BARIUM SOLIDIFICATION/STABILIZATION OF LEGACY FLY ASH
The Graduate Faculty of The University of Akron
In Partial Fulfillment
Of the Requirements for the Degree
Master of Science
Ariya Reza Fathi
BARIUM SOLIDIFICATION/STABILIZATION OF LEGACY FLY ASH
Ariya Reza Fathi
Advisor Dean of the College Dr. Stephen Duirk Dr. Donald P. Visco Jr.
Faculty Reader Dean of the Graduate School Dr. Christopher Miller Dr. Chand Midha
Faculty Reader Date Dr. David Roke
Department Chair Dr. Wieslaw Binienda
A legacy fly ash pile in Ashtabula County has 20,000 – 30,000 cubic yards of
material and contains barium and other various heavy metals. The concentrations of
barium in the fly ash are below Regional Screening Levels (RSLs) for soil, but if the
barium leached out it would pose a threat to human health. Solidification/Stabilization
(S/S) of the ash was investigated with concrete and sulfate. After physical/chemical
characterization, the fly ash was identified as class F fly ash, meaning it has no self- stabilizing/cementing characteristics, and had barium concentrations ranging from 0-
1,500 mg/kg. Fly ash was used as a replacement for either cement or fine aggregate at
10-50%. Cement replacement did not achieve a compressive strength above 3,000 psi when using a 6,500 psi concrete mixture when cement to waste ratios exceeded 20% replacement. When fine aggregate replacement was investigated, fine aggregate to waste ratios up to 40% was effective at achieving above 3,000 psi concrete. X-ray diffraction
(XRD) revealed the barium present was already in the insoluble barium sulfate form.
Therefore, the Toxicity Characteristic Leaching Procedure (TCLP) and accelerated leaching procedure were used to access the potential for barium to leach form the ash and from the concrete mixtures. Barium concentrations never exceeded the U.S. EPA drinking water maximum contaminated level (MCL) for 2 mg/L. Concrete made with ash spiked with 1,500 mg/kg had increased barium leached but still below the primary drinking standard with the highest concentration being 1.79 ± 0.44 mg/L. Therefore, the ash can be handled as solid waste if no beneficial use can be identified.
I would like to express my sincere gratitude to my advisor, Dr. Stephen Duirk, for his support, direction, and instruction throughout my master’s studies. Thank you to my committee for their support and advice: Dr. Stephen Duirk, Dr. David Roke, and Dr.
Christopher Miller. Furthermore, thank you to all my associates, and friends, for all of the collaboration and support during this research: Elizabeth Crafton, Nana Ackerson,
Deepak Aryal, and George Carleton. Finally, my deepest appreciation is felt for my parents, Alireza and Cecelia Fathi, and my sister, Sussian Quackenbush for all their support and assistance.
TABLE OF CONTENTS
ABSTRACT ...... iii
INTRODUCTION ...... 1
1.1 Background ...... 1
1.2 Problem Statement ...... 3
1.3 Specific Objectives ...... 4
LITERATURE REVIEW ...... 6
2.1 Background and Regulations ...... 6
2.2 Exposure to Heavy Metals and Health Effects ...... 8
2.3 Fly Ash Classification and Composition ...... 10
2.4 Barium Chemistry and Health Effects...... 11
2.5 Concrete Overview ...... 13
2.6 Solidification/Stabilization of Heavy Metals ...... 14
3.1 Chemicals and Basic Laboratory Equipment ...... 18
3.2 Fly Ash Characterization ...... 19
3.2.1 Fly Ash Homogenization ...... 19
3.2.2 Sieve Analysis ...... 20
3.2.3 X-Ray Diffraction (XRD) Analysis ...... 22
3.2.4 Characterization of Heavy Metals in the Fly Ash ...... 24
3.3 Experimental Methods ...... 24
3.3.1 Mortar Strength ...... 25
3.3.2 Concrete Compressive Strength ...... 27
3.3.3 Toxicity Characteristic Leaching Procedure (TCLP) ...... 29
3.3.4 Accelerated Leach Test ...... 30
3.3.5 Leached Barium with Sulfate Additions ...... 31
RESULTS AND DISCUSSION ...... 33
4.1 Introduction ...... 33
4.2 Mortar compressive strength with effects of ash ...... 33
4.3 Concrete compressive strength with ash replacement ...... 37
4.4 Leaching Tests results ...... 40
4.4.1 Acid Digestion ...... 40
4.4.2 TCLP results ...... 42
4.4.3 Accelerated Leaching Test ...... 54
4.5 Sulfate Stabilization ...... 60
CHAPTER V ...... 64
CONCLUSIONS AND RECOMMENDATION ...... 64
5.1 Introduction ...... 64
5.2 Conclusions ...... 65
5.3 Recommendations ...... 66
BIBLIOGRAPHY ...... 68
APPENDIX A ...... 72
A.1 Turbidimetric Barium Analysis ...... 72
A1.1 TCLP Turbidimetric Barium Results ...... 72
A.1.2 Accelerated Leaching Test Turbidimetric Barium Results...... 73
A.2 Compressive Strength ...... 77
ii A.2.1 Concrete Compressive Strength ...... 77
LIST OF TABLES Table Page
3.2: Mortar composition of cement replacement (C/W) by weight ...... 26
3.3: Concrete Composition - Fine Aggregate Replacement (S/W) ...... 27
3.4: Concrete Composition - Cement Replacement (C/W)...... 28
A.1 cement replacement (C/W) compressive strength ...... 77
A.2 fine aggregate replacement (S/W) compressive strength ...... 77
LIST OF FIGURES
3.1: Sieve analysis of fly ash from the Cristal processing plant in Ashtabula, OH...... 21
3.2: Sieve analysis of the same fly ash performed by Geotechnical Testing Services Inc 22
3.3: XRD results for barian celstine and celestine from the Cristal fly ash...... 23
Figure3.4: XRD results of calcite in the Cristal fly ash...... 23
4.1: Compressive strength of mortar cubes with cement replacement (C/W) 10– 50% on day 3 and 7 compared to the control group...... 35
4.2: Compressive strength of mortar cubes with fine aggregate replacement (S/W) 10– 40% on day 3 and 7 compared to the control group...... 36
4.3: Compressive strength of concrete cylinders with cement replacement (C/W) 10–50% ...... 38
4.4: Compressive strength of concrete cylinders with fine aggregate replacement (S/W) 10-50% ...... 40
4.5: Barium (Ba), lead (Pb), and aluminum (Al) concentration of ash from acid digestion...... 42
4.6: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from C/W 10% ...... 44
4.7: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from C/W 20% ...... 45
4.8: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from S/W 10% ...... 46
4.9: TCLP concentration of leached aluminum (Al), and barium (Ba) from S/W 20% ... 47
4.10: TCLP concentration of leached aluminum (Al), and barium (Ba) from S/W 30% . 48
4.11: TCLP concentration of leached aluminum (Al), and barium (Ba) from S/W 40% . 49
4.12: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from spiked S/W 10% ...... 50
4.13: TCLP concentration of leached aluminum (Al), and barium (Ba) from spiked S/W 20% ...... 51
4.14: TCLP concentration of leached aluminum (Al), and barium (Ba) from spiked S/W 30% ...... 52
4.15: TCLP concentration of leached aluminum (Al), and barium (Ba) from spiked S/W 40% ...... 53
4.16: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 10% ...... 55
4.17: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 20...... 56
4.18: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 30...... 58
4.19: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 40...... 59
4.20: Leached barium concentration from ash, with a blank (BL) with no sodium sulfate, 1 mole (1 mol) of sodium sulfate, 5 moles (5 mol) of sodium sulfate, and 11 moles (11 mol) of sodium sulfate from 0.25 days to 10 days...... 61
4.21: Sulfate concentration from day 1 and day 11 of accelerated leaching test. 62
A.1: Barium concentration C/W 10-20% and S/W 10-40% from TCLP analyzed with the turbidimetric method...... 72
A.2: Barium concentration S/W 10% from accelerated leaching test analyzed with the turbidimetric method...... 73
A.3: Barium concentration S/W 20% from accelerated leaching test analyzed with the turbidimetric method...... 74
ii A.4: Barium concentration S/W 30% from accelerated leaching test analyzed with the turbidimetric method...... 75
A.5: Barium concentration S/W 40% from accelerated leaching test analyzed with the turbidimetric method...... 76
Anthropogenic activities such as industrialization, natural resource extraction/utilization, and agriculture have resulted in both organic and inorganic environmental pollution. Heavy metals are inorganic elements generally defined as metals with relatively high densities, atomic weights, or atomic numbers. Heavy metals are moderately scarce in the Earth's crust but are present in many aspects of modern life.
However, human exposure to heavy metals at greater then naturally occurring concentrations can pose a serious threat to human health. People are mainly exposed to heavy metals through ingestion from drinking water, food, or contact with contaminated soil (Martin et al., 2009). Mining and industrial wastes can contain high concentrations of heavy metals, and in some cases, have been abandoned without treatment due to the lack of environmental regulations. In 1980, congress passed the Bevill amendments to the Resource Conservation and Recovery Act (RCRA), which can exclude special wastes as hazardous until the waste is officially analyzed by a state regulator(USEPA, 2017c).
