CHARACTERIZATION OF TRACE ELEMENTS IN DRY FLUE GAS DESULFURIZATION (FGD) BY-PRODUCTS
DISSERTATION
Presented in Partial Fulfillment of the Requirements for
the Degree Doctor of Philosophy in the Graduate
School of The Ohio State University
By
Panuwat Taerakul, B.ENG., M.B.A., M.S.
*****
The Ohio State University 2005
Dissertation Committee: Approved by
Professor Harold Walker, Adviser
Professor Linda Weavers
Dr. Tarunjit Butalia Adviser
Professor Yu-Ping Chin Civil Engineering Graduate Program
ABSTRACT
This study investigates the amount, distribution, variation and fate of trace elements in FGD by-product. Dry FGD by-product including lime spray dryer (LSD) ash have possible uses in agricultural and construction applications. However, the variability in the chemical properties, especially levels of trace elements, of LSD ash is a concern due to the possible release to the environment. Changes in levels of inorganic constituents in LSD ash collected from the McCracken Power Plant were measured. Variability of elemental composition in leachate and bulk chemical properties (i.e. available lime index
(ALI), calcium carbonate equivalent (CCE), and total neutralization potential (TNP)) were also examined. Little variability over different time periods (e.g., daily to yearly) and little variability between samples collected from different particle collection hoppers were observed. Trace elements including As, Se, and Hg in LSD ash and in the leachate did not surpass limits for land application (EPA 503 Rule) or RCRA. Therefore, LSD ash is a consistent material which can be beneficially re-used in an environmentally sound manner.
ii Further study of LSD ash samples was conducted to examine the distribution of
trace elements in different fractions. LSD ash was fractionated using a 140-mesh (106
µm) sieve into two fractions: fly ash/unburned carbon-enriched fraction (>106 µm) and a
calcium-enriched fraction (<106 µm). A lithium heteropolytungstate solution with a specific gravity of 1.84 g/mL was further used to separate unburned carbon and fly ash- enriched fractions from the >106 µm fraction. The results show that the concentration of
As was consistently greater in the calcium-enriched fraction, while the Hg concentration was significant in all fractions. Specific surface area was found to be a factor controlling the levels of mercury in LSD ash fractions. The As concentration was also significant in economizer ash while Hg was undetected, suggesting that sorption of As on the surface of fly ash occurs at high temperature (600 °C) while sorption of Hg does not. As and Hg were also found to be more stable in the calcium-enriched fraction possibly due to the significant level of Ca and greater pH of the leachate. Results suggest that As and Hg are stable in LSD ash due to significant amounts of Ca. However, when LSD is disposed in a landfill, dissolution of Ca may lower the pH and calcium concentration in the leachate which may facilitate the release of As and Hg.
The study of trace elements in LSD ash was used as a base-line for investigating trace elements in dry FGD by-product collected from the Ohio State Carbonation and
Ash Reactivation (OSCAR) process. The OSCAR process is a pilot-plant of a new dry
FGD system used on a slip stream of flue gas from the McCracken Power Plant. Two different sorbents (i.e., regenerated sorbent and supersorbent) were tested using the
OSCAR process. Levels of trace elements, particularly As, Se and Hg, were found at greater levels compared to LSD ash. Higher concentrations of trace elements were found
iii in OSCAR baghouse samples, compared to cyclone samples, due to lower temperature and high surface area of baghouse particles. Operational parameters, such as sorbent injection rate and flue gas flow rate, influenced levels of As and Se in the cyclone samples. For example, the As concentration increased as flue gas flow rate increased while the Se concentration increased as the sorbent injection rate increased. Based on the
RCRA limits, leaching tests indicated that all cyclone samples are not hazardous. Results suggest that OSCAR sobents are effective for capturing trace elements, and the OSCAR by-product can be beneficially reused.
iv ACKNOWLEDGMENTS
I would like to thank, Dr. Harold Walker, for his advice and encouragement during my PhD research. He helped me stay focused on producing quality work. His patience, sincerity and kindness are greatly appreciated.
I also thank Dr. Linda Weavers, Dr. Taranjit Butalia and Dr. Yu-Ping Chin for serving me as my committee members. I am grateful to Dr. Linda Weavers who gave suggestions and comments for my research and Dr. Taranjit Butalia who provided me a lot of information about coal combustion products.
It has been a pleasure and I am thankful to work with the following lab members:
Mikko, Sunny, Jason, Yi-Fang, Jing, Eun-Kyoung, Maggi, Julie, Mike, Eric, Dong, Li-
Mei, Ziqi, Yu-sik, Aaron, Hiong, Mustafa, Vibhash, Laura, Sindhu, Dawn, Clayton, and
Ya-ning. Everybody has been very helpful to me and I have learned so much from working with others. Our lunch break is always fun for many of us to go out and experience new food. Getting together outside of the lab allows me to connect to our personal lives.
I thank Dr. Danold Golightly for his kind suggestions on the operation of ICP-
OES. I appreciate Doug Beak and Kevin Jewel for providing me acid digestion techniques and Ray Hunter for his design of a tool used for sampling in the power plant. I
v also would like to thank staff at the McCracken Power Plant for their assistance during the sampling of LSD ash and OSCAR samples.
Finally, I am fortunate to have a strong support and an encouragement from my family and my girlfriend. Especially my parents, I could not imagine how much love and support that they give me. Without them, it would be so hard for me to bounce back when
I am in tough times.
vi VITA
December 5, 1971 ………………………... Ratchaburi, Thailand
1993 ………………………………………. B.ENG., Environmental Engineering,
Chulalongkorn University, Thailand
1993-1994………………………………… Environmental Engineer, Progress
Technology Consultant, Bangkok, Thailand
1997 ………………………………...... M.B.A., Finance, Wright State University ,
Dayton, Ohio
2000………………………………………. M.S., Environmental Engineering, The
Ohio State University
1998-present ……………………………… Graduate Research Associate, The Ohio
State University
PUBLICATIONS
Research Publication
1. Taerakul, Panuwat, Lamminen, Mikko, He, Yongtian, Walker, Harold W.,
Tranina, Samuel J., Whitlatch, Earl, “Long-Term Behavior of Fixated Flue Gas
Desulfurization Material Grout in Mine Drainage Environments”, Journal of
Environmental Engineering 130(7), 816, (2004).
vii 2. Harold W. Walker, Panuwat Taerakul, Tarunjit Butalia, William E. Wolfe, and
Warren A. Dick, Minimize and Use of Coal Combustion By-Products (CCBs): Concept and Applications. In: Ghassemi A. editors. The Handbook of Pollution Control and
Waste Minimization, New York: Marcel Dekker, 2002.
FIELDS OF STUDY
Major Field: Civil Engineering
viii TABLE OF CONTENTS
Page
Abstract……………………………………………………………………………………ii
Acknowledgments ………………………………………………………………………...v
Vita ……………………………………………………………………………………....vii
List of Figures …………………………………………………………………………...xiv
List of Tables ……………………………………………………………………………xvi
Chapters
1. INTRODUCTION AND BACKGROUND...... 1
1.1. Flue Gas Desulfurization (FGD) Process...... 1
1.2. Production and Utilization of FGD By-Product...... 5
1.3. Trace Elements in FGD By-Product...... 6
1.4 Trace Element Characterization of Dry FGD Materials ...... 9
1.5. Dissertation Overview...... 11
2. VARIABILITY OF INORGANIC CONSTITUENTS IN LIME SPRAY DRYER ASH15
2.1 Abstract ...... 15
2.2 Introduction ...... 16
2.3 Experimental...... 18
2.3.1 Sample Collection ...... 18
ix 2.3.2 LSD Ash Characterization...... 19
2.3.3 Inorganic Analysis ...... 20
2.3.4 Coal Sample Analysis ...... 21
2.4 Results and Discussion...... 21
2.4.1 Variability in Inorganic Composition of LSD Ash...... 21
2.4.2 Variability in Leaching of Inorganic Elements and Bulk Chemical Properties
of LSD Ash...... 28
2.5 Conclusions ...... 30
3. DISTRIBUTION OF ARSENIC AND MERCURY IN LIME SPRAY DRYER ASH41
3.1 Abstract ...... 41
3.2 Introduction ...... 42
3.3 Experimental...... 43
3.4 Results and Discussion...... 46
3.4.1 General Characteristics of LSD Ash Fractions...... 46
3.4.2 Distribution of Arsenic in LSD Ash...... 48
3.4.3 Distribution of Mercury in LSD Ash ...... 50
3.4.4 Factors Impacting the Distribution of Arsenic and Mercury...... 52
3.4.5 Leaching Characteristics of LSD Ash Fractions...... 55
3.5 Conclusions ...... 57
4. CHARACTERIZATION OF BY-PRODUCTS GENERATED FROM THE OHIO
STATE CARBONATION AND ASH REACTIVATION (OSCAR) PROCESS...67
4.1 Abstract ...... 67
4.2 Introduction ...... 68
x 4.3 Materials and Methods...... 70
4.3.1 Sample Collection ...... 70
4.3.2 General Characterization...... 70
4.3.3 Elemental Composition Analysis...... 71
4.3.4 Bulk Chemical Properties...... 71
4.4 Results and Discussion...... 71
4.4.1 Bulk Mineral and Surface Characterization of OSCAR Sorbents and By-
Products...... 71
4.4.2 Inorganic Element Concentrations of OSCAR Sorbents ...... 72
4.4.3 Inorganic Element Concentrations of OSCAR By-Products...... 73
4.4.4 Reuse Applications...... 80
4.5 Conclusions ...... 81
5. CONCLUSIONS AND FUTURE WORK ...... 89
5.1. Conclusions ...... 89
5.2. Future work ...... 91
APPENDIX A. LONG-TERM BEHAVIOR OF FIXATED FLUE GAS
DESULFURIZATION MATERIAL GROUT IN MINE DRAINAGE
ENVIRONMENTS...... 93
A.1 Abstract ...... 94
A.2 Introduction ...... 95
A.3 Site Description...... 97
A.4 Materials and Methods...... 99
A.4.1 Sampling Locations and Techniques...... 99
xi A.4.2 Chemical Analysis of Water Samples ...... 101
A.4.3 Analysis of Grout Core Samples...... 102
A.5 Results and Discussion...... 103
A.5.1 Long-Term Water Quality Trends...... 103
A.5.2 Geochemical Stability of Fixated FGD Material Grout ...... 106
A.5.3 Water Quality in Vicinity of Fixated FGD Material Grout ...... 108
A.6 Conclusions ...... 111
A.7 Acknowledgment ...... 112
APPENDIX B. MINIMIZATION AND USE OF COAL COMBUSTION BY-
PRODUCTS (CCBS): CONCEPTS AND APPLICATIONS...... 122
B.1 Abstract...... 123
B.2 Introduction and Background ...... 124
B.3 Federal Regulations Influencing CCB Generation and Use...... 125
B.4 Physical, Chemical and Engineering Properties of CCBs...... 127
B.4.1 Physical and Engineering Properties of CCBs...... 128
B.4.2 Chemical Properties of CCBs...... 129
B.5 Factors Affecting CCB Generation...... 132
B.5.1 Boiler Technology...... 132
B.5.2 Air Pollution Control Technology...... 135
B.5.3 Types of Coal...... 137
B.6 Strategies for Minimization of CCBs...... 139
B.6.1 Reduction at Source...... 139
B.6.2 Use of Coal Combustion By-Products...... 140
xii B.6.3 Treatment and Disposal...... 148
B.7 Life Cycle Assessment (LCA) Model for Minimization of CCBs ...... 148
B.8 Barriers to CCB Utilization ...... 150
B.9 Conclusions...... 151
B.10 Acknowledgment ...... 152
xiii LIST OF FIGURES
Figure Page
2.1. Unit Operation Diagram of Lime Spray Drying System at McCracken Power Plant34
2.2. X-Ray Diffraction Patterns of LSD Material from McCracken Power Plant ...... 35
2.3. SEM Images of LSD Material from McCracken Power Plant...... 36
2.4. Relative Standard Deviations of Elements in LSD Ash Collected at Different Time
Periods...... 37
2.5. Profiles of Hg, As and Se Concentrations in LSD Material Collected from
McCracken Power Plant...... 38
2.6. Relative Standard Deviations of Elements in Leachates from LSD Ash at Different
Time Periods...... 39
2.7. Profiles of As and Se Concentrations in Leachate Solution from TCLP Test of LSD
Material Collected from McCracken Power Plant...... 40
3.1 X-Ray Diffraction Patterns of LSD Ash Fractions of Sample 4...... 60
3.2 Scanning Electron Microscopic Images of LSD Ash Fractions of Sample 4...... 61
3.3 Distribution of Arsenic in LSD Ash Fractions...... 62
3.4 Distribution of Mercury in LSD Ash Fractions...... 63
3.5 Arsenic and Mercury Concentrations in LSD Ash Fractions as Functions of
Calcium, Specific Surface Area and Organic Carbon...... 64
xiv 3.6 Arsenic Concentrations in Leachates of LSD Ash Fractions...... 65
3.7 Mercury Concentrations in Leachates of LSD Ash Fractions...... 66
4.1 Operation Diagram of OSCAR Process...... 85
4.2. X-Ray Diffraction Patterns of LSD Ash, OSCAR Sorbents, and OSCAR By-
Product. RS is regenerated sorbent. SS is supersorbent. 4CY and 33CY are cyclone
samples. 4BH and 33BH are baghouse samples...... 86
4.3 Impact of Operating Condition to the Levels of As and Se in Cyclone Samples ....87
4.4 Relationships of Trace Element Concentrations and Ca Concentrations. Fig. 4.4a:
As and Se in cyclone samples. Fig. 4.4b: As, Se and Hg in baghouse samples...... 88
A.1 Representative geological cross-section of the Roberts-Dawson site...... 116
A.2 Site location and description...... 117
A.3 Flow rate of mine drainage water at surface water site 5...... 118
A.4 Flux of acidity, sulfur, iron, aluminum, calcium and boron at surface water site 5,
before and after placement of fixated FGD material grout...... 119
A.5 X-ray powder diffraction pattern of core sample collected from site 9906...... 120
A.6 X-ray powder diffraction pattern of core sample collected from site 2002...... 121
B.1 CCBs Production in million metric tons in the United States in 1998 (88)...... 159
B.2 Approximate Ash Distribution as a Function of Boiler Technology (4,56,103).....160
xv LIST OF TABLES
Table Page
2.1 Elemental Composition of Lime Spray Dryer Ash, Quick Lime, and Coal Collected
from McCracken Power Plant ...... 31
2.2 Results of TCLP Test of LSD Ash from McCracken Power Plant ...... 32
2.3 Results of ALI, CCE and TNP of LSD Ash from McCracken Power Plant...... 33
3.1 Elemental Compositions of Lime Spray Dryer Ash Fractions and Economizer Ash58
3.2 Specific Surface Area and Organic Carbon Content of LSD Ash Fractions ...... 59
4.1 Specific Surface Area of Raw Material (Lime and LSD Ash) and OSCAR Sorbents
and Samples...... 82
4.2 Elemental Composition of Sorbents, OSCAR By-Product and LSD Ash...... 83
4.3 Concentrations of Trace Elements from TCLP Tests...... 84
A.1 Concentrations of contaminants with either a primary maximum contaminant level
and/or secondary maximum contaminant level in the Clarion sandstone aquifer and
adjacent reservoir...... 113
A.2 Water quality in wells in the downdip area of the mine, installed either before
(9719) or after (9901, 9904, 2002) grouting operations. Water quality data for two
wells (9903 and 9906) installed in the upper mine works after grouting are also
shown...... 115
xvi B.1 Amount of CCBs Disposed of in Landfills in the United States in 1998 Compared
to Disposal of Municipal Solid Waste (MSW)...... 153
B.2 Amount of CCBs Produced in the United States in 1998 Compared to Traditional
Non-Fuel Mineral Commodities...... 154
B.3 Summary of Physical Characteristics and Engineering Properties of Fly Ash,
Bottom Ash, Boiler Slag and FGD Material (4,44,45,56,96,97,98,99)...... 155
B.4 Trace Elemental Composition of Fly Ash, Bottom Ash, Boiler Slag (56) and FGD
Material (101)...... 156
B.5 Range of Values Observed for TCLP Analysis of Dry FGD Materials (44,101) and
Ash (97)...... 157
B.6 Major CCB Applications and Environmental Benefits of CCB Use...... 158
xvii CHAPTER 1
INTRODUCTION AND BACKGROUND
1.1. Flue Gas Desulfurization (FGD) Process
Acid rain is a major environmental problem which occurs when sulfur dioxide
(SO2) and nitrogen oxides (NOx) in the atmosphere interact with water and oxygen. These acidic gases are mostly produced from the combustion of coal, especially high sulfur coal, in electric utilities [1]. Nitrogen in acid rain is also a concern. It results in the deposition of nitrate in aquatic systems. In the U.S., approximately 1000 Mt of coal is mined each year, 90% of which is used for electricity generation [2]. Approximately 49
% of total coal mined in the U.S. is obtained from Appalachia and Interior regions which contain high sulfur coal [3].
To reduce acid gas in the atmosphere, flue gas desulfurization (FGD) processes are required in coal-fired utilities to remove SO2 and NOx from the flue gas [4]. Based on phase II of the sulfur dioxide requirement in Title IV of the 1990 Clean Air Act, an allowance for the emission of SO2 was lowered from 2.5 to 1.2 lb/mmBtu, effective since
January 1, 2000 [5]. In addition, an emission requirement of NOx for all kinds of utility
1 boilers was established on January 1, 1997 [6]. On March 10, 2005, The Clean Air
Interstate Rule (CAIR) was established in order to rule the state implement plans for
emissions of SO2 and NOx in 28 eastern states of the U.S [7]. Based on CAIR, by 2015,
SO2 and NOx emissions are expected to be reduced by 70% and 60%, respectively [7].
Therefore, current regulations and expected future lower allowances of emission of SO2 and NOx will result in a significant increase in flue gas desulfurization at coal-fired utilities.
Currently, there are a large number of FGD processes used in coal-fired utilities.
Based on the water content of the residual solid, they can be generally categorized into
two different types; wet and dry systems. Wet FGD systems are currently the most
common FGD systems used in coal-fired utilities [8]. In general, wet FGD systems use an
alkaline slurry to interact with flue gas in a scrubber to remove acid gas. The scrubber or
absorber unit may come in different forms [8]. Based on the sorbent used, wet FGD
processes may be separated into two different subcategories, regenerated and non-
regenerated sorbents. Regenerated sorbents include sodium sulfite, magnesium oxide,
sodium carbonate, and amines. These sorbents are effectively used in the wet FGD
process and can be reused again after a regeneration process, however, the operational
costs of these wet FGD processes are high. Therefore, most utilities now use non-
regenerated sorbent systems [8].
For non-regenerated sorbent systems, lime and limestone are commonly used.
Lime or limestone is first ground into a powder to maximize dissolution and mixed with
water. Then, the sorbent slurry is pumped into the reactor tank or scrubber to interact
2 with the acid gas [9]. The reaction of sulfur with the lime or limestone slurry can be shown as;
+ + → • SO 2 CaO 0.5H 2O CaSO3 0.5H 2O (1)
+ + + → • SO 2 CaO 2H 2O 0.5O2 CaSO4 2H 2O (2)
and
+ + → • + SO 2 CaCO3 0.5H 2O CaSO3 0.5H 2O CO2 (3)
+ + + → • + SO 2 CaCO3 2H 2O 0.5O2 CaSO4 2H 2O CO2 (4)
The production of calcium sulfite or calcium sulfate depends on available oxygen. SO2 removal from the flue gas in wet FGD systems is high (>90%) and the amount of sorbent used in the wet system can be determined from the stochiometry of sorbent and SO2. In modern wet FGD systems, the mole ratio of calcium sorbent and SO2 ranges from 1.01 to
1.1 [9]. A mole ratio close to one indicates highly effective removal of SO2 by the sorbent with virtually complete conversion of lime or limestone to calcium sulfite or calcium sulfate. However, because the residual solid in a wet FGD system is in the form of a slurry, a dewatering process is necessary. In addition, scaling is also a concern in the
FGD process. Formation of calcium sulfate in the residual slurry results in significant deposition of particles in the reactor [9]. Additional space and equipment (i.e., pipe, air blower) are required to prevent scaling by calcium sulfate and dewater solids, which increases the capital cost for wet FGD systems.
For dry FGD systems, a dry sorbent or wet slurry is used to interact with SO2 in the flue gas. However, the resulting by-product is dry. As a result, there is no dewatering process involved in a dry FGD system. Current dry FGD technologies include lime spray
3 dryer (LSD), and dry sorbent injection systems. In a LSD system, slaked lime (Ca(OH)2) is injected into a spray dryer absorber through either a rotary atomizer or two fluid nozzle injectors to absorb SO2 from the flue gas [10]. Lime reacts with SO2 forming calcium sulfite (CaSO3) or calcium sulfate (CaSO4). The moisture in the slaked lime (reacted or unreacted) evaporates by the heat from the flue gas and SO2 may diffuse through the surface of unreacted lime causing a gas-solid reaction [10]. However, most SO2 is captured when the sorbent is wet [10]. Dry particulates, including calcium sulfate/sulfite, unreacted lime, and fly ash, are collected by a particle collection system; a baghouse
(fabric filter) or an electrostatic precipitator (ESP). The reaction of slaked lime and SO2 can be shown as;
+ → • + SO 2 Ca(OH ) 2 CaSO3 0.5H 2O 0.5H 2O (5)
+ + → • SO 2 Ca(OH ) 2 H 2O CaSO4 2H 2O (6)
The stochimetry of calcium to sulfur in the residual from an LSD system is about 2 for removal of SO2 at 85% [11]. Compared with a wet FGD system, an LSD system shows lower efficiency, thereby requiring more calcium sorbent for removing SO2 which results in higher operating costs. The SO2 removal efficiency for a dry system is lower than for a wet FGD system, however, the LSD technology has a lower capital investment [11].
The different kinds of dry sorbent injection systems vary by the temperature and location of sorbent injection into the system; furnace injection (1800-2400 ºC), economizer injection (800-1100 ºC) and post air heater injection (<350 ºC) [12]. In the furnace injection system, sorbent (Ca(OH)2 or CaCO3) decomposes or calcines to CaO and reacts with SO2 as shown;
→ + CaCO3 CaO O2 (7) 4 → + Ca(OH) 2 CaO H 2O (8)
+ + → SO2 CaO 0.5O2 CaSO4 (9)
The reaction (sulfation reaction) of SO2 with CaO occurs at a temperature range from 900 to 1200 ºC [11]. Reaction at a temperature below 900 ºC is too slow for effective capture by CaO. For the economizer injection system, sorbent (Ca(OH)2) directly captures SO2 without dehydration or calcination process of the sorbent at lower temperature. Similar to the sulfation reaction in a furnace injection system, the capture of SO2 drops significantly as the temperature decreases below 900 ºC [13]. For the post air heater injection system, a slaked lime slurry is injected by a rotary or twin-fluid slurry atomizer [11]. SO2 is captured by the moisture in the sorbent and then reacts with Ca(OH)2. The moisture is later vaporized by heat in the flue gas which is similar to the mechanism of an LSD system [11].
1.2. Production and Utilization of FGD By-Product
Since establishment of Title IV of the Clean Air Act in 1990, the production of
FGD by-product has significantly increased [4]. In 2002, 26.5 Mt of total FGD by- product (i.e., wet and dry systems) was produced in the U.S. [14]. In 2002, approximately
70% of the total production of FGD by-product was sent to landfills [14]. The disposal of
FGD by-product consumes landfill space and increases the operating cost of FGD due to the additional costs for handling and disposal. Therefore, it is a challenge to minimize the production and disposal of FGD by-product, and still reduce acid gases emitted from coal-fired power plants.
As mentioned earlier, wet FGD systems have a high efficiency for SO2 removal, having a mole ratio of calcium to sulfur close to one in residual solids [9]. However, the
5 construction of wet FGD systems involves a lot of processes including the particle collection system, scrubber, sorbent preparation, and dewatering systems. Thus, wet FGD systems may not be suitable for small coal-fired utilities due to high capital investments.
A simpler process such as a dry sorbent injection system may be more convenient to install in a small utility [11]. However, a typical mole ratio of calcium to sulfur for a dry
FGD system is close to two indicating that twice as much calcium must be used to capture sulfur, resulting in greater operating costs and greater production of FGD by- product [11].
There is significant interest in minimizing landfill space by utilizing FGD by- product in a variety of civil engineering applications. Studies have demonstrated the use of FGD by-product in construction, agriculture, and mine reclamation [15,16,17,18,19].
In construction applications, FGD by-product was found to be beneficial for use in embankments for highways, and for flowable fill [17, 20 ]. Other coal combustion products such as fly ash have also been used such as a replacement of cement in concrete
[21]. In agricultural applications, FGD by-product can be used as a substitute for lime and as a soil amendment [ 22 ]. FGD by-product also has been tested for use in the construction of livestock feeding and hay storage pads [23]. In mining applications, high alkalinity FGD by-product was used to reduce the production of acid mine drainage
(AMD) [16]. FGD by-product has been tested as a material for grouting, capping or filling in several projects to reduce the production and discharge of AMD [16,20].
1.3. Trace Elements in FGD By-Product
Concern about trace elements in FGD by-product (e.g., As, Se and Hg) limits the utilization of FGD by-product. A previous study found that the potential for release of
6 toxic materials is a barrier impacting the utilization of FGD by-product [20]. Butalia and
Wolfe state that there is a concern about the inconsistency in the composition of FGD by- product and lack of information regarding the potential release of trace elements from the
FGD by-product [20].
During the coal combustion process, trace elements (e.g., As, Se and Hg) are decomposed by heat and travel with the flue gas. These trace elements may be captured during the desulfurization process and become part of FGD by-product. In addition, trace elements are captured by fly ash. For example, As can be adsorbed by Ca on the surface of fly ash forming calcium arsenate [24]. Volatile As is also found to condense to arsenic oxide (As2O3) at a temperature below 1000 ºC [24]. Trace elements, such as As, Cd, Ni and Zn are carcinogenic [25]. Se is an essential element but the release of high levels of
Se was found to severely impact the environment [26,27]. Se in fly ash disposed in settling pond was found to be a source of Se that impacts surface water. Dissolved selenium can be accumulated in fish tissues and cause deleterious effects to fishes and birds [27].
Hg is toxic to the nervous system and has been linked to several neurological diseases such as Alzhheimer’s, Dementia, Autism, and Parkinson’s [28]. Hg was also found to accumulate in sediment and be converted to methyl mercury by bacteria which is extremely toxic [29]. In addition, Hg in sediment was found to be a source of Hg that can be transferred and accumulated in tissue of aquatic organism [29].
Approximately 48 tons of Hg is released from coal-fired utilities in the U.S every year [30]. To reduce the emission of Hg from coal fired utilities, the U.S. Environmental
Protection Agency (EPA) issued The Clean Air Mercury Rule to reduce the emission of
7 Hg from coal-fired utilities to 15 tons per year by 2018 [30]. This rule regulates the Hg emissions from new and existing coal-fired utilities. A trading permit system called cap- and-trade is used to create an incentive of continuous innovation of Hg reduction technology. There are two phases for the mercury reduction program. In the first phase, the Hg emission is expected to be reduced to 38 tons by taking advantage of “co-benefit” reduction of SO2 and NOx in CAIR. The second phase of The Clean Air Mercury Rule will reduce the Hg emission to 15 tons per year by 2018 [30].
