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1 IMMOBILIZATION OF MERCURY IN CO NTAMINATED SOILS USING ALUMINUM DRINKING WATER TREATMENT RESIDUALS By ANNA HOVSEPYAN A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLOR IDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2008
2 2008 Anna Hovsepyan
3 To my mother, Tatyana Chukhrukidze.
4 ACKNOWLEDGMENTS I especially thank Dr. Jean-Claude J. Bonzongo for his guidance, support, kindness, hum or, and many other great qualities that ma de him an exceptional advisor. I thank my committee members, Dr. Gabriel Bitton, Dr. Joseph J. Delfino, and Dr. Willie G. Harris, for their valuable expertise, advice, and support that helped me bri ng this project to completion. My special thanks go to Mr. and Mrs. Glick, founde rs of the Sally and William Glick Graduate Research Award, for their generosity and intere st in environmental research. It was my honor and privilege to be the recipient of this award in 2006. I thank Dr. Paul J. Lechler of the Nevada Bureau of Mines and Geology for providing mercury contaminated soil samples. I also thank Dr. Xinde Cao and Lisa Stanley for their help with soil characterization. I thank Gill Brubaker and Gary Scheiffele of the Particle Engineering Resear ch Center of the University of Florida for their help and advice in Al-WTRs charact erization. I am grateful to the students in the Department of Environmental Engineering Sciences for their fri endship and support. My special thanks go to my family and friends. I greatly appreciate their love, support, and encouragement.
5 TABLE OF CONTENTS page ACKNOWLEDGMENTS...............................................................................................................4 LIST OF TABLES................................................................................................................. ..........7 LIST OF FIGURES.........................................................................................................................8 ABSTRACT...................................................................................................................................10 CHAP TER 1 INTRODUCTION: MERCURY IN SOILS AND REMEDI ATION.................................... 12 Problem Statement.............................................................................................................. ....12 Overview of Soil Remediation Techniques and Rationale for this Study.............................. 15 Drinking Water Treatment Residuals (WTRs)....................................................................... 19 Research Objectives and Dissertation Outline....................................................................... 21 2 ALUMINUMBASED DRINKING WATER TREATMENT RESIDUALS (AlWTRs) AS A SORBENT FOR MERCURY: MA TRIX CHARACTERIZATION AND DETERMINATION OF SORPTION CAPACITIES............................................................ 27 Introduction................................................................................................................... ..........27 Materials and Methods...........................................................................................................28 Collection and Characterization of Al-WTRs................................................................. 28 Determination of the Maximum Sorpti on Capacity of AlWTRs and Sorption Isotherms......................................................................................................................31 Mercury Desorption from Al-WTRs...............................................................................32 Kinetics of Hg Sorption on Al-WTRs and Effect of pH................................................. 33 Results and Discussion......................................................................................................... ..34 Characterization of Al-WTRs..........................................................................................34 Determination of the Maximum Sorpti on Capacity of AlWTRs and Sorption Isotherms......................................................................................................................35 Mercury Desorption from Al-WTRs...............................................................................37 Kinetics of Hg Sorption on Al-WTRs............................................................................. 37 Conclusions.............................................................................................................................41 3 IMMOBILIZATION OF MERCURY IN CONT AMINATED SOILS BY ALUMINUMBASED DRINKING WATER TR EATMENT RESIDUALS (Al-WTRs)... 56 Introduction................................................................................................................... ..........56 Materials and Methods...........................................................................................................59 The Al-WTRs and Soil Sample Collection, Ch aracterization, and Site Description ...... 59 Column Leaching Studies................................................................................................62