This has resulted in waste from mines and ash piles being left in place instead of treated or disposed. RCRA has also identified 8 heavy metals (i.e., arsenic, barium, cadmium, chromium, lead, mercury, selenium, and silver) and set maximum allowable concentrations in soil and water, in order to minimize potential health effects due to human exposure (Linak et al., 1994). However, some legacy waste streams are stuck “in a limbo” recognized as hazardous but no remedial action is required as long as the waste pile is not disturbed. The main hypothesis proposed research is that hazardous waste can be modified through chemical stabilization, solidification or both, into a sustainable value-added engineered product for both industrial and residential uses.
Leaching of heavy metals into groundwater is the major concern of managing heavy metal laden waste streams. Heavy metals mixtures elicit confound toxicological responses even at concentrations below individual maximum contaminant concentrations.
Exposure to high concentrations of metal mixtures can cause higher blood pressure, paralysis and an irregular heartbeat (Martin et al., 2009). To prevent metal leaching, chemical stabilization chemically alters the metal laden waste streams by reducing the solubility of heavy metals(Y. M. Wang et al., 2001). Therefore, stabilized waste streams could possibly be disposed of at municipal waste landfills or used for other beneficial uses. Physical solidification/chemical stabilization utilizes a chemical binder that encapsulates the heavy metals, which also reduces the solubility of the metals. Both processes have been used to turn waste into value-added products.
A legacy ash pile located in Ashtabula county will be the main focus of this research. It is located on industrial property in between two buildings and is estimated to be between 20,000 – 30,000 cubic yards of material laden with barium, strontium, and
2 other heavy metals. The material was deposited there over 60 years ago, and no action
has been taken to remediate the waste. Barium concentrations in the ash pile range from
non-detect to over 1500 mg/kg, which is below the U.S. Environmental Protection
Agency (EPA) Regional Screening Level (RSL). Other heavy metals were found (i.e.,
cobalt, copper, lead, nickel and zinc), however they are below the RSL for either
industrial or residential purposes. However, all the heavy metals are leaching out of this
exposed ash pile and contaminating the regional aquifer, and a nearby stream. The
maximum allowable concentrations of these metals are significantly lower in drinking
water sources("Secondary Drinking Water Standards: Guidance for Nuisance
Chemicals," 2016). The land owner would like to remove the ash pile to expand current
production capacity. Therefore, remediation of the ash pile would benefit the current
owner and protect the community’s groundwater supply.
1.2 Problem Statement
Anthropogenic wastes containing heavy metals are a serious threat to human
health in surrounding communities. Legacy wastes containing heavy metals are in a
limbo where the waste is sent to a hazardous waste landfill with the potential for heavy
metal leaching exists, or the waste is left in place to continue leaching. Specifically an
ash pile containing barium will be investigated for beneficial use with solidification with
cementitious material, and stabilization through precipitation of barium sulfate (Vaidya et al., 2010).
3 1.3 Specific Objectives
During the research, these specific objectives were considered:
1 Identifying the physical/chemical characteristics of the ash. The physical
characteristics will be measured to determine the materials structural
properties. A sieve analysis will be performed to obtain information on the
gradation of the ash, and x-ray diffraction (XRD) will be used to observe the
crystalline structure of the ash. The physical/chemical characterization will
then dictate which stabilization technique will be the most suitable for the raw
2 Investigating the use of cementitious binders to increase the structural
properties to perform solidification. Selecting a binder that can increase the
unconfined strength is necessary to effectively use the ash for capping waste
sites and enables it for varying uses. Binders such as Portland cement, lime,
slag cement and bentonite can be combined with the ash and cured in order to
achieve solidification. Cementitious materials have an advantage with wet
materials since water is needed for the chemical reaction to occur (Singh et
al., 2006). This may be advantageous since ash samples provided have been
exposed to the elements for multiple decades
3 Chemical stabilization will be investigated to prevent the heavy metals
leaching from the material through the addition of sulfate, or through the
sulfate originally present in the ash. The objective of chemically stabilizing
the waste is to make the heavy metals in the waste less soluble creating stabile
insoluble chemical species through complexation reactions (Crannell et al.,
4 2000). The formation of insoluble complexes will be promoted through the addition of an inexpensive sulfate salt.
2.1 Background and Regulations
Solid and hazardous waste generation has been a part of industrial and agricultural
anthropogenic activities prior to the recent introduction of more sustainable practices.
The Solid Waste Disposal Act (SWDA) was introduced in 1965, to enable states to manage waste, from numerous sources. SWDA created a set of safety requirements that landfills need to meet to ensure the protection of the local community (SCDHEC, 2008).
The issue was that a considerable amount of the waste was still not properly regulated.
At the time SWDA was created, the chemical industry was producing more than four million chemicals that elicited aquatic and humanological toxic responses. In 1960 88 million tons of solid waste was generated a year, by 2009 it had increased to 243 million tons per year (USEPA, 2014). The increase of waste required further management to properly handle it. In 1970, Congress created the Environmental Protection Agency
(EPA) which amended the outdated SWDA and established the Resource Conservation and Recovery Act (RCRA) in 1976. RCRA had three goals when it was created: 1) ensure that waste is managed in a manner that protects human health and the environment, 2) reduce or eliminate, as expeditiously as possible, the amount of waste
6 generated, including hazardous waste, and 3) conserve energy and natural resources
through waste recycling and recovery (USEPA, 2002, p. 2)
Congress noted at the time of establishing RCRA that 30-35 million tons of hazardous waste was disposed of in an unregulated manner each year (USEPA, 2002).
Unregulated disposal of hazardous waste could lead to unwanted ecological and human toxicity.
RCRA regulated eight heavy metals. A heavy metal is any relatively dense metal or metalloid compared to water that is noted for its potential toxicity, especially in human or environmental context. These metals are: arsenic (As), barium (Ba), cadmium (Cd), chromium (Cr), lead (Pb), mercury (Hg), selenium (Se), and silver (Ag). Heavy metals have been used in numerous industries that have increased the exposure of these RCRA metals. Arsenic has been used in semi-conductors, barium in spark plugs, cadmium in cigarettes, chromium with electroplating, lead in batteries, mercury in thermometers, selenium in glass, and silver used in jewelry (Martin et al., 2009). These metals can affect the cellular organelles in humans damaging cellular function in the body
(Tchounwou et al., 2012).
In order to establish if a solid waste is hazardous, a procedure must first be established that can simulate the potential leaching of heavy metals from the waste under prolonged environmental conditions. RCRA established the Toxicity Characteristic
Leaching Procedure (TCLP) to observe the metals leaching from solidified wastes under accelerated environmental conditions. The regulatory maximum concentrations for
RCRA metals is considerably different from the national primary drinking standards, which are approximately 100 times less in concentration. RCRA sets the allowable
7 leaching concentrations of arsenic at 5 mg/L, barium at 100 mg/L, cadmium at 1 mg/L,
chromium at 5 mg/L, lead at 5 mg/L, mercury at 0.2 mg/L, selenium at 1 mg/L, and silver
at 5 mg/L (USEPA, 2005). The primary drinking standards maximum contaminant limits are stricter in most cases, with arsenic at 0.01 mg/L, barium at 2 mg/L, cadmium at 0.005 mg/L, chromium at 0.1 mg/L, lead at 0.015 mg/L, mercury at 0.002 mg/L, and selenium
at 0.05 mg/L (USEPA, 2017a). RCRA has a high maximum allowable concentration of contaminants of concern (COCs) since the material is a waste product. If placed in a solid waste landfill, the waste will leach into leachate collection systems and be sent for treatment. Regional Screening Levels (RSLs) also set risk base criteria establishing maximum allowable concentrations of COCs for residential and industrial soil concentrations. For the residential and industrial soil limits, arsenic is 0.68 and 3.0 mg/kg respectively, barium has 15,000 and 220,000 mg/kg, cadmium has 71 and 982
mg/kg, chromium (III) has 120,000 and 1,800,000 mg/kg, chromium (VI) 0.3 and 6.3
mg/kg, lead has 400 and 800 mg/kg, mercury has 11 and 46 mg/kg, selenium has 390 and
5800 mg/kg, and silver has 390 and 5800 mg/kg (USEPA, 2017b).
2.2 Exposure to Heavy Metals and Health Effects
Heavy metals are a difficult group to classify as there is no agreed upon
definition. In most cases it is accepted that metals with a specific density of 5 g/cm3 or
higher, but there are metals that fall below that specific density and are still classified as
heavy metals. Barium for example is considered a heavy metal, but only has a specific
density of 3.62 g/cm3("Barium," 2017). Lead, cadmium, mercury, and arsenic are
considered carcinogenic heavy metals that pose the largest threat to human health. These
8 metals were used in a variety of products from batteries, metal coatings, light bulbs, and
cigarettes. Heavy metals have been used in multiple industries for thousands of years,
but from the middle of the 19th century for over a hundred years the production and use of metals increased drastically. Pollution related to heavy metals was reduced in the 1990s in developed nations as many wastes were regulated and managed for the first time.
Humans are exposed to heavy metals in multiple pathways, either through inhaling from
burning material, ingestion of plants that have bioaccumulated the metals, drinking water
from runoff from heavy metal sites (Järup, 2003).