Previous studies suggest that volatile trace elements, which are released during the combustion of coal, sorb onto the surface of fly ash at temperatures as high as 1000
ºC [24]. Calcium species, iron oxide, aluminum oxide and other metal oxides on the surface of fly ash oxidize trace elements, resulting in chemical bonding onto the surface of fly ash [24,31]. Unburned carbon has also been suggested to be an effective material for capturing volatile trace elements from flue gas [32,33,34]. Marato-Valer et al. indicated that oxygen functionality and halogen species promote the mercury capture on the carbon-rich fly ash [31]. Therefore, in a system which has no particle collection prior to the desulfurization process, trace elements associated with fly ash and unburned carbon are mixed with the sorbent material and collected at the particle collection system.
For dry injection systems, sorbent especially calcium sorbents (e.g., Ca(OH)2 or
CaCO3) can be calcinated at high temperature forming CaO [12]. CaO can react with volatile arsenic oxide in the flue gas forming calcium arsenate at high temperature (600-
1000 ºC) [35,36]. Se as selenium oxide was also reported to react with CaO forming
CaSeO3 at 600 ºC [37]. Therefore, dry calcium sorbents may remove trace elements in the flue gas at high temperature. However, removal of trace elements in dry injection
8 systems may be limited by the level of acid gas. For example, the reaction of CaO with volatile selenium oxide competes with the reaction of CaO and SO2 resulting in a significant reduction in Se removal [38,39]. The formation of CaSO4 may also inhibit the reaction of CaO with volatile selenium oxide. Thus, removal of trace elements depends on both the temperature during sorbent injection and the level of acid gas in the system.
Trace elements may be also captured by the lime slurry used in the wet FGD system or LSD system. Volatile arsenic was reported to be effectively captured by hydrated lime forming calcium arsenate [35]. Mercury as HgCl2 vapor was also reported to be captured by hydrated lime [40]. Moisture from the sorbent may also help to remove trace elements by capturing particulates. For example, As was reported to condense into fine As2O3 particles [41,42], and subsequently captured by droplets of Ca(OH)2 injected in the wet FGD system or LSD system.
1.4 Trace Element Characterization of Dry FGD Materials
The focus of this dissertation is on trace elements in dry FGD by-product. A review of the literature indicates very little work has been conducted on this topic. In one of the few studies in this area, dry FGD by-product including LSD ash was examined in order to investigate possible use as a substitute material in land reclamation [43,44,45]. In this study, LSD ash was collected from the McCracken Power Plant in 1991. It was found that LSD ash primarily consisted of portlandite (Ca(OH)2) and hannebachite
(CaSO3.0.5H2O). Portlandite in LSD ash suggests that there is a sufficient amount of alkalinity in LSD ash for neutralizing acid.
Elemental analysis of LSD ash indicated that concentrations of Ca and S are significant in LSD ash as the primary constituents of by-product from the desulfurization
9 process and excess sorbent used. Other elements such as Fe, Si and Al are primarily associated with fly ash. Trace elements including As, Be, Co, Cr, Cu, Fe, Ni, P, and Pb were also found to partition primarily to fly ash. Mass concentrations of these elements tend to increase as the LSD ash weathered in an open environment.
Leaching analysis using toxicity characteristic leaching procedure (TCLP) test
(EPA method 1311) indicated that concentrations of trace elements were all below the
RCRA limits and drinking water standards. A long-term leaching test was also conducted with different mixings of acidic mine spoils and LSD ash. It was found that adding of
LSD ash decreased the leachings of As, Cd, Cr, Cu and Ni which were below the maximum contaminant levels (MCL’s). However, Se concentrations were greater than the MCL in most samples and there was no correlation between Se concentrations and mixing of LSD ash. In this previous study, engineering tests and a field study also demonstrated that LSD ash can be successfully used as a structural fill material for a truck ramp.
Another previous study also investigated characteristics of LSD ash from two different coal-fired power plants [46]. They found that trace elements were mostly associated with fly ash suggesting that the lower emission levels of trace elements from these two power plants to the atmosphere may relate to the high ash content of coal used
[46].
Xu et al. suggested that the lower the operating temperature in LSD system not only increases the efficiency of desulfurization in LSD system but also increases the removal of vapor-phase trace elements. This previous study also indicated that the quality
10 (size, amount, and types) of the sorbent used in sorbent injection process had a significant impact to the sorption capacity of trace elements [47].
Examining of an impact of trace elements to the utilization of dry FGD by- product in agriculture applications were also found in previous studies [48,49]. Clark et al. indicated that trace elements (e.g., Ag, As, B, Cd, Cr, Cu, Hg, Mo, Ni, Pb, Se, and Zn) potentially constrained its use on land due to a risk of contamination in water and plant.
However, levels of these trace elements in soil amendment or plant grown were usually below regulated standards [48]. Wright et al. indicated that with proper use of dry FGD by-product, there was no impact of wheat seedling root growth from trace elements.
However, they found that levels of B, Se, As and Mo in rye glass grown increased when dry FGD by-product was used for soil treatment [49].
1.5. Dissertation Overview
This research focuses on the characterization of trace elements in FGD by-product and examines the possible impact to the environment due to the release of these trace elements during utilization. In chapter 2, the variability of inorganic constituents in LSD ash collected from the McCracken Power Plant was examined. Changes in the properties of FGD by-product over time may impact reutilization due to the added costs and risk associated with working with a variable material. In addition, significant changes in levels of trace element may increase the risk of environmental impact given that high levels may go undetected by routine monitoring. A hypothesis of this study is that variability of the properties of LSD ash may be influenced by longer sampling period and properties of coal and lime. Variation of other properties such as bulk chemical properties
(i.e., available lime index, calcium carbonate equivalent and total neutralization potential)
11 were also examined along with the characterization of inorganic elements. The variation of LSD ash at the McCracken Power Plant was also used as a base-line test for the study of dry FGD by-product from the Ohio State carbonation and ash reactivation (OSCAR) process which is discussed in chapter 4.
LSD ash is a heterogeneous material with multiple mineral components. To better understand the presence and fate of trace elements in LSD ash, it is useful to investigate the distribution of elements between different LSD ash fractions. LSD ash from the
McCracken Power Plant was fractionated to examine the distribution of trace elements in different LSD ash components. This study is discussed in chapter 3. At the McCracken
Power Plant, fly ash is partially removed at the economizer. However, a major portion of the fly ash enters the spray dryer absorber and is collected with the lime residual at the baghouse. Possible sorption of trace elements by fly ash, carbon and Ca(OH)2 was reported [24,31,33,34,35,36,40]. Therefore, the levels of trace elements in LSD ash may vary in different fractions (e.g., lime enriched or unburned carbon fractions). In addition, release of trace elements in LSD ash may also be controlled by the distribution of trace elements in the ash. The hypothesis that differences in the levels and characteristics of trace elements in different LSD ash fractions may result in differences in release trace elements is examined.
In chapter 4, trace elements in dry FGD by-products collected from the OSCAR process were examined. The OSCAR process is a dry FGD process, and was installed as a slip-stream process at the McCracken Power Plant. New OSCAR dry sorbents made from lime or LSD ash were shown at the lab-scale to improve sulfur capture when compared with a conventional dry FGD system [50,51]. Increasing the sulfur removal
12 efficiency decreases the use of lime resulting in lower operating costs which eliminates the main disadvantage of dry systems. It is then hypothesized that increasing of sulfur removal efficiency in OSCAR system may also increase removal of trace elements and increase levels of trace elements in by-product. Although the OSCAR process may lower the production of by-product, this study also examined the possibility for the utilization of this material by comparing the levels of trace elements with land application limits
(EPA 503 Rule) and comparing the leaching results with the Resource Conservation and
Recovery Act (RCRA) limits. Levels of trace element in dry FGD by-product from the
OSCAR process was compared with the base-line testing of LSD ash collected from the
McCracken Power Plant.
In appendix A, a long-term field study on the impact of the utilization of FGD by- product on the water quality at an underground coal mine was examined. In this study, a fixated FGD material grout consisting of fly ash, filter cake and lime was injected into the down-dip portions of the Roberts-Dawson mine. Injected grout material was expected to coat exposed pyritic surfaces and create an impermeable barrier to AMD. After grouting, water was expected to fill the mine and eliminate the exposure of pyritic surfaces and reduce AMD formation. Fixated FGD material grout also provides a source of alkalinity to precipitate AMD constituents. The water quality of surface water and groundwater and surface at the mine and at the receiving stream was monitored to determine the impact from the fixated FGD material.
In appendix B, this study reviewed factors, including specific unit operations, operating conditions, and the source and type of coal which may impact the generation of coal combustion by-products (CCBs). A number of applications currently utilize CCBs
13 have been demonstrated, including agricultural amendment, construction of low permeability liners and road sub-base, and mine reclamation. Other factors impacting the utilization of CCBs including chemical composition and engineering properties were also described. Case studies demonstrate that with proper understanding of the properties of
CCBs, effective utilization of these materials results in decrease of solid materials entering landfills, decrease of greenhouse gas emissions, and conservation of existing natural resources.
14 CHAPTER 2
VARIABILITY OF INORGANIC CONSTITUENTS IN LIME SPRAY DRYER
ASH
(Modified from a version submitted to Fuel, In Press)
2.1 Abstract
Flue gas desulfurization (FGD) by-products, including lime spray dryer (LSD) ash, have many demonstrated uses. However, concern about the temporal variability in the chemical properties of this material has limited widespread utilization. To determine the variability in inorganic constituents, this study measured elemental composition, leaching properties, available lime index (ALI), calcium carbonate equivalent (CCE), and total neutralization potential (TNP) for a representative LSD ash. All parameters investigated showed little variability over different time periods (e.g., daily to yearly) and little variability between samples collected from different particle collection hoppers.
Metal concentrations including As, Se, and Hg in LSD ash and in the leachate did not surpass limits for land application (EPA 503 Rule) or limits for the determination of
15 hazardous waste as specified in the Resource Conservative and Recovery Act (RCRA).
The low variability in ALI, CCE, TNP, and inorganic and organic composition suggests that LSD ash is a consistent and environmentally benign material for agricultural and other engineering applications.
2.2 Introduction
Flue gas desulfurization (FGD) by-product is a residual material generated from processes used to remove sulfur dioxide from flue gas following coal combustion.
Approximately 26 Mt (metric) of FGD by-product are produced in the United States every year [4,52,53,54], with more than 18 Mt (72% of total production) sent to landfills.
To minimize landfilling, numerous studies have been conducted to examine beneficial uses of FGD by-product in a variety of applications including construction, agriculture and mine reclamation [15,16,19,23,48,55].
The lime spray dryer (LSD) system is the most common dry FGD technique used in utility coal fired power plants [4,56]. In 1998, the total capacity of boilers equipped with LSD systems in the U.S. was 11,315 MW, which is 80.4% of the total capacity equipped with dry FGD technology, or 11.8% of total capacity equipped with FGD technology [9]. During the LSD process, a fine spray of slaked lime (Ca(OH)2) is injected into the scrubber and reacts with sulfur oxides, resulting in the formation of calcium sulfate or calcium sulfite. Moisture in the reacted lime is lost due to the heat from the flue gas. The resulting dry calcium sulfite/sulfate mixture, along with fly ash, is collected as “LSD ash” by an electrostatic precipitator or a baghouse [4,56]. A number of researchers have studied the re-use of by-product produced from dry FGD systems, 16 including LSD systems, and associated environmental impacts [15,19,20,23,44,48,55].
The United States Environmental Protection Agency (USEPA) recommends that LSD ash be exempt from physical and chemical tests for hazardous material if used in a few specific applications. Concern remains, however, regarding the use of these materials due to the presence of trace elements [4].
An additional issue limiting the re-use of LSD ash, as well as other coal combustion by-products, is the perceived temporal variability of this material. A number of factors may influence the chemical composition of LSD ash over time including chemical compositions of lime and coal, and changes in plant operations. For example, an increase in un-reacted lime in the LSD ash may result in an increase in pH facilitating the release of some trace elements (e.g. arsenic) [57]. Abrupt changes in levels of trace inorganic and organic compounds are not detected by periodic monitoring and potentially could result in leachate values above regulatory levels, thus affecting human health.
Long-term changes in bulk properties (e.g., available lime index) discourage utilization due to the added investment required for effective use. Despite the concern over the variability of FGD by-product, little or no long-term data is available to evaluate this issue.
This study characterized Hg, As, Se, other inorganic elements in LSD ash from the McCracken Power Plant at The Ohio State University in Columbus, Ohio. Elemental composition and leaching measurements were conducted along with tests of lime availability, calcium carbonate equivalence, and total neutralization potential. Samples were collected over daily, weekly, and monthly time intervals to characterize different
17 scales of variability. Inorganic data were compared with measurements from a previous study at McCracken Power Plant conducted in 1991-1992 to explore long-term variability.
2.3 Experimental
2.3.1 Sample Collection
LSD ash was obtained from the McCracken Power Plant located on the main campus of The Ohio State University. The power plant uses bituminous coal as a fuel source in a single spreader-stoker boiler (boiler #8) which generates a steam capacity of
115,000 lbs per hour. A unit operation diagram of the McCracken Power Plant is shown in Figure 2.1. The McCracken Power Plant has a LSD system for removing sulfur dioxide from the flue gas. Particulates, including fly ash and calcium-rich residue from the LSD process, are collected by woven fiberglass filter bags in a pulse-jet baghouse. The baghouse contains six hoppers for collection of particulate material. A small amount of solids is collected from the economizer, combined with material from the baghouse, and stored in the ash silo.
LSD ash, coal, and lime (unslaked) samples were collected. One week of daily sampling followed by four weeks of weekly sampling were conducted during May and
June 2001, and monthly sampling was conducted from May 2001 to February 2002.
Samples were not composited, but instead each sample represents a distinct sampling event. LSD ash was collected in hoppers F and A in the fabric filter baghouse. Lime samples were taken from the falling stream of a transfer belt just prior to slaking. Coal samples were obtained from the release point of the feed belt into the combustion
18 chamber. LSD ash, coal and lime were transferred into clean (EPA procedures), cylindrical high-density polyethylene (HDPE) sample containers, approximately 75% full for storage prior to inorganic analysis. Certified clean 950-mL brown glass bottles
(cleaned by EPA procedures) were used to store samples for organic analysis. All samples subsequently were stored in an environmental room (4 to 12 °C) until the appropriate chemical measurement procedures were performed.
Both coal and lime samples required initial comminution to obtain particle sizes small enough for subsequent analyses. Samples were effectively homogenized by tumbling [58]. Subsamples for analyses were isolated by a 24-chute stainless steel riffle.
2.3.2 LSD Ash Characterization
Mineralogical analysis of LSD ash samples was accomplished using a Philips X-
Ray Diffraction (XRD) instrument (Philips Analytical, Natick, MA) with CuKα radiation at 35kV and 20mA. The XRD step-scanned measurements were carried out from 3 to
60°2θ with a fixed time of 3 seconds per 0.05o2θ. Data were analyzed by semi- quantitative data reduction software (WinJade, version 2.0). Prior to XRD analysis, samples were dried in an oven at 60°C for 24 hrs.
Scanning electron microscopic (SEM) images were taken using a Philips XL-30
ESEM. Samples for SEM analysis were prepared by ejecting approximately 50 mg of sample through a straw onto an ultra-pure aluminum SEM stub using portable ultra clean compressed air. Double-sided high purity carbon tape was applied to the SEM stub to
19 allow the powdered sample to attach on the surface. Samples were then gold-coated and kept in a desiccator until analysis.
Bulk chemical characteristics relevant to agricultural applications determined by titration (ASTM C 25-96a), included available lime index (ALI) and calcium carbonate equivalence (CCE). Total neutralization potential (TNP) was measured by ASTM method
C1318-95.
2.3.3 Inorganic Analysis
Complete elemental analyses for LSD ash and lime were accomplished by digesting approximately a 300-mg sample by microwave heating with a combination of
10-mL deionized water, 6-mL nitric acid, 2-mL hydrochloric acid, and 2-mL hydrofluoric acid. This was followed by a second microwave heating with 20-mL boric acid (EPA method 3052) and dilution with high-purity water to 100 mL. Coal fly ash, 1633b, provided by the National Institute of Standards and Technology (NIST) was digested along with LSD ash samples for method validation. Recovery percentages between 80%-
140% were obtained for every reported inorganic element. Leachate analyses were conducted by using the toxicity characteristic leaching procedure (TCLP) test (EPA method 1311). The pHs of the leachate were measured by using a pH meter, Orion Model
525A.
A Vista Pro simultaneous inductively-coupled plasma-optical-emission spectrometer (ICP-OES) system (Varian, Walnut Creek, CA) was used to determine Ag,
Al, B (only leaching tests), Ba, Ca, Cd, Cr, Cu, Fe, Mg, Mn, Mo, Na, P, Pb, S, Si, Sr, and
Zn in sample solutions (EPA method 6010B). As and Se were determined by a SpectrAA 20 880Z Zeeman graphite furnace atomic absorption (AA) spectrometer (Varian, Walnut
Creek, CA), and Hg was determined by AA with a vapor generation accessory (EPA method 7060A, EPA method 7740 and EPA method 7470A). Anions were measured by a
DX-500 ion chromatography system (Dionex, Sunnyvale, CA) (EPA method 300.0). All analyses included controls (duplicate, blank, and check standards) for every fifteen samples or less.
2.3.4 Coal Sample Analysis
Coal samples were sent to Elemental Analysis Corporation (Lexington, KY) for elemental analyses using Proton Induced X-ray Emission or “PIXE” technique. PIXE analysis provided concentrations of Ag, Al, As, B, Ca, Cr, Mg, Pb, S, Se, Si, and Sr. Hg was determined by cold vapor atomic fluorescence spectrometry (CV-AFS) following acid digestion by microwave heating. CV-AFS was performed by Dolan Chemical
Laboratory at American Electric Power (Groveport, OH).
2.4 Results and Discussion
2.4.1 Variability in Inorganic Composition of LSD Ash
XRD, SEM, and inorganic elemental analyses indicated that the LSD ash was generally a mixture of hannebachite, fly ash, unreacted lime (i.e. portlandite), and other minor constituents. In XRD patterns (Figure 2.2) from samples collected from May 17,
2001 to February 26, 2002, portlandite (Ca(OH)2) and hannebachite (CaSO3.0.5H2O) were typically the only two major minerals found. Although a small peak of calcite
(CaCO3) was observed for a sample collected on June 14, 2001, in general, X-Ray diffraction pattern results indicated little or no difference among the weekly and monthly
21 samples. These results also agreed with mineralogy results found in samples collected in
1991 [44]. From SEM images shown in Figure 2.3, there were flake-like crystals, identified as hannebachite, forming over bulk solid spherical fly ash particles. In addition, there were needle-like crystals distributed over sheet-like particles which suggests the formation of ettringite (Ca6Al2(SO4)3(OH)12.26(H2O)) on top of the hannebachite surface
[59]. Ettringite, however, was not observed by X-Ray diffraction due to the low mass fraction of ettringite compared to hannebachite and portlandite. Major element concentrations in LSD ash were consistent with XRD and SEM results, indicating that the majority of material was present as hannebachite, portlandite, and fly ash.
To examine the temporal variability in the inorganic elemental composition of
LSD ash, samples were collected over a period of 10 months. The inorganic elemental compositions of LSD ash from Hoppers F and A during 2001-2002, and from a previous study (1991-1992) are shown in Table 2.1 The relative standard deviation (RSD) for each element was used as a measurement of the variability. The RSD of an element i in Table
2.1 was calculated as
SD = i RSDi (1) AVEi where AVEi is the average concentration of element i and SDi is the standard deviation of element i. Generally, RSDs in the elemental composition of LSD ash during 2001-2002 were small (<23%) except for As, P, and Pb which had RSDs of 30%, 44%, and 203%, respectively (in Hopper F). During the period 1991-1992, RSDs for inorganic elements were generally larger (~ 20-50%) than those for the 2001-2002 interval. Considered in
22 total, however, relatively little variability in the inorganic composition was observed for this material over the 11-year period.
2.4.1.1 Sources of Temporal Variability in Inorganic Composition
Assuming plant operating conditions were constant, one hypothesis that may explain the low variability in inorganic composition of LSD ash is that the variability in lime and coal properties was low over the period of study. To verify this hypothesis, the elemental composition of feed lime and coal were examined, and the results are shown in
Table 2.1 The impact of coal and lime variability on the inorganic composition of LSD ash was assessed by comparing the RSDs in elemental composition of the feed lime and coal to the RSDs of LSD ash using propagation error analysis. In order to perform this analysis, the following assumptions were made: 1) Inorganic constituents in LSD ash originated from the feed lime and feed coal. 2) The coal to lime feed rate was consistently at 12:1. 3) The contribution of Ca on LSD ash from coal was insignificant compared to the contribution from lime. 4) Inorganic components originating from the feed lime were completely captured in the baghouse. 5) Losses of inorganic elements from feed coal (i.e., mass not accounted for in the baghouse) occurred through removal in the economizer or loss out the stack due to volatility of certain inorganic constituents. Based on assumptions
1, 4 and 5, an equation of the elemental concentrations in LSD ash as a function of the concentrations in feed lime and coal constituents can be written as,
= + ' Ci,LSD Ci,lim e f lim e Ci,coalf i,coal (2)
23 where Ci,LSD is concentration of element i in LSD ash, Ci,lime is concentration of element i in lime, and Ci,coal is concentration of element i in coal. flime is the fractional mass of feed
' lime per unit mass of LSD ash production, and f i,coal is the fractional mass of element i contributed from the feed coal per unit mass of LSD ash production in the baghouse. It
' should be noted that f i,coal is expected to vary with each element due to differences in losses of each element in the economizer and out the stack. flime, on the other hand, remains constant since the major components of lime are non-volatile and lime is added after the economizer. The values of Ci,LSD, Ci,lime and Ci,coal were determined by elemental analysis of collected LSD ash, lime and coal samples. Based on the amount of each
' element retained in the baghouse hoppers, flime and f i,coal were determined. From assumptions 3 and 4, and looking at Table 2.1, CCa,coal was assumed to be negligible.
Then, using the concentration of Ca in LSD ash and lime, flime was calculated as 0.47. To validate the previous assumption, the amount of Ca predicted to be in the LSD ash, including the contribution from the feed coal (assumption 2) was calculated as 33.14%
(0.47×69.3% + 0.47×12×0.1% = 33.14%). Thus, including the contribution of Ca from the feed coal showed only a 2% difference compared to the measured Ca in LSD ash
(32.43%) indicating a reasonable value of flime.
For elements contributed primarily from the coal, some may be lost at the
' economizer or through the stack (assumption 5). Therefore, f i,coal will be different for
' each element. By using equation 2, and assuming flime is constant at 0.47, f i,coal for each
' element was calculated. With values from flime and f i,coal, the RSDs of inorganic elements in LSD ash were estimated using
24 = ()2 2 + ()' 2 2 RSD i,LSD f lim e RSD i,lim e f i,coal RSD i,coal (3)
RSDi,lime and RSDi,coal were calculated from the inorganic results of composite samples of lime and coal. The estimated RSDs for all inorganic elements in LSD ash over the 10- month period were calculated by equation 3 and are summarized in Table 2.1.
The results show that the RSDs estimated by propagation error analysis were similar to the measured RSDs for most major elements (S, Si, Fe, Al, K and Mg), which suggests that the variability of these major elements in feed material was largely responsible for the variation in LSD ash. However, the estimated RSD of Ca (1%), which is mainly contributed from lime, was smaller than the measured RSD (6%), which indicates that other factors besides Ca variability in lime might have affected the Ca variability in LSD ash. One source of Ca variability is from changes in plant operating conditions, such as the lime feed rate. Because the magnitude of the variability potentially arising from changes in Ca feed rates is low (6%), it likely represents only a small fraction of the variability in the other inorganic elements. Results of estimated
RSDs of major inorganic elements in Hopper A showed similar trends to Hopper F.
For the trace elements, Sr and Cu, measured RSDs and estimated RSDs were similar with differences within a factor of three. For P, Mn, Ni and As, greater differences were found. For As, variability in the capture efficiency of volatile As species may provide an additional source of variability. In addition, errors from measurements of elements at near detection limits by different approaches (i.e., ICP-OES and PIXE or AA and PIXE) may also introduce differences in RSDs for these trace elements.
25 2.4.1.2 Variability of Inorganic Composition in Different Collection Hoppers
The McCracken facility utilizes two parallel flue gas channels, with each channel supplying three hoppers in sequence for the collection of solids from the bag house.
Because light and small particles may tend to fly higher and travel farther, this leads to preferential removal of small particles in the downstream hoppers, which may influence the elemental compositions of LSD ash from different hoppers. To examine this possibility, samples were collected from two different hoppers. Hopper F is the third collection hopper from one channel, while Hopper A is the first one in sequence of the other channel. From the results in Table 2.1, differences between mean concentrations of samples in Hopper F and A during year 2001-2002 were small (15%) for most elements, and within a range of a factor of two for all elements. Ca, S, Mg, Se, Cu, Sr, As, P, and
Hg were larger in Hopper F, while Al, K, Fe, Si, Mn, Be, Li, and Co were found larger in
Hopper A. Our results suggest that Ca-containing solids (e.g., hannebachite, portlandite, and ettringite) made up a greater fraction of LSD ash from Hopper F, and the major constituents of fly ash were greater in Hopper A. However, the differences between them were within 15%, which is comparable to the variability over time.
2.4.1.3 Variability in Inorganic Composition over Different Time Scales
Variability (as RSD) in the elemental composition of LSD ash also was examined over different time scales. RSDs of elements in LSD ash for samples collected over daily, weekly, monthly and yearly intervals are shown in Figure 2.4. Elements on the x-axis were arranged from the lowest value to the highest value of RSD based on the yearly data. Results suggest that RSDs over the yearly time scale were the largest for most 26 elements. There was little change in RSDs among daily, weekly, and monthly sampling periods, except the RSD of Pb on the daily period that was higher than on the yearly period. These results indicate that the lowest variability of elements is observed over time scales of a year or less, while longer time periods introduces slightly greater variability.
2.4.1.4 Inorganic Composition in Relation to Regulatory Limits
Figure 2.5 shows the concentrations of Hg, As, and Se in LSD ash samples from
January 1991 to February 2002 in comparison with regulatory limits. Hg data were only available from May 2001. The 99% confidence t-test intervals for these three elements were calculated to determine the variations of the mean concentrations to compare with limits for the land application of sewage sludge or EPA 503 Rules [60]. The EPA 503
Rules regulates an acceptable level of inorganic elements that are potentially hazardous in soil. Although LSD ash is not listed as a material regulated in EPA 503 Rule, this rule is appropriate for comparing with the level of inorganic elements in LSD ash to determine potential risk following land application. For Hg, the upper bound of the 99% confidence t-test interval was 0.47 mg/kg, which is greatly below the limit of 57 mg/kg for the EPA
503 Rules. The upper bounds of the 99% confidence t-test interval of As and Se were
39.7 and 30.6 mg/kg which were below the limit at 75 and 100 mg/kg, respectively. In addition, concentrations of As and Se in all samples collected during this 11-year period did not violate the EPA 503 Rules. Due to their significance, only Hg, As, and Se were reported in Figure 2.5. However, the 99% confidence t-test interval for other elements,
Cr, Cu, Pb, Mo, and Ni, also were also below the EPA 503 limits. It should be noted, however, that because the 99% confidence t-test values were close to regulatory limits in
27 some cases (e.g., As), the finding that concentrations were consistently below EPA 503 limits may not be representative of LSD ash obtained from other facilities. In addition, although, the EPA 503 Rule provides a mean for comparison, the land application of coal combustion by-products like LSD ash is not regulated by this rule.
2.4.2 Variability in Leaching of Inorganic Elements and Bulk Chemical Properties of LSD Ash
To examine whether significant variability occurs in the leaching properties of
LSD ash, TCLP tests were carried out on all samples. Variability in ALI, CCE, and TNP also were determined. Mean concentrations and RSDs of elements from the leachates
(TCLP test) of the samples collected in 2001-2002 and in 1991 are shown in Table 2.2.