6 Assessing the Bioavailability of Mercury in Al-WTRs Treated Soils ............................ 65 Chemical Fractionation of Mercury in Used Soils.......................................................... 66 Results and Discussion......................................................................................................... ..68 Column Leaching Studies: General Characteristics........................................................ 68 Florida Sandy Soil...........................................................................................................69 Nevada Soil.....................................................................................................................73 Conclusions.............................................................................................................................76 4 DISTRIBUTION OF MERCURY SORBE D ONTO ALUMINUM-BASED DRINKING WATER TREATMENT RESIDU ALS (Al-WTRs) AND IMPLICATIONS FOR LONG-TERM STABILITY OF FORMED MERCURY-Al-WTRs COMPLEXES.........................................................................................................................89 General Introduction........................................................................................................... ....89 Part-1: Chemical Fractionation of Mercur y in Al-WTRs by Sequential Extraction .............. 89 Introduction................................................................................................................... ..89 Material and Methods......................................................................................................91 Preparation of aged me rcury-spiked Al-WTRs ........................................................ 91 Sequential extraction procedure...............................................................................92 Results and Discussion.................................................................................................... 94 Part-2: Mercury Incorporation in Micropores of Al-W TRs and Modeling of Intraparticle Diffusion.............................................................................................................................98 Introduction................................................................................................................... ..98 Materials and Methods....................................................................................................99 Aged mercury-spiked Al-WTRs.............................................................................. 99 Determination of SSA-N2 and SSA-CO2 of Al-WTRs before and after mercury sorption................................................................................................................. 99 Determination of the activation energy (Ea) of mercury sorp tion onto Al-WTRs particles...............................................................................................................101 Results and Discussion.................................................................................................. 102 Determination of SSA-N2 and SSA-CO2 of Al-WTRs before and after mercury sorption............................................................................................................... 102 Modeling of intraparticle diffusion a nd prediction of m ercury diffusivity............ 104 General Conclusions............................................................................................................ .106 5 CONCLUSIONS AND RECOMME NDATIONS............................................................... 117 Conclusions...........................................................................................................................117 Recommendations................................................................................................................ .118 LIST OF REFERENCES.............................................................................................................120 BIOGRAPHICAL SKETCH.......................................................................................................135