Barium is not known to cause cancer in humans, though it does have some
adverse effects. Short term exposure to barium can lead to vomiting, diarrhea, problems
breathing, change in blood pressure and muscle weakness (Ramanathan, 2007). Long
term or bulk amounts of barium exposure cause heart rhythm changes, paralysis, and
death. Barium is plentiful in the earth’s crust and is used for a multitude of purposes.
Barium sulfate is used in the medical field for diagnostics (Martin et al., 2009). Arsenic
is a carcinogenic heavy metal that is used in wood preserving, paints, and semi- conductors. In small concentrations it can cause vomiting, decreased production of blood cells, and nausea. Long term/high exposure can lead to death, darkening of skin, and warts (Council, 1999). Cadmium is another carcinogen metal used in electroplating, cigarettes, and is naturally occurring in soil and coal. Consuming large quantities of cadmium can cause vomiting and diarrhea. Long term exposure to cadmium will cause damage to kidneys, weakened bones, and lung damage (Järup et al., 1998). Lead is a well-known carcinogenic heavy metal that has been used in many industries in the past.
Lead is released into the air from coal mining and was used in paints and water pipes.
9 High concentrations can have varying effects on different genders, causing miscarriages
with woman and damage to organs that produce sperm for men (Jaishankar et al., 2014).
Finally, mercury is a metal that is also carcinogenic to humans and is released from coal
burning into the air. It has been used in thermometers, light bulbs, and batteries in past.
Mercury can cause damage to the nervous system at even miniscule doses. At high concentrations exposure can cause permanent brain damage that can result in memory issues, change in personality, damage to hearing or vision, and birth complications
(Martin et al., 2009).
2.3 Fly Ash Classification and Composition
Fly ash properties are depending mainly on the source of the coal, the
combustions mechanism (i.e., how the coal was burned), and how the ash is collected.
Fly ash mainly contains fused inorganic material generated by burning coal to produce
steam at electrical generation power plants. Two main types of ash classified from
burning coal are class F and C. Class F is from bituminous coals that have a low
concentration of calcium compounds, usually less than 10%. Class C ash from
subbituminous coal contains a higher concentrations of calcium compounds, regularly
greater than 20%. Fly ash in general contains 20-40% SiO2, 10-30% Al2O3, 3-10%
Fe2O3, 0.5-8% MgO, 1-8% SO3, 0.5-2% TiO2, 0.5-2% C, 0.33-3% H2O, 10-35% CaO,
0.5-4% K2O, and 0.5-6% Na2O. Class C ash can be considered self-cementing and
doesn’t require a binder to solidify the material (Mackiewicz et al., 2005). Class F is
typically used in testing and studies for practical use with cement binders, while class C
is focused with self-cementing (Shehata et al., 2000). Fly ash strengthens concrete by
10 reacting with sulfate compounds present in cement and excess lime. Class C already has excess calcium compounds and can self-stabilize unlike class F fly ash (Styron, 2001).
Currently only a small percentage of fly ash is used beneficially, a majority of the ash goes to landfills. A study observing the effects of fine aggregate replacement with class F fly ash tested found that replacing 10-50% of fine aggregate material showed a positive effect on concrete. The study found that the compressive strength, splitting tensile strength, flexural strength, and the modulus of elasticity increased when fine aggregate was replaced at any percentage with fly ash (Siddique, 2003). The maximum compressive strength was achieved with a batch of concrete with 50% of the fine aggregate replaced with class F fly ash. Cement replacement with fly ash was also studied and found that delayed early strength was the result. The study concluded that the class F fly ash could be used in structural concrete at larger percentages than typically used in concrete (Siddique, 2003).
2.4 Barium Chemistry and Health Effects
Barium is an alkali earth metal with an atomic mass of 137.327 g/mole. It has a melting point of 727˚C, and a boiling point of 1845˚ C ("Barium," 2017). Barium salts are soluble in water, with barium chloride having a high solubility. Some compounds have very low solubility such as, barium sulfate, carbonate, and sulfide being the least soluble compounds. Barium chloride has a solubility of 375 g/L at 26˚C, compared to barium sulfate which has a low solubility of 1.6 mg/L at 20˚C. This difference in solubility is the basis of stabilizing barium. By introducing sodium sulfate (Na2SO4) to
11 the barium chloride it is possible to greatly reduce the mobility of barium and make it
nearly insoluble. Equation 1 shows the resulting chemical reaction (Kucher et al., 2006).
2 + 2 4 4 + 2 (1)
𝐵𝐵𝐵𝐵𝐵𝐵𝐵𝐵 𝑁𝑁𝑁𝑁 𝑆𝑆𝑆𝑆 → 𝐵𝐵𝐵𝐵𝐵𝐵𝐵𝐵 𝑁𝑁𝑁𝑁𝑁𝑁𝑁𝑁 Barium is most commonly found in the earth’s crust as barium sulfate. Barium sulfate is
used currently with the oil and gas industry to create lubricant muds for drilling
operations. Barium is also used in making dyes, bricks, glass, and rubber. Barium
sulfate is used regularly in medical diagnostic radiology with no toxic effects. Many
public water systems contain concentrations of barium ranging from 1-172 g/L. In the
United States a mean concentration of 43 g/L has been observed in cites, and a maximum
concentration of 380 g/L (Ramanathan, 2007).
When ingested, barium is distributed over 93% of the body, mainly in the bones and teeth. Autopsies have shown that human bones contain nearly 7 ppm of barium in children, and 8.5 ppm in adults, showing no accumulation. Mainly animal studies have been used to observe barium exposure (Kravchenko et al., 2014). After ingestion 72% of barium is excreted through feces. Since barium is an alkaline earth metal, it is considered a bone seeking element, but provides no known purpose in bones. Barium is not considered to be a cancer-causing element but can cause cardiovascular and kidney problems. When rats and mice were exposed to 5 ppm of barium for their lifetimes, an autopsy showed no signs of tumors in any major organ (Ramanathan, 2007).
12 2.5 Concrete Overview
Concrete is a material that is made with cement which acts as a binder to a filler
that is both fine and coarse aggregates. Cement is made from burning limestone and clay
together at temperatures between 1400-1600˚C. When water is mixed with cement a chemical process called hydration begins that hardens the cement with the aggregates to create concrete. Mortar is similar to concrete but does not contain a coarse aggregate.
During hydration compounds in the cement form hydrates with water that will eventually
increase the compressive strength of the material (Jeff Thomas, 2014).
The main factors in making concrete is the water to cement ratio, workability, and
the expected compressive strength. Selecting a proper water to cement ratio is key to
producing a concrete that has optimal strength and workability. Too much water and the
concrete will lose compressive strength, but too little water and placing the concrete
correctly will be difficult to mold, and the workability will be poor. The composition of
ordinary Portland cement (OPC) used commonly in concrete is 50% tricalcium silicate,
25% dicalcium silicate, 10% tricalcium aluminate, 10% tetracalcium aluminoferrite, and
5% gypsum (Beth Chamberlain, 1995). Aggregates are used to lower the cost of concrete
by reducing the amount of cement required. By adding a chemically inert aggregate it is
possible to reduce the amount of the most expensive component, cement. The aggregate
is a mix of fine sand and coarse stones.
After introducing water to the cement and aggregate, the cement begins to
hydrate and form a paste. As the cement hydrates, it hardens and gains compressive
strength. If there is not enough water the hydration process will be incomplete, and the
concrete will be weaker than that of concrete with excess water. Ensuring that all of the
13 tricalcium silicate hydrates is important to attaining the full value of the material. The compounds that provide the strength to concrete are the tricalcium silicate and dicalcium silicate. The tricalcium silicate mainly provides the early strength for the concrete while curing. Equation 2 shows the reaction of tricalcium silicate and water to produce a calcium silicate hydrate (C-S-H).
2 3 5 + 7 2 2 2 + 3 ( )2 + 173.6 (2)
𝐶𝐶𝐶𝐶 𝑆𝑆𝑆𝑆𝑂𝑂 𝐻𝐻 𝑂𝑂 → 3𝐶𝐶𝐶𝐶𝐶𝐶 ∙ 2𝑆𝑆𝑆𝑆𝑆𝑆 ∙ 4𝐻𝐻 𝑂𝑂 𝐶𝐶𝐶𝐶 𝑂𝑂𝑂𝑂 𝑘𝑘𝑘𝑘 The dicalcium silicate reacts slower than the tricalcium silicate, and it provides an increase in strength over a longer period. The dicalcium silicate may react for years if the conditions are right. Equation 3 shows the reaction of dicalcium silicate with water (Beth
2 2 4 + 5 2 2 2 + ( )2 + 58.6 (3)
𝐶𝐶𝐶𝐶 𝑆𝑆𝑆𝑆𝑆𝑆 𝐻𝐻 𝑂𝑂 → 3𝐶𝐶𝐶𝐶𝐶𝐶 ∙ 2𝑆𝑆𝑆𝑆𝑆𝑆 ∙ 4𝐻𝐻 𝑂𝑂 𝐶𝐶𝐶𝐶 𝑂𝑂𝑂𝑂 𝑘𝑘𝑘𝑘 The strength of the concrete is dependent on equations 2 and 3, and the amount of water available during the reaction. After seven days the concrete is assumed to have reached at least 70% of its expected fully cured strength. After 28 days the concrete is expected to be at its relative full strength. During the curing process the concrete should be kept wet to ensure that the tricalcium silicate fully reacts and that the concrete does not dry
(Jeff Thomas, 2014).