RSDs for concentrations of elements from the leachates shown in Table 2.2 were found to be larger than those observed in the elemental composition data. However, variations in elements were similar for ash collected from Hopper F and A, within a factor of two, except for Si. Comparing the data in 1991 to the 2001-2002 data, concentrations of elements in leachates were similar. Variations were within a factor of two except for Na,
Mo, Li, Pb, Zn, Se, Al, Fe, Cu, Cr, and Cd. Changes of concentrations of these latter elements were within an order of magnitude, unless they were under the method detection limit. Most concentrations in leachate of LSD ash collected in 1991, except that for Al, were similar or lower than the results in 2001-2002. To examine temporal variability of elements in leachate from LSD ash, RSDs were examined over different time scales.
RSDs of elements in leachate over daily, weekly, monthly, and yearly periods are presented in Figure 2.6. Results indicate that RSDs were highest over yearly periods.
28 RSDs over monthly periods were slightly lower than yearly periods but higher than over daily and weekly periods. There was little difference in RSDs between daily and weekly periods. These results suggest that the variability of elements in leachates increased with increasing sampling period. These results are consistent with the data on the temporal variability of inorganic composition of this material.
Data from leaching analyses were compared with RCRA limits. Variability in As and Se concentrations are shown in Figure 2.7. 99%-confidence t-tests indicated that As and Se mean concentrations were under 2.7 and 13.4 µg/L, and, therefore, did not exceed the RCRA limits for As and Se at 5000 and 1000 µg/L, respectively. The 99%- confidence intervals of other regulated elements such as Ag, Ba, Cd, Cr, and Pb also did not exceed the RCRA limit. These data indicated that the concentration of contaminants in leachate produced over this 11-year time period were consistently lower than RCRA limits, thus supporting the characterization of LSD ash as a consistently non-hazardous material.
From Table 2.3, RSDs of ALI, CCE, and TNP in 2001-2002 were relatively small especially for CCE. For the results of samples collected over 10 months, the variations as
RSDs of ALI, CCE, and TNP in hopper F were calculated as 22%, 6%, and 16%, respectively. Compared with data in 1991, ALI decreased from 20.0 to 14.1 % as CaCO3 in 2001-2002. This smaller value of ALI indicates better efficiency of the spray dryer process to capture SO2 while minimizing the unreacted free lime in the by-product. There was only a small change in CCE from 66.2 to 65.4 % as CaCO3 in 2001-2002. The data above indicate that the LSD ash would perform in a consistent fashion as a substitute for
29 lime over a range of time scales from daily to yearly, without significant alteration of loading rates.
2.5 Conclusions
In this study, low variability in elemental composition of LSD ash was observed over an 11-year time period. The small variability observed was shown to be due largely to the low variability in the chemical properties of coal and lime. Changes in plant operating conditions may also have contributed to the variability, at least for Ca.
Although larger variability in elemental composition and leachates was observed over longer time scales, results of trace element analyses (e.g. Hg, As, Se) in LSD ash and in the leachates observed over the 11-year period did not violate regulatory limits. ALI,
CCE and TNP results also indicated long-term stability in bulk properties of this material indicating a reliable material for utilization.
Results in this study suggest that LSD ash produced from a typical lime spray dryer process can be beneficially re-used in an environmentally sound manner.
30 Lime Spray Dryer Ash Quick Lime Coal Feed Hopper F (2001- 2002) Hopper A (2001) 1991-1992 n=13 n=13 Elemental Units n=15 n=5 n=5 Compositions PROP PROP AVE RSD ERR AVE RSD ERR AVE RSD AVE RSD AVE RSD RSD RSD Ca % 32.45 6% 1% 34.3 6% 1% 31.4 12% 69.3 1% 0.1 20% S % 12.46 9% 8% 13.0 4% 8% 8.4 27% 0.038 11% 2.3 8% Si % 3.44 17% 11% 3.0 14% 11% 2.9 27% 0.45 7% 0.78 12% Fe % 2.51 17%13% 1.8 11% 13% 2.8 37% 0.202 4% 1.1 13% Al % 1.32 10%11% 1.1 14% 11% 1.9 28% 0.122 7% 0.64 12% 31 K % 0.38 16% 15% 0.37 11% 15% 0.23 29% 0.02 4% 0.04 15% Mg % 0.35 8% 10% 0.36 3% 10% 0.92 8% 0.9 8% 0.009 6% Sr mg/kg 327.7 17% 19% 334.2 9% 19% 253.2 3% 312 4% 46.3 35% P mg/kg 190.3 44% 6% 237.2 33% 6% 86.6 46% 51.8 15% 31.6 7% Mn mg/kg 154.6 13% 25% 150.2 4% 23% 48.4 24% 217 2% 6.9 75% Ni mg/kg 42.5 12% 20% 38.6 11% 20% 23.1 36% 11.5 7% 13.0 23% Cu mg/kg 38.6 16% 18% 40.7 7% 19% 19.0 24% 14.4 9% 6.0 23% As mg/kg 37.6 30% 73% 39.0 22% 74% 36.7 22% 0.4 15% 6.5 75% Se mg/kg 28.6 23% N/A 34.7 8% NA 10.9 46% UDL 1.6 13% Li mg/kg 22.1 20% N/A 20.2 18% NA 23.1 23% 1.6 3% NA Cr mg/kg 20.9 9% N/A 18.6 10% NA 16.5 38% UDL 11.6 35% Mo mg/kg 13.0 18% N/A 12.5 6% NA 7.9 29% UDL UDL Pb mg/kg 5.9 203% N/A UDL 19.4 43% UDL 21.0 57% Hg µg/kg 429 13% N/A 441 14% NA NA UDL 199 17% UDL – Under detection limit, N/A - No data available Table 2.1 Elemental Composition of Lime Spray Dryer Ash, Quick Lime, and Coal Collected from McCracken Power Plant
2001-2002 1991 RCRA Element Units Hopper F Hopper A n=14 n=5 n=1 Limit Ave RSD Ave RSD Ca mg/L 3557 7% 3552 3% 3224 S mg/L 364 19% 407 5% 206 Na mg/L 22.9 16% 23.9 11% 3.6 K mg/L 17.0 10% 19.9 6% 22.1 B mg/L 3.6 22% 3.8 7% 4.7 Sr mg/L 2.3 10% 2.6 8% N/A Mo µg/L 325 19% 383 6% 88 Li µg/L 244 11% 286 6% 90 Ba µg/L 225 39% 177 21% 348 100000 Si µg/L 138 47% 61.1 74% 140 P µg/L 83.5 48% 79.0 10% <120 Pb µg/L 53.5 17% 61.1 10% 17 5000 Mg µg/L 35.1 30% 37.0 8% 50 Zn µg/L 15.7 27% 14.3 13% <6 Se µg/L 11.3 34% 13.6 30% 4 1000 Ag µg/L 8.8 6% 8.9 6% <24 5000 Al µg/L 6.5 17% 7.2 25% 200 Ni µg/L 5.9 164% 2.5 40% <10 Fe µg/L 4.9 26% 6.3 34% <29 Cu µg/L 3.4 20% 3.7 7% <13 As µg/L 2.3 35% 1.5 39% <5 5000 Cr µg/L 1.8 36% 2.3 12% 9 5000 Be µg/L 1.0 0% 0.9 25% <2 Mn µg/L 0.2 190% 0.6 70% <1 Cd µg/L <1 <1 <3 1000 Hg µg/L <0.2 <0.2 <0.2 200 Cl- mg/L 41.3 21%40.9 6% 45.5 2- SO4 mg/L 100422% 1130 4% 460 pH* 12.6 1% 12.5 2% N/A N/A – No data available. *This pHs were measured after 18 hours of agitation
Table 2.2 Results of TCLP Test of LSD Ash from McCracken Power Plant
32
2001-2002 1991 Tests Units Hopper F n=4 n=14 Ave RSD Ave RSD
ALI % as CaCO3 14.1 22% 20.0 28% CCE % as CaCO3 65.4 6% 66.2 3% TNP % as CaCO3 17.1 16% N/A N/A – No data available.
Table 2.3 Results of ALI, CCE and TNP of LSD Ash from McCracken Power Plant
33 Coal Slaked Stack Lime Bag- House Economizer
Spray Dryer Boiler Absorber
Ash Silo
Ash To Disposal
Figure 2.1. Unit Operation Diagram of Lime Spray Drying System at McCracken Power
Plant
34 P - Portlandite, syn - Ca(OH)2
H - Hannebachite, syn - CaSO30.5H2O C - Calcite - CaCO H 3 P H P P P P H P 05/17/01 HH H H H HH H H
05/24/01
05/31/01
06/07/01
C 06/14/01
01/25/02
02/26/02
0 102030405060 θ 2
Figure 2.2. X-Ray Diffraction Patterns of LSD Material from McCracken Power Plant
35 3µm 3µm
05/17/01 05/24/01
3µm 3µm
06/07/01 06/14/01
Figure 2.3. SEM Images of LSD Material from McCracken Power Plant
36 180 160 Daily 140 Weekly 120 Monthly 100 Yearly 80 60 40 20
Relative Standard Deviation (%) Relative Standard 0 S P K Si Sr Li Ni Cr Al Fe Se Pb As Be Ca Cu Co Hg Mo Mg Mn
Figure 2.4. Relative Standard Deviations of Elements in LSD Ash Collected at Different
Time Periods
(elements were ordered from lowest to highest RSDs during the yearly period)
37 120 100 Se limit 80 As limit 60 Hg limit 40 20
1 Hg As Se
Concentration (mg/kg) EPA 503 0 01/91 - 01/92 05/01 - 06/01 01/02-02/02 Figure 2.5. Profiles of Hg, As and Se Concentrations in LSD Material Collected from
McCracken Power Plant
38 350
300 Daily Weekly 250 Monthly 200 Yearly 150
100 50
Relative Standard Deviation (%) Relative Standard 0 P S B K Si Sr Li Fe Se Cr Ni Al Pb Sn As Be Ca Zn Ba Cd Cu Na Ag Mn Mg Mo Hg Hg
Figure 2.6. Relative Standard Deviations of Elements in Leachates from LSD Ash at
Different Time Periods
(elements were ordered from lowest to highest RSDs during the yearly period)
39 16 14 g/L)
µ 12 10 8 6 As Se 4 2 Concentration ( 0 04/91 05/01 - 06/01 01/02-02/02
Figure 2.7.Profiles of As and Se Concentrations in Leachate Solution from TCLP Test of
LSD Material Collected from McCracken Power Plant
40 CHAPTER 3
DISTRIBUTION OF ARSENIC AND MERCURY IN LIME SPRAY DRYER ASH
3.1 Abstract
In this study, lime spray dryer (LSD) ash samples were collected from the
McCracken Power Plant on the Ohio State University campus and fractionated using a
140-mesh (106 µm) sieve into two fractions: a fly ash/unburned carbon-enriched fraction
(>106 µm) and a calcium-enriched fraction (<106 µm). The > 106 µm fraction was further separated by density fractionation into an unburned carbon fraction and a fly ash- enriched fraction using a lithium heteropolytungstate solution with a specific gravity of
1.84 g/mL. The results show that the concentration of As was consistently greater in the calcium-enriched fraction, while the Hg concentration was significant in all fractions.
Further analysis suggested that specific surface area was an important factor controlling the distribution of mercury in different LSD fractions. Comparing the LSD ash data to samples collected from the economizer demonstrated that arsenic was effectively captured by fly ash at 600 ºC, while mercury was not. Results from leaching tests
41 suggested that arsenic and mercury were more stable in the calcium-enriched fraction possibly due to the greater pH of the leachate.
3.2 Introduction
The lime spray dryer (LSD) system is the most common dry flue gas desulfurization (FGD) technology used in coal combustion utilities ranging from <10 to
500 MW [61]. A fine spray of slaked lime (Ca(OH)2) is injected into the scrubber, which reacts with sulfur oxides resulting in the formation of calcium sulfate or calcium sulfite.
The moisture in the reacted lime is lost due to heat from the flue gas. The resulting dry calcium sulfite/sulfate mixture, along with fly ash, is later collected as “LSD ash” by electrostatic precipitation or a fabric filter baghouse.
LSD ash can be re-used in a variety of applications. For example, LSD ash may be used in agricultural applications to replace limestone which is commonly used as a soil conditioner for controlling soil pH [44]. A previous study showed that LSD ash can also be successfully used a structural fill in transportation applications [44]. However, concerns about the release of trace elements limits further increases in re-use of LSD ash and other coal combustion by-products. For example, a previous study showed that mercury may be removed from flue gas in wet FDG processes [62]. However, the stability of mercury in these materials is still unknown. Thus, the successful reduction in mercury emissions from coal-fired power plants requires an understanding of the fate of mercury in coal combustion by-products like LSD ash.
The stability of mercury and other trace elements like arsenic in LSD ash depends on the associations of these trace elements with different LSD ash components
(e.g., fly ash, calcium solids, and unburned carbon). Previous studies suggest that volatile
42 arsenic in the flue gas adsorbs on the surface of fly ash at temperatures as high as 1000 ºC
[24]. High carbon content material such as unburned carbon has also been suggested to be an effective material for capturing volatile trace elements (e.g., arsenic and mercury) from flue gas [31,32,33,34]. In a lab scale, volatile arsenic was reported to be effectively captured by hydrated lime through the formation of calcium arsenate [35]. Mercury as
HgCl2 vapor was also reported to be captured by hydrated lime in wet FGD systems
[40,62]. The studies above suggest that mercury and arsenic may associate among a variety components comprising LSD ash.
In this study, a fractionation procedure previously developed by Maroto-Valer et al. [63] was adapted to examine the distribution of mercury and arsenic in different components of LSD ash. The fractions investigated included a fly ash-enriched fraction, a carbon-enriched fraction, and a calcium-enriched fraction. The LSD ash was collected at a fabric filter baghouse following a LSD system at the McCracken Power Plant.
Elemental composition and the leaching analyses of each fraction were conducted to determine distributions of arsenic and mercury in the solid by-product and also the mobility of arsenic and mercury in each fraction.
3.3 Experimental
LSD ash sampling. Samples were collected from the McCracken Power Plant on
The Ohio State University campus in Columbus, Ohio. The plant utilizes a single spreader stoker boiler and eastern bituminous coal as a fuel source. LSD ash was collected by woven fiberglass filter bags in a pulse jet baghouse. LSD ash samples were pulled from a collection hopper (hopper F) using a 250-mL polytetrafluoroethylene
(PTFE) beaker attached to a 2-m PTFE-coated steel rod and transferred into a 1-L HDPE
43 bottle. Multiple grab samples were collected in order to obtain approximately 750 mL of sample. Samples were then stored in an environmental chamber (4-12 oC) until fractionation and other analyses were performed. Four LSD ash samples were collected;
LSD1 (May 2001), LSD2 (June 2001), LSD3 (January 2002), and LSD4 (August 2003).
LSD ash fractionation. The fractionation of LSD ash was carried out in two steps; size fractionation and density fractionation [63]. Size fractionation was accomplished by using a 140-mesh (106 µm) sieve. After the sieving, two fractions were obtained; a calcium-enriched (<106 µm) fraction (~85% mass) and a fraction (>106 µm) enriched in unburned carbon and fly ash. This latter fraction (~15% mass) was further fractionated by density separation using a lithium heteropolytungstate (LST) solution with a specific gravity of 1.84 g/mL. 10 mL of LST solution and 0.5 g of the unburned carbon/fly ash-enriched fraction were mixed in a centrifuge tube and centrifuged at 8000 rpm for 30 minutes. After density fractionation, the settled particles were collected as the fly ash-enriched fraction, and the particles remaining in solution were collected as the unburned carbon fraction. LST solution and the unburned carbon fraction were poured on to a 0.7 µm fiber glass filter paper. Approximately 10 mL deionized water was added into the centrifuge tube which was then shaken to loosen the settled particles, representing the fly ash-enriched fraction. The fly ash-enriched fraction in deionized water was then poured on to 0.7 µm fiber glass filter paper. Both fractions were washed with 50 mL deionized water to remove the LST solution and dried in an oven at 60 ºC for 12 hours prior to analysis.
General characterization. Mineral phases of fractionated samples and LSD ash were analyzed by using a Philips X-Ray Diffraction (XRD) instrument (Philips
44 Analytical, Natick, MA) with CuKα radiation at 35kV and 20mA. Scanning electron microscopic (SEM) images were obtained by using a Philips XL-30 ESEM. The specific surface area (SSA) was measured by using the BET (Brunauer-Emmett-Teller) surface area technique using a manually controlled Micromeritics FlowSorb 2300 volumetric system (Micromeritics, Norcross, GA). A nitrogen (30% v/v) and helium mixture was used as the adsorbate gas.
Measurement of the organic carbon content of solid samples was obtained by subtracting the total inorganic carbon content (TIC) from the total carbon content (TC).
TC was analyzed by using a ThermoQuest carbon/nitrogen analyzer (ThermoQuest,
Waltham, MA) and TIC was determined by carbon coulometry (UIC Inc., Joliet, IL).
Elemental composition. The elemental composition of LSD ash and different ash fractions were determined by digesting 100 to 300 mg samples by a microwave heating method (EPA 3052) with a combination of 10 mL deionized water, 6 mL nitric acid, 2 mL hydrochloric acid, and 2 mL hydrofluoric acid. This was followed by a second microwave heating with 20 mL boric acid. Samples were diluted to 100 mL by adding deionized water. A coal fly ash standard (NIST 1633b) was digested along with the samples for method validation. The recovery of trace elements from the coal fly ash standard ranged from 80 to 110%. Better dissolution of the fly ash/unburned carbon- enriched fraction, the fly ash-enriched fraction and the unburned carbon fraction was accomplished by grinding the sample using an agate mortar prior to acid digestion.
Elemental analyses of solutions from the digestion were analyzed using a Vista
Pro simultaneous inductively coupled plasma optical emission spectrometer system
(Varian, Walnut Creek, CA). Arsenic was analyzed by a SpectrAA 880Z Zeeman
45 graphite furnace atomic absorption (AA) spectrometer (Varian, Walnut Creek, CA). A vapor generation accessory was attached to the AA for Hg analysis. All analyses included controls such as duplicates, blanks, and check standards for every fifteen samples or less.
Leaching experiments. Leaching of LSD ash was conducted by using the toxicity characteristic leaching procedure (TCLP, EPA method 1311). However, due to the small amount of sample available after the fractionation process, a modified TCLP method was used for the fractionated samples. Approximately 0.2 to 1 g of each fractionated sample was transferred into a 30 mL polypropylene bottle. Then, leaching solution (acetic acid solution prepared at pH 2.88±0.05) was added to samples with a volume to mass ratio of 20 to 1. All samples were agitated for 18 hours at a rotation speed of 30 revolutions per minute. Samples were then filtered using 0.45 µm filters and acidified to 5% HNO3 by volume for elemental analysis as described above.
3.4 Results and Discussion
3.4.1 General Characteristics of LSD Ash Fractions
The elemental composition of LSD ash and the separated fractions is shown in
Table 3.1. It was found that after separation using the 140-mesh sieve, Ca and S were enriched in the <106 µm fraction. The level of Ca in the <106 µm fraction increased up to
34% which is similar to the levels of Ca in LSD ash. The level of organic carbon in the calcium-enriched fraction was as high as 5% (see Table 3.2), which is much lower than the organic carbon in the LSD ash (up to 13.6%). Compared with LSD ash, levels of Al and Si were also low in the calcium-enriched fraction, indicating lower levels of fly ash particles in this fraction (see Table 3.2).
46 The >106 µm fraction was found to consist of both fly ash and unburned carbon.
As shown in Table 3.1, Al, Fe, Si and organic carbon content in the >106 µm fraction were significantly enriched compared to the LSD ash and calcium-enriched fraction.
Further density separation by using LST solution was conducted to separate low density
(<1.84 g/mL) unburned carbon from higher density fly ash (>1.84 g/mL) in the >106 µm fraction. Results (see Table 3.1) indicated that levels of all major elements in the unburned carbon fraction were much lower than other fractions. The organic carbon content in the unburned carbon fraction was measured as high as 92.6%. The high density fraction (>1.84 g/mL) showed a high level of fly ash constituents (i.e., Al, Fe, Si), so this fraction is called the fly ash-enriched fraction. However, the organic carbon content in the fly ash-enriched fraction was measured as high as 40.7%
In Figure 3.1, the XRD results show that LSD ash is a combination of portlandite
(Ca(OH)2) and hannebachite (CaSO3.0.5H2O). In Figure 3.2A, SEM images of LSD ash also showed needle-like crystals, which is an indication of ettringite
(Ca6Al2(SO4)3(OH)12.26(H2O)), forming over flake-like particle (i.e., portlandite and hannebachite). However, there was no ettringite or fly ash mineral forms (e.g., mullite and quartz) in the XRD pattern in LSD ash, due to the low levels of these minerals in comparison with portlandite and hannebachite. The calcium-enriched fraction showed similar XRD and SEM results. Portlandite and hannebachite were found to be the major mineral phases in the calcium-enriched fraction. This result agrees with the elemental composition data which showed that Ca and S were mostly separated into this fraction. In addition, needle-like crystals (an indication of ettringite) and flake-like particles (an indication of portlandite and hannebachite) were also found in the SEM image of the
47 calcium-enriched fraction. For the >106 µm fraction enriched with both fly ash and unburned carbon, quartz (SiO2) and mullite (Al6Si2O13) were observed by XRD patterns in addition to portlandite and hannebachite.
The fly ash-enriched fraction was found to have quartz and mullite as major mineral phases (see Figure 3.1). This result agrees with the elemental composition results which show the enrichment of Al and Si in this fraction. The SEM image of the fly ash- enriched fraction (Figure 3.2D) also shows material with a fine porous structure (~1 µm) suggestive of the carbon present in this fraction, which probably results in the high specific surface area (up to 15.4 m2/g) as shown in Table 3.2. Quartz and mullite were also detected in the unburned carbon fraction. The XRD result for the unburned carbon fraction shows a broad peak and a high level of noise indicating an amorphous phase in this fraction. This result agrees with the carbon and inorganic element analyses which indicate that the organic carbon content in the unburned carbon fraction was as high as
92.6% (see Table 3.2).
3.4.2 Distribution of Arsenic in LSD Ash
For arsenic, only size fractionation could be accomplished, due to the solubility of arsenic in the LSD solution. The distribution of arsenic in LSD ash, the calcium-enriched fraction, and the fly ash/unburned carbon-enriched fraction is shown in Figure 3.3. The arsenic concentration in LSD ash ranged from 31.9 to 44.9 mg/kg. After size fractionation, the arsenic concentration in the calcium-enriched fraction was consistently higher in all samples than LSD ash, ranging from 47.1 to 66.2 mg/kg. On the other hand, lower levels of arsenic were generally found in the fly ash/unburned carbon-enriched fraction (30.4 to 39.6 mg/kg).
48 This result indicates that arsenic is captured by slaked lime in the spray dryer absorber, thereby accounting for the enrichment of arsenic in the calcium-enriched fraction. Previous studies indicate that volatile arsenic as arsenic oxide reacts with calcium species on the surface of the fly ash forming calcium arsenate at high temperature [24]. However, due to the slow kinetics of this surface reaction, vapor-phase arsenic oxide in the flue gas largely condenses resulting in an enrichment of arsenic oxide
(As2O3 and As2O5) in the fine particle fraction [41,42]. Based on studies by Mahuli et al.
[35] and Jadhav and Fan [36], arsenic is effectively captured by hydrated lime within the optimal temperature range of 300 to 1000 ºC. Within this range of temperature, the calcination reaction of Ca(OH)2 provides CaO which can react with arsenic oxide forming calcium arsenate. However, the operating temperature within the lime spray dryer absorber at the McCracken Power Plant typically ranged from 140-190 ºC, which is lower than the optimal range for the calcination reaction. These fine arsenic oxide particles may then be captured in liquid Ca(OH)2 droplets, and subsequently captured at the baghouse.
Based on the data in Table 3.1 and Figures 3.1 and 3.2, the >106 µm fraction also contained calcium in the form of portlandite, hannebachite and ettringite. As a result, arsenic may also appear in the fly ash/unburned carbon-enriched fraction due to this contribution of calcium. However, there was no relationship observed between the level of arsenic and calcium in the >106 µm fraction. This suggests that another mechanism was responsible for the capture of arsenic in this fraction. Sorption of volatile arsenic on the surface of fly ash may contribute to the level of arsenic in the >106 µm fraction. As mentioned above, previous studies indicate that arsenic oxide can react with calcium
49 species (e.g., calcium silicates and calcium oxide) and oxygen in fly ash forming calcium arsenate at temperatures <1000 ºC [24,36]. Organic carbon in the >106 µm fraction may also be responsible for the sorption of arsenic oxide from the flue gas. A previous study indicated that vapor-phase arsenic oxide can be effectively absorbed on the carbon surface [32].
Performing a mass balance of arsenic in LSD ash, considering the mass of arsenic in the calcium-enriched fraction and the fly ash/unburned carbon-enriched fraction, results in a total arsenic content of 131 to 144 %, compared to arsenic in LSD ash. The lower level of arsenic measured in LSD ash may be due to incomplete digestion of arsenic associated with carbon particles. After the acid digestion of LSD ash, some carbon particles were observed in the solution. Lager particle size (~2 mm) of carbon fraction may a slower dissolution rate due to less surface area to contact with the acid.
Better digestion of the fly ash/unburned carbon-enriched fraction was accomplished by grinding the material before the digestion. However, it is difficult to grind the carbon mixture in the LSD ash without separation.
3.4.3 Distribution of Mercury in LSD Ash
Figure 3.4 shows that significant levels of mercury were found in all separated fractions of LSD ash. In the calcium-enriched fraction, mercury concentrations ranged from 486 to 602 µg/kg. The presence of mercury in this fraction suggests the sorption of mercury onto hydrated lime in the LSD ash. The combustion process transforms mercury into volatile elemental mercury (Hg0), which is oxidized to Hg2+ at a temperature around
450-700 ºC [64]. Mercuric chloride (HgCl2), which is commonly found to be a major oxidized mercury specie in the flue gas [65], may exist prior to the spray dryer absorber
50 and be captured in the slaked lime droplet, which is later dehydrated by the heat of the flue gas [10]. Dry calcium particle may adsorp HgCl2 on the surface. A previous study indicated that Ca(OH)2(s) can effectively capture HgCl2 by physisorption [40]. This is consistent with the presence of mercury in the calcium-enriched fraction.
To determine the distribution of mercury between the fly ash and unburned carbon fractions, further density fractionation was conducted. It was found that levels of mercury in the fly ash-enriched fraction ranged from 533 to 867 µg/kg, indicating that mercury was also greatly sorbed to fly ash. Previous studies suggested that iron oxide on the surface of fly ash interacts with Hg0 leading to the adsorption of mercury via chemisorption [64,66,67]. It should be noted that mercury may also interact with carbon, which was also found at high levels (26.2 to 40.7 %) in this fraction.