7 LIST OF TABLES Table page 1-1 Selected Hg standards for different environmental compartments.................................... 23 1-2 Comparison of chemical and physical prope rties of WTRs wi th typical soil levels (from Dayton and Basta, 2001).......................................................................................... 24 1-3 Comparison of WTR nutrient leve ls with soil nutrient leve ls adequate for growth of most crops (from Dayton and Basta, 2001)....................................................................... 25 1-4 Average concentrations of 15 metals in Al-W TRs determined using the EPA method 3050....................................................................................................................................26 2-1 Physicochemical properties of Al-WTR s collected from the Bradenton Drinking Water Treatment Facility (Florida, USA).......................................................................... 43 2-2 Determination of Hg maximum so rption capacity (m g/g) at pH 6.5................................. 44 2-3 Sorption isotherm parameters fo r Hg sorption on Al-WTRs at pH 6.5 ............................. 44 2-4 Kinetic Parameters for Hg Sorption on Al-WTRs at pH 6.5 and initial Hg concentration of 40 m g/L................................................................................................... 44 3-1 Experimental design for column studies with soil weight in each colum n of 300 g......... 77 3-2 Physicochemical characteristics of soils used in colum n leaching studies........................78 3-3 Toxicity of Floridas sandy so il leachate collected from the 1st pore volume as determined by the MetPLATETM toxicity test................................................................... 79 3-4 Mass balance of Hg in column studies using Floridas sandy soil. ................................... 80 3-5 Mass balance of Hg in column studies using Nevadas soil.............................................. 81 4-1 Mercury, Al, Fe, and Si concentrations a nd distribution (%) in different fractions of Hg-spiked Al-WTRs ........................................................................................................ 108
8 LIST OF FIGURES Figure page 2-1 SEM micrographs of the orig inal Al-WTRs collected from the Bradenton Drinking Water Treatment Facility (Florida, USA) showing irregular size and nonhomogeneity of particles.................................................................................................... 45 2-2 Elemental spectra (EDS) of the orig inal Al-W TRs before use in Hg sorption experiments and of the recovered Al-WTRs after the sorption experiments confirming Hg sorption on Al-WTRs................................................................................ 46 2-3 Mercury sorption isotherm with initia l Hg concentrations of 10, 20, 40, and 80 m g/L and pH 6.5..........................................................................................................................47 2-4 Langmuir plot for Hg sorption on Al-WTR s at initial Hg concentrations of 10, 20, 40, and 80 mg/L and pH 6.5.. ............................................................................................48 2-5 Effect of contact time on Hg sorp tio n on Al-WTRs at 40 mg/L initial Hg concentration and pH 6.5...................................................................................................49 2-6 Lagergren, pseudo-first or der kinetic plot for Hg sorption on A l-WTRs at initial Hg concentration of 40 mg/L and pH 6.5................................................................................ 50 2-7 Pseudo-second order kinetic plot for Hg sorption on Al-W TRs at initial H g concentration of 40 mg/L and pH 6.5................................................................................ 51 2-8 Weber and Morris intraparticle diffusion plot for Hg sorption on Al-W TRs at initial Hg concentration of 40 mg/L and pH 6.5.......................................................................... 52 2-9 Bangham intraparticle di ffusion plot for Hg sorption on Al-WTRs at initial Hg concentration of 40 m g/L and pH 6.5................................................................................ 53 2-10 Effect of pH on Hg sorption on Al-WTR s at initial Hg concentration of 40 mg/L. The values are averages (n=3) and error bars represent standard errors of the m ean........ 54 2-11 Effect of pH on Zeta Potential of Al-WTRs......................................................................55 3-1 Mercury concentrations in Florida sandy soil leachates of control (no Al-WTRs) and sam ples treated with Al-WTRs at app lication rates of 2.5%, 5%, and 10%...................... 82 3-2 Mercury concentrations in Florida sandy soil leachates collected from the 5th to the 13th pore volume in control (no Al-WTRs) and samples treated with Al-WTRs at application rates of 2.5%, 5%, and 10%............................................................................ 83 3-3 Cumulative mass of Hg leached per kg of soil from the contro l (no Al-WTRs) and Al-WTRs treated columns of Florid a sandy soil after 13 pore volumes........................... 84