2.6 Solidification/Stabilization of Heavy Metals
Solidification/Stabilization (S/S) is commonly used for remediating heavy metals in soils, due mainly to the short time and low cost. S/S is considered a Best
14 Demonstrated Available Technology (BDAT) for RCRA wastes. By using readily
available materials such as cement, lime, and fly ash the cost of remediation can be
lowered. Solidification encapsulates waste to a soil/solidified material to keep harmful
materials from being transported into the aqueous phase. Stabilization, chemically reacts
with a compound, altering it to a form that is unlikely to leach from solidified material.
The objective using S/S is to prevent pollutants from leaching out of the solidified material and decrease the bioavailability of the pollutant (USEPA, 2012).
S/S has been used on multiple wastes containing heavy metals with variable results, that depend on the selected binder, heavy metals present, and the conditions of the material. Lead S/S has proven to achievable using OPC. With just OPC it was possible to reduce lead concentrations, though the compressive strength of the material was reduced, and the setting time was extended. The lead was observed to be physically encapsulated in the C-S-H gel when using OPC as a binder (Y.-S. Wang et al., 2018). In general, heavy metal waste needs S/S even before it can be disposed of in a landfill.
Beneficially reusing the material provides the opportunity to avoid the use of landfills and generate an economic net gain. Cement has been proven to be the most adaptable binder currently available and reduce the leaching of multiple heavy metals at once. The
C-S-H gel is crucial to encapsulating the heavy metals with solidification. Heavy metals accelerate the hydration of the tricalcium silicate in cement, speeding up increase of early strength the concrete gains. When C-S-H gel binds to metals, the carbonation of tricalcium silicate increase the capacity of bonding to heavy metal cations, entrapping more metals in the process (Chen et al., 2009).
15 Hazardous barium wastes have been treated in situ S/S and found satisfactory results using sodium sulfate to stabilize the barium with a cement binder. Besides stabilizing the barium in the waste, the sodium sulfate appeared to increase the compressive strength of concrete compared to the control group. The barium waste was mixed only with cement without the addition of sulfate and the compressive strength of the concrete decreased compared to the mixture with sulfate. In tests with no sulfate present barium leached out at in excess when cement was used as a binder. When sulfate was introduced with 1.6 times the stoichiometric amount required, the barium was effectively stabilized with the cement binder (Vaidya et al., 2010).
Other options for stabilization is using self-cementing ash that has higher concentration of calcium compounds. Class C fly ash does not require the addition of a binder such as cement or limestone. By hydrating the fly ash and immediately molding and compacting the fly ash it is possible to stabilize the material. In order to achieve this, it is necessary to know the optimum moisture content. The maximum strength was achieved with 1-7% below the optimum moisture content, and then not delaying compaction. Once the water is added to the ash the hydration begins to form bonds with soil particles. Once compaction starts later the bonds are broken and the compaction is breaking the cementitious bonds, effectively weakening the solidification of the material.
Limiting the delay of compaction is key to stabilizing the fly ash (Mackiewicz et al.,
2005). The issue with self-cementing fly ash is the initial conditions of the fly ash containing heavy metals is not a controllable factor. Most fly ashes are class F and show little potential for self-cementing.
16 Besides adding cement as a binder to heavy metal waste, reagent grade stabilizers can be used. In one study Ca(H2PO4) and calcium carbonate (CaCO3) were used to stabilize a soil contaminated with Cd, Cu, Ni, Pb, and Zn. The TCLP concentrations of metals were reduced more than 87% in lab studies. The field concentrations decreased greatly after 30 days, with a reduction of 95% of Cd, Cu, Pb, and Zn. By using both a mix of the two reagents it was capable to immobilize the heavy metals in the waste without using a cement binder (Y. M. Wang et al., 2001).
MATERIALS AND METHODS
3.1 Chemicals and Basic Laboratory Equipment
Acetic acid was purchased from Riedel-de Haën (Hanover, Germany). High-
density polyethylene (HDPE) wide mouth bottles in volumes of 100, 250, and 500 ml and the non-sterile syringe filters (25mm) were purchased from Fisher Scientific (NJ, USA).
The BariVer 4 Barium Reagent Powder Pillow and SulfaVer 4 Sulfate Reagent Powder
Pillow were purchased from HACH (CO, USA). Sodium sulfate and barium chloride
were purchased from Fisher Scientific (Pittsburgh, PA, USA) The bio-cylinder concrete
molds were purchased from Deslauriers (IL,USA). A 2 cubic ft. portable electric cement
mixer made by Klutch, was used to mis the concrete (Northern Tool and Equipment,
Burnsville, MN, USA). Deionized (DI) water was used in a multitude of experiments
and was produced with a Barnstean ROPure Infinity/NANOPure system (Barnstead-
Thermolyne Corp. Dubuque, IA, USA). A Burrell wrist-action shaker was used to shake
solutions (Pittsburgh, PA, USA). The experimental pH was measured with an Orion 5
Star pH Meter equipped with Ross ultra-combination electrode (Thermo Fisher
Scientific, Pittsburgh, PA, USA). The pH calibration was performed with Orion application solution, and Orion reference electrode filling solution (Thermo Fisher
Scientific, Pittsburgh, Pa, USA). An Agilent Technologies 700 Series Inductively
18 Coupled Plasma Optical Emission Spectrometer (ICP-OES) was used to analyze inorganic samples (Santa Clara, CA, USA). The IV-ICPMS-71A was purchased from
Inorganic Ventures as the standard for the ICP-OES (Christiansburg, VA, USA). A chlorine bath was used to clean all glassware and HDPE bottles for 24 hours before use, and then rinsed with DI water and dried.
3.2 Fly Ash Characterization
The fly ash comes from a company called Cristal, located in Ashtabula, OH.
Generation of ash relatively was unknown since it was in existence prior to purchase of the facility from Proctor and Gamble and has been left undisturbed since it was found.
The ash pile has been in place for over 60 years. The ash pile was relative homogenous, except for the contaminants of concern: barium (Ba) and strontium (Sr). A characterization of the fly ash pile was done previously and found the presence of barium at high concentrations and traces of other heavy metals.
3.2.1 Fly Ash Homogenization
Prior to characterization, to the fly ash sample was homogenized to ensure even distribution of barium and strontium. Homogenization was accomplished using a method known as quartering, where the sample was split into 4 even quarters (Simmons, 2014).
After being split each quarter was then thoroughly mixed with hand shovels separately.
Then two of the quarters are mixed together to form half of the sample, and then finally the halves are mixed together by shaking thoroughly the container the sample was placed in. The end product was a homogenous sample, which was then stored in a sealed bin.
19 The process was repeated multiple times to ensure the sample was kept homogenous and
did not separate over time.
3.2.2 Sieve Analysis
American Landfill Management Inc. provided the waste material to identify the
characteristics of the fly ash. ASTM C136 Standard Test Method for Sieve analysis of
Fine and Coarse Aggregates was performed to size distribution of particles within the ash pile ("Standard Test Method for Sieve Analysis of Fine and Coarse Aggregates," 2014).
The ash was dried in an oven at 110 ± 5°C to remove water content, and then placed into set of standard sieves. The sieves were arranged in decreasing opening size order, and each sieve was weighed before the material was introduced. The ash was placed in the top sieve and then mechanically shaken until the ash no longer moved to different sieves.
The sieves then are separated and weighed individually. Figure 3.1 shows the results from the sieve analysis performed at the University of Akron. For comparison, the sieve analysis performed by Geotechnical Testing Services is shown in Figure 3.2. Both analyses have a similar shape; however, Geotechnical Testing Services used larger sieve sizes to further characterize the ash.
Figure 3.1: Sieve analysis of fly ash from the Cristal processing plant in Ashtabula, OH.
Figure 3.2: Sieve analysis of the same fly ash performed by Geotechnical Testing Services Inc.
3.2.3 X-Ray Diffraction (XRD) Analysis
Further characterization of the ash was performed using x-ray diffraction (XRD).
XRD points x-rays at a sample and rotate around the sample. The x-rays were reflected from the crystalline structure, and the waves are recorded and used to identify the material (Clair N. Sawyer, 2003). Finley crushed samples of the ash were dried in a 110
± 5°C oven overnight to remove the water content before being submitted to XRD analysis. Four samples were analyzed, and three of the four samples showed signs of celestine, barian celestine ((Sr, Ba) SO4), and Calcite (CaCO3). The results from sample
#2 are shown in Figure 3.3 and 3.4.
Figure 3.3: XRD results for barian celstine and celestine from the Cristal fly
Figure3.4: XRD results of calcite in the Cristal fly ash.
23 3.2.4 Characterization of Heavy Metals in the Fly Ash
Method 3050B Acid Digestion of Sediments, Sludges, and Soils was used to
prepare the ash for analysis with inductively coupled plasma atomic emission
spectrometry (ICP-AES). To properly analyze the ash, it was necessary to remove any
interferences by performing an acid digestion. 1 gram of wet ash was digested first with
hydrochloric acid (HCl) and then with multiple additions of nitric acid (HNO3) and
hydrogen peroxide (H2O2). During the digestion the sample was heated and refluxed at a temperature of 95 ± 5˚C, but boiling was never achieved. The samples were then filtered using a 0.45 μm syringe filter, and then made to a volume of 100 ml with DI water.