The levels of mercury in the unburned carbon fraction (< 1.84 g/mL) ranged from
262 to 594 µg/kg. In addition, the organic carbon content ranged from 76.9 to 92.6% suggesting the interaction of mercury on the carbon surface is a primary mechanism for mercury sorption onto this fraction. Elemental composition results of the unburned carbon fraction also showed low levels of calcium, iron and other elements compared with other fractions suggesting little impact of mercury sorption by other mechanisms
0 (e.g., sorption by Ca(OH)2 or iron oxide surface). Previous studies indicate that both Hg
0 and HgCl2 are captured by unburned carbon [68,69,70]. Hg was also found to be associated with sulfur impregnated in activated carbon [70]. Dunham et al. indicated that the carbon surface may also be responsible for the oxidation of Hg0 [66]. However,
0 HgCl2 was reported to be captured by carbon at a greater level than Hg [68].
51 During size fractionation, the mass loss was less than 0.2% of total mass of LSD ash. A mass balance of mercury in the LSD ash fractions indicated that the total mass of mercury in the calcium-enriched and fly ash/unburned carbon-enriched fraction ranged from 92 to 129% of total mass of mercury in LSD ash. The variation of mass may be due to errors from the measurement of mercury near the detection limit of atomic absorption.
However, for the density separation, the total mass of mercury in the fly ash-enriched and unburned carbon fractions ranged from 31 to 41% of total mass of mercury in the >106
µm fraction indicating of mercury loss during density separation. The total mass of the fly ash-enriched fraction and unburned carbon fraction ranged from 48 to 63% of the total mass of solid particles in >106 µm fraction. Therefore, the loss of mercury may be due to the dissolution of the calcium fraction in the >106 µm fraction during density separation and washing.
3.4.4 Factors Impacting the Distribution of Arsenic and Mercury
As mentioned earlier, sorption of arsenic by the calcium residue was suggested to be a major mechanism in capturing arsenic from the flue gas. The relationship between arsenic and calcium concentrations in LSD ash, the calcium-enriched fraction and the fly ash/carbon fraction were determined as shown in Figure 3.5A. A comparison of arsenic and specific surface area was also conducted (see Figure 3.5B). No statistically significant relationship was found between arsenic and calcium or surface area. This result suggests that sorption of arsenic on fly ash was also significant and the arsenic in the fly ash/unburned carbon-enriched fraction was not solely due to the calcium in that fraction.
52 For mercury, the effect of calcium concentration was also investigated. From
Figure 3.5C, mercury concentration in each fraction was plotted as a function of calcium concentration. Results indicated that there was no relationship between calcium and mercury concentrations, even for the LSD ash and the calcium-enriched fraction. The relationship between mercury and organic carbon is shown in Figure 3.5D. Here, there was also no relationship between mercury and organic carbon for LSD ash and the calcium-enriched fraction. However, an overall decrease in mercury concentration as the organic carbon content increased was observed for the fly ash/unburned carbon-enriched fraction, fly ash-enriched fraction and unburned carbon fraction (r2=0.53). Previous studies indicate that either fly ash or carbon catalyzes the oxidation of Hg0 to Hg2+ and enhances the absorption of mercury on the surface [64,66]. There was also a report for the sorption of Hg0 on the fly ash surface that oxidizes Hg0 suggesting a chemisorption of mercury on the surface of fly ash [66]. However, oxidized mercury may bind better on the surface of fly ash particles resulting in a decrease of mercury as the organic carbon content increases and fly ash constituents decrease. From Figure 3.2, the porous structure of the fly ash-enriched fraction is much finer than the unburned carbon, which results in greater specific surface area (see Table 3.2) and greater capture of mercury. Pavlish et al.explained this phenomenon as a mass transfer controlling system when removal of mercury can be improved by extending residence time, increasing amount of sorbent or reducing the particle size to increase surface area [62]. A previous study indicated that capture of mercury by fly ash was found to increase as a function of specific surface area
[71]. From the result in Figure 3.5E, an increase in mercury was found with increasing
53 specific surface area (without considering outliers). However, the correlation of the data is poor (r2=0.54), possibly due to the small number of samples.
Ash from an economizer located prior to the LSD system was also collected. The economizer ash had a high organic carbon content of 49.4%. Levels of fly ash constituents (e.g., Al, Fe, Si) were also high and comparable to the fly ash-enriched fraction (See Table 3.1). The operating temperature at the economizer was around 600 ºC which is much higher than in the LSD system (190 ºC) and the baghouse where the LSD ash was collected. From the results, the arsenic concentration in economizer ash was 33.6 mg/kg which was comparable to the range of arsenic in the fly ash/unburned carbon- enriched fraction. This result suggests that arsenic sorbs onto the fly ash surface even at high temperature. As mentioned earlier, arsenic was reported to react with calcium silicates or calcium oxide on the surface of fly ash at temperatures as high as 1000 ºC
[24,35].
Unlike arsenic, no mercury was detected in ash particles collected from the economizer. For mercury, Senior et al. reported that under equilibrium conditions, Hg0 starts to oxidize as the temperature drops below 677 ºC. Therefore, at the economizer, there might be available oxidized mercury which can be adsorbed on the surface of fly ash more effectively than Hg0 [64]. Our result suggests, however, that high volatility of mercury at the high temperature (600 ºC) of the economizer diminished the sorption of mercury onto fly ash or carbon particles. This result agrees with a previous study which found that all mercury is in the vapor phase at a temperature above 500 ºC [72].
Therefore, capture of mercury in all LSD ash fractions except the lime-enriched fraction likely occurred after the economizer where the temperature drops to less than 500 ºC.
54 3.4.5 Leaching Characteristics of LSD Ash Fractions
The leaching of arsenic from different LSD ash fractions is shown in Figure 3.6.
Results indicated that leaching of arsenic from all fractions was low. Less than 0.2% of arsenic present in the solid phase was found in the leachate of LSD ash and the calcium- enriched fraction, and less than 5% of arsenic leached from the fly ash/unburned carbon- enriched fraction. The fly ash/ unburned carbon fraction had the highest leachate concentration ranging from 22.6-76.3 µg/L. The parent LSD ash and calcium-enriched fractions had similar leachate concentrations (2.0-4.0 µg/L). Although there was more arsenic in the calcium-enriched fraction (47.1-66.2 mg/kg), less arsenic leached out from this fraction compared to the fly ash/unburned carbon-enriched fraction which had 30.4-
39.6 mg/kg arsenic concentration.
In the calcium-enriched fraction, arsenic may be present as fine particulate arsenic oxide (As2O3 and As2O5) associated with the hydrated lime which is highly soluble.
However, previous studies indicated that soluble arsenic can also form low solubility calcium arsenate (Ca3(AsO4)2 or CaHAsO3) [73,74,75], which may explain the low levels of arsenic leached from this fraction. For the fly ash/unburned carbon-enriched fraction, arsenic was likely captured on the surface of calcium oxide or calcium silicate forming calcium arsenate. Therefore, the low solubility of calcium arsenate was also expected to
3- 2- provide low levels of arsenic (as AsO4 or HAsO3 ) in the leachate from this fraction.
However, higher levels of arsenic were found in the leachate of the fly ash/unburned carbon-enriched fraction. This result, however, can be explained by examining the pH of the final leached solutions. In the fly ash/unburned carbon-enriched fraction, the pH value measured at the end of the leaching tests ranged from 6.6 to 7.1, while the pH value in the
55 leachate of the LSD ash and calcium-enriched fraction ranged from 12.3 to 12.7.
However, previous studies indicated that arsenic tends to leach more from fly ash at high pH [76 ]. Lecuyer et al. indicated that arsenic leaching is affected by the calcium concentration in the sample. The presence of CaCO3 may sorb arsenic on the surface at high pH [77]. In addition, ettringite was also found to adsorb arsenic under alkaline conditions [78].
For mercury, leaching from different LSD ash fractions is shown in Figure 3.7.
Due to the small amount of sample available after the fractionation process, there was no fly ash-enriched fraction available for the leaching tests. From the results, all mercury measurements in leachates of LSD ash and the calcium-enriched fraction were less than method detection limit (0.2 µg/L) indicating that less than 1% of mercury was released from the LSD ash. Other studies also indicated that releases of mercury from FGD material during leaching analyses are generally below detection levels [62]. However, detectable levels of mercury were leached from the fly ash/unburned carbon-enriched and unburned carbon fractions. Less than 4% of mercury was released from the carbon- enriched fraction and less than 15% of mercury was released from the unburned carbon fraction. These results indicate that mercury was more stable in the parent LSD ash and calcium-enriched fraction, again possibly due to the higher pH of the leachate (12.3-
12.7). The lower pH of the leachate from the fly ash/unburned carbon-enriched (6.6 to
7.1) and unburned carbon (3.2-3.7) fractions possibly allowed more mercury release. A previous study suggested that adsorption of mercury increases as the pH increases [79].
In addition, a previous study on the absorption and desorption of mercury in solution on fly ash indicates that increasing pH in solution in the presence of calcium enhances
56 pozzolanic reactions resulting in greater stability of mercury in solid phase [ 80 ].
However, in order to confirm the impact of the pH of the leachates from different LSD ash fractions, additional leaching experiment of LSD ash fractions at a constant pH is needed.
Based on these results, both arsenic and mercury release occurs at greater levels in fractions enriched with unburned carbon and fly ash. High levels of calcium and high pH of the leached solution, on the other hand, reduces the concentrations of arsenic and mercury in the leachate. Therefore, dissolution of the calcium fraction in LSD ash upon disposal may lower the pH, which may facilitate the release of arsenic and mercury.
3.5 Conclusions
In LSD ash, calcium-enriched fraction consistently contains higher level of arsenic. Therefore, arsenic was mainly captured at the spray dryer absorber. However, less arsenic released from the calcium-enriched fraction possibly due to the formation of calcium arsenate and the adsorption of arsenic by CaCO3 and ettringite at high pH.
Mercury concentration was significant in all fractions, suggesting that multiple mechanisms share responsibilities for mercury capture in LSD ash. Lower level of mercury also released from the calcium-enriched fraction possibly due to greater adsorption of mercury at high pH. The significant level of calcium in the leachate at high pH may cause a pozzolanic reaction which stabilizes mercury adsorption.
57
Samples Al Ca Fe K Mg S Si % % % % % % % LSD1 1.3 30.9 2.7 0.39 0.33 11.4 4.7 LSD2 1.4 32.8 2.6 0.47 0.37 12.8 4.2 LSD3 1.4 33.8 2.4 0.49 0.38 11.1 4.0 LSD4 2.5 25.2 1.5 0.40 0.38 11.3 4.2 CA1 0.7 32.0 2.1 0.10 0.43 14.3 0.5 CA2 0.9 31.5 2.6 0.09 0.43 13.9 0.8 CA3 0.9 34.0 2.4 0.11 0.47 12.2 0.8 CA4 2.0 29.4 1.4 0.34 0.39 12.0 3.3 FA/UC1 2.5 16.6 7.5 0.17 0.30 7.4 3.7 FA/UC2 2.8 11.7 6.3 0.22 0.27 3.6 3.8 FA/UC3 2.9 17.0 6.0 0.18 0.30 7.2 3.9 FA/UC4 4.6 10.8 3.1 0.58 0.27 4.3 8.1 FA1 3.5 15.9 11.1 0.21 0.15 8.9 4.9 FA2 4.5 13.3 6.2 0.24 0.14 6.8 5.5 FA3 4.7 8.7 9.6 0.35 0.21 3.5 5.7 FA4 6.3 4.8 3.2 0.66 0.16 2.6 9.6 UC1 1.7 0.9 1.7 0.09 0.05 0.7 1.8 UC2 1.3 0.2 1.1 0.06 0.03 0.4 0.9 UC3 0.9 0.2 1.0 0.07 0.03 0.5 0.2 UC4 1.9 0.9 0.8 0.20 0.04 0.7 2.8 ECON 5.3 0.1 3.1 0.6 0.3 0.8 9.6 LSD – lime spray dryer ash, CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction, FA – fly ash-enriched fraction, UC – unburned carbon fraction, ECON – economizer ash
Table 3.1. Elemental Compositions of Lime Spray Dryer Ash Fractions and Economizer Ash
58
LSD1 LSD2 LSD3 LSD4 Samples Surface Org. Surface Org. Surface Org. Surface Org. Area Carbon Area Carbon Area Carbon Area Carbon (m2/g) (%) (m2/g) (%) (m2/g) (%) (m2/g) (%) LSD 7.5 7.1 7.1 7.0 7.0 13.6 6.5 13.4 CA 5.1 2.1 4.3 2.0 5.8 2.6 7.3 5.0 FA/UC 7.3 40.6 6.5 48.7 7.0 55.8 7.6 47.1 FA N/A N/A 15.0 26.2 15.4 30.4 11.9 40.7 UC 7.7 86.1 7.5 76.9 9.4 81.4 7.5 92.6
N/A – no data available LSD – lime spray dryer ash, CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction, FA – fly ash-enriched fraction, UC – unburned carbon fraction
Table 3.2. Specific Surface Area and Organic Carbon Content of LSD Ash Fractions
59 P - Portlandite, syn - Ca(OH)2
H - Hannebachite, syn - CaSO30.5H2O
M - Mullite - Al6Si2O13
Q- Quartz - SiO2
P,H P P HP P H H H H H H P LSD Ash
P P,H CA P Q,M PP H H H H H H H P
Q,M FA/UC H M Q M M M MMM MQ M M Q,M FA M Q MM M M M M MQ M M UC
0 10203040506070 θ 2 Figure3.1 X-Ray Diffraction Patterns of LSD Ash Fractions of Sample 4.
CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction, FA – fly ash-enriched fraction, UC – unburned carbon fraction
60 A B
4µm 4µm
C D
80µm 4µm
E
30µm
Figure 3.2 Scanning Electron Microscopic Images of LSD Ash Fractions of Sample 4.
A - LSD ash, B – lime-enriched fraction, C – fly ash/unburned carbon-enriched fraction, D – fly ash-enriched fraction, E – unburned carbon fraction
61 Arsenic LSD ash Ca 70.0 FA/UC 60.0 50.0 40.0 30.0 20.0 10.0
Arsenic concentration (mg/kg) 0.0 LSD1 LSD2 LSD3 LSD4
Figure 3.3 Distribution of Arsenic in LSD Ash Fractions.
CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction.
62 Mecury LSD ash Ca 1000 FA/UC 900 FA g/kg) g/kg) UC µ 800 700 600 500 400 300 200 100
Mercury concentration ( concentration Mercury 0 LSD1 LSD2 LSD3 LSD4
Figure 3.4 Distribution of Mercury in LSD Ash Fractions.
CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction, FA – fly ash-enriched fraction, UC – unburned carbon fraction
63
70 70 A B 60 60
50 50
40 40 Arsenic (mg/kg) Arsenic (mg/kg) 30 30
20 20 0 10203040 0246810 2 Calcium (%) SSA (m /g)
900 900 C D
800 800 700 700 g/kg) g/kg) µ µ 600 600 500 500 400 400 Mercury ( Mercury Mercury ( Mercury 300 300 200 200 0 10203040 0 20406080100 Calcium (%) Org. C (%) 900 E 800 LSD ash 700 CA g/kg)
µ 600 FA/UC 500 FA 400 UC
Mercury ( Outlier 300 200 0 2 4 6 8 10 12 14 16 18 2 SSA (m /g) Figure 3.5 Arsenic and Mercury Concentrations in LSD Ash Fractions as Functions of Calcium, Specific Surface Area and Organic Carbon. A – arsenic vs calcium, B – arsenic vs specific surface area, C – mercury vs calcium, D – mercury vs organic carbon, E – mercury vs specific surface area. CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction, FA – fly ash-enriched fraction, UC – unburned carbon fraction. 64 Arsenic 90 LSD ash 80 Ca
g/L) g/L) 70 FA/UC µ 60 50 40 30 20 10 Arsenic concentration ( concentration Arsenic 0 LSD1 LSD2 LSD3 LSD4
Figure 3.6 Arsenic Concentrations in Leachates of LSD Ash Fractions.
CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction.
65 Mercury 5.0 LSD ash Ca
g/L) g/L) FA/UC
µ 4.0 UC 3.0
2.0
1.0 MDL 0.0 Mercury concentration ( concentration Mercury LSD1 LSD2 LSD3 LSD4
Figure 3.7 Mercury Concentrations in Leachates of LSD Ash Fractions.
CA – calcium-enriched fraction, FA/UC – fly ash/unburned carbon-enriched fraction, UC – unburned carbon fraction
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CHAPTER 4
CHARACTERIZATION OF BY-PRODUCTS GENERATED FROM THE OHIO
STATE CARBONATION AND ASH REACTIVATION (OSCAR) PROCESS
4.1 Abstract
Two different sorbents (i.e. regenerated sorbent and supersorbent) were tested at the Ohio State Carbonation and Ash Reactivation (OSCAR) process, a pilot-plant of a new dry flue gas desulfurization system. Trace elements, particularly As, Se and Hg, were found to be more effectively captured in the solid by-products compared to traditional lime spray dryer (LSD) ash. Experiments using either regenerated sorbent or supersorbent indicated that increasing Ca improved the capture of Se at the cyclone but had no impact on the capture of trace elements at the baghouse. The concentrations of most trace elements were below the limits regulated in the EPA 503 Rule except As and
Se. However, leaching tests indicated that all OSCAR cyclone samples were not hazardous. Results suggest that OSCAR cyclone samples can be utilized in construction, reclamation, and agricultural applications.
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4.2 Introduction
Calcium-based flue gas desulfurization (FGD) processes are characterized as either “wet” or “dry” depending on the moisture content of the resulting residual material.
Dry processes typically have lower capital costs of construction, but because of the low efficiency of sulfur capture in these systems, the consumption of lime is higher than in wet systems [9]. Although wet systems operate at close to 100% capture efficiency, the capital costs associated with these systems are higher than for dry systems [9].
Recently, a new dry FGD process was developed, called the Ohio State
Carbonation and Ash Reactivation (OSCAR) Process, which improves the sulfur capture efficiency of dry FGD calcium-based sorbents [50]. The OSCAR sorbent used in this process is generated by reacting a calcium-containing material (e.g., lime or lime spray dryer ash) with CO2 in the presence of surfactants [51]. The resulting sorbent is highly porous, thereby increasing sulfur capture efficiency. Laboratory testing indicates that this synthesized sorbent improves sulfation by 20% compared to hydrated lime sorbents [81].
In this paper, by-products produced from a demonstration pilot-plant of the
OSCAR process built at the McCracken Power Plant on the Ohio State University campus, were investigated. The effectiveness of the OSCAR process for removing SO2 and volatile inorganic pollutants (e.g. Hg, As, Se) is presented elsewhere [82]. The pilot- scale OSCAR process utilized a 1-MW equivalent slipstream from the McCracken Power
Plant. Approximately 6% of the flue gas from a coal-fired boiler at the power plant was utilized. A simplified process flow diagram of the OSCAR process is shown in Figure 4.1.
OSCAR sorbent was injected into a riser reactor allowing the sorbent to react with constituents of the flue gas at high temperature before entering a cyclone, the first particle
68
collection system. After traveling through a heat exchanger, particulates escaping from the cyclone were collected by a baghouse filter [50].
Two different sorbents were tested with this pilot-plant: “supersorbent (SS)” and
“regenerated sorbent (RS)”. SS was made by mixing fresh lime, water and CO2 obtained from the McCracken power plant flue gas (~15% CO2) in a slurry bubble carbonator [50].
RS was made by the same process as SS, except lime spray dryer (LSD) ash collected from the McCracken Power Plant was used as a source of calcium instead of fresh lime.
Following carbonation, the sorbent was sent to a powder dryer and then injected into the
OSCAR riser reactor.
In this study, we examined the concentrations of trace elements and PAHs, bulk chemical properties (e.g., available lime index), and the physical and engineering properties of OSCAR by-products. Previous laboratory work has demonstrated that the
OSCAR sorbent provides greater capture of As and Se, compared to a typical dry FGD system like a lime spray dryer (LSD), due to the higher surface area and pore volume
[50]. However, greater capture of these trace elements poses greater potential for release upon utilization or landfill disposal. Therefore, examination of trace elements in OSCAR by-product is necessary to understand the potential environmental impacts of using or disposing of this material. The effects of operating conditions (e.g., sorbent feed rates, flue gas flow rates, and boiler temperatures) on the levels of trace elements in OSCAR by-product were also investigated. Trace element concentrations in OSCAR by-product and results from leaching tests were compared with relevant environmental regulations to determine appropriate disposal and /or re-use options.
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4.3 Materials and Methods
4.3.1 Sample Collection
All experiments were conducted using flue gas generated from the combustion of eastern bituminous coal in a single spreader stoker boiler (boiler #8) at the McCracken
Power Plant. Details of the McCracken Power Plant and LSD system at this facility are found elsewhere [82, 83 ]. Two sets of experiments are presented here, one set of experiments was conducted with RS (experiments 1-5 and 8) and the other with SS
(experiments 32-35 and 37-39). Results from additional experiments are not presented here due to mechanical problems which occurred during operation. Lime and LSD ash
(from the McCracken ash silo), used as raw material for sorbent preparation, were also collected to determine background concentrations of elements in the sorbent raw materials. Samples for elemental composition analysis were collected in 1-L HDPE bottles. The collected samples were stored in an environmental chamber (4 to 12 °C) until analysis.
4.3.2 General Characterization
For mineralogical analysis, samples were dried by heating in an oven at 60°C for
24 hrs. X-Ray Diffraction (XRD) data were obtained by using a Philips XRD instrument
(Philips Analytical, Natick, MA) with CuKα radiation at 35kV and 20mA. A Philips XL-
30 environmental scanning electron microscopic (SEM) was used for taking SEM images. BET specific surface area (SSA) analysis was conducted using a controlled
Micromeritics FlowSorb 2300 volumetric system (Micromeritics, Norcross, GA) with a nitrogen (30% v/v) and helium (70% v/v) mixture.
70
4.3.3 Elemental Composition Analysis
Detailed methods for elemental composition analyses are described elsewhere
[83]. Briefly LSD ash, lime, and OSCAR sorbent and by-product were digested by using a microwave-assisted acid digestion method (SW-846 method 3052). Concentrations of the elements Ag, Al, As, Ba, Ca, Cd, Cr, Cu, Fe, Mg, Mn, Mo, Na, P, Pb, S, Si, Sr, and
Zn in sample solutions were measured by a Vista Pro simultaneous inductively coupled plasma optical emission spectrometer (ICP-OES) system (Varian, Walnut Creek, CA)
(SW-846 method 6010B). Low levels of As, Se, and Hg were examined by a SpectrAA
880Z Zeeman graphite furnace atomic absorption spectrometer (AAS) (Varian, Walnut
Creek, CA) (SW-846 method 7060A, 7740, and 7470A). Controls (duplicate, blank, and check standards) were included during these measurements for every fifteen samples or less. The toxicity characteristic leaching procedure (TCLP) test (EPA method 1311) was used to determine leachability of OSCAR by-product and LSD ash.
4.3.4 Bulk Chemical Properties
Characterizations of bulk chemical properties relevant to agricultural applications were determined including available lime index (ALI) (ASTM C 25-99), calcium carbonate equivalence (CCE) (ASTM C 25-99) and total neutralization potential (TNP)
(ASTM C1318-95).
4.4 Results and Discussion
4.4.1 Bulk Mineral and Surface Characterization of OSCAR Sorbents and By-
Products.
As shown in Figure 4.2, XRD analyses indicated that calcite (CaCO3) was the major mineral phase in RS, resulting from the conversion of available lime during sorbent
71
activation. Hannebachite (CaSO3·0.5H2O) and quartz (SiO2) were also found in RS as mineral contributions from the LSD ash used as raw material. Calcite was the only major mineral phase found in SS, indicating near complete conversion of slaked lime to calcite.
In by-product samples collected from the cyclone and baghouse during experiments with either RS or SS, calcite, gypsum (Ca(SO4).2H2O), and lime (CaO) were the primary mineral phases detected. The presence of gypsum in the OSCAR by-product indicates the capture of SO2, while the presence of lime results from the decomposition of calcite in the riser reactor. SEM images of by-product collected from the cyclone consisted mainly of flake-like particles, suggestive of Ca particles (e.g., calcite, gypsum), mixed with fly ash and unburned carbon (not shown). For the samples collected from the baghouse, smaller particles were observed (generally less than 3 µm). Generally, samples collected from the baghouse had larger specific surface area than samples collected from the cyclone (see Table 4.1). Portlandite (Ca(OH)2) and hannebachite were the major mineral phases observed in LSD ash.
4.4.2 Inorganic Element Concentrations of OSCAR Sorbents
Inorganic element concentrations of lime, LSD ash, SS and RS were measured for later comparison to levels determined on the solid by-products. From the results in Table
4.2, Ca, S, Si, Al, Fe, and organic carbon were found to be the major elements in RS. Fly ash constituents (Al, Si and Fe) and organic carbon in RS originated from the LSD ash.
Compared to the raw material (LSD ash), concentrations of organic carbon, Al, and Si were lower in RS, while Ca concentrations were higher. The dilution of these elements in the RS may have occurred due to settling of the high specific gravity and low solubility
72
fly ash particles during sorbent preparation. For most trace elements, the levels in LSD ash and RS were similar within a factor of five.
Ca, S, Si, Al, Fe, and organic carbon were also the major constituents in SS.
Lower levels of fly ash constituents, Al, Fe, S, and Si, were detected compared to RS.
Elemental analysis of raw material (lime) showed much lower concentrations of Al, Fe, S,
Si and organic carbon than those observed in SS. In addition, the flue gas used as a source of CO2 during sorbent activation was free of particles. Thus, fly ash constituents and organic carbon were likely introduced into the SS by LSD or RS residual in the pilot- scale slurry bubble carbonator. Trace element levels in SS were similar or less than RS except Hg, which was found to be higher in SS. The higher Hg concentration in the SS may be due to Hg capture from the McCracken flue gas during the sorbent carbonation phase. Volatile Hg as HgCl2 is effectively captured by lime [34].
4.4.3 Inorganic Element Concentrations of OSCAR By-Products
4.4.3.1 Comparison of Trace Elements on OSCAR Samples and LSD Ash
Average elemental compositions of by-product samples collected from the
OSCAR cyclone and baghouse using RS and SS are shown in Table 4.2. These data are compared to levels in LSD ash determined in a previous study which demonstrated low variability in chemical characteristic of LSD ash collected over an 11-year period from the McCracken Power Plant [83]. The average concentrations of most major elements, except Ca and S, in samples collected from the cyclone were higher than in the LSD ash.
In by-product samples collected from experiments using RS, average concentrations of
Al, Fe, K, Mg and Si in cyclone samples were at least a factor of two greater compared to those in LSD ash. The amount of fly ash constituents in LSD ash were lower than found
73
in OSCAR cyclone samples possibly due to the partial removal of fly ash prior to the spray dryer and the greater amount of Ca sorbent used in the LSD system. For the samples collected from the OSCAR baghouse, average concentrations of most major elements except S were similar to samples collected from the cyclone. The Ca/S ratio in by-product from the OSCAR baghouse was lower than in LSD ash, due to more effective capture of S by the sorbent in the OSCAR system [82].
Samples collected from the cyclone were also found to have higher concentrations of most trace elements when compared to the LSD ash, except for Ag, Be, Hg, Mn and
Se. In experiments with SS, Hg concentrations in cyclone samples were much lower than in the sorbent, possibly due to the release of Hg in the riser reactor at high temperature
(~600 °C). Heebink and Hassett reported that Hg starts to decompose and release as elemental mercury at a temperature of around 400°C [84]. Concentrations of most trace elements in baghouse samples were significantly higher by as much as an order of magnitude than those in the LSD ash. For example, the average concentration of As in by-product samples collected from the baghouse was 741±106 mg/kg in experiments with SS, while the As concentration in LSD ash was only 35±7 mg/kg. Hg concentration in the baghouse samples was 2.9±0.6 mg/kg in experiments with SS, while the Hg concentration in LSD ash was 0.4±0.06 mg/kg. For Se, the concentration in baghouse samples was 142±26 mg/kg in experiments with SS, while the Se concentration in LSD ash was 30±3 mg/kg.
