Zn, Mn, Pb, Cr, Cu, Se, Be, Co, Tl, Hg and C d, although these results were not observed in all CCR samples from Florida.

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159 BIOGRAPHICAL SKETCH Xiaoling Dong, the only child of her parents, was born and brought up in Mianchi, China. She went to University of Science and Technology of China to pursue her life science Following this she switched her major to environmental science and obtained her master degree in Earth and Space Science D epartment at the same universit y. In 2009, she came to University of Florida to purse a PhD degree in Soil and Water Science. She was a research assistant and worked on biochar and Florida coal combustion residues during her study at University of Florida.