Method 3050B is not a total digestion but can dissolve most elements. Silicates are difficult to dissolve with method 3050B, but they should not interfere (USEPA, 1996).
The acid digestion samples, as well as all of the leaching test samples were analyzed with an Inductively Coupled Plasma Atomic Emission Spectrometry (ICP-
AES). An ICP-AES uses argon gas and concentric quartz tubes that are water cooled by
a radio frequency generator to make a magnetic field. The argon is then ignited to 6000 –
8000 K, and then the sample is introduced. The sample dissociates, and certain elements
are measured with a monochromator or polychromator (Clair N. Sawyer, 2003). The
ICP-AES was used to identify all of the RCRA metals present, as well as various other
3.3 Experimental Methods
One approach to utilize the fly ash is to use it in concrete to substitute for either
cement of fine aggregate. Therefore, the ash was used to substitute for either cement or
24 fine aggregate as percentage of the mass of either constituent would have been added according to standard mortar and concrete matrices. The following describes each procedure for testing both mortar and concrete compressive strength after the appropriate curing time. Upon completion of these compressive strength tests, two different leach procedures were performed to ensure the heavy metals are retained bound in the concrete matrix and will not seep out.
3.3.1 Mortar Strength
The Standard Test Method for Compressive Strength of Hydraulic Cement
Mortars was modified by supplementing a percentage of either sand/fine aggregate or cement with the ash ("Standard Test Method for Compressive Strength of Hydraulic
Cement Mortars (Using 2-in. or [50-mm] Cube Specimens)," 2016). The composition of the mortar was originally just cement, sand, and water. Portland type 1 cement was used with construction sand and mixed as specified in ASTM C305 in a mixer ("Standard
Practice for Mechanical Mixing of Hydraulic Cement Pastes and Mortars of Plastic
Consistency," 2014). Table 3.1 and 3.2 show the original mixture as well as the four sand/fine aggregate replacements (S/W) and five cement replacements (C/W). Please note since the sand/fine aggregate is a larger percentage of each mix, more mass of fly ash can be utilized for S/W than for C/W.
Table 3.1: Mortar composition of fine aggregate replacement (S/W) by weight
Control 10% 20% 30% 40% Mixture S/W S/W S/W S/W # of cubes 6 6 6 6 6 Cement, g 500 500 500 500 500 Sand, g 1375 1237.5 1100 962.5 825 Water, ml 242 242 242 242 242 Waste Material, g 0 137.5 275 412.5 550
Table 3.2: Mortar composition of cement replacement (C/W) by weight
Control 10% 20% 30% 40% 50% Mixture C/W C/W C/W C/W C/W
# of cubes 6 6 6 6 6 6 Cement, g 500 450 400 350 300 250 Sand, g 1375 1375 1375 1375 1375 1375 Water, ml 242 242 242 242 242 242 Waste Material, g 0 50 100 150 200 250
The mixtures in Table 3.1 and 3.2 were placed in two molds of three 2” cubes each, for a total of six cubes per batch. These mixtures were then placed into a moisture room for a day before being removed from the molds and placed back into the moisture room. After three days three of the cubes were removed and dried. Then the compressive strength was measured using a press with a loading rate of 200 lbs./s. The
26 three remaining 2” cubes were tested by a press after seven days in the moisture room
and the compressive strength was recorded ("Standard Test Method for Compressive
Strength of Hydraulic Cement Mortars (Using 2-in. or [50-mm] Cube Specimens),"
3.3.2 Concrete Compressive Strength
Similar to the mortar compressive strength tests, the Standard Test Method for
Compressive Strength of Cylindrical Concrete Specimens was used to test concrete
cylinders that had been modified with ash. Table 3.3 and 3.4 show the distribution of
materials used to make 12 3” X 6” cylinders (Harvey, 2017).
Table 3.3: Concrete Composition - Fine Aggregate Replacement (S/W)
Control 10% 20% 30% 40% 50% Mixture S/W S/W S/W S/W S/W
# of Cylinders 12 12 12 12 12 12
Cement, g 3588 3588 3588 3588 3588 3588 Fine Aggregate, g 9118.5 8206.7 7294.8 6383.0 5471.1 4559.3 Coarse Aggregate, g 11824.5 11824.5 11824.5 11824.5 11824.5 11824.5 Water, ml 2005.5 2005.5 2005.5 2005.5 2005.5 2005.5 Waste Material, g 0 911.9 1823.7 2735.6 3647.4 4559.3
27 Table 3.4: Concrete Composition - Cement Replacement (C/W)
Control 10% 20% 30% 40% 50% Mixture C/W C/W C/W C/W C/W # of Cylinders 12 12 12 12 12 12 Cement, g 3588 3229.2 2870.4 2511.6 2152.8 1794.0 Fine Aggregate, g 9118.5 9118.5 9118.5 9118.5 9118.5 9118.5 Coarse Aggregate, g 11824.5 11824.5 11824.5 11824.5 11824.5 11824.5 Water, ml 2005.5 2005.5 2005.5 2005.5 2005.5 2005.5 Waste Material, g 0 358.8 717.6 1076.4 1435.2 1794.0
The values presented in Table 3.3 and 3.4 produced 15 cylinders; however, only 4
cylinders where used for each of the 3, 7, 28 day compressive strength tests. The larger
quantities of materials were used instead to keep the cylinders consistent. During mixing
some of the mixture would stick to the inside walls of the mixer. The paste that would
form on the walls after mixing was avoided, as it had a different consistency than the
majority of the concrete. After mixing was complete the concrete was placed in each
cylinder in two layers, with each layer rodded 25 times. The cylinders were then placed
on a concrete vibrating table until the consolidation was complete ("Standard Practice for
Making and Curing Concrete Test Specimens in the Laboratory," 2016). After the
vibration, the cylinders were placed in the moisture room for 24 hours, before being
removed from the molds and then returned to the moisture room. In groups of four on
curing days 3, 7, and 28 the cylinders were removed and dried before being tested for
28 compression strength with a press following ASTM C39/C39M ("Standard Test Method for Compressive Strength of Cylindrical Concrete Specimens," 2017).
3.3.3 Toxicity Characteristic Leaching Procedure (TCLP)
To ensure the stabilization of barium and strontium inside of the concrete matrix, the TCLP was performed to observe the movement of barium from the solid to the aqueous phase. The modified TCLP procedure used ten grams of the sample that passed a 3/8 in sieve but not a No. 4 sieve. The samples are then placed in a 250 ml high density polyethylene (HDPE) bottle with 200 ml of extraction fluid #2. The extraction fluid, which was an acetic acid solution, had a pH of 2.88 ± 0.05 (Miller et al., 2000). Samples of C/W and S/W were tested with the TCLP. Fine aggregate samples were spiked with
1500 mg/kg of barium and also tested with the TCLP. All the concrete samples came from multiple cylinders from the same batch.
Nine ash samples were tested as well with the TCLP to observe the amount of barium that would leach from the ash. All of the concrete samples were performed in triplicate as per the procedure, but the nine ash samples were done individually since they all were from the homogenous sample. After the 200 ml of extraction fluid was added to the ten grams of sample the bottle was sealed and placed on a wrist action shaker for 18 hours. Once the shaking was complete, the solution was decanted into eight 14 ml centrifuge tubes. The tubes were only filled to 12 ml, for a total of 96 ml being centrifuged for 15 minutes or until the solution appears clear. The centrifuge tubes were then decanted once again before being filtered with a 0.45 μm syringe filter. The solution is filtered into a 125 ml HDPE bottle, and then the pH was recorded. Then concentrated
29 nitric acid is added to lower the pH to below 2.0 to acidify and preserve the
sample(USEPA, 1992). The samples were then stored in a fridge until it was time they
Method 8014 was used to analyze the samples for the concentration of barium.
10 ml of the acidified samples will be placed into a sample cell for a 2100Q portable
turbidimeter. A 5 N sodium hydroxide solution was used to raise the to 5, and then the
sample cell is placed in the turbidimeter and the zero value is recorded. The sample cell
was removed, and the BariVer 4 Barium Reagent Powder Pillow was added. The cell is
then swirled until mixed thoroughly before being set down for five minutes. After the
five minutes the cell is placed back into the turbidimeter and the final concentration was
recorded(HACH, 2014). Beside method 8014 an ICP-AES was used to measure the
3.3.4 Accelerated Leach Test
The Standard Test Method for Accelerated Leach Test for Diffusive Releases
from Solidified Waste ASTM C1308 was performed to observe the leach rates of barium
from the concrete mixtures that had ash supplementing a component. The concrete
mixtures selected to be samples for the test were the 10-40% S/W. These were selected
due to their performance in previous testing, and due to the long duration, the test takes to perform. The four mixtures were made again using the same method as when initially creating the concrete for the compressive strength testing. The difference was that instead of 3” X 6” cylinder molds, 35 mm camera film containers were used instead.
This made cylinders with a diameter of 1.1875” and a height of 2”. These cylinders were
30 attached to the lids of 500 ml HDPE bottle with monofilament string and placed into 450 ml of DI water. The caps were sealed, and the container was placed into a fume hood for storage. After 24 hours the container was removed from the hood and the lid was unscrewed with the cylinder and placed onto a new 500 ml HDPE bottle with another 450 ml of DI water. The new bottle was sealed and placed into the fume hood. The previous bottle with contaminated DI water was quickly stirred before being filtered with a 0.45
μm syringe filter. 100 ml were taken and acidified with nitric acid to a pH less than 2.0.