9 3-4 Concentrations of me thyl-Hg produced in Florida sandy soil after 10 days of incubation u nder water saturation and anoxic conditions.................................................. 85 3-5 Mercury concentrations in Nevada so il leachates of control (no Al-WTRs) and samples treated with Al-WTRs at app lication rates of 2.5%, 5%, and 10%...................... 86 3-6 Cumulative mass of Hg leached per kg of soil from the contro l (no Al-WTRs) and Al-WTRs treated columns of Neva da soil after 11 pore volumes..................................... 87 3-7 Concentrations of met hyl-Hg produced in Nevada soil after 10 days of incubation under water saturation and anoxic co nditions.................................................................... 88 4-1 Mercury, Al, Fe, and Si distribution (%) in the v arious fractions of Al-WTR................ 109 4-2 The N2 gas sorption on Al-WTRs with no Hg and Hg-spiked........................................ 110 4-3 The CO2 gas sorption on Al-WTRs with no Hg and Hg-spiked...................................... 111 4-4 Mean BET (N2) and micropore (CO2) surface area of Al-WTRs with no Hg (black bars) and Hg-spiked (white bars) Al-WTRs.................................................................... 112 4-5 Micropore volume based on pore size di stribution of AlWTRs with no Hg (black bars) and Hg-spiked Al-WTRs (grey bars)...................................................................... 113 4-6 Mean micropore volume of Al-WTRs with no Hg (control) and Hg-spiked AlWTRs. ..............................................................................................................................114 4-7 Arrhenius plot of Hg sorption on Al-WTRs at 110 C, 250 C, and 350 C..........................115 4-8 Diffusivities ( D ) of Hg in Al-WTRs as a function of varying values (m ean distance between sorption sites).....................................................................................................116
10 Abstract of Dissertation Pres ented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy IMMOBILIZATION OF MERCURY IN CONT AMINATED SOILS USING ALUMINUM DRINKING WATER TREATMENT RESIDUALS By Anna Hovsepyan August 2008 Chair: Jean-Claude J. Bonzongo Major: Environmental Engineering Sciences The goal of this project is to evaluate the potential of aluminum drinking water treatment residuals (Al-WTRs) as sorbent for the immobiliza tion of mercury (Hg) in contaminated soils. The Al-WTRs are readily available and non-h azardous waste by-products of the municipal drinking water treatment process composed primarily of amorphous aluminum (hydr)oxides and are produced daily in very larg e quantities at both national and global scales. While numerous studies have evaluated the use of Al-WTRs in controlling wate r pollution from agricultural phosphorus runoffs, the potential of Al-WTRs for the immo bilization of metals in contaminated soils remains poorly studied. To our knowledge, no data on the interaction between Hg and AlWTRs exists in the peer-reviewed literature. To evaluate the use of this waste material in remediation of Hg-contaminated soils, the follo wing tasks were completed: (1) characterization of Al-WTRs and assessment of Al-WTRs potential to sorb Hg from aqueous solutions under varying conditions such as pH and contact time (2) assessment of the ability of Al-WTRs to sorb and immobilize Hg in contaminated soils using colu mn studies mimicking the effect of acid rain (3) assessment of the ability of Al-WTRs to redu ce/eliminate the bioavailab ility and toxicity of Hg from treated soils using microbial Hg methylation as a surrogate for its bioavailability and MetPLATE toxicity test, and (4) evaluation of mech anisms of Hg immobilization on Al-WTRs
11 with implications for long-term stability. The ove rarching hypothesis of th is work is that both chemical and physical characteristics of Al-WTR s would favor the sorption, and therefore, the immobilization of Hg by formation of Hg-[Al-WTRs] complexes that are highly stable under environmental conditions. Results suggest that Al -WTRs have an average specific surface area of 48 m2/g and an internal micropore surface area of 120 m2/g. Obtained sorption isotherms in aqueous solutions indicated a strong affinity of Hg for Al-WTRs. Using the Langmuir adsorption model, a relatively high maximum sorption capac ity of 79 mg Hg/g Al-WTRs was determined. Sorption kinetic data was best fit to a pseudo-first order model, while the use of the WeberMorris and Bangham models suggested that the in traparticle diffusion coul d be the rate-limiting step. Also, Al-WTRs effectively immobilized Hg in the pH range of 3 to 8. The results showed that the leachability, toxicity, and methylation potential of Hg wa s reduced by the addition of AlWTRs to soils, pointing out the ability of Al-WTRs to immobili ze Hg in soils. However, the effect was more pronounced in newly-contamin ated soil compared to well-aged soil. The findings of this study suggest that it is likely that Hg was inco rporated into micropores of AlWTRs as shown by the porosity and chemical fr actionation data. Finally, future studies are needed to determine the exact binding sites of Hg on Al-WTRs. Overall the results suggest that the addition of Al-WTRs to Hg-contaminated so ils could be an efficient remediation method.