Previous studies indicated that volatile As, Se, and Hg species can be effectively captured by lime (CaO) [34,36,37]. In the OSCAR process, the desulfurization relies on a calcination reaction which provides CaO to react with SO2 in the flue gas [85], which can
74
also react with trace elements. Previous studies also demonstrate that capture of trace elements increases with decreasing temperature, and increasing surface area and organic carbon content of the by-product [66,86]. However, organic carbon in baghouse samples were less than in cyclone samples for experiments with both SS and RS. Thus, the greater capture of trace elements in the baghouse is more likely due to the lower operating temperature (~300 °C) compared to the cyclone (~600 °C) and high specific surface area of particles in the baghouse. In addition, the baghouse is located at the far-end of the
OSCAR process, thus increasing the solids residence time compared to capture at the cyclone. The longer residence time of particles in the flue gas may increase the interaction between volatile trace elements and the sorbent resulting in more sorption of trace elements in the baghouse. Accumulation of particles on the fabric filter between cleanings also increases solid residence time.
4.4.3.2 Effects of Operating Parameters on Elemental Composition on Solid By-
Products
Operational parameters (e.g., sorbent injection rate and flue gas flow rate) were varied to optimize the desulfurization process. These operational parameters may also result in changes in the levels of inorganic elements in by-products. The range of operational parameters used in OSCAR experiments are shown in Figure 4.3. Previous studies indicate the levels of volatile trace elements on solid by-products are significantly influenced by flue gas temperature [86]. However in this study, the temperature of the flue boiler remained relatively constant, varying from 629 to 681 °C and the temperature of the baghouse was approximately 300 °C. Therefore, changes in the composition of
75
OSCAR by-product presented here for a given unit process were not due to changes in temperature.
Under a constant sorbent injection rate, the higher flow rate of flue gas results in more flue gas constituents contacting the injected sorbent, and thus an increase in the solid-phase concentration of volatile trace elements was expected in cyclone samples.
From the results, higher flue gas flow rates did indeed result in higher levels of As in solid by-products from the cyclone. For example, in experiments 34, 37 and 38 the sorbent injection rate was constant at 16 kg/hr. However, As concentration was the highest in experiment 37 and the lowest in experiment 38. Experiment 37 had the highest flow rate at 5479 m3/hr while experiment 38 had the lowest (5055 m3/hr).
Se concentrations, on the other hand, significantly decreased when the flue gas flow rate increased. Previous work indicates that volatile arsenic reacts with available calcium in the flue gas forming a solid-phase calcium arsenate deposit on the surface on fly ash at a temperature range of 600-1000 °C [24]. The majority of Se (97%), on the other hand, was reported to remain in the vapor phase even at the temperature as low as
327 °C [38] which is much lower than the temperature at the riser reactor. Therefore, due to the high volatility of Se, the reaction of Se with the sorbent in the riser reactor may be low. In addition, the reaction of Se with CaO was also reported to be significantly reduced in the presence of SO2 [51]. Therefore, at the higher flue gas flow rate, the removal of Se which has a slow reaction rate with the sorbent [51] may decrease due to decreased the residence time. Also, Se concentrations were only 0.9 and 2.4 mg/kg in samples collected from experiments without sorbent injection. Therefore, an increase in flue gas flow rate increases the fraction of fly ash in the by-product resulting in a dilution
76
of Se concentration. It should be noted that other factors such as the short run times or unsteady sorbent mass flow rates may also have influenced the results. For example, experiment 35 had the second highest Se and lowest As concentrations despite the high flue gas flow rate.
In experiments (4 and 5) the sorbent injection rate was varied while the flue gas flow rate was held constant. For this set of experiments, we expected that the concentration of trace elements would decrease with increasing sorbent injection rate, due to dilution of fly ash content and lower capture efficiency per mass of sorbent. Results showed that as the sorbent injection rate increased, As concentration decreased as expected by this hypothesis. Compared with experiment 5 (8 kg/hr), As concentration was lower in experiment 4 which had a significantly higher sorbent injection rate (44 kg/hr).
For Se, the opposite result as shown for As was observed. Se concentration was lower in experiment 5 which had the lowest sorbent injection rate. As mentioned above,
Se is highly volatile and should remain mostly in the vapor phase prior entering the riser reactor. As mentioned above, the reaction of Se with sorbent in the riser reactor may be very low due high volatility of Se. Instead of diluting Se concentration in sample, additional sorbent may remove more volatile Se due to reduced kinetic limitations at the higher sorbent concentrations in the flue gas. Additional sorbent also lowers the low Se- content fly ash fraction resulting in an increase of Se concentration in the by-product.
4.4.3.3 Effects of Major Elements on Trace Elements on OSCAR Solid By-Products.
As explained earlier, Ca sorbent was shown at the lab-scale to effectively sorb trace elements from the flue gas [24,37]. Therefore, it is likely that trace elements in
77
OSCAR samples may associate with the Ca fraction in the samples. Figure 4.4a shows the As and Se concentrations in comparison with Ca in cyclone samples collected from both RS and SS experiments and fly ash (no sorbent injected). Results indicate that As concentrations remained constant at around 90 mg/kg in fly ash and RS samples and decreased as the Ca concentrations increased in SS samples. Because the level of As in fly ash was also approximately 90 mg/kg, this suggests that the Ca fraction in the RS by- product was enriched with As. The level of As decreased with increasing Ca in SS by- product, indicating dilution of As concentration upon the injection of the SS. Lower level of As in the Ca fraction in the SS by-product may be due to the lower specific surface area of SS, compared with RS.
A positive linear relationship between Se and Ca concentrations in cyclone samples was observed in Figure 4.4a. This result indicates that Se captured in the cyclone was associated with the Ca fraction of OSCAR by-product, more so than the fly ash fraction and agreed with a previous study which showed that that sorption of Se by Ca sorbent involves the chemical reaction of CaO and SeO2 which occurs at a temperature around 600 °C [37].
For the OSCAR baghouse samples, the relationships between levels of trace elements (As, Se and Hg) and Ca concentrations are shown in Figure 4.4b. Unlike cyclone samples, samples collected from the baghouse had negative slopes with R2 of
0.75, 0.81 and 0.52 for As, Se and Hg, respectively. The low temperature in the baghouse resulted in more sorption of trace elements on the surface [87]. The longer residence time also increased the contact time between the volatile trace elements and the sorbent particles resulting in greater removal of trace elements in the baghouse. Therefore,
78
additional Ca sorbent increased Ca concentrations, but diluted the fly ash constituents including trace elements in the baghouse.
4.4.3.4 Trace Elements in Comparison to Regulatory Limits
In Table 4.4, leaching tests of OSCAR by-products collected from the cyclone are presented. Leaching tests were not conducted for baghouse samples due to inadequate amounts of sample. The results show that concentrations of all trace elements in leachate were low. In fact, no concentration surpassed RCRA limits suggesting that the OSCAR by-products collected from the cyclone were not hazardous. Compared to LSD ash, most of the trace elements in the leachates of the OSCAR cyclone samples were either similar or lower. However, concentrations of As and Se in leachates of the OSCAR cyclone samples were higher by as much as an order of magnitude possibly due to greater capture of As and Se by the OSCAR sorbent resulting in more As and Se concentrations in the leachates. For example, concentration of As in the leachate of sample 5 was 0.556 mg/L, while it was only 0.004 mg/L in the leachate of LSD ash. Although, TCLP test simulates leaching environment in a municipal solid waste landfill, this method is still appropriate for evaluating the toxicity for OSCAR by-products due to a simplicity and common use of this method.
To examine whether the OSCAR by-product can be safely utilized, trace elements in the OSCAR by-products were compared to the land application limits for sewage sludge (EPA 503 Rule). As shown in Table 4.2, concentrations of As in all cyclone samples from RS experiments (i.e., 2, 4, 5, 8) were above the limit of 75 mg/kg.
However, there was no violation of As concentrations in cyclone samples with SS. Other trace elements in cyclone samples were also below the limits for both RS and SS
79
experiments. For the baghouse samples, As concentrations were much higher than the
EPA 503 limits by as much as an order of magnitude. Se concentrations also surpassed the limit of 100 mg/kg in all samples except one sample from experiment 8A. However, other elements in the baghouse samples were well below the limits. Therefore, it may be necessary to mix this material with conventional materials, such as lime, to lower the trace element concentrations below the EPA 503 Rule limits in order to utilize the
OSCAR by-products for land application.
4.4.4 Reuse Applications
4.4.4.1 Bulk Chemical Properties of OSCAR By-Products
The ALI of the OSCAR cyclone samples (ranging from 0.08±0.01 % to
0.33±0.002 % as CaCO3) were low compared to the LSD ash (e.g., 14.5±0.4 % as
CaCO3) indicating that unreacted lime was not the major constituent in the OSCAR cyclone samples. However, CCE and TNP of the OSCAR cyclone samples showed a wide range of results possibly due to the variation in the amount of Ca sorbent added during the testing. The range of CCE results of the OSCAR cyclone samples were from
9.43±0.09 % to 63.0±0.3 % as CaCO3, and CCE in LSD ash was measured as 63.0±0.3 % as CaCO3. TNP results in the OSCAR cyclone samples ranged from 4.1±0.2 % to
34.9±0.6 % as CaCO3, and TNP in LSD ash was 16.3±0.4 % as CaCO3. Although the
ALI in the OSCAR cyclone samples was lower, CCE and TNP results showed that there is CaCO3 in the samples which could be used as a source of alkalinity for neutralizing acid soil. Based on bulk chemical properties, alkalinity properties may be used for neutralizing acid soil as lime substitute.
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4.5 Conclusions
Most trace elements were found at greater levels in OSCAR by-product than in
LSD ash. Levels of trace elements in OSCAR baghouse samples were much higher than in the cyclone samples possibly due to smaller size of particles and lower temperature in the baghouse. Increase of flue gas flow rate was found to improve As concentration in the baghouse, while addition of sorbent in the riser reactor improve Se concentration. Based on the leaching tests, all OSCAR cyclone samples were not hazardous. However, As concentrations in cyclone samples from RS injection experiments were higher than the
EPA 503 Rule limit. Most baghouse samples were also found to have As and Se above the EPA 503 Rule limits.
81 Specific surface area Sample ID Cyclone Baghouse (m2/g) (m2/g) 2 22.0 N/A
4 20.9 32.8
5 26.5 29.5 8 21.6 N/A 32 N/A N/A 33 10.9 22.3 34 12.5 33.1 35 N/A N/A
37 23.0 18.9
38 11.4 19.5 39 13.8 19.0 N/A – not available
Table 4.1. Specific Surface Area of Raw Material (Lime and LSD Ash) and OSCAR Sorbents and Samples
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Constituents RS Samples with RS injection SS Samples with SS injection
Units LSD ash Sorbent Sorbent EPA RM Cyclone Baghouse Cyclone Baghouse RM 503 (LSD (lime) ash) n=5 n=4 n=9 n=8 n=15 Al % 4.4 2.2 6.7 ± 1.6 6.4 ± 1.4 0.2 0.9 4 ± 1.6 6.2 ± 1.1 1.3 ± 0.1 Ca % 18 26 11±8 14±6 54 32 22±7 11±5 33±2 Fe % 3.9 1.2 6.2 ± 1.4 6 ± 3.1 0.1 1.4 3.6 ± 1.8 3.7 ± 0.9 1.3 ± 0.1 K % 0.7 0.5 0.9 ± 0.2 1 ± 0.04 0.05 0.2 0.6 ± 0.2 1 ± 0.1 0.4 ± 0.06 Mg % 0.4 0.3 0.4 ± 0.1 0.4 ± 0 0.5 0.3 0.3 ± 0 0.3 ± 0 0.3 ± 0.02 S % 7.8 8.7 3.6 ± 2.2 9.2 ± 2.6 0.04 2.1 2.1 ± 0.7 7.6 ± 0.2 12.7 ± 0.8 Si % 8.1 4.2 12.3 ± 2.4 9.4 ± 2.1 0.5 1.6 7 ± 2.7 8.3 ± 3.6 3.5 ± 0.6 Org. C % 20.2 6.1 25.4 ± 9.8 14.2 ± 4.3 0 7.7 12.4 ± 3.9 11.7 ± 2.1 5.9 ± 3.8 Ag mg/kg 8.5 UDLUDL UDL 0.05 0.4 0.4 ± 0.2 0.9 ± 0.2 N/A As mg/kg 43 57 81 ± 16 557 ± 72 UDL 5.9 62 ± 24 741 ± 106 35 ± 7 75.0 Ba mg/kg 124 1774 1342 ± 112 1270 ± 120 19 37 152 ± 55 273 ± 190 N/A Be mg/kg 0 1.6 0.8 ± 0.3 3.5 ± 3.5 0 0.2 2 ± 0.9 6.6 ± 1.2 5.7 ± 0.9 Cd mg/kg 0.1 UDL 2.3 ± 1.1 3.6 ± 1.2 UDL 0.8 1.2 ± 1 4 ± 0.4 1.2 ± 0.1 85.0 Co mg/kg 22 16 38.4 ± 6.6 59.2 ± 8.2 1.5 3.8 15.1 ± 7.2 38.5 ± 6.8 9.8 ± 1.8 Cr mg/kg 44 43 224 ± 168 379 ± 82 2.2 26 56 ± 26 327 ± 127 21 ± 1.8 3000.0 Cu mg/kg 26 62 53 ± 14 153 ± 15 5.4 20 26 ± 8 132 ± 20 39 ± 5 4300.0 Hg mg/kg 0.4 0.3UDL 3.1 ± 0.4 UDL 2.1 0.3 ± 0.2 2.9 ± 0.6 428 ± 57 57.0 Li mg/kg 46 19 74±14 106±36 2.9 11 45±15 94±11 23±4 Mn mg/kg 84 115 146 ± 24 534 ± 285 18 72 59 ± 14 230 ± 45 157 ± 19 Mo mg/kg 1.5 3 19.2 ± 14.4 48 ± 11 UDL 1.8 5.1 ± 2.6 42 ± 13N/A 75.0 Na mg/kg 1109 4810 3090 ± 1247 3771 ± 1629 373 1107 999 ± 230 2062 ± 133 13.0 ± 1.1 Ni mg/kg 29 39 137 ± 124 300 ± 69 4 19 38 ± 16 235 ± 91 42 ± 4 420.0 P mg/kg 168 134 278 ± 24 826 ± 151 26 83 222 ± 78 916 ± 143 182 ± 81 Pb mg/kg 16 26 51 ± 4 179 ± 26 16 14 29 ± 9 206 ± 74 3.3 ± 6.5 840.0 Se mg/kg 14 8 10 ± 2 109 ± 18 UDL UDL 13 ± 5 142 ± 26 30 ± 3 100.0 Sn mg/kgUDL 8 7±2 22±1 UDL 3 2±1 19±4 N/A Sr mg/kg 227 205 325 ± 34 370 ± 9 469 295 342 ± 20 353 ± 16 326 ± 58 Zn mg/kg 23 N/AN/A 45 UDL 45 57 ± 39 685 ± 191N/A 7500 RS – regenerated sorbent, SS – supersorbent, RM – Raw material, UDL – under detection limit, N/A – not available
Table 4.2. Elemental Composition of Sorbents, OSCAR By-Product and LSD Ash.
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Sample Sorbents Ag As Ba Cd Cr Pb Se Hg ID mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L 2 RS 0.033 0.148 0.076 0.002 0.006 <0.006 0.352 <0.0001 4 RS 0.037 0.251 0.109 <0.001 0.005 0.005 0.467 <0.0001 5 RS 0.034 0.556 0.205 0.002 0.006 0.005 0.23 <0.0001 8 RS 0.037 0.209 0.114 0.001 0.004 0.002 0.264 <0.0001 32 SS 0.002 0.151 0.098 <0.001 <0.001 <0.006 0.342 <0.0003 33 SS 0.002 0.106 0.069 <0.001 <0.001 <0.006 0.206 <0.0003 34 SS 0.002 0.176 0.072 <0.001 <0.001 <0.006 0.264 <0.0003 35 SS 0.002 0.12 0.062 <0.001 <0.001 <0.006 0.279 <0.0003 37 SS 0.002 0.119 0.105 <0.001 <0.001 <0.006 0.083 <0.0003 38 SS 0.002 0.118 0.074 <0.001 <0.001 <0.006 0.331 <0.0003 39 SS 0.002 0.147 0.079 <0.001 <0.001 <0.006 0.797 <0.0003 LSD ash Lime 0.031 0.004 0.081 <0.001 0.004 0.016 0.022 <0.0001 RCRA 5.0 5.0 100 1.0 5.0 5.0 1.0 0.2 RS – regenerated Sorbent, SS – supersorbent
Table 4.3. Concentrations of Trace Elements from TCLP Tests.
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Stack Baghouse Cyclone Heat Riser Reactor Exchanger
Sampling Sampling location location
CO2, Exit Flue Gas Slurry Bubble Powder Dryer LSD ash or Carbonator lime, Make- up Water Flue gas ~6% from McCracken coal boiler Sorbent Generator
Figure 4.1 Operation Diagram of OSCAR Process
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P - Portlandite, syn - Ca(OH)2 L - Lime - CaO G - Gypsum - CaSO H - Hannebachite, syn - CaSO30.5H2O 4 C - Calcite - CaCO Q- Quartz - SiO2 3
P P, P H H H P C H LSD Ash L, L, Q L G G C C G C CC C RS 4CY 4BH
SS 33CY 33BH
20 30 40 50 60 θ 2 Figure 4.2. X-Ray Diffraction Patterns of LSD Ash, OSCAR Sorbents, and OSCAR By-
Product. RS is regenerated sorbent. SS is supersorbent. 4CY and 33CY are cyclone samples. 4BH and 33BH are baghouse samples.
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25 160 80
140 20 120 60
15 100 80 40
10 60 Arsenic (mg/kg) Selenium (mg/kg)
40 20 Operating Conditions 5 20
0 0 0 245832333435373839
As, mg/kg Se, mg/kg Temperature , (x30) °C Flue Gas Flow Rate (x300) m3/hr Sorbent Inj. Rate (kg/hr)
Figure 4.3 Impact of Operating Condition to the Levels of As and Se in Cyclone Samples
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35 a As SS 100 As RS As FA 30 Se SS Se RS 80 Se FA 25 Outlier
20 60
15 Se (mg/kg) As (mg/kg) 40 10
20 2 R =0.78 5
0 0 0 5 10 15 20 25 30 35 Ca (%)
900 8 b As SS 800 As RS 7 2 Se SS 700 R =0.75 Se RS 6 600 Hg SS Hg RS 5 500 Outliers 4 400
3 (mg/kg) Hg As, Se (mg/kg) 300 2 2 200 R =0.52
2 1 100 R =0.68 0 0 4 9 14 19 24 Ca (%)
Figure 4.4. Relationships of Trace Element Concentrations and Ca Concentrations. Fig.
4.4a: As and Se in cyclone samples. Fig. 4.4b: As, Se and Hg in baghouse samples.
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CHAPTER 5
CONCLUSIONS AND FUTURE WORK
5.1. Conclusions
This dissertation conducted studies on trace elements in dry FGD by-product. In chapter 2, characterization data of LSD ash collected from the McCracken Power Plant was examined over an 11-year period to access the variation in the levels of trace elements. The impact of changes in coal and lime properties and operating conditions were examined. This study shows that LSD ash had low variability in elemental composition over the 11-year time period. The low variability in the chemical properties of coal and lime were primarily responsible for the low variability in LSD ash. Larger variability in elemental composition and leachates was observed over longer time scales, however, trace elements (e.g. Hg, As, Se) in LSD ash and in the leachates observed over the 11-year period did not violate regulatory limits. Results suggest that LSD ash can be beneficially utilized without an adverse impact to the environment.
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In chapter 3, LSD ash was fractionated to determine the distribution of trace elements. It was found that the calcium-enriched fraction consistently contained higher levels of As when compared with the parent LSD ash or the fly ash/unburned carbon- enriched fraction. However, less As was released to leaching solution from the calcium- enriched fraction, possibly due to formation of calcium arsenate or adsorption of arsenic by calcium carbonate or ettringite at high pH. Hg concentration was significant in all
LSD ash fractions. However, like As, Hg released from the calcium-enriched fraction was lower than other LSD ash fractions possibly due to greater adsorption of Hg on the surface of particle at high pH. In addition, significant level of Ca in the leachate at high pH may result in pozzolanic reaction which stabilizes Hg adsorption.
Testing of trace elements in by-products generated by the OSCAR process was conducted to determine the impact from the different sorbent materials and changes in operational conditions. Compared with LSD ash, trace elements were found at greater levels in OSCAR by-product especially in the OSCAR baghouse samples. Smaller size of particles and lower temperature in the baghouse probably caused greater capture of trace elements in by-product samples. Changes in operational condition such as increase of flue gas flow rate improved As captured in the cyclone samples. Increase of sorbent feed rate also improves Se concentration in the cyclone. All trace elements in the leachate of
OSCAR cyclone samples were below RCRA limits. Compared with EPA 503 Rule, As concentrations in cyclone samples from RS injection experiments violate land application limits. As and Se concentrations in most baghouse samples were also found to violate the
EPA 503 Rule limits.
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In summary, LSD ash is a consistent material and can be used as a reliable substitute material in a variety of civil engineering application. Significant amount of Ca in LSD ash was found to be responsible in controlling the release of As and Hg from
LSD ash. OSCAR samples were found to have significant levels of trace elements.
However, both LSD ash and OSCAR samples are not hazardous material and may be utilized in civil engineering applications.
5.2. Future work
The variability study of LSD ash indicated that coal and lime properties have a correlation with the changes in elemental composition of LSD ash. Therefore, significant changes in coal properties may result in significant changes in the elemental composition of LSD ash. However, a consistency in coal properties was observed in this study possibly due to use of coal from the same source. Significant changes in coal properties
(e.g. levels of S) may result in changes in LSD ash properties. In addition, operational conditions change the Ca concentration in LSD ash due to control of the sorbent injection rate to achieve low SO2 levels at the stack. Then, changes in S concentration in coal would result in significant changes in Ca and trace element concentrations in LSD ash.
The LSD ash fractions showed a correlation of Hg and surface area of particles.
However, there were not enough samples to establish a strong relationship. In LSD system, there is significant amount of fly ash and unburned carbon which can oxidize Hg0 to Hg2+. In addition, Cl level is also significant in bituminous coal used at the McCracken
Power Plant. Then, oxidized mercury as HgCl2 may be predominant resulting in favorable Hg capture by physisorption. Therefore, it is interesting to examine in more
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detail of the removal of Hg as a function of surface area with and without the existent of fly ash or unburned carbon.
A stoker boiler at the McCracken Power Plant did not operate at full capacity, and therefore, the flue gas was at a temperature lower than 700 °C to the OSCAR system.
Lower temperature affects the rate of calcinations of sorbent to CaO which may have a significant impact on the removal of trace elements from the flue gas. Further testing of sorbent as designed for the OSCAR process is expected to be conducted at the Shand power plant in Saskatchewan, Canada. A study of trace elements in by-product samples would be interesting to evaluate the efficiency of OSCAR sorbent and the impact from the operational conditions to the levels of trace elements at high temperature.
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APPENDIX A
LONG-TERM BEHAVIOR OF FIXATED FLUE GAS DESULFURIZATION
MATERIAL GROUT IN MINE DRAINAGE ENVIRONMENTS
Panuwat Taerakul1, Mikko Lamminen2, Yontian He3, Harold W. Walker4, Samuel J.
Traina5, and Earl Whitlatch6, Member
Journal of Environmental Engineering 130(7), 816, (2004)
1 Grad. Res. Asst., Dept. Civil and Environmental Engineering and Geodetic Science, The Ohio State University, 470 Hitchcock Hall, 2070 Neil Avenue, Columbus, OH 43210; Phone (614) 292-7340, Fax: (614) 292-3780, E-mail: [email protected]. 2 Grad. Res. Asst., Dept. Civil and Environmental Engineering and Geodetic Science, The Ohio State University, 470 Hitchcock Hall, 2070 Neil Avenue, Columbus, OH 43210; Phone (614) 292-7340, Fax: (614) 292-3780, E-mail: [email protected]. 3 Grad. Res. Asst., School of Nat. Res., The Ohio State University, 2021 Coffey Road, Columbus, OH 43210; Phone (614) 292-2265, Fax: (614) 292-7432, E-mail: [email protected]. 4 Asst. Prof., Dept. Civil and Environmental Engineering and Geodetic Science, The Ohio State University, Columbus, 470 Hitchcock Hall, 2070 Neil Avenue, OH 43210; Phone (614) 292-8263, Fax: (614) 292- 3780, E-mail: [email protected]. 5 Director, Sierra Nevada Research Institute, University of California, P.O. Box 2039, Merced, CA, 95344; Phone (209) 724-4311, Fax: (209) 724-4424, E-mail: [email protected]. 6 Assoc. Prof., Dept. Civil and Environmental Engineering and Geodetic Science, The Ohio State University, 470 Hitchcock Hall, 2070 Neil Avenue, Columbus, OH 43210; Phone (614) 292-8155, Fax: (614) 292-3780, E-mail: [email protected]. 93
A.1 Abstract
In this research, we examine the long-term (~ 4 years) behavior of fixated flue gas desulfurization (FGD) material grout following placement within the Roberts Dawson underground coal mine. Surface water and groundwater samples were collected to examine the impact of grouting on water quality, and core samples were obtained to assess the geochemical stability of the grout material. Surface water samples collected from the main seep at the Roberts Dawson mine indicated that four years after grout placement the long-term fluxes of acidity, iron, sulfur and calcium were slightly elevated compared to pre-grout conditions. The long-term discharge of these constituents was likely due to continued dissolution of grout material (for Ca and S) as well as changes in flow paths and subsequent solubilization of metal salts accumulated within the mine voids (for acidity, Fe, Al and S). Although the fluxes of these elements were elevated, no measurable deleterious impact was observed for the underlying groundwater or adjacent surface water reservoir. Groundwater samples collected from monitoring wells installed within the grout material indicated that acid mine drainage waters were neutralized by the grout material. Mineralogical analyses demonstrated minimal penetration of mine drainage water into the high strength fixated FGD material grout, and little weathering of the material was observed. These data indicate that the high strength fixated FGD material grout injected into the Roberts-Dawson mine was geochemically stable and could locally neutralize mine drainage waters. However, more complete grouting and more extensive mine flooding is likely needed in order to bring about significant improvements in seep water quality.
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A.2 Introduction
The removal of sulfur oxides from coal combustion flue gas results in the production of over 25 million metric tons of flue gas desulfurization (FGD) material every year (ACAA 2001). FGD material is the residual product of an FGD process with varying physical and chemical characteristics depending on the FGD process used. In a lime-based, wet FGD process, the resulting product consists of a wet thixotropic sludge composed primarily of calcium sulfite, calcium sulfate and water (ACAA 2003). For handling purposes, this material is often dewatered and stabilized with fly ash and lime, and the resulting material termed “fixated FGD material.”
Although a number of beneficial uses for FGD material are available (for a review, see Walker et al. 2002a), including the production of FGD-gypsum for wallboard
(Drake 1997), amendment of minespoil for abandoned mine land (AML) reclamation
(Stehouwer et al. 1995a; Stehouwer et al. 1995b), and the replacement of clay in low- permeability liners (Butalia and Wolfe 1997; Wolfe and Butalia 1998) and pads for animal feed lots (Wolfe and Cline 1995), the majority of this material (82%) is disposed of in landfills [Kalyoncu 1999]. Because coal mines are often located near coal-fired power plants and available landfill space is declining, there is interest in the disposal of
FGD material in deep mines following removal of coal. Furthermore, placement of FGD material and other coal combustion by-products (CCBs) in deep mine environments may potentially reduce the production of acid mine drainage if the conditions that lead to acid mine drainage (AMD) are reduced or eliminated (e.g., exposed pyrite and the presence of water and oxygen).