The samples were then stored in a fridge until they were to be tested. Each mixture was tested in triplicate, and the replacing of water went on for 11 days("Standard Test Method for Accelerated Leach Test for Diffusive Releases from Solidified Waste and a Computer
Program to Model Diffusive, Fractional Leaching from Cylindrical Waste Forms," 2017).
To analyze for the concentration, method 8014 was used, which was the same method used for the TCLP. An ICP-AES was used as well as method 8014 to measure the leached metals.
Samples from the accelerated leaching test were also analyzed for the sulfate concentration to observe the change in leached sulfate from the initial day to the final day with method 8051 (HACH, 2018). Similar to method 8014 method 8051, the procedure was the same except that the powder added was the SulfaVer 4 Reagent Powder Pillow.
3.3.5 Leached Barium with Sulfate Additions
A final test to observe the amount of barium leached from the ash when exposed to DI water with the addition of sodium sulfate mixed into the ash. 10 g of sample were added to a 500 ml HDPE bottle and then 500 ml of DI water was added. Four different
31 batches were tested in triplicate. A blank with only ash present, ash with 1 mole of sulfate mixed in, ash with 5 moles of sulfate mixed in, and ash with 11 moles of sulfate added in. The samples had aliquots of 10 ml taken after 6 hours, 1 day, 5 days, and 10 days. The samples were then acidified with nitric acid to lower the pH to below 2, and then stored before analysis with an ICP-AES.
RESULTS AND DISCUSSION
This chapter will present and discuss the data observed from the experiments performed using the fly ash obtained from Ashtabula, OH. The initial experiments identified the possible beneficial uses of the waste with cementitious products (i.e., mortar and concrete). Upon achieving initial compressive strength results for certain mixtures of concrete, leaching experiments were performed to determine if solidification was an acceptable treatment technology. The TCLP and the accelerated leaching test were used to observe the leaching of barium from the waste and concrete, and the
SulfaVer 4 method was used to analyze the concentration of sulfate present during the accelerated leaching test. Finally, barium stabilization was investigated through the addition of sodium sulfate.
4.2 Mortar compressive strength with effects of ash
The mortar cubes were tested for compressive strength after curing for 3 and 7 days. The control group contained zero waste ash, and the compressive strength at 3 days was 891 ± 270 psi, and at 7 days it was 878 ± 101 psi. The compressive strength was
33 expected to increase with time, but the compressive strength decreased from the 3 day to the 7 day. The confidence interval from the 3 day to the 7 day decreased but the compressive strength remained stagnant. Figure 4.1shows the observed compressive strength of the mortar cubes from the control group used for comparison with cement replacement (C/W).
The control group mixture was then altered with 10–50% of the C/W with 6 cubes made in each of the five batches. The results of the C/W in mortar cubes are shown in
Figure 4.1. Interpreting the compressive strength results, it appears that no pozzolanic effect is occurring over the first 7 days, and the removal of any percentage of cement is having a negative effect on the mortar. 10% C/W obtained the highest compressive strength at day 7 with 635 ± 500.1 psi, with any higher percentage replacement reducing the compressive strength. Comparing the C/W to the control group, the C/W had a
27.7% decrease in compressive strength when replacing 10% cement. 20, 30, 40, and
50% C/W compressive strength had a decline of 38.6% or more compared to the control group. The waste material provided no beneficial effects to the mortar and reduced the compressive strength for all C/W batches.
Figure 4.1: Compressive strength of mortar cubes with cement replacement (C/W) 10– 50% on day 3 and 7 compared to the control group .
The fine aggregate replacement (S/W) in mortar had the opposite effect compared to C/W. S/W 10–40% was replaced in four batches with the fly ash. The results of the compressive strength tests are shown in Figure 4.2. A significant compressive strength increase was noticeable with all the batches compared to the control group. All four batches of S/W at day 7 had a compressive strength of 1400 psi or more. Observing
Figure 4.2 it appears that the 10 and 20% S/W were on an upward trend, but when 30% or more was replaced the optimal compressive strength had been passed. 30% and more was detrimental to the compressive strength, but still obtained higher compressive strength than the control group. The 40% S/W, which had the lowest compressive
35 strength of the S/W, still had a compressive strength 59.5% higher than the control group.
The advantages of S/W compared to the cement replacement was the amount of waste ash used in the batches. At equal percent replacement, S/W used 2.75 times more waste ash than the C/W by weight. S/W provided a positive property to the mortar by increasing the compressive strength of the mortar and reducing the availability of a larger amount of waste than the C/W. While the results from the compressive strength tests are no clear indication of the uses of the ash with cementitious materials, it was an optimistic result that led to further investigation with the S/W.
Figure 4.2: Compressive strength of mortar cubes with fine aggregate replacement (S/W) 10 –40% on day 3 and 7 compared to the control group .
4.3 Concrete compressive strength with ash replacement
Four 3” X 6” concrete cylinders were tested for compressive strength on day 3, 7,
and 28. For the control group on day 3 the observed compressive strength was 4663.6 ±
283.6 psi and then increased to 5769.8 ± 171.9 psi at day 7. After 28 days the control
group was found to have a compressive strength of 6456.5 ± 174.2 psi. Contrasting the
mortar control group, the concrete control group has an increase in compressive strength
over the curing period. Figure 4.3 and 4.4 each show the compressive strength of the
control group compared to the cement replacement (C/W) and fine aggregate replacement
(S/W), respectively. The control group was used to compare the effects of the C/W and
C/W in concrete of 10–50% resulted in similar results as to C/W in mortar, with a
decrease in the compressive strength compared to the control group. With 10% C/W
with ash the compressive strength of the concrete at 28 days was 4816.8 ± 392.2 psi,
which was a 25.4% decrease in compressive strength compared to the control group. The
20, 30, 40, and 50% batches continued to have decreasing compressive strengths, as the
percentage replacement increased. At 20% C/W the compressive strength decreases by
37.7% compared to the control group. 30% and higher decrease the compressive strength by 57% or more. Figure 4.3 shows the results of the C/W compared to the control group.
40 and 50% C/W detrimentally effected the compressive strength of the concrete, with a decrease in compressive strength of 69.6% compared to the control group, and the concrete not reaching 2000 psi.
Figure 4.3: Compressive strength of concrete cylinders with cement replacement (C/W) 10 –50% .
S/W in concrete of 10-50% ash, was observed to have varying results but follow a similar trend to C/W in concrete. Figure 4.4 shows the results of the S/W compared to the control group. While none of the S/W achieved a higher compressive strength than the control group, the effect on the compressive strength was minimal compared to the
C/W. At 10 and 20% S/W the compressive strength of the concrete was observed to
6051.3 ± 472.6 psi and 5220.0 ± 284.2 psi, respectively. 10 and 20% S/W decreased the compressive strength by 6.25% and 19.2%, respectively compared to the control group.
At 40% S/W the compressive strength was 3817 ± 187.5 psi which was a decrease of
38 40.8% from the control group. 50% S/W had a compressive strength of 2836.8 ± 186.0
psi, which was a decrease of 56.1% compared to the control group.
The S/W did not provide any benefits to the concrete compared to the control
group, but the concrete is minimally affected by the ash compared to the C/W. S/W used
2.54 times more ash at equal percent with the C/W. Both 10 and 20% S/W had a
compressive strength over 5000 psi, while none of the C/W achieved more than 4816 psi.
The C/W provided no advantage compared to the S/W that reduced the compressive
strength less and used more of the ash. C/W did not achieve a compressive strength above
3,000 psi when the percent replaced exceeded 20%. When S/W was investigated, replacement up to 40% was effective at achieving above 3,000 psi concrete.
Figure 4.4: Compressive strength of concrete cylinders with fine aggregate replacement (S/W) 10 -50%.
4.4 Leaching Tests results
4.4.1 Acid Digestion
An acid digestion was performed on 1 gram of the ash, and the sample was analyzed for all the RCRA metals and aluminum. The analysis of the acid digestion was to find the possible leaching concentrations of metals present in the ash. Figure 4.5 show the concentrations of barium, lead, and aluminum that were leached from the ash. In
Figure 4.5 the barium concentration of the ash was 13.5 ± 6.7 mg/L or 1,350 mg/kg in soil. The soil concentration was well below the RSLs standard for residential and
40 industrial soil, which are 15,000 and 220,000 mg/kg, respectively. The 13.5 mg/L of
barium was above the primary drinking standard of 2 mg/L, but well below the RCRA
standard of 100 mg/L of barium (USEPA, 2005, 2017a). The concentration of aluminum
that leached out of the ash from the acid digestion was 102.9 ± 28.6 mg/L or 10,290
mg/kg. There is no primary drinking water standard or RCRA standard for aluminum,
but the concentration was noticeable. The lead concentration shown in Figure 4.5 from
the ash was 1.3 ± 2.3 mg/L and 130 mg/kg. The 1.3 mg/L of lead was higher than the
primary drinking water standard which set the limit at 0.015 mg/L, but was still below the
RCRA standard of 5 mg/L. RSL standard was also met, with residential and industrial limits being 400 and 800 mg/kg, respectively. Strontium was found to be present in the ash, and the aqueous concentration of strontium was over 100 mg/L or 10,000 mg/kg.