12 CHAPTER 1 INTRODUCTION: MERCURY IN SOILS AND REMEDIATION Problem Statement Heavy m etals found in the environment orig inate from both anth ropogenic and natural sources. Unlike organic pollutants, metals ar e not biodegradable and their persistence, transformation, and transfer to th e food chain can lead to negativ e effects on ecological functions and human health (Steinnes 1997). Currently, the United States Environmental Protection Agency (US-EPA) Priority Pollutant List incl udes Ag, As, Be, Cd, Cr, Cu, Hg, Ni, Pb, Sb, Se, Tl, and Zn as metals of serious environmenta l concern (Adriano, 2001). Additionally, more than 50% of the reported contaminated soil sites list ed on the US National Priority List are impacted by heavy metals (Mulligan et al. 2001). Unfort unately, the remediation of metal contaminated soils remains one of the most intractable problems of environmental restor ation (Mulligan et al. 2001), and this is both a national and international concern. It re quires the development of costeffective and efficient remedial approaches fo r metals and other contaminants, while avoiding adverse effects on the treated systems. Mercury (Hg) is a naturally oc curring element in the earths crust, with natural background concentrations in soils ranging from a few part s per billion (ppb) to le vels in hundreds of ppb (Adriano 2001; Steinnes 1997). Natu ral sources of Hg in the environment include volcanoes, hot springs, and the weathering of ro cks (DItri 1972; Drasch et al. 2004). However, concerns related to the effects of Hg on the environment and human health arise primarily from the anthropogenic inputs of this pollutant to diffe rent environmental compartments Anthropogenic activities that release Hg into the environment include, but ar e not limited to, the combustion of fossil fuels, mining, waste incineration, rubber processing, the improper disposal of commercial and industrial products containing Hg, and industrial ac tivities such as the paper and pulp industry,
13 and chlor-alkali plants (Chen et al 1999; DItri 1972; Drasch et al 2004). With regard to soils, the major sources of Hg pollution are chlor-alkali plants, mine tailings, and diffuse sources such as the burning of fossil fuels in coal-fired power pl ants (DItri 1972; Drasch et al. 2004; Steinnes 1997). Concerns over the biogeochemical cycling of Hg in the environment have been driven primarily by the toxicity of its alkyl-compounds, namely met hyl-Hg. However, both inorganic and organic Hg species represent a hazard to living organisms, even when present at trace levels in the environment (Drasch et al. 2004). The toxi city of Hg species has been the subject of abundant research (DItri 1972; Kazantzis 1980; Magos and Clar kson 2006; Patra and Sharma 2000), and although not the focus of this disser tation, it is noteworthy that studies on Hg methylation dominate the curr ent literature on Hg biogeochemistry. Despite some ongoing controversies on the role of diffe rent microbial groups and abiotic factors, processes leading to Hg methylation and methyl-Hg accumulation in so ils and sediments are quite well-understood. Methyl-Hg is a lipid-soluble organometallic compound which passes easily through cell membranes to then bioaccumulate and biomagnify in food chains (Baldi 1997; Boudou and Ribeyre 1997). As a neurotoxin, methyl-Hg attacks the central nervous syst em (Grieb et al. 1990; Lindqvist et al. 1991), and the princi pal pathway for human exposure to this toxic pollutant is the consumption of contaminated fish in which the majo rity, if not all, of Hg bioaccumulates in the methylated form. Accordingly, regul atory agencies focus primarily on fish as the target organism to protect human health and other sensitive organi sms. In fact, Hg found in aquatic systems and which ultimately accumulates in fish originates predominantly from watershed soils (Babiarz et al. 1998; Hurley et al. 1995; St. Louis et al. 1996). This is due to soils high capacity to sequester Hg and to behave as both sink and source of Hg over long periods of time (Johansson et al. 1991;
14 Piao and Bishop 2006). This is also linked to the high ratio of dry land surfaces to water covered surfaces within watersheds where atmospheric deposition is the main source of Hg contamination. Therefore, in addi tion to erosion proce sses that introduce Hg from top soils to surface waters via runoffs, hydrologic and biogeoche mical perturbations are the principal agents for Hg mobilization accumulated in surface a nd vadose zone soils and contamination of waterways. As a result of strict regula tions by the US-EPA, emissions of Hg in the US have been decreasing since the 1990s (Hylander and Meili 2003; Jackson et al. 2000; Pirrone et al. 1996), and research on the development of new soluti ons for pollution prevention as well as for remediation of sites already contaminated with Hg has been stimulated. Table 1-1 summarizes selected Hg standards enforced by either th e US-EPA or the Floridas Department of Environmental Protection (FDE P 2005; US-EPA 2003a; US-EPA 2003b). On a worldwide scale, several sites remain heavily cont aminated with Hg from both hi storic and present anthropogenic activities (Adriano 2001; Hylander and Goodsite 2006; Nriagu and Wong 1997). For instance, soil samples collected from gold mining impacted soils in North-West USA exhibit total-Hg (THg) concentrations that range from 132 mg/ kg to 635 mg/kg (Kim et al. 2003), while soils from areas surrounding a chlor-alkali plant in the Netherlands had THg concentrations ranging from 4.3 mg/kg to 1150 mg/kg (Bernaus et al. 2006). Unfortunately, the negative effects of certain pollutants such as Hg become known only several years after their initial intentional (e.g. landfills) and/or non-intentional disposal (e.g. diffuse sources) (Pacyna and Keeler 1995). Cons equently, reactive appr oaches such as the development of efficient and cost-effectiv e remediation methods remain crucial.