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In a 1999 Report to Congress (USEPA 1999), the United States Environmental
Protection Agency (USEPA) recommended that the disposal of fixated FGD material and other CCBs be exempt from regulation as a hazardous waste under the Resource
Conservation and Recovery Act (RCRA), Subtitle C. However, the recommendation specifically excluded the placement of CCBs in deep mine environments, indicating that regulation as a hazardous waste under Subtitle C may be warranted for minefilling applications based on potential risks. The report noted that acid mine drainage in deep mine environments may consume the acid neutralizing capacity of the CCBs and result in prolonged release of contaminants. Further, placement of CCBs in deep mines located beneath a regional water table could result in contamination of drinking water supplies.
Prior to a final determination, the USEPA recommended that more information be gathered related to the risks associated with the placement of CCBs in deep mine environments.
In a previous paper (Lamminen et al. 2001), we described the short-term (~ 1 year) impacts associated with the injection of fixated FGD material grout at the Roberts-
Dawson underground coal mine. The fixated FGD material grout consisted of fixated
FGD material with added water. Immediately following grout injection, increases in the concentration of acidity, Al, B, Ca, Co, K, Li, Fe, Mg, Mn, Ni, Pb, S, Si, Sr, and Zn were observed in surface water seeps as well as groundwater wells installed within the coal layer. Geochemical equilibrium speciation modeling suggested that the increase in concentration of a number of these parameters was due to re-routing of mine drainage water within previously inaccessible voids and the subsequent dissolution of accumulated solids, as well as the dissolution of fixated FGD material grout. Following this initial
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increase, however, the levels of these elements began to decline. While these data indicated placement of fixated FGD material grout at the Roberts Dawson mine resulted in short-term degradation of water quality, the long-term behavior of this material at the site remained unknown.
In this paper, we report on the long-term impacts related to the placement of fixated
FGD material grout at the Roberts-Dawson mine. Water quality monitoring was carried out for four years following the placement of fixated FGD material grout to assess the impact of grouting on the groundwater in the immediate vicinity of fixated FGD material grout, groundwater within the regional aquifer, and surface water surrounding the site. In addition, core samples were collected to examine the physical properties and geochemical stability of fixated FGD material grout after prolonged exposure to AMD. These water quality and mineralogical data provide important new information for assessing the risk associated with the placement of fixated FGD material grout in deep mine environments.
A.3 Site Description
This study was carried out at the Roberts-Dawson mine, a site spanning an area of
0.059 km2 (14.6 acres) located in central-eastern Ohio. The mine was closed in the
1950’s following the removal of approximately 6×104 m3 of coal. The hydrogeology of the site was extensively characterized (Bair and Hammer, 1999) and consists of a perched aquifer in the Freeport Sandstone overlying the middle Kittanning (#6) coal layer which is 1-2 m thick. The middle Kittanning #6 coal layer forms a second perched water table overlying the regional water table within the Clarion Sandstone (see Figure A1). The
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strike of the face and butt cleats of the Kittanning coal are N10-20°E and N70-80°W, receptively (ver Steeg 1942).
Figure A2 shows a map of the known mine voids in relation to the main seep and adjacent receiving stream at the Roberts-Dawson site. Fixated FGD material grout was injected into the down-dip portions of the Roberts-Dawson mine between October 1997 and January 1998 (Walker et al. 1999, Walker et al. 2002b). Two types of fixated FGD material grout were injected into the mine: a high strength grout and a low strength grout.
The high strength grout was injected into the shaded regions shown in Figure A2, while the lower strength grout was injected into the unshaded areas. Borehole cameras used at the time of injection indicated that the high strength grout effectively filled mine voids, at least within the vicinity of the injection wells. The lower strength grout, on the other hand, flowed over a more extensive area, coating exposed pyritic surfaces but not filling mine voids. The design strengths of the low strength and high strength FGD material grouts were 520 kPa (75 lbs/in2) and 1000 kPa (145 lbs/in2) after 91 days, respectively
(Damian and Mafi 1999). Laboratory testing of actual grout mixes after 90 days yielded strengths of 1180 ± 500 kPa and 1960 ± 651 kPa for the low and high strength grouts, respectively (Wolfe and Butalia, 1999); roughly twice the required design strength.
Fixated FGD material grout was also injected into limited areas of the unmapped portion of the mine. A total of 18,182 m3 of grout was injected through 317 boreholes (Damian and Mafi 1999).
The fixated FGD material grout injected at the Roberts Dawson site was a 1.25:1 mixture, on a dry mass basis, of class F fly ash and dewatered scrubber sludge with an additional 5% lime (CaO). The grout consisted primarily of calcium, silicon, iron, sulfur,
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and aluminum (Laperche and Traina 1999a, 1999b). Minor elements also present in the grout included Sb, As, Ba, Be, B, C, Cr, Cd, Cu, Pb, Mn, Ni, K, Na, Se, St, and Zn. The primary mineral phases of the unweathered grout, detected by X-ray diffraction, were hannebachite (CaSO3 • ½H2O), mullite (Al6Si2O13), quartz (SiO2), hematite (Fe2O3), magnetite (Fe3O4), glass, and ettringite (Ca6Al2(SO4)3(OH)12 • 26H2O) (Laperche and
Traina 1999a). The high and low strength grouts varied only in water content. The higher strength grout had lower water content to produce a slump of 1.6 to 2.4 cm (4-6 inches), while the lower strength grout had greater water content and a slump of 3.1 to
3.9 cm (8-10 inches) (Damian and Mafi 1999). Preliminary laboratory studies indicated that the grout could neutralize acid mine drainage from the Roberts Dawson site (Walker et al. 1999).
A.4 Materials and Methods
A.4.1 Sampling Locations and Techniques Sampling of surface water and groundwater was carried out at the Roberts-
Dawson site before and after grouting. The major seep discharging AMD from the known mine voids at the Roberts Dawson site is shown as site 5 in Figure A2. There was an additional seep (not shown on Figure A2) just south of site 5 which drained the unmapped portion of the mine. The seeps discharged into a receiving stream and flowed into a collection pond. AMD exited the collection pond and discharged through a culvert to Wills Creek Reservoir. Flow at site 5 consisted of seepage over a large area, so flow measurements were calculated as the difference in the upstream and downstream flow rates in the adjacent stream. Stream flow rates were measured using the “bucket and
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stopwatch” technique for low flow rates and weirs installed at the site for high flow rates.
Water samples were taken at site 5 by collecting the most significant flow of drainage emerging in the vicinity of the original mine opening and seep. Water samples were collected on the opposite side of the reservoir (sampling site 12) relative to the Roberts-
Dawson AMD discharge point to assess water quality impacts to Wills Creek Reservoir.
Surface water samples were collected using disposable 60 mL Luer-Lok syringes
(Becton Dickinson, Franklin Lakes, NJ) and placed in 60 mL polypropylene bottles.
Samples were filtered on-site with 0.45 µm disposable sterile cellulose acetate membrane filters (Corning, Wilmington, NC) to analyze for dissolved constituents.
Groundwater monitoring locations are also shown in Figure A2. Well 9727 was installed prior to grouting operations in the lower Clarion sandstone layer, 55 m below the ground surface and 29 m below the Kittanning #6 coal layer. Well 9719 was also installed prior to grouting operations. This well was located in a pillar of the Kittanning
#6 coal layer.
Core samples were collected at the site of wells 9901, 9904, 9906 and 2002, 2-3 years after injection of fixated FGD material grout. After collection of core samples, monitoring wells were installed. Wells 9901, 9904, and 2002 were located in the downdip portions of the mine voids near the main seep (surface water site 5). Well 9906 was located on the northwest edge of the Roberts Dawson site, near a grout injection hole that took 840 m3 of low strength grout. Only core samples collected from wells 9906 and
2002 were confirmed to contain fixated FGD material grout.
Wells 9719, 9727, 9901, 9904, and 9906 were constructed with 5.1 cm (2 in.) schedule 40 PVC pipe in conformance with ASTM standard D 5787-95. Monitoring well
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2002 was installed with 20 ft of 4 inch diameter PVC casing seated into the top of rock with the remaining hole completed as an open bedrock well.
Water levels in groundwater wells were measured with a Heron Water Level
Probe (Hamilton, OH). Prior to sample collection, wells were purged using either dedicated submersible Redi-Flow™ pumps (Ben Medows Company, Canton, GA) or by using a Reel E-Z portable well pump (Redmond, WA). Some wells were purged manually using disposable one-liter, high-density polyethylene bailers (Timco
Manufacturing, Prairie Du Sac, WI).
Filtered and unfiltered samples for metals analysis, surface water and groundwater, were acidified to 10% (volume/volume) acid concentration using ultra pure nitric acid. After collection, samples were stored either in a cooler or cold room (4 °C) until analysis.
A.4.2 Chemical Analysis of Water Samples
Surface water and groundwater samples were analyzed for pH, conductivity, sulfate, arsenic, chloride, alkalinity, metals and other inorganic constituents. pH was measured using a Model 525A pH meter (Thermo Orion, Beverly, MA), either in the field or in the laboratory. Conductivity was measured in the laboratory using a digital conductivity meter (Fisher Scientific, Suwanee, GA). Alkalinity was determined by titration or by using a Lachat Quickchem AE Autoanalyzer (Milwaukee, WI). Chloride and sulfate were determined using either the Autoanalyzer or an Ion Chromatograph
(Dionex Corporation, Sunnyvale, CA).
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The concentration of arsenic in surface water and groundwater was determined using a Perkin Elmer Graphite Furnace Atomic Absorption (GFAA) Spectrometer
(Norwalk, CT, model 4100XL) or a Varian SpectraAA 880 zeeman GFAA Spectrometer
(Walnut Creek, CA). Analyses for Al, Ba, Be, B, Cd, Cr, Ca, Cr, Co, Cu, Fe (total and dissolved), Pb, Li, Mg, Mn (total and dissolved), Mo, Ni, P, K, Si, Na, Sr, S, and Zn were carried out using an Inductively Coupled Plasma Optical Emission Spectrometer (ICP-
OES) at the Ohio Agricultural Research and Development Center (OARDC) STAR
Laboratory in Wooster, OH, or in the Environmental Engineering Laboratory at the
Columbus Campus of Ohio State University using a Varian Vista Pro ICP-OES (Walnut
Creek, CA). The error in analyses was less than 5% based on duplicate samples, and the percent difference in the anion-cation balance was generally less than 10%.
Based on pH and the metals data, acidity (mg/L as CaCO3) was calculated as;
+ + + + Acidity = 50,000×(2[Fe2 ]+ 3[Fe3 ]+ 3[Al3 ]+ 2[Mn 2 ]+1[H ]) (1)
where the concentrations of iron, aluminum, manganese, and protons are in moles/L.
Only soluble concentrations of these metals were used in the calculations. Because speciation of iron was not carried out, calculations of acidity assumed that all the iron was in the Fe3+ oxidation state.
A.4.3 Analysis of Grout Core Samples
Grout core samples were collected from the site approximately 2 and 3 years after the grouting operation. Upon collection in the field, core samples were immediately
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transported to Ohio State University for mineralogical analysis by x-ray diffraction
(XRD). XRD analysis was carried out using a x-ray diffractometer (Phillips Analytical,
Natick, MA) using Cu Kα radiation at 35kV and 20 mA. Measurements were made using a step scanning technique with a fixed time of 4 s/0.05º2 Θ , from 8 to 55 or
60º2Θ . Prior to analysis, grout core samples were air dried and ground with a synthetic sapphire mortar and pestle to ≤ 250 µm. Crystalline phase assignments were made based on comparative analyses of reference samples, searches of the ICDD (International
Center for Diffraction Data), and data in the published literature.
pH values for the borehole cores were obtained in the following fashion. For each depth increment, 5-g samples of ground and dried core material were placed into 40-mL, screw-cap centrifuge tubes along with 20-mL of HPLC grade H2O. These tubes were then capped and shaken on an oscillating shaker to facilitate equilibration of the solution phase with the solids present in each sample. After a reaction period of 2 h, each sample was centrifuged at 4000 rpm to separate the solid and solution components. The pH of the resulting centrifugate was then measured with an Orion-Ross pH electrode.
A.5 Results and Discussion
A.5.1 Long-Term Water Quality Trends
Surface water and groundwater monitoring was carried out to characterize long- term impacts on water quality. Trends in the flow rate of mine drainage and the flux of major AMD and grout constituents at the main seep (site 5) of the Roberts Dawson site are shown in Figures 3 and 4, respectively. The bar in each graph represents the grouting period. As can be seen in Figure A3, the flow rate at site 5 decreased to less than 1.7× 10-
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4 m3/s (10 liters per minute) at the end of grouting operations. However, shortly after grouting was completed, a diffuse seep emerged approximately 50 m down slope toward the receiving stream. As a result, the net contribution of flow to the receiving stream was not reduced at this site, despite sealing of the original seep. After the new seep emerged, the flow rate was seasonal, with high flow rates typically in January through April and lower flow rates during the summer months.
As seen in Figure A4, significant fluxes of acidity, sulfur, iron, aluminum, calcium and boron were observed at this new seep. Acidity, iron and sulfate are all common constituents of acid mine drainage, while the fixated FGD material grout contained calcium, sulfur and boron. For all constituents shown in Figure A4, there was a large increase in flux immediately after grout injection, which reflected both an increase in the concentrations of these parameters and elevated flow rates from January through
April. A similar trend was observed for electrical conductivity, sulfate, Co, K, Li, Mg,
Mn, Na, Ni, Pb, Sr, and Zn. Previously, it was demonstrated that the large initial increase in the concentrations of acidity, iron, sulfur and aluminum was due to re-routing of mine drainage flow, and the subsequent dissolution of accumulated iron and aluminum sulfate salts and ferrihydrite (Lamminen et al. 2001). The initial increase in calcium concentration after grouting was attributed to the dissolution of fixated FGD material grout and/or exchange of calcium from soil material due to the elevated ion concentrations in the mine drainage water.
For all the chemical parameters shown in Figure A4, except boron, the initial sharp increase in flux was followed by a slower, long-term decrease. For acidity, iron and sulfur, the long-term (after July, 1999) fluxes of these constituents stabilized at levels
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slightly higher than levels observed prior to the injection of fixated FGD material grout.
Assuming Fe3+, Al3+, H+, and Mn2+ were the dominant species present, these constituents accounted for, on average, 68%, 21%, 9% and 2% of the total flux of acidity, respectively. Little variation in flow rate was observed before and after grouting.
Therefore, the elevated fluxes were primarily a result of elevated concentrations of acidity, iron and sulfur in the mine drainage waters. It is unlikely that the fixated FGD material grout was responsible for the elevated levels of acidity and iron. More likely, continued dissolution of soluble metal salts resulted in the increased levels of acidity, iron and sulfur in the mine drainage water. For aluminum, the long-term flux was similar to flux values measured prior to FGD by-product injection, possibly reflecting lower amounts of accumulated aluminum salts within the mine voids. As of the last sampling date (September, 2001), the flux of calcium remained elevated compared to pre-grout levels, but continued to decrease. The decrease in calcium flux indicates that available calcium in the grout is also decreasing, due to the prolonged dissolution and/or changes in the strength, permeability and mineralogical properties of the material. The flux of boron increased immediately at the end of grouting but quickly returned to near pre-grout levels, perhaps due to the lack of attenuation of this element during transport.
While the concentrations of some contaminants at the main seep (site 5) were higher during the last year of the monitoring program compared to pre-grout levels, little or no deleterious impact on water quality was observed for either the surface water reservoir (site 12) or the underlying Clarion sandstone aquifer (well 9727). In Table A1, a list of parameters analyzed at the Roberts Dawson site for which either a primary or secondary maximum contaminant level (MCL) has been established are shown, along
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with the MCL value. As can be seen, all analytes were within the range of acceptable values based on the MCLs, except for manganese. In both the adjacent reservoir and the
Clarion sandstone aquifer, the concentration of manganese was significantly higher than the established MCL. However, for both the groundwater and surface water control sites, the concentrations of manganese observed in April of 2001 were comparable or lower than values detected prior to grouting operations.
A.5.2 Geochemical Stability of Fixated FGD Material Grout
To determine the geochemical stability of the fixated FGD material grout upon exposure to acid mine drainage, mineralogical analyses were carried out on well cores obtained from the Roberts-Dawson site. The core sample obtained at the site of well 9906 in 1999 was only 21 cm long and had a fluid, paste-like consistency, indicating a high moisture content (moisture content was not measured). A representative diffraction pattern for this core is shown in Figure A5. The diffraction pattern shown in Figure A5 represents a subfraction of the core at a depth of 15-18 cm from the top of the core.
It is apparent from the presence of hannebachite 2CaSO3·(H2O) and ettringite
(Ca6Al2(SO4)3(OH)12⋅26H2O), as well as pH values > 9, that this core consisted entirely of fixated FGD material grout material. The x-ray diffraction pattern obtained from the
15-18 cm depth increment (Figure A5) is strikingly similar to those reported by Laperche and Traina (1999a) for unweathered fixated FGD material grout. Laperche and Traina
(1999a) did not observe any Fe phases in fixated FGD material grout, unless the material had reacted with mine drainage waters. Thus, the presence of ferrihydrite clearly indicates the reaction of the fixated FGD material grout with Fe-containing mine fluids.
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Indeed, the ambient pH values present in the acidic mine drainage fluids should lead to the formation of the mineral schwertmanite and not ferrihydrite (Bigham et al. 1996).
Whereas, it cannot be determined if this reaction occurred during or post grout emplacement, the former seems most likely due to the low permeability one typically observes in solidified fixated FGD material grouts. Also, it should be noted that although schwertmanite was not detected by powder x-ray diffraction, its presence cannot be rule out. Also, the pH of this core was 9.6 which is at the lower limit of the stability field for ettringite (Myneni et al. 1998). Thus, a decrease in pH below this point would likely lead to significant weathering of this material.
The core collected in 2000 at the site of well 2002 was significantly longer in length and contained a more varied mineral assemblage than the core at 9906. The core from 2002 was well consolidated and was much harder than the material from the earlier sampling. A subsample within the central portion of the core was dominated by hannebachite and ettringite followed by lesser quantities of quartz and muscovite (see
Figure A6). The pH values in these samples were all > 9.00. The elevated pHs (> 9) and the presence of hannebachite and ettringite are diagnostic of fixated FGD material grout.
While muscovite is present throughout this core, it is not commonly found in fixated
FGD materials (Laperche and Traina, 1999a) nor is it likely to remain intact during coal combustion. Thus, the presence of muscovite in the grout sections of this core was either a result of micaceous materials from the mine over-burden and/or underclay or it was physically incorporated into the matrix of the grout during grout injection.
The presence of hannebachite in the core from 2002 is particularly noteworthy in that this phase is from the FGD filter-cake (Laperche and Traina, 1999a). Its long-term
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persistence in these borehole samples indicates minimal altering of the fixated FGD material grout in this particular location within the mine. Apparently, there was little if any intrusion of acidic mine waters into these samples when they were in-place in the field. This contention is also supported by dramatic change in pH observed in the subsample containing fixated FGD material grout (pH = 9.35) and the subsample immediately adjacent (pH = from 2.34). These data indicate that the fixated FGD material grout in this region of the mine maintained low permeability and has not reacted with local acidic mine drainage water to any significant extent. It should be noted, however, that the lower strength grout used to coat pyritic materials likely had greater exposure to mine drainage waters, and therefore, greater potential for weathering. As a result, a significant reduction in the potential for grout weathering may occur if all the mine voids were completely filled with grout.
A.5.3 Water Quality in Vicinity of Fixated FGD Material Grout
To better understand the interactions between acid mine drainage and fixated
FGD material grout, monitoring wells were installed following the collection of grout core samples. These monitoring wells were located directly within mine voids, and in some cases, directly within fixated FGD material grout. The core samples collected during well construction were analyzed in order to confirm the presence of fixated FGD material grout. Wells installed prior to grouting operations, on the other hand, were located within coal pillars of the remaining Kittanning #6 coal layer, in order to maintain hydraulic performance of the wells after inundation of the mine voids with grout.
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The data in Table A2 show the concentrations of important inorganic constituents in wells installed within the mine void layer, both prior to (well 9719) and after (wells
9901, 9903, 9904, 9906, and 2002) placement of fixated FGD material grout. Refer to
Figure A2 for the locations of these wells at the Roberts Dawson site. Core samples collected from sites 9906 and 2002 were confirmed to contain fixated FGD material grout. Grout was not detected in core samples collected at well sites 9901, 9903, and
9904, but these wells were confirmed to be within the mine void layer. For wells located in the downdip portion of the mine voids, the concentrations of most constituents was higher in wells installed prior to grouting (9719) compared to water quality in wells installed at least a year after grouting operations were completed. For example, the average pH values in wells 9901, 9904, and 2002 were 5.4, 6.4, and 5.9 respectively, compared to an average pH value of 4.1 for well 9719. Also, wells installed in the downdip portions of the mine after grouting all had measurable alkalinity (average alkalinity from 8.7 to 101.7 mg/L as CaCO3) while alkalinity was not detected in well
9719. Acidity, As, Al, Be, Cd, Cu, Fe, Si, Pb, and Zn were all similar or lower in concentration in downdip wells installed after grouting compared to wells installed prior to grouting. Calcium and boron, both constituents present in the fixated FGD material grout [Laperche and Traina, 1999a, 1999b], were found at higher concentrations in wells installed in the downdip portions of the mine after grouting, with the exception of well
2002.
The generally higher pH, higher alkalinity, and lower minor and trace element concentrations in wells installed after grouting in the downdip portions of the site reflects more extensive interactions of AMD drainage waters with fixated FGD material grout.
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These data indicate that AMD waters were partially neutralized. This hypothesis is consistent with the lower iron and aluminum in these wells, as these constituents would be precipitated as iron and aluminum hydroxides as a result of interaction with calcium hydroxide in the fixated FGD material grout and the increased pH values. The levels of calcium and boron in these wells also generally support this hypothesis, although the processes controlling these elements are not as straightforward as for iron and aluminum.
For wells 9901 and 9904 elevated levels of both calcium and boron were observed, indicating dissolution of fixated FGD material grout upon reaction with AMD waters.
Interestingly, calcium and boron concentrations in well 2002 were lower than in both
9901 and 9904, as well as 9719 which was installed prior to grouting operations. The lower calcium and boron in well 2002 indicates minimal penetration of AMD waters into the grout, or possibly dilution from seepage into the open bedrock borehole from the upper Freeport sandstone and fractured roof shale. For well 2002, neutralization of AMD waters likely occurred on the outer edge of the water/grout interface, removing iron and aluminum (as hydroxides), calcium and sulfur (as gypsum), prior to arriving at the monitoring well.
Well 9906 was located on the northwest section of the mapped portion of the mine. The core sample collected from this site indicated the presence of fixated FGD material grout, however, the sample had little or no strength. This well had high pH and high conductivity, with low levels of acidity, aluminum, cadmium, calcium, chromium, iron, manganese and sulfur. Arsenic levels in this well were high (above 60 ppb) for all sampling dates except one (July 2000). The high arsenic levels in this well may be to the enhanced solubilization of arsenic at high pH values, and also a reduction in the
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adsorption capacity of iron oxide solids for arsenic anions under these conditions. Water quality in well 9903, also located on the northern side of the site, was similar to 9906, except that water from 9903 had lower pH, arsenic, aluminum, and chloride, and higher calcium.
In comparing the water quality in wells installed prior to and after grouting it should be noted that the grouting process altered hydraulic flow paths within the mine voids. Therefore, samples collected in wells installed directly within the fixated FGD material grout may reflect water following a different flow path than samples collected from wells installed in the coal pillars. Despite the differences in well construction, however, the water quality and mineralogical data clearly show significant neutralization of mine drainage waters in the immediate vicinity of the fixated FGD material grout and minimal physical and chemical altering of the grout in the downdip portions of the
Roberts-Dawson mine.
A.6 Conclusions
Based on the water quality data and mineralogical analyses conducted, placement of fixated FGD material grout within the Roberts Dawson mine resulted in little or no deleterious impacts to water quality of the surrounding surface water or underlying groundwater. Although significant neutralization of AMD waters was observed in the immediate vicinity of the fixated FGD material grout, no reduction in the concentration or flux of major elements in mine seepage, including acidity, iron, calcium, and sulfur was observed, largely due to changes in water flow paths and subsequent dissolution of accumulated metal salts within the mine voids. Little penetration of AMD waters within
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the high strength grout occurred, and subsequently, little grout weathering. This latter result indicates that the fixated FGD material grout was geochemically stable over the period of study, despite the low pH of the mine drainage waters. Weathering of the lower strength grout used to coat mine surfaces was not determined but would be expected to be greater.
A.7 Acknowledgment
This project was funded in part by the Ohio Coal Development Office (OCDO),
Ohio Department of Development, under OCDO Grant No. D-95-17. The authors thank
Jackie Bird and Howard Johnson at OCDO for their role in establishing a coalition of funding for the project from federal, state and local agencies as well as private industry.
Additional support was provided by American Electric Power, Ohio Environmental
Protection Agency, Ohio Department of Natural Resources, US Department of Energy,
Dravo Lime Company, Office of Surface Mines, Corps of Engineers, US Environmental
Protection Agency, and The Ohio State University. The authors also thank Yu-Ping Chin and James Wood at Ohio State University who participated in Phase I of this project, and
John Massey-Norton, American Electric Power, for serving as project manager.
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Parameter Primary Secondary Clarion Reservoir
MCL MCL (Site 9727) (Site 12) pH (pH units) 6.5-8.5 7.28 8.06
TDS 500 275 261
Sulfate 5007 250 36 108
As (ppb) 10 5.5 1.0
Al 0.05-0.2 <0.001 <0.001
Ba 2 0.068 0.004
Be 0.004 <0.001 <0.001
Cd 0.005 <0.001 <0.001
Cl 250 3 12
Cr 0.1 <0.001 <0.001
Cu 1.38 1.0 <0.002 <0.002
Fe 0.3 <0.007 <0.007
Mn 0.05 0.981 0.120
Ni 0.1 <0.004 <0.004
Pb 0.015 <0.006 <0.006
Zn 5 <0.002 <0.002
Table A.1. Concentrations of contaminants with either a primary maximum contaminant level and/or secondary maximum contaminant level in the Clarion sandstone aquifer and adjacent reservoir. Data are for samples collected in April, 2001. All concentrations are in mg/L, unless noted.