The acid digestion provided the initial conditions and the concentrations in the ash.
Figure 4.5: Barium (Ba), lead (Pb), and aluminum (Al) concentration of ash from acid digestion .
4.4.2 TCLP results
The TCLP was performed on the nine ash samples as well as on two batches of
C/W in concrete 10–20%, in triplicate, and four batches of S/W in concrete 10 – 40%, in
triplicate. Not all C/W samples were tested due to the fact that high than 20% C/W
yielded less than a 3,000 psi concrete. Another set of S/W in concrete was made with 10
– 40% fly ash that was spiked with a concentration of 1500 mg/kg of barium added into the ash, in triplicate. The TCLP barium concentration from the ash was 0.26 ± 0.32 mg/L, which was below the primary drinking standards of 2 mg/L. The TCLP concentration of
42 aluminum was 4.5 ± 4.1 mg/L, which was different from the 102.9 mg/L of aluminum in the acid digestion. Lead was not observed in the samples of ash from the TCLP.
Strontium was observed again at higher concentrations compared to the other metals present, with more than 100 mg/L of strontium in all ash samples. Comparing the TCLP ash samples to the acid digestion of the ash in all cases the TCLP leached metals at lower concentrations. Observing the results, it can be conjectured that barium was being stabilized to keep it from moving into the aqueous phase.
Figure 4.6presents the results of the TCLP on the 10% C/W, and Figure 4.7shows
20% C/W. The 10% and 20% C/W had a barium concentration of 0.72 ± 0.09 mg/L and
0.67 ± 0.40 mg/L, respectively. The C/W results showed that barium was effectively prevented from transporting into the aqueous phase. The concentration of barium increased compared to the TCLP of the ash, but this could be due to the variability of the barium concentration in the ash samples. The resulting barium concentration was still below the primary drinking water standards and the RCRA standards. Strontium concentration shown in Figure 4.6 and 4.7 were 30.5 ± 27.1 mg/L for 10% and 56.5 ±
15.2 mg/L for 20% C/W, respectively. The acid digestion and TCLP of the ash had strontium concentrations over 100 mg/L, so by introducing the ash into the concrete the concentration of strontium was lowered considerably. The concentration of aluminum in
Figure 4.6 and 4.7, which followed the same trend as strontium, with C/W having lower concentration of aluminum compared to the acid digestion. 10% C/W had a concentration of 1.1 ± 0.8 mg/L and 20% C/W had 1.9 ± 1.2 mg/L of aluminum. Similar to strontium, the increase in percent replacement increased the concentration of aluminum.
Figure 4.6: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from C/W 10%.
Figure 4.7: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from C/W 20%.
The S/W of 10–40% TCLP results are shown in Figure 4.8, 4.9, 4.10, and 4.11
which present barium and aluminum concentrations leached during the TCLP. The
concentration of barium in the S/W was still below the primary drinking standards and
RCRA standards for TCLP results. The highest observed concentration of barium was
0.9 ± 0.3 mg/L at the 10% S/W. The variability of the concentration of barium in the ash made it difficult to interpret the results, though it appeared that the concrete was preventing the barium from leaching. Aluminum was at its highest concentration with
30% S/W which was 2.6 ± 1.1 mg/L. The concentration was still less than the TCLP of the ash and the acid digestion, which may indicate the concrete was affecting the leaching
45 of the aluminum. Strontium was at its lowest concentration with 10% S/W shown in
Figure 4.8, with a concentration of 78.7 ± 35.3 mg/L. 20, 30, and 40% S/W all had strontium concentrations above 100 mg/L. Lead was not observed in any of the samples.
S/W with the ash showed to be affective at preventing the leaching of barium, and possibly aluminum, and strontium.
Figure 4.8: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from S/W 10%.
Figure 4.9: TCLP concentration of leached aluminum (Al), and barium (Ba) from S/W 20%.
Figure 4.10: TCLP concentration of leached aluminum (Al), and barium (Ba) from S/W 30%.
Figure 4.11: TCLP concentration of leached aluminum (Al), and barium (Ba) from S/W 40%.
1500 mg/kg of barium was introduced to the ash and used with same fine aggregate batches as prior. 10–40% S/W was repeated with the spiked ash and the barium and aluminum concentrations from the TCLP are reported in Figure 4.12, 4.13,
4.14, and 4.15. All samples of barium remained below the standard, but the confidence interval does surpass the 2 mg/L allowed in Figure 4.12, 4.13, and 4.15. The highest observed concentration of leached barium from the spiked samples was 1.79 ± 0.44 mg/L from 40% S/W. The concentrations leached do not approach the 100 mg/L of barium the
RCRA standard allows. It appears that the concrete is effectively preventing the leaching of barium, even when the material was spiked with 1500 mg/kg of barium. Aluminum had similar concentrations to that of the leached aluminum from the original S/W. There
49 does not appear to be a trend in the percentage of ash added to the amount of aluminum
as it increases and decreases at random. Strontium concentrations were smallest in the
10% S/W with a concentration of 18.9 ± 67.8 mg/L and are shown in Figure 4.12. 20, 30,
and 40% S/W had strontium concentrations above 100 mg/L. Increasing the percentage
of ash above 10% in both the original ash and the spiked ash had leached strontium
concentration higher than 100 mg/L.
Figure 4.12: TCLP concentration of leached aluminum (Al), barium (Ba), and strontium (Sr) from spiked S/W 10 %.
Figure 4.13: TCLP concentration of leached aluminum (Al), and barium (Ba) from spiked S/W 20%.
Figure 4.14: TCLP concentration of leached aluminum (Al), and barium (Ba) from spiked S/W 30%.
Figure 4.15: TCLP concentration of leached aluminum (Al), and barium (Ba) from spiked S/W 40%.
Similar to Vaidya et al. (2010) the TCLP showed that barium can be lowered to below the RCRA standard of 100 mg/L, and in most cases below the primary drinking water standard of 2 mg/L. When the fly ash was spiked with 1500 mg/kg of barium, concentrations of leached barium increased compared to the leached barium concentrations from the C/W and S/W that used the original ash. S/S can be inferred to both have occurred due to the concentrations of leached barium from the ash. The acid digestion leached 13.5 ± 6.7 mg/L of barium, but the TCLP only leached out 0.26 ± 0.32 mg/L. The leached concentration of barium should be at a higher concentration from the ash, but it appeared that something was preventing the barium from leaching.
53 Stabilization seems to be occurring, which lowered the leached concentration in the
TCLP, and could be due to the presence of sulfate in the ash.
4.4.3 Accelerated Leaching Test
The accelerated leaching test was performed to determine the diffusivity of barium, strontium, and aluminum through the cementitious matrix. Also, all the RCRA metals were monitored in case lead or another metal was to slowly diffuse out the concrete samples. Therefore, S/W 10–40% were used in testing in triplicate for 11 days.
Cement replacement samples were not examined since the fly ash does not demonstrate any self-cementing properties. The 10% S/W results from the accelerated leaching test are shown in Figure 4.16 for barium, aluminum, and strontium. Barium was not observed on any day except for day 8, with a concentration of 0.06 ± 0.27 mg/L. The concrete effectively inhibited the barium from leaching out. Strontium concentrations were observed on each day and never exceeded7 mg/L. There was no observable trend of the concentration changing over time. The aluminum concentrations were observed on all days with the highest analyzed concentration being1.1 ± 2.7 mg/L. A trend does seem to occur from day 1 to 8 where the concentration increases, but from day 9 to 11 the amount of aluminum flatlines.
Figure 4.16: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 10 %.
The 20% S/W had similar results to the 10% S/W, but did have an increase in days where barium was observed. Barium concentrations shown in Figure 4.17 are observed from day 6 to 11while from day 1 to 5 there was no barium that leached from the samples. The highest concentration of barium leached was 0.09 ± 0.03 mg/L. Still it appeared that the barium was prevented from leaching at high concentrations. A trend is noticeable that after day 6 the concentration of barium increases continuously. The amount of barium observed was well below the primary drinking water and the RCRA standard. Strontium concentrations were similar to the 10% S/W, and had a lower
55 observed maximum concentration of 4.9 ± 3.9 mg/L (Figure 4.17). The highest observed concentration of aluminum was 0.54 ± 0.03 mg/L shown in Figure 4.17. The concentrations found were less than those found in the 10% S/W. Aluminum and strontium were observed for all 11 days. At 20% S/W, barium did leach out at a higher concentration compared to the 10% S/W, but still at a level below the primary drinking water standards.
Figure 4.17: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 20 .
30% S/W had the leached metal concentrations increase compared to both 10% and 20% S/W, shown in Figure 4.18. For 30% S/W barium was observed earlier than
56 before, starting on day 5 and continuing until day 11. Again, the concentration of barium increased compared to 20% S/W. The highest observed concentration was 0.16 ± 0.21 mg/L of barium. The concentration of leached barium was still below the primary drinking water standards and the RCRA standard. Strontium was found to be in the samples for all 11 days, and the concentration leached out increased to 8.0 ± 20.4 mg/L and was observed on day 2. Aluminum appeared for all 11 days and increased from the
20% S/W, with the highest concentration of 0.63 ± 0.57 mg/L. It appeared there was a trend with aluminum leaching out at increased rates the more time passed until day 5 where the concentration became stagnant. Again, the 30% S/W prevented barium from leaching at higher concentrations.
Figure 4.18: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test wit h S/W 30.
40% S/W followed similar trends compared to the 10 – 30% S/W, and the metals
leached are shown in Figure 4.19. The barium leached out earlier than the other batches,
on day 3 and continued till day 11. Barium was also observed at the highest
concentration from the accelerated leaching test on the final day of testing with 0.23 ±
0.01 mg/L on day 11. From day 3 to day 11 the barium increased continually. Barium
appears to follow a trend whereas time increases the amount of barium leached does as
well. Still the concentrations observed are below the primary drinking water standards
and RCRA standard. The highest concentration of strontium was 7.2 ± 6.0 mg/L, which
58 was a decrease compared to the 30% S/W. The decrease in leached strontium could be
due to the variability of strontium concentration in the ash. Aluminum concentrations
similarly to strontium concentrations decreased compared to 30% S/W. The highest
observed aluminum concentration was 0.56 ± 0.03 mg/L on day 9. The leached
aluminum concentration followed a trend where the leached concentration increased with time for the first 5 days and then fluctuated up and down afterwards. The decreased aluminum concentrations compared to the 30% S/W could be due to variability of the aluminum concentration throughout the ash. The barium was again prevented from leaching at high concentrations effectively even at 40% S/W
Figure 4.19: Barium (Ba), aluminum (Al), and strontium (Sr) concentrations leached from the accelerated leaching test with S/W 40.
The accelerated leaching results showed consistent data that barium was being prevented from leaching from the cementitious matrix. While leached barium concentration appears to increase over time from the 20, 30, and 40% S/W, the concentrations observed do not approach the primary drinking water standard, and even less so for the RCRA standard of 100 mg/L. The highest observed concentration of leached barium was 0.23 ± 0.01 mg/L, from 40% S/W. At 10% S/W barium was only observable for one day and no trend could be observed. Strontium and aluminum concentrations appeared to increase minutely for 10 – 30% S/W and decreased slightly at
4.5 Sulfate Stabilization
500 ml of deionized (DI) water was used to submerge 10 g of ash mixed with sodium sulfate at a concentration of 1 mole, 5 moles, and 11 moles of sulfate, with a blank included to compare to. Then barium leaching was observed at 6 hours, 1 day, 5 days, and 10 days. Figure 4.20 shows the leached concentration of barium from the ash samples. The amount of barium leached appeared to decrease as time went by, but even the blank ash with no addition of sulfate only leached 0.13 ± 0.12 mg/L, well below the primary drinking water standard and the RCRA standard. With such low concentrations of leached barium in the blank, it seems the ash already contains sulfate that has formed barium sulfate. The blank concentration of leached barium stayed level throughout and no changes occurred over the 10 days. The ash with 1, 5, and 11 moles of sulfate added
60 all had decreased concentrations of leached barium over time. As time passed more
barium sulfate formed and fell out of solution.
Figure 4.20: Leached barium concentration from a sh, with a blank (BL) with no sodium sulfate, 1 mole (1 mol) of sodium sulfate, 5 moles (5 mol) of sodium sulfate, and 11 moles (11 mol) of sodium sulfate from 0.25 days to 10 days.
The accelerated leaching test samples were analyzed with a SulfaVer 4 method.
The initial and final days were analyzed for sulfate concentrations to observe if there was
excess sulfate was present. The initial concentration of sulfate on day 1 was 1.46 ± 2.06
61 mg/L, and on the final day the concentration was 0.37 ± 0.02 mg/L, shown in Figure
Figure 4.21: Sulfate concentration from day 1 and day 11 of accelerated leaching test.
The change of sulfate from the initial to the final day could correlate to the increase in leached barium observed from the accelerated leaching test. Each day the water was changed in the accelerated leaching test, and it appeared that the concentration of sulfate leached decreased, while the barium leached increased. The amount of sulfate observed also indicates that barium did not leach in the TCLP of the ash due to the sulfate naturally present in the ash, that stabilized it by forming barium sulfate. S/S may still be
62 required though to keep the barium concentration from rapidly increasing when the sulfate concentration present decreases.
CONCLUSIONS AND RECOMMENDATION
The study first identified the physical and chemical composition of the fly ash and
then observed the barium leaching. An acid digestion was performed on the ash to detect
the heavy metals present in the ash aside from barium. Solidification/Stabilization (S/S)
was used to investigate a change in the amount of leached barium from the ash by mixing
the ash with a binder. The Toxicity Characteristic Leaching Procedure (TCLP) was used
on the cement replacement (C/W) and fine aggregate replacement (S/W) by shaking 200
ml of an acetic acid solution with a pH of 2.88 with 10 grams of the sample. The
accelerated leaching test was performed by submerging concrete cylinders in deionized
(DI) water for 11 days. The accelerated leaching test samples were analyzed for sulfate
concentrations to check the amount of sulfate present in the ash. Finally, the ash was submerged in DI water for 10 days with various amounts of sulfate added to observe stabilization.
64 5.2 Conclusions
1. Through x-ray diffraction (XRD) analysis the presence of barium, strontium,
calcium, and sulfate were confirmed to be present in the ash. The sulfate was key,
as it indicated that some form of chemical stabilization was already occurring in
2. Solidification with a cementitious binder proved to limit the leaching of barium to
below both RCRA and primary drinking water standards. 10% S/W reduced the
compressive strength minimally with only a reduction of 6.25% compared to the
control group of concrete. With 10% C/W the compressive strength was
decreased by 25.4% compared to the control group. The larger decrease in
compressive strength with the same 10% replacement is noticeable since the C/W
incorporates less ash into its mixture than the S/W. 10% S/W contains 2.54 times
more ash than 10% C/W, and has less of a detrimental effect on the compressive
strength. This is true for each S/W and C/W, and shows the advantage of S/W
compared to C/W for application.
3. Leached barium from ash that the TCLP had been performed on, was below the
primary drinking water standard of 2 mg/L with 0.26 ± 0.32 mg/L. When 1500
mg/kg of barium was mixed into the ash and used in concrete with the TCLP, the
leached barium concentration remained miniscule. The leached barium
concentration being so low indicates the presence of sulfate, forming barium
sulfate, meaning the ash was stabilized.
4. Sulfate analysis on the accelerated leaching test samples showed a trend that
sulfate leached out in higher concentrations on day 1 versus day 11 where the
65 concentration had decreased. The decrease in sulfate as time continues was
inverse to the leached barium concentration. At later days the leached barium
concentration increased as the leached sulfate concentration decreased. With 40%
S/W the barium concentration started from below detection and increased to 0.23
mg/L, which was still below the primary drinking water standard.
5. When sulfate was mixed with the ash and soaked in DI water for 10 days, the
leached barium concentrations were decreased, but the ash with no addition of
sulfate had a concentration of 0.13 ± 0.12 mg/L. The low concentration found
just from the ash indicates the presence of sulfate that was forming barium sulfate,
self-stabilizing the ash.
1. Further research into the total sulfate concentrations in the ash, to determine the
ability of the ash to self-stabilize itself.
2. Additional experiments should be performed on fine aggregate replacement with
another control group to compare the effect of the ash on the two different
3. More investigation into other possible beneficial uses such as capping material or
drying material for dredged soil. Discovering another use for the ash may help
find additional funding for other projects.
4. Diffusivities of aluminum, barium, and strontium should be modeled and
compared quantitatively to one another.
66 5. Continuing the accelerated leaching test to past 11 days to observe the change in
barium concentration over a longer period of time.
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A.1 Turbidimetric Barium Analysis
A1.1 TCLP Turbidimetric Barium Results
Figure A.1: Barium concentration C/W 10 -20% and S/W 10-40% from TCLP analyzed with the turbidimetric method .
72 A.1.2 Accelerated Leaching Test Turbidimetric Barium Results
Figure A.2: Barium concentration S/W 10% from accelerated leaching test analyzed with the turbidimetric method .
Figure A.3: Barium concentration S/W 20% from accelerated leaching test analyzed with the turbidimetric method .
Figure A.4: Barium concentration S/W 30% from accelerated leaching test analyzed with the turbidimetric method .
Figure A.5: Barium concentration S/W 40% from accelerated leaching test analyzed with the turbidimetric method .
76 A.2 Compressive Strength
A.2.1 Concrete Compressive Strength
Table A.1 cement replacement (C/W) compressive strength
Day 10% C/W 20% C/W 30% C/W 40% C/W 50% C/W (psi) (psi) (psi) (psi) (psi) 3 3447 ± 26 2935 ± 103 1914 ± 177 1335 ± 196 866 ± 31 7 4214 ± 136 3565 ± 240 2513 ± 87 1883 ± 141 1214 ± 86 28 4816 ± 392 4017 ± 217 2750 ± 120 1961 ± 166 1369 ± 42
Table A.2 fine aggregate replacement (S/W) compressive strength
Day 10% S/W 20% S/W 30% S/W 40% S/W 50% S/W (psi) (psi) (psi) (psi) (psi) 3 4798 ± 509 4274 ± 195 3492 ± 206 3025 ± 254 2233 ± 131 7 5542 ± 336 4748 ± 187 4012 ± 253 3447 ± 329 2693 ± 326 28 6051 ± 472 5220 ± 284 4311 ± 330 3817 ± 187 2836 ± 186