7 Maximum contaminant level goal (MCLg) 8 Action level 113
Parameter 9719 9901 9903 9904 9906 2002
(n=5) (n=5) (n=5) (n=2) (n=6) (n=4)
Acidity (mg/L as CaCO3) 284 ± 34 124 ± 134 2.7 ± 2.5 99.2 ± 48.2 9.1 ± 4.6 21.1 ± 36.7
Alkalinity (mg/L as CaCO3) nd 102 ± 83 296 ± 35 67.8 ± 12.1 262 ± 98 8.7 ± 3.1
pH (pH units) 4.1 5.4 6.9 6.4 10.2 5.9
Conductivity (µS/cm) 1431 ± 135 1352 ± 381 1134 ± 132 2200 ± 296 1786 ± 186 330 ± 109
TDS 956 ± 91 905 ± 268 756 ± 85 1450 ± 221 1194 ± 124 217 ± 77
± ± 114 As (ppb) 5.1 ± 2.1 5.1 ± 2.9 4.6 ± 3.7 2.4 0.9 61.2 17.5 nd
Al 3.75 ± 1.11 1.1 ± 2.3 0.021 ± 0.030 0.041 ± 0.039 1.501 ± 0.864 0.216 ± 0.419
B 0.297 ± 0.030 0.414 ± 0.089 0.457 ± 0.112 0.798 ± 0.112 0.398 ± 0.057 0.073 ± 0.049
Ba 0.004 ± 0.001 0.015 ± 0.009 0.033 ± 0.011 Nd 0.020 ± 0.010 0.021 ± 0.012
Be 0.001 ± 0.001 nd nd Nd nd nd
Cd 0.021 ± 0.028 0.004 ± 0.008 nd 0.007 ± 0.010 nd nd
Ca 184 ± 14 216 ± 63 193 ± 34 365 ± 31 78 ± 48 36 ± 12
Cl 27.6 ± 9.1 52.9 ± 19.7 23.4 ± 15.4 217.8 ± 87.4 409.4 ± 64.6 10.2 ± 5.3
Cr 0.001 ± 0.002 nd nd 0.007 ± 0.009 0.002 ± 0.003 nd
Cu 0.093 ± 0.089 0.027 ± 0.061 nd 0.028 ± 0.040 nd 0.007 ± 0.010
Fe (dissolved) 94.2 ± 10.7 42.3 ± 44.4 0.873 ± 0.844 32.5 ± 17.5 0.246 ± 0.525 6.9 ± 12.4
Mg 40 ± 5 43 ± 19 41 ± 5 84 ± 6 7.6 ± 16.5 10.2 ± 4.6
Mn (dissolved) 2.74 ± 0.44 2.37 ± 1.85 0.392 ± 0.105 6.31 ± 0.71 0.077 ± 0.148 0.59 ± 0.61
Na 25 ± 3 21 ± 5 27 ± 6 62 ± 28 93 ± 40 4.1 ± 1.7
Ni 0.045 ± 0.015 0.036 ± 0.050 nd 0.046 ± 0.022 0.003 ± 0.006 0.008 ± 0.011
Pb 0.010 ± 0.022 0.005 ± 0.011 0.013 ± 0.024 Nd 0.010 ± 0.013 nd 115
S 284 ± 24 232 ± 111 123 ± 128 343 ± 23 71 ± 17 44 ± 24
Si 17.15 ± 1.82 9.68 ± 5.32 8.22 ± 3.45 11.30 ± 3.50 2.86 ± 0.83 6.84 ± 2.07
Zn 0.101 ± 0.009 0.120 ± 0.080 0.001 ± 0.002 0.048 ± 0.016 nd 0.020 ± 0.017
Table A.2. Water quality in wells in the downdip area of the mine, installed either before (9719) or after (9901, 9904, 2002) grouting operations. Water quality data for two wells (9903 and 9906) installed in the upper mine works after grouting are also shown. Concentrations are in mg/L, unless noted. All concentrations correspond to average values over the period April 2000 to September 2001. The number of sampling dates (n) recorded for each well during this period is shown in parentheses.
Freeport 0-60
Kittanning Coal 1-2 Clay, Siltstone, 9-12
Clarion Sandstone 12-18
Not to
Figure A.1. Representative geological cross-section of the Roberts-Dawson site.
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9903
9906
Direction of Groundwater Fl
Q
9727 9719 Receiving S 2002
9901 Site 5
AMD 9904 Collection Unmapped Pond Mine Voids Scale 0 15
To Wills Creek Res.
Figure A.2. Site location and description.
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0.008 /s)
3 0.006
0.004
0.002
Flow Rate (m 0.000
Jul-99 Oct-96 Apr-97 Jun-98 Jan-00 Feb-01 Nov-97 Dec-98 Aug-00 Sep-01
Figure A.3. Flow rate of mine drainage water at surface water site 5.
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5 4
4 3 3 2 2
1 1 Flux Sulfur (g/s) Sulfur Flux (g/s) Acidity Flux 0 0
Jul-99 Jul-99 Oct-96 Apr-97 Oct-96 Apr-97 Jun-98 Jan-00 Jun-98 Jan-00 Feb-01 Feb-01 Nov-97 Dec-98 Nov-97 Dec-98 Aug-00 Sep-01 Aug-00 Sep-01
2.0 0.20
1.5 0.15
1.0 0.10
0.5 0.05
Flux Iron (g/s) 0.0 0.00 Flux Aluminum (g/s) Aluminum Flux
Jul-99 Jul-99 Oct-96 Apr-97 Oct-96 Apr-97 Jun-98 Jun-98 Jan-00 Jan-00 Feb-01 Feb-01 Nov-97 Dec-98 Nov-97 Dec-98 Aug-00 Sep-01 Aug-00 Sep-01 1.00 6.0E-03
0.75 4.5E-03
0.50 3.0E-03
0.25 1.5E-03
Flux Boron (g/s) Boron Flux (g/s) Calcium Flux 0.00 0.0E+00
Jul-99 Jul-99 Oct-96 Apr-97 Oct-96 Apr-97 Jan-00 Jan-00 Jun-98 Jun-98 Feb-01 Feb-01 Nov-97 Dec-98 Nov-97 Dec-98 Aug-00 Sep-01 Aug-00 Sep-01
Figure A.4. Flux of acidity, sulfur, iron, aluminum, calcium and boron at surface water
site 5, before and after placement of fixated FGD material grout.
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1400 Q H: hannebachite 1200 Q: quartz E: ettringite 1000 G: gismondine F: ferrihydrite 800 H
600
Intensity E H 400 E Q H G H H F H H E F 200
0 10 20 30 40 50 60 70 °2θ Figure A.5. X-ray powder diffraction pattern of core sample collected from site 9906.
120
1600
H H: hannebachite 1400 Q: quartz E: ettringite 1200
1000 H
800 H
Intensity 600 E Q H H H H H H E H H H H HH 400 H H E 200 0 10 20 30 40 50 60 70 °2θ Figure A.6. X-ray powder diffraction pattern of core sample collected from site 2002.
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APPENDIX B
MINIMIZATION AND USE OF COAL COMBUSTION BY-PRODUCTS (CCBS):
CONCEPTS AND APPLICATIONS
A Book Chapter for
The Handbook of Pollution Control and Waste Minimization
by
Harold W. Walker1, Panuwat Taerakul1, Tarunjit Butalia1, William E. Wolfe1,
and Warren A. Dick2
1Department of Civil and Environmental Engineering
and Geodetic Science
470 Hitchcock Hall
2070 Neil Avenue
Columbus, OH 43210
2School of Natural Resources 135 Williams Hall 1680 Madison Ave. Wooster, OH 44691 The Ohio State University
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B.1 Abstract
Federal regulations on emissions from coal burning power plants and efforts to improve air quality have increased the generation of solid by-products at coal-fired electric power plants. Approximately 97.7 million metric tons of coal combustion by- products (or “CCBs”) are generated annually in the United States, with 70% of these materials entering landfills. As a result, there is increasing need to modify existing coal burning processes to reduce the generation of CCBs while at the same time maintaining effective air pollution control. The generation of CCBs in coal burning power plants depends on a number of factors, including specific unit operations, operating conditions, and the source and type of coal. In addition to minimizing the generation of CCBs, it is important to develop beneficial uses for existing and future by-product streams. A number of uses for CCBs have been demonstrated, including agricultural amendment, construction of low permeability liners and road sub-base, and mine reclamation.
Successful utilization of CCBs depends, however, on detailed knowledge of the physical, chemical and engineering properties of CCBs. Important chemical properties to consider include elemental composition, mineralogy, and leaching behavior. Important physical and engineering properties include particle size and shape, density, strength, permeability, compaction characteristics and swell potential. Case studies demonstrate that with proper understanding of the properties of CCBs, effective utilization of these materials can be accomplished. Use of CCBs greatly reduces the amount of solid materials entering landfills, reduces greenhouse gas emissions, and conserves existing natural resources.
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B.2 Introduction and Background
During coal-fired electric power production four main types of coal combustion by-products (or CCBs) are produced: fly ash, bottom ash, boiler slag, and flue gas desulfurization (FGD) material (88,89). In 1998, 97.7 million metric tons of CCBs were produced in the United States (see Figure B1). Fly ash was generated in the largest quantity (57.1 million metric tons) with FGD material the second most abundant CCB
(22.7 million metric tons). Roughly 15.1 million metric tons of bottom ash were generated and 2.7 million metric tons of boiler slag were produced. Although the majority of CCBs produced currently enter landfills and surface impoundments, there is great potential for the effective and environmentally sound utilization of these materials.
Currently, the amount of CCBs entering landfills and surface impoundments is greater than half of the total municipal solid waste (MSW) disposed of in the United
States (see Table B1). Of the 97.7 million metric tons of CCBs generated in 1998, 69.4 million metric tons of CCBs (or 70%) were disposed of in landfills or surface impoundments (88). In 1997, the most current year for which data are available, the total
MSW disposed of in landfills was 119.6 million tons (90). The amount of CCBs disposed each year is greater than the amount of paper (37.4 million metric tons), plastic
(15.5 million metric tons), wood (8.4 million metric tons) and glass (6.9 million metric tons) discarded.
Recently, the American Coal Ash Association (ACAA) proposed that CCBs be considered a product, and therefore, they recommend these materials be referred to as coal combustion products (CCPs). Considered as a commodity, CCBs are ranked as the third largest non-fuel mineral commodity produced in the United States (88, 91). As
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shown in Table B2, the amount of CCBs generated every year exceeds the amount of
Portland cement generated in the United States, is significantly greater than the production of iron ore, and falls just behind the production of crushed stone, sand and gravel.
The purpose of this chapter is to review the current state-of-the-art in technology for minimizing CCB generation, maximizing CCB use, and reducing the disposal of
CCBs in landfills and surface impoundments. This chapter will first present a review of important federal regulations influencing the generation and utilization of CCBs in the
United States. Next, the physical, chemical and engineering properties of CCBs will be discussed and the operational factors affecting CCB generation will be presented. The chapter will conclude with a discussion regarding strategies for minimizing CCB production and maximizing the utilization of CCBs. Potential barriers to utilization and minimization in the future will also be discussed.
B.3 Federal Regulations Influencing CCB Generation and Use
Governmental regulations of emissions from electric power plants combined with efforts to improve air quality has had a profound effect on the amount and type of CCBs produced in the United States over the past 25 years. The Clean Air Act of 1967 was the first legislation to establish the authority of the federal government to promulgate air quality criteria (92). It set the groundwork for future “technology-forcing legislation”; i.e., legislation that set standards unattainable utilizing existing technology. This regulatory approach required industry and utilities to develop new technologies to meet promulgated standards.
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The Clean Air Act Amendments of 1970 established Natural Ambient Air Quality
Standards (NAAQS) and set specific pollutant removal requirements (New Source
Performance Standards or NSPS) for both stationary and mobile sources (92). NSPS, which are applicable to coal-fired utilities, were written in part 60, subpart D, Da, Db, and Dc of 40 CFR (Code of Federal Regulation) (93). NSPS in 40 CFR, part 60, subpart
D, set air pollutant levels for coal-fired steam generators with heat input rates over 73 megawatts (MW), constructed or substantially modified after August 17, 1971.
Amendments to the Clean Air Act in 1990 added new provisions to reduce the formation of acid rain by decreasing sulfur and nitrogen oxide emissions. Key to these provisions was the requirement to reduce annual SO2 emissions by 10 million tons below 1980 levels, and to reduce NOx emissions by 2 million tons below 1980 levels. To achieve these emission reductions, the Clean Air Act Amendments of 1990 promulgated NOx and
SO2 performance standards and set up an innovative emission trading system for SO2 reduction. In phase I of the SO2 reduction program, the legislation required 110 identified utilities to reduce SO2 emissions to 2.5 lb/mmBTU by January 1995. Phase II mandated further reductions in emissions to 1.2 lb/mmBTU for all utilities generating at least 25 MW of electricity. It is estimated that phase II requirements will affect 2,128 utilities in the United States (5). The NOx reduction program was also separated into two phases. In phase I, Group 1 boilers (dry-bottom wall and tangentially fired boilers) were required to meet NOx performance standards by January 1996 (6). Phase II set lower
NOx emission limits for Group 1 boilers and established initial NOx emission limitations for Group 2 boilers (cell burner technology, cyclone boilers, wet bottom boilers, and other types of coal-fired boilers) (5).
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To meet these federal regulations, coal-fired utilities have switched to alternative fossil fuels or installed air pollution control technologies such as electrostatic precipitators, baghouses, and wet or dry SO2 scrubbing systems. Currently, CCBs generated as a result of air pollution control processes are regulated under subtitle D of the Resource Conservation and Recovery Act (RCRA) which pertains to non-hazardous solid wastes (94). In 1988, and then again in 1999, USEPA issued a "Report to
Congress" examining the environmental impacts associated with CCB use and disposal
(4,56). Reports in both 1988 and 1999 concluded that CCBs were non-hazardous and non-toxic materials. In early 2000, based on its own findings in the Report to Congress as well as input from environmental groups, the USEPA maintained its previous ruling that CCBs will continue to be regulated under subtitle D of RCRA. As a result, the use and/or disposal of CCBs is regulated at the state level. For example, regulations in the state of Ohio consider fly ash, bottom ash, boiler slag and FGD generated from coal or other fuel combustion sources to be exempt from regulation as hazardous waste (95).
B.4 Physical, Chemical and Engineering Properties of CCBs
Information regarding the physical, chemical and engineering properties of CCBs is required before these materials can be safely and effectively utilized. The physical and engineering properties, in particular, are important parameters affecting the behavior of
CCBs in various engineering applications. Information regarding the chemical composition is important for addressing potential environmental impacts associated with
CCB utilization and disposal. Chemical data is also useful for explaining physical properties when pozzolanic or cementitious reactions take place.
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As mentioned above, the four main types of CCBs are fly ash, bottom ash, boiler slag, and FGD material. Fly ash is a powdery material removed from electrostatic precipitation (ESP) or baghouse operations while bottom ash is a granular material removed from the bottom of dry-bottom boilers. Boiler slag is a granular material that settles to the bottom of wet-bottom and cyclone boilers. It forms when the operating temperature in the boiler exceeds the ash fusion temperature. Boiler slag exists in a molten state until it is drained from the boiler. The majority of FGD material is a mixture of fly ash and dewatered scrubber sludge. Scrubber sludge is produced when flue gases are exposed to an aqueous solution of lime or limestone. The wet scrubber sludge is dewatered and stabilized with fly ash and extra lime. Alternatively, the scrubber sludge can be oxidized to calcium sulfate (CaSO4) to produce synthetic FGD gypsum. Dry FGD processes are widely used in which limestone is injected directly into the boiler or flue gas stream. Dry FGD by-products are removed from the flue gas by electrostatic precipitation or baghouse operations.
B.4.1 Physical and Engineering Properties of CCBs.
A number of the physical and engineering properties of fly ash, bottom ash, boiler slag and FGD material are summarized in Table B3 (4,44,45,56,96,97). Fly ash is usually spherical with a diameter ranging from 1 to 100 microns. Fly ash has the appearance of a gray cohesive silt and has low permeability when compacted. Bottom ash and boiler slag are granular in shape with sizes ranging from 0.1 to 10.0 millimeters.
Boiler slag has a glassy appearance. Bottom ash has a permeability higher than fly ash while boiler slag has permeability similar to course sand. Fly ash, bottom ash and boiler
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slag have dry densities that range between 40-100 pounds per cubic foot (4,56,98,99).
Fly ash has greater shear strength than both bottom ash and boiler slag.
The physical characteristics of FGD material depend on type of FGD system used; wet or dry (see Table B3). Wet FGD systems generate by-products with diameters ranging from 0.001 to 0.05 millimeters. Dry FGD systems produce by-product with diameters ranging from 0.002-0.074 millimeters. FGD material generally has low permeability, ranging from 10-4 to 10-7 centimeter per second. The unconfined compressive strength is affected by the moisture content of FGD, the percentages of fly ash and lime. For example, wet FGD scrubber sludge is similar to toothpaste in consistency and has little unconfined compressive strength. However, the strength of wet
FGD is greatly improved when FGD sludge is stabilizing by mixing with lime and fly ash.
B.4.2 Chemical Properties of CCBs.
The chemical characteristics of fly ash, bottom ash, and boiler slag depend greatly on the type of coal used and the operating conditions of the boiler (4,56). Over 95% of fly ash consists of oxides of silicon, aluminum, iron and calcium with the remaining 5% consisting of various trace elements (4,56). The chemical composition of fly ash is affected by the operating temperature of the boiler because the operating temperature influences the volatility of certain elements. For example, sulfur may be completely volatilized at high temperature and removed during lime scrubbing, thus reducing the amount in the fly ash, bottom ash and boiler slag (4,56).
Table B4 shows the trace element content of fly ash, bottom ash, boiler slag, and
FGD material (4,56,101). The elemental composition of fly ash from two types of
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collection methods is shown. Mechanical collection methods generally collect larger particles from the flue gas while finer ash particles are collected by ESPs or baghouses.
However, similar ranges of most trace elements are found in both types of collection methods. Some exceptions to this are arsenic, boron, lead and selenium which may be found at slightly higher fractions in fly ash collected by ESPs or baghouses. Cadmium and fluorine may be present at higher levels in ash collected by mechanical methods.
The chemical characteristics of FGD by-product depend on type of absorbent used and the sulfur content of the coal. In the United States, approximately 90% of FGD systems use lime or limestone as a sorbent (100). In lime-based FGD processes, the absorbent reacts with sulfur in the flue gas and forms a calcium compound; either calcium sulfite or calcium sulfate, or a calcium sulfite-sulfate mixture (4,56). In systems that use dual-alkali scrubber technology, sodium hydroxide, sodium sulfite, or lime is used as absorbent solution. These types of systems generate calcium sulfite and sodium salts (4,56). In spray-drying scrubber systems, sodium sulfate and sodium sulfite are produced with sodium-based reagents. When fly ash is added to FGD the quantity and characteristics of the fly ash will also affect FGD chemical characteristics.
The most significant components in FGD include calcium and sulfur, with lesser amounts of silica, aluminum, iron, and magnesium if fly ash is added. The elemental composition of dry FGD materials has been determined based on data from a variety of dry scrubber technologies, including spray dryer systems, duct injection, lime injection multistage burner (LIMB) processes, and a number of fluidized bed combustion (FBC) processes (i.e., bed-ash process and cyclone ash process) (44,45). The calcium content of dry FGD material varies in the range from 10% to 30% depending on the particular
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scrubber technology. The sulfur content of dry FGD material typically varies between
4% and 11%. The silicon content of dry FGD may range from 2% to 11% while the aluminum content can vary from 1% to 7%. Table B4 shows the trace element content of dry FGD materials (101). Although detectable amounts of arsenic, cadmium, chromium copper, lead, molybdenum, nickel, selenium, and zinc are present in dry FGD materials, levels of these constituents are typically lower than USEPA land application guidelines for sewage sludge.
For many CCB applications, it is important to understand the leaching behavior of these materials. The United States Environmental Protection Agency (USEPA) Toxicity
Characteristic Leaching Procedure (TCLP) is a method commonly used for characterizing the leaching potential of organics, metals and other inorganic constituents from CCB matrices (102). Table B5 shows the results of TCLP analyses of dry FGD materials and ash produced from various air pollution control technologies. Typically, very low levels of organic materials are found in CCBs, and therefore, TCLP tests focus on examining the leaching behavior of inorganic constituents. The TCLP values for FGD shown in
Table B5 were determined for a variety of dry scrubber technologies. TCLP leachate typically meets most primary and secondary drinking water standards. Levels of silver, arsenic, barium, cadmium, copper, iron, mercury, manganese, nickel, phosphorus, antimony, and zinc in leachate are typically below the limit of detection. For all FGD materials shown in Table B5, high pH values are observed, thus making FGD an attractive product for applications requiring alkaline materials. Typically, with the exception of sulfur and calcium, higher levels of most inorganic elements are found for
TCLP tests carried out with ash than for FGD. It should be noted that the acidic
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conditions and high liquid to solids ratio of the TCLP test are perhaps more favorable for leaching than conditions typically observed in field applications.
B.5 Factors Affecting CCB Generation
The physical and chemical properties of CCBs and the quantity of CCBs produced will depend on the mechanical design and operation of the combustion process, the type of air pollution control equipment utilized, as well as the characteristics of the coal used in the combustion process (4). In order to minimize CCB generation, it is important to understand how these factors impact the type and amount of solid by- product produced. In all cases, however, efficient energy production and low pollutant air emissions must be maintained.
B.5.1 Boiler Technology.
The boiler used in an electric power plant is a closed vessel that is heated from the combustion of coal to produce hot water or stream. There are four major types of boiler technologies in current commercial application; pulverized coal (PC) boilers, stokers, cyclones, and fluidized bed combustion systems. Figure B2 shows the approximate distribution of ash and slag produced by different kinds of boiler technology.
The most widely used boiler technology is the PC boiler. The coal used in PC boilers is finely ground prior to combustion. The large effective surface area of finely ground coal used in PC boilers increases combustion efficiency. The greater efficiency of combustion reduces the total volume of ash by-products produced. There are two types of pulverized coal boilers; wet bottom and dry bottom boilers. The larger sized ash that falls to the bottom in a dry-bottom process remains dry and becomes bottom ash.
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For the wet-bottom process, ash is removed as a flowing slag. Large ash particles fall to the bottom of the furnace and flow out of the furnace in a molten state which later solidifies as slag (4,56). As seen in Figure B2, dry-bottom PC boilers produce 80% fly ash and 20% bottom ash. PC boilers with a wet-bottom design produce 50% fly ash and
50% slag. The predominance of fly ash in these two types of boilers is primarily a result of the small particle size of ground coal used in the combustion process.
Stoker boiler technology is typically used in smaller utility plants. A stoker boiler is classified based on the location of the stoker, the method of coal feeding, and the method of grating coal in the furnace. Spreader stokers are the most widely used of all stoker technologies (4,56). The bottom ash generated by spreader stokers ranges from free-flowing ash to fused slag, while the bottom ash created from other types of stokers is normally slag (4,56,103). Figure B2 shows that spreader stokers produce 35-60% fly ash and 40-65% bottom ash and slag. Other types of stokers produce about 10% fly ash and
90% bottom ash and slag.
Cyclone boilers are used for coal combustion and are designed to circulate air to enhance the combustion of fine coal particles in suspension. This design helps to reduce erosion and fouling problems in the boiler. Larger ash particles stick to the molten layer of slag and flow out. Combustion occurs in a horizontal cylindrical vessel attached to the boiler. This kind of design facilitates the flow of molten slag and also reduces the cost of particulate collection (4,56). Most of the by-product from the cyclone design is in the form of slag. Figure B2 shows that cyclones produce 30% fly ash and 70% slag.
There are also other technologies that are used for coal combustion. These alternative boilers may also aid in controlling air emission. Fluidized bed combustion is
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a boiler technology that can be used with a variety of fuels (103). This technology has a high combustion efficiency at low operating temperatures (4). The fluidized bed combustion system consists of a blower that injects pre-heated air into the fluidization vessel, and a bed material that can be sand or a reactive solid. Injection of air into the vessel fluidizes the bed material and aids in combustion. The amount of CCBs produced from FBC is based on the type of FBC. Two types of FBC include bubbling fluidized bed systems and circulating fluidized bed systems.
Bubbling fluidized bed systems have gas velocities of 5-12 feet per second. Gas flow passes through the bed and causes the bed material to “bubble”. In bubbling FBC systems, the particle size of bottom ash in the bed is usually larger and is packed denser
(~45 pounds per cubic foot) than in circulating FBC systems (4,103). CCBs generated from bubbling FBC systems include ash, sand, and other inert bed material. Lime or limestone may be added directly to the bed to aid in sulfur emission control (4,103). As a result, by-products from FBC boilers may also contain unreacted lime, calcium sulfate and/or calcium sulfite.
Circulating fluidized bed combustion systems have higher gas velocities of about
30 feet per second. In circulating systems, some of the bed material is recovered from the gas phase and re-injected into the fluidized bed vessel. Bottom ash in the bed of circulating FBC systems is usually finer and more densely packed (~35 pounds per cubic foot) than in bubbling FBC systems (4,103). Ash generated from circulating FBC systems consists mainly of fly ash with lesser amounts of bottom ash (103).
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B.5.2 Air Pollution Control Technology.
The type of technology used for controlling pollutants released to the atmosphere during coal combustion influences the generation and characteristics of CCBs. There are two main categories of air pollution control technologies that generate CCBs during coal combustion; particulate control and gaseous emission control technologies.
Particulate control technologies during coal combustion capture fly ash from the flue gases before they are released to the atmosphere. The processes most often used for particulate control are electrostatic precipitation, fabric filtration, scrubbers, and mechanical collectors. The electrostatic precipitator (ESP) is the most common process used for capturing fine ash particles in coal-fired utilities (4,103). ESPs capture ash by applying an electrical charge to the ash particles. The charged particles are subsequently attracted to oppositely charged collector surfaces in an intense electrical field. Following collection, the particles are sent to a hopper. This technology is appropriate for capturing fly ash from coal with high sulfur content. In fact, sulfur oxides in the flue gas may increase the efficiency of particle capture in the ESP (4,56,103). The capability of ESP to capture fly ash in the flue gas is more than 99% when this process is properly operated and maintained (4,56).
A fabric filter unit, or baghouse, is an appropriate technology for particulate control in combustion processes that use coal with low sulfur content. This technology operates by forcing the flue gas through a fine mesh filter. Fly ash is trapped and builds up on the filter surface. The ash on the filter forms a cake which is then periodically removed. The efficiency of the filter increases as ash forms a thick layer on the filter surface. However, thick cake formation also leads to greater head losses in the process.
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Fabric filters can remove over 99% of fly ash from the flue gas in coal-fired utilities
(4,56).
Scrubbers can also be used for particulate control and operate by applying water to contact the fly ash in the flue gas in a spray tower. This technology also can remove over 99% of large ash particles but less than 50% for particles with size smaller than 1 or
2 micrometers (4,103). Mechanical collectors are instruments used for removing primarily large ash particles. They operate by forcing the ash particles against a collector wall where the dry ash by-product is collected. The efficiency is lower than 90% for small particles.
Desulfurization technology is used for capturing gaseous sulfur oxides from flue gas in coal-fired utilities. The use of desulfurization technology results in the generation of FGD material. There are two major types of FGD systems; non-recovery and recovery systems. 95% of FGD systems in the United States are non-recovery systems (4,103).
Non-recovery systems produce by-product material, mainly calcium sulfate or sulfite, that has to be disposed or used. The non-recovery FGD process can be separated into two types; wet and dry systems. Wet systems operate by contacting the flue gas with a slurry of water and sorbent. Examples of wet scrubber systems include direct lime, direct limestone, alkaline fly ash and dual-alkali. As mentioned earlier, approximately 90% of
FGD systems in the United States use lime or limestone as a sorbent (100). Typically, the calcium sulfite/sulfate sludge produced in wet systems is dewatered and mixed with fly ash and lime to produce “stabilized” FGD. Examples of dry non-recovery FGD systems include spray drying and dry sorbent injection (4). Wet FGD systems produce more FGD per pound of coal than that of dry systems because of the use of water in the
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process. Recovery systems produce materials that can be used again in the FGD process because the sorbent can be recycled. Recovery FGD processes also have wet and dry systems. Examples of recovery FGD systems include Wellman-Lord and Magnesium oxide systems and aluminum sorbent and activated carbon sorbent systems (4).
B.5.3 Types of Coal.
Different types of coal have different heating values and also different ash contents. The highest ranked coal with respect to heating value is anthracite while the lowest ranked coal is lignite (103). The generation of ash and slag from the combustion process is affected by the ash content which is determined by the rank of the coal.
Anthracite coal in the United States generates the greatest amount of ash, about 30%.
Bituminous coal ash content ranges from 6-12%, and subbituminous and lignite coals have ash contents ranging from 6-19 % (4,56). The rank of coal is affected by the specific region of coal mining, mine, seam and production method (4,103). The use of low ash content coal reduces the management cost for removing particulate matter. In the United States, the average ash content of coal used decreased from 13.5% in 1975 to
9.22% in 1996 (4,103,104). In addition, coal can be cleaned before the combustion process to reduce the quantity of CCBs. The cleaning process as a pretreatment for coal can reduce the ash content by 50 to 70% (103).
The generation of FGD depends on sulfur content of the coal. The sulfur content varies from region to region. Some coal produced in Iowa has sulfur content as high as
8% by weight while coal from Wyoming may have an average sulfur content of less than
1% by weight (103). Coal cleaning processes that reduce the ash content of coal can reduce sulfur emissions by removing pyrites and other metal sulfides from coal.
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Technologies for pre-combustion coal desulfurization are characterized as physical, chemical or biological. Currently, the physical cleaning processes are the most widely used. Physical processes use density differences to separate out pyrites from the coal (105). Chemical processes use a chemical agent to desulfurize coal. Chemical desulfurization processes that use chemical reagents such as ferric salts, chlorine, and ozone are currently not cost effective due to high chemical recovery costs. In addition, chemical processes are also energy intensive due to high operating pressures (600-1000 psi) and temperatures (100-500 °C) (106,107,108,109). Recently, new chemical agents have been developed that may provide a cheaper approach for the chemical cleaning of coal (110).
In biological processes, microorganisms are used to remove sulfur from coal.
Biological processes can be operated at room temperature and atmospheric pressure, and therefore, have lower costs than some physical and chemical coal cleaning techniques. In addition, biological pretreatment of coal does not reduce the BTU value of the coal, but instead, may increase coal energy content due to the remaining biomass
(106,107,108,111). Biological coal cleaning processes can remove both inorganic and organic sulfur. Thiobacillus ferrooxidans can oxidize inorganic pyrite (FeS2) in coal while microorganisms such as Rhodococcus rhodochrous, Sulfologus brierleyi, and
Sulfolobus acidocaldarius remove organic sulfur compounds (112,113,114,115,116).
Although pre-combustion cleaning can reduce flue gas desulfurization requirements, additional by-products may be produced and energy may be required.
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B.6 Strategies for Minimization of CCBs
The established hierarchy for minimization of waste materials in any process consists of the following: reduction>recycle/re-use>treatment>disposal. In this hierarchy, strategies for reduction and recycle/use are favored over "end-of-pipe" treatment and disposal options for waste material.
B.6.1 Reduction at Source.
Strategies for the reduction of CCBs at the source include process modifications, feedstock improvements, improvements in efficiency of equipment, better management practices, and recycling of material within or between processes. Possible process modifications may include changes to boiler operating conditions, selection of wet versus dry scrubbing technologies, or addition of pre-combustion coal cleaning, to name a few.
Changes in feedstock may also aid in CCB minimization. Recent advances are providing more efficient technologies for removal of pollutants from flue gas in coal combustion facilities. For example, new sorbents with high efficiency for SO2 capture do a better job of removing SO2 from flue gas while at the same time reducing the amount of solid by- products produced during the process. Better management of the coal combustion process may also lead to improved CCB minimization. Internal audits provide opportunities to optimize process operations thus maximizing energy production and minimizing CCB production. For example, one factor limiting use of CCBs is the variability in CCB properties. Better management practices may reduce this variability, lower total volume of CCBs produced, and enhance utilization.
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B.6.2 Use of Coal Combustion By-Products.
A number of applications of CCBs have been developed and demonstrated in order to reduce the amount of CCBs disposed of in landfills. The first column in Table
B6 shows demonstrated applications of CCBs. In 1998, the American Coal Ash
Association (ACAA) reported that 28 million metric tons of CCBs were used in the
United States. This represented 30% of all CCBs generated in that year. One demonstrated application is the use of CCBs in cement, concrete and grout. CCBs may also be used in flowable and structural fill applications. Demonstrated applications of
FGD material include use in the production of wallboard or in mine land reclamation.
The use of CCBs reduces the need for landfill space and also reduces the utilization of natural resources. Table B6 presents some of the environmental benefits associated with CCB use. Approximately 10 million metric tons of fly ash and bottom ash were used as a replacement for cement in non-fill applications. This was the largest application of fly ash and represented roughly half of all the fly ash used in 1998. It has been estimated that every ton of cement replaced with fly ash eliminates the emission of approximately 1 ton of CO2 to the atmosphere (88). Based on this figure, the use of fly ash in cement applications reduced the release of CO2 by 10 million tons in 1998. Also, in 1998, 0.36 million metric tons of fly ash were used for flowable fill. This application also replaced cement and represented approximately 1.8% of all fly ash used. The use of fly ash in flowable fills reduced the emission of CO2 by 0.36 million metric tons in 1998.
If all of the fly ash generated were used at current utilization percentages (50% for cement replacement and 1.8% of flowable fill), CO2 emissions would be reduced by roughly 32 million metric tons (31 million tons as a result of cement replacement and 1
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million tons due to the use of fly ash in flowable fill). Structural fill is an application in which CCBs, and in particular fly ash, are used to replace natural soil. In 1998, the use of fly ash and bottom ash for structural fill saved 3.6 million metric tons of soil and could save up to 11.2 million metric tons of soil if 100% of all CCBs were used at the current utilization rate for structural fill (13.2%).
In 1998, 2.2 million metric tons of FGD were used which represented only 8% of all FGD produced. 1.6 million metric tons of FGD were used in the production of wallboard. This represented the single largest use of FGD and 73% of all FGD used in that year. As a result, approximately 1.6 million metric tons of natural gypsum were saved by using FGD gypsum in the wallboard industry. FGD could replace up to 18 million tons of gypsum if all FGD in the United States were reused, 73% of which for wallboard manufacturing. This amount of FGD wallboard could supply a good fraction of the 1.6 million new houses built in the United States every year. About 0.6 million tons of clay used for mining applications were saved in 1998 by replacement with CCBs and up to 2.2 million tons could be saved if all CCBs were used (6.8% for mining applications). The use of CCBs in 1998 reduced landfill space consumption by about 17 million m3. If all CCBs were used, landfill space consumption would be reduced by 59 million m3 each year.
B.6.2.1 Cement/Concrete/Grout Application.
Using fly ash in cement and concrete increases the strength, workability and resistance to alkali-silica reactivity and sulfate, and reduces permeability, bleeding and heat of hydration (96). The Federal Highway Administration (FHWA) and ACAA have reported that fly ash enhanced concrete has lower strength than pure portland cement in
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early periods, but provides higher strength in the long-term (96). In 1998, 10.2 million metric tons of CCBs were used in concrete applications. The amount of fly ash used with cement varies from 15 to 20% of the total weight. Typically, 1 to 1.5 kg of fly ash is used for every 1 kg of cement replaced in concrete applications (96). Lowering the cement content reduces the emission of CO2 caused by the calcination of limestone and fuel burning during cement production.
Future efforts to reduce nitrogen oxide emissions from coal combustion facilities may negatively impact the utilization of fly ash in concrete. To control emission of NOx, many coal-fired utilities may utilize new low NOx emission technologies. Some of these technologies operate at lower temperatures than traditional boilers. Operation at lower temperature may result in an increase in carbon and ammonia content of fly ash (96).
Carbon content, which is typically measured by determining the weight loss on ignition
(LOI), affects concrete strength development and so is restricted by industry standards.
In the United State, the specification of LOI of fly ash is between 3% and 5% for use in ready mix-concrete (117,118). The use of low NOx technology may increase the LOI above 5% and thus reduce the utilization of fly ash in concrete applications (20). For example, in 1998 ACAA reported that 19 out of 20 coal-fired utilities in Ohio may be affected by the low NOx emission technologies and 46% of all coal-fired utilities in the
United States may be required to control NOx (20).
B.6.2.2 Flowable Fill.
Flowable fill is another application in which fly ash can be used in place of cement. Flowable fill is defined as a self-leveling, self-compacting cementitious material that is in a flowable condition at the time of placement and has a compressive strength of
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1,200 pounds per square inch (psi) or less at 28 days (119). Flowable fill is also known as Control Density Fill (CDF), Controlled Low-Strength Material (CLSM), unshrinkable fill, flowable mortar, plastic-soil cement slurry, K-Krete, and/or Flash Fill (119).
Flowable fill may contain a mixture of fly ash, bottom ash, water, and portland cement.
This application is especially suitable for filling in void spaces that are difficult to reach.
The use of fly ash and bottom ash in flowable fill applications is increasing.
Recently, a technical guidance manual for flowable fill applications has been written; the
ACI229 Committee Report on Controlled Low Strength Material (119). However, the current use of CCBs in this application is still relatively small. As mentioned above, only
360,000 metric tons of fly ash and bottom ash were used in flowable fill applications in the United States which represented only 1.8% of all fly ash and 0.3% of all bottom ash used.
B.6.2.3 Embankment/Structural Fill.
Embankment and structural fill is currently the second largest application of
CCBs. CCBs are used to replace conventional soil in structural fill applications. CCBs have many advantages over natural soil for use as structural fill, including lower unit weight, high shear strength to unit weight ratio, and good availability in bulk (96).
These benefits lead to significant savings in material costs when CCBs are used. For example, in 1993, FGD material from pressurized fluidized bed combustors (PFBC) was used as embankment to repair part of Ohio State Route (SR) 541 near Coshocton, Ohio
(120,121,122). The cost of this project was $77,000 while the estimated cost of using conventional materials would have been between $105,000 and $120,000. This
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represented a savings of between 26% and 36%, before counting the environmental benefits associated with saving existing natural resources.
The use of CCBs for structural fill has increased in recent years. ASTM
(American Society for Testing and Materials) Standard E1861 defines the appropriate guidelines for the use of CCBs for structural fill (88,123). In 1998, approximately 4 million tons (12.7%) of CCBs were used as structural fill. Fly ash and bottom ash were used the most, 2.5 and 1.1 million metric tons respectively. Although only 18,000 metric tons of FGD were used in 1998, this material has demonstrated excellent strength over a wide range of moisture contents compared to natural soils (120,121,122). FGD could be an excellent alternative for use in embankment and structural fill applications in the future (120,121,122).
B.6.2.4 Stabilized Base/Sub-Base.
Mixing fly ash with lime and aggregate can produce a good quality road base and sub-base. This material is also called lime-fly ash –aggregate (LFA) or pozzolanic- stabilized mixture (PSM) bases. The fly ash content in the material for road base typically varies from 12% to 14%. Lime content also varies from 3% to 5% (96). The proportion of lime can be replaced with portland cement or cement kiln dust. The advantages of using LFA or PSM for road base and sub-base applications include increased strength and durability of the mixture, lower cost, autogenous (self-generating) healing, and less energy consumption (96). Use of LFA reduces the energy to produce cement. In addition, it does not require heat like an asphalt base.
In 1998, there were 3.6 million tons of CCBs used in road base and sub-base applications. The CCBs used most commonly for these applications were fly ash and
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bottom ash (1.4 million tons of fly ash and 1.6 million tons of bottom ash). For example, ten municipal and commercial projects in and around the City of Toledo used approximately one million tons of LFA from 1970 to 1985 (124). In 1998, Hunt et al.
(125) developed an economic analysis of using LFA compared with other pavement materials. It was found that LFA base pavement was 20% cheaper than aggregate base pavement and 15% cheaper than bituminous base pavement (125).
B.6.2.5 Mining Applications.
Use of CCBs in mining applications can aid in the abatement of acid mine drainage (AMD), reduce subsidence, and reduce offsite sedimentation control. Acid mine drainage is an environmental problem caused by water drainage from abandoned mines and coal refuse piles. Water in abandoned mines react with pyrite and other metal sulfides in the presence of oxygen and produces acidity (126). Fly ash and/or FGD can be used to minimize the exposure of pyrite to water and oxygen. Also, the alkalinity in
FGD can neutralize AMD already generated. In 1998, 2 million metric tons of CCBs in the United States were used for mining applications. This represented 6.8% of all CCBs used. Most of material used in this application was fly ash (1.9 million tons).
At mine sites, refuse waste materials such as soil, rock, slate and coal are commonly found and can pose serious environmental problems. These refuse waste piles, commonly called “gob piles”, contain pyrite and produce acidity. FGD material may be used as a liner to construct ponds to collect runoff from gob piles. The low permeability of FGD may also be used as a cap to minimize the amount of water entering the gob pile. For example, the Rock Run Reclamation site at New Straitsville, Ohio had approximately 14 acres of gob piles prior to reclamation utilizing CCBs (20). The
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drainage from these gob piles to Rock Run had a pH of 2.27. At this site, two feet of stabilized FGD from the Conesville coal-fired power plant in Ohio was used to cover the gob piles. Utilization of FGD as a cap material resulted in improvements in water quality at the site. Estimated cost saving from using FGD for gob pile reclamation, instead of clay, ranged from $8,350 to $12,600 (20).
Another potential application of FGD is in the reclamation of abandoned surface mines (20,127). An example of such an application was carried out at the Fleming Site located in Franklin Township of Tuscarawas County, Ohio (128,129). The Fleming site was an abandoned clay and coal mine. In the past, flooding downstream of the Fleming
Site resulted in offsite soil sedimentation at an estimated rate of 450 tons/acre/year. In
1994, several AMD treatment approaches were developed at this site utilizing limestone, a dry FGD material from an PFBC plant, and a 2.5:1 mixture of FGD and yard waste.
All treatments resulted in neutralization of mine drainage. Trace metal analysis showed water quality improved, and in fact, met all drinking water standards.
B.6.2.6 Wallboard Manufacture.
Instead of being stabilized with fly ash and sent to landfill, FGD material can be used as a material to manufacture wallboard. The calcium sulfite in FGD can be oxidized to calcium sulfate and dewatered to produce synthetic FGD gypsum (20,130). In 1998,
1.6 million tons of FGD were used in wallboard manufacturing (88). This represented
72.9% of all FGD used. There will be a $20 million investment in oxidation and dewatering equipment to produce synthetic FGD gypsum at the Zimmer Plant (located in
Moscow, Ohio) by Cinergy, American Electric Power, and Dayton Power and Light (131,
132, 133). The synthetic FGD gypsum produced at the Zimmer plant will supply a
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wallboard plant at Silver Grove, Kentucky. Upon completion, the Silver Grove wallboard plant will have the highest wallboard production capacity in the world at 900 million square feet per year.
B.6.2.7 Agricultural Applications.
There are many advantages of using CCBs in agricultural applications, and addition of CCBs can improve the properties of soil and increase plant growth. For example, the lime in CCBs can raise the pH of acidic soil to neutral levels, and increase yields of alfalfa above those observed using agricultural limestone (45). Trace elements in FGD can be utilized by plants and may aid in plant growth (105). In 1998, 920,000 metric tons of CCBs were used in agricultural applications. It has been estimated that
25% of the agricultural lime used in Ohio could be replaced by FGD (20). Assuming every ton of agricultural lime is equivalent to 1.67 tons of FGD, the potential use of FGD for replacing agricultural lime in Ohio alone would be 365,000 tons per year (43).
CCBs can also be used in animal production facilities as a base for livestock feeding and hay storage (23). The moisture from mud can deteriorate the quality of hay bales and decrease animal yields. Utilizing CCBs as a base material can reduce muddy conditions. In 1997, 24 livestock feeding and hay storage pads (ranging size from 1,500-
15,000 square feet) were built in eastern and southern Ohio. Over 150 FGD pads were constructed in 12 counties in Ohio in 1998 (23). An economic analysis of the construction of FGD pads performed for Gallia County, Ohio in 1997 showed that FGD pads were 26% cheaper than aggregate pads and 65% cheaper than concrete pads (134).
FGD can also be used for constructing low permeability liners for water holding ponds and manure storage. Normally, the construction of manure facilities in the United
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States utilizes compacted clay. The successful utilization of FGD requires that the material provide sufficiently low permeability and does not degrade groundwater quality.
A study has been conducted to evaluate the permeability and the water quality of the leachate from an FGD liner (135, 136, 137). It was found that the FGD liner has a permeability as low as 10-7 centimeters per second. Moreover, trace element concentrations in the leachate were generally lower than the drinking water standards
(137). It has been estimated that replacing clay or geomembranes with FGD material for pond liners could save construction costs by as much as $2-$3 per square foot (135, 136,
137).
B.6.3 Treatment and Disposal.
Options for treatment and disposal are important considerations for evaluating the ultimate fate CCBs. For example, in wet scrubbing technologies a calcium sulfite slurry is produced that is de-watered prior to disposal. The de-watering step reduces the total volume of material going to final disposal, but also generates a liquid waste stream. The de-watered scrubber sludge is typically mixed with equal amounts of fly ash in order to improve the handling characteristics of the material. The exact proportion of fly ash to de-watered scrubber sludge has a significant impact on the performance of these materials during disposal and/or use.
B.7 Life Cycle Assessment (LCA) Model for Minimization of CCBs
A life cycle assessment model (138) can be used to determine optimum strategies for minimizing the environmental impacts associated with coal combustion processes.
The benefit of this approach is that it does not trade gains made in minimizing the amount
148
of CCBs with other environmental impacts. In a life cycle assessment model, the coal combustion process can be broken down into the following elements: resource extraction, electricity generation, electricity transmission, and electricity use. In developing an LCA model, each element of the coal combustion process must be evaluated with respect to the following variables: materials choice, energy use, solid residues, liquid residues, and gaseous residues. Once this has been carried out, an LCA matrix can be developed which includes a numerical score (from 1 to 4) for each variable and element. The rank of the entire process can then be determined as,
= R ∑∑M ij (1) ij where R is the rank of the process and Mij is the numerical score summed over each variable (i) and element (j). It should be noted that in certain cases it might not be possible or appropriate to develop a single score defining a process. In such cases, ISO
14040 provides criteria by which a comparison can be made between two or more potential options using the LCA framework (139).
The LCA model provides a framework for analyzing different process alternatives and determining the most environmental sound option. For example, it was mentioned above that one strategy to minimize CCB generation is to utilize a low sulfur, low ash coal source. However, utilizing low sulfur, low ash coal may not be the most environmentally sound alternative in all cases. If a preferable coal source is located far from the coal-fired utility, significant environmental impacts may be associated with the transportation of these materials.
149
Recently, a comparison between the use of fly ash and natural soil for structural fill was conducted using the LCA framework (140). This study considered a number of factors, such as natural resource use, energy consumption, air emissions, and solid waste.
It was concluded that the use of fly ash as structural fill resulted in less use of natural raw materials, significant reductions in solid waste disposal, and was more energy efficient than the use of natural resources. In this particular study, however, the impact of fly ash on water quality was not quantitatively assessed due to variations in soil and fly ash leaching behavior.
B.8 Barriers to CCB Utilization
Currently, the cost of disposing and managing CCBs in landfills is relatively low and this reduces the incentive to utilize CCBs. For example, landfill costs for coal-fired utilities in Ohio range from $2-$40 per ton (20,45). A lack of standards for using CCBs is also a barrier for CCB utilization. For example, the use of CCBs in structural fill has typically fluctuated in the past. Increasing use of CCBs for structural fill is expected in the future as a result of ASTM Standard E1862 (123). New, large volume, cost effective, applications of CCBs are needed.
Because CCBs are secondary products of energy production, few controls on their generation are in place. As a result, the physical, chemical, and engineering properties of
CCBs may vary. Significant variations can be observed in CCB properties within a given plant, as well as variations between plants. Future emission standards for coal-fired power plant will also affect the utilization of CCBs. As mentioned above, NOx control technologies influence the carbon content of fly ash. The increased carbon content of fly
150
ash generated from low NOx boilers could reduce the potential for use in cement applications, the biggest current CCB market.
Although extensive testing has shown that CCBs are non-toxic and non- hazardous, public concern regarding the environmental impacts of CCBs may limit utilization. Continued research on the environmental impacts associated with CCB use and public education are needed. As is the case for any product or process, efforts should be focused primarily on reducing CCB generation, while maintaining efficient energy production and air pollution controls.
B.9 Conclusions
Fly ash, bottom ash, boiler slag and FGD material are by-products from the combustion of coal and are considered to be solid wastes from federal regulatory perspective. Currently, 100 million tons of CCBs are produced in the United States every year. Approximately 70 million tons of CCBs are disposed of in landfills and surface impoundments.
The two primary strategies for minimizing CCBs include reduction at source and effective utilization. A life cycle assessment model should be used to determine the most environmentally beneficial approach for minimizing CCB generation and disposal. A number of applications have been developed for using CCBs, including the use of fly ash as a substitute for cement in concrete and grout applications, the use of fly as in flowable and structural fill, the use of calcium sulfate rich FGD scrubber sludge as a replacement of natural gypsum in wallboard manufacturing, and a variety of mine reclamation applications. The utilization of CCBs reduces the consumption of natural resources,
151
reduces emissions of greenhouse gases to the atmosphere, and reduces the need for new landfill construction. Potential barriers to CCB use include the low cost of landfilling, the lack of available large-volume or high value applications, and variations in CCB material properties.
B.10 Acknowledgment
The authors would like to thank the Ohio Coal Development Office (OCDO) for its support of much of the research cited herein. We would also like to thank the reviewers for their helpful comments.
152
6 Material Metric Tons × 10 Reference Total MSW 119.6 (90) CCBs 69.4 (88) Paper 37.4 (90) Plastic 15.5 (90) Wood 8.4 (90) Glass 6.9 (90)
Table B.1. Amount of CCBs Disposed of in Landfills in the United States in 1998 Compared to Disposal of Municipal Solid Waste (MSW)∗.
∗ Data for disposal of MSW are for 1997, the most current year for which data are available. 153
6 Commodity Metric Tons × 10 Reference Crushed Stone 1,500e (91) Sand & Gravel 1,020e (91) CCBs 97.7 (88) Cement 85.5e (91) Iron Ore 62e (91)
Table B.2. Amount of CCBs Produced in the United States in 1998 Compared to Traditional Non-Fuel Mineral Commodities.
e Estimated 154
FGD Material Bottom Ash Physical Characteristics Fly Ash / Boiler Slag Wet Dry
Particle Size (mm) 0.001-0.1 0.1-10.0 0.001-0.05 0.002-0.074
Compressibility (%) 1.8 1.4
Dry Density (lb/ft3) 40-90 40-100 56-106 64-87
Permeability (cm/sec) 10-6-10-4 10-3-10-1 10-6-10-4 10-7-10-6 Shear Strength
Cohesion (psi) 0-170 0 Angle of Internal Friction (degree) 24-45 24-45 Unconfined Compressive Strength 0-1600 41-2,250 (psi)
Table B.3. Summary of Physical Characteristics and Engineering Properties of Fly Ash, Bottom Ash, Boiler Slag and FGD Material (4,44,45,56,96,97,98,99).
155
Element Fly Ash Bottom Ash/Boiler Slag Dry FGD Material (mg/kg) Mechanical ESP/Baghouse Range Median Range Median Range Median Range Median Arsenic 3.3-160 25.2 2.3-279 56.7 0.50-168 4.45 44.1-186 86.5 Boron 205-714 258 10-1300 371 41.9-513 161 145-418 318 Barium 52-1152 872 110-5400 991 300-5789 1600 100-300 235 Cadmium 0.40-14.3 4.27 0.10-18.0 1.60 0.1-4.7 0.86 1.7-4.9 2.9 Cobalt 6.22-76.9 48.3 4.90-79.0 35.9 7.1-60.4 24 8.9-45.6 26.7 Chromium 83.3-305 172 3.6-437 136 3.4-350 120 16.9-76.6 43.2 Copper 42.0-326 130 33.0-349 116 3.7-250 68.1 30.8-251 80.8 156 Fluorine 2.50-83.3 41.8 0.4-320 29.0 2.5-104 50.0 ------Mercury 0.008-3.0 0.073 0.005-2.5 0.10 0.005-4.2 0.023 ------Manganese 123-430 191 24.5-750 250 56.7-769 297 127-207 167 Lead 5.2-101 13.0 3.10-252 66.5 0.4-90.6 7.1 11.3-59.2 36.9 Selenium 0.13-11.8 5.52 0.6-19.0 9.97 0.08-14 0.601 3.6-15.2 10.0 Silver 0.08-4.0 0.70 0.04-8.0 0.501 0.1-0.51 0.20 ------Strontium 396-2430 931 30-3855 775 170-1800 800 308-565 432 Vanadium 100-377 251 11.9-570 248 12.0-377 141 ------Zinc 56.7-215 155 14-2300 210 4.0-798 99.6 108-208 141
Table B.4. Trace Elemental Composition of Fly Ash, Bottom Ash, Boiler Slag (56) and FGD Material (101).
Chemical Constituent FGD Ash (mg/L) pH 9.58-12.01 --- 11,840- TDS --- 13,790 Ag <0.024 0.0-0.05 Al 0.12-0.20 --- As <0.005 0.026-0.4 B 0.543-2.17 0.5-92 Ba <0.002 0.30-2.0 Be 0.141-0.348 <0.0001-0.015 Ca 1,380-3,860 --- Cd <0.003 0.0-0.3 Co <0.014-0.026 0.0-0.22 Cr <0.005-0.028 0.023-1.4 Cu <0.013 0.0-0.43 Fe <0.029 0.0-10.0 Hg <0.0002 0.0-0.003 K 1.3-22.1 --- Li 0.04-0.18 --- Mg <0.04-1,360 --- Mn <0.001 0.0-1.9 Mo 0.025-0.088 0.19-0.23 Na 1.32-9.82 --- Ni <0.01 0.0-0.12 P <0.12 --- Pb <0.001-0.017 0.0-0.15 S 132-979 --- Sb <0.24 0.03-0.28 Se <0.001-0.005 0.011-0.869 Si 0.10-0.33 --- Sr 0.83-3.38 --- V <0.019-0.024 --- Zn <0.006 0.045-3.21 Cl- 19.6-67.8 --- 2- SO3 <1.0-43.2 --- 2- SO4 236-2,800 ------Table B.5. Range of Values Observed for TCLP Analysis of Dry FGD Materials (44,101) and Ash (97).
157
Reduction in Emission or Natural Resource Utilization CCB Utilization per Year Current Utilization Rate 100% CCB Utilization9
Cement/Concrete/Grout 10×106 tonscement 32×106 tonscement × 6 10 × 6 10 10 tons CO 2 32 10 tons CO 2
Flowable Fill × 6 × 6 0.36 10 tons CO 2 1.08 10 tons CO 2
Structural Fill 3.7 ×106 tons Soil 11 11.2×106 tons Soil
Wallboard Gypsum 1.6×106 tons gypsum 16.5×106 tons gypsum
Mining Applications 0.6×106 tons clay 2.2×106 tons clay
Total CCB Reduction in Landfills 17 ×106 m 3 59×106 m 3
Table B.6. Major CCB Applications and Environmental Benefits of CCB Use.
9 The environmental benefit for 100% CCB utilization is calculated assuming the percent of material used for a particular application (e.g., cement/concrete/grout) is independent of overall CCB utilization. 10 Assumes 1 lb CO2/1 lb cement not used, and 1 lb of cement replaced/1 lb fly ash used. 11 Assumes 1 lb Soil/1 lb of fly ash used. 158
120
97.7 100
80
57.1 60
40 Million Metric Tons 22.7 20 15.1 2.7 0 Fly Ash Bottom Boiler FGD Total Ash Slag CCPs
Figure B.1. CCBs Production in million metric tons in the United States in 1998 (88).
159
Dry PC Boiler
Wet PC Boiler
Spreader Stoker
Other Stoker
Cyclone
0 20406080100 % Ash Proportion Fly Ash Bottom Ash/Boiler Slag
Figure B.2. Approximate Ash Distribution as a Function of Boiler Technology
(4,56,103).
160
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