15 Overview of Soil Remediation Techniques and Rationale for this Study Although much is now known on the biogeochem istry of metals in soils and sediments, research on the development of cost-effec tive and environmental-friendly remediation techniques remains challenging. Various ex situ and in situ technologies have already been proposed for the remediation of Hg contaminated soils. These technologie s include conventional engineering-based methods employing techniqu es such as physical separation, thermal processes, washing, stabilization and solidification, (DIt ri 1972; Suer and Allard 2003; Wasay et al. 1995), and other emerging technologies such as phytoextraction (More no et al. 2005). While these methods can be efficient in removing Hg fr om soil, only a few of these techniques have been tested commercially, and their widespread use remains limited due to the prohibitive costs and other disadvantages discussed below. Excavation is the most traditional method for soil remediation and it involves the removal and transport of contaminated soils for disposal in hazardous landfills or for off-site treatment (McGowen et al. 2001). However, through this process there is a risk of spreading the contaminated soil and dust. Excavation could be very efficient in removing metal contaminants from very small contaminated areas. However, it becomes very expensive and impracticable when dealing with very large contaminated areas (Basta and McGowen 2004). Also, excavation of large areas should be followed by restoration of the soil, which includes backfilling of the area with clean soil and landscaping. The major disadvant age of excavation is hi gh cost, therefore for areas with no immediate risks of metal exposur e to living organisms, other methods should be considered. Soil washing involves physical separation or ch emical extraction to remove metals from contaminated soil (Dermont et al. 2008; Tsang et al. 2007). Physical separation processes that could be used specifically for Hg involve techniqu es such as screening an d gravity concentration
16 and separation (US-EPA 1995). However, the efficiency of physical separation depends on various factors that include metal speciation, soil te xture, moisture content, heterogeneity of soil matrix, etc. (Dermont et al. 2008). The treatment is not effective if there is a high variability of metal chemical forms or if there is a high si lt/clay or high humic c ontent in treated soils (Dermont et al. 2008; Tsang et al. 2007). Chemical extraction uses fluids to solubilize metals from soils and to allow the transfer of released metals from the solid soil phase to the aqueous phase (Dermont et al. 2008). Various extracting fluids such as aci ds/bases, surfactants, chelating agents, salts, or redox agents have been proposed (Dermont et al 2008; Ehsan et al. 2006). Acid solutions are extracting agents that could be used for chemical extraction of Hg (Vanbenschoten et al. 1994). Soil leachin g with strong oxidizing agents such as Cl2 and OClcould be used to destroy organic complexes of Hg and to transf er Hg to the liquid phase (Szpyrkowicz et al. 2007). However, these strong oxidizing agents may al so oxidize minerals that are in the reduced form and are part of the soil structure (Tessier et al. 1979). Similar to the physical separation, the effectiveness of this technique depends on meta l speciation and soil quali ty including texture, buffering capacity, and organic matter conten t (Hamby 1996; Tsang et al. 2007). Another disadvantage of this process is the production of large volumes of hazardous liquid wastes that require further treatment and safe disposal (US-EPA 1997a). Electro-remediation involves movement and subsequent removal of ionic metal forms from the soil in an imposed electric fiel d (Altin and Degirmenci 2005; Cox et al. 1996; Hakansson et al. 2008; Page and Page 2002). For Hg remediation, as a pretreatment iodide/iodine complexing agent could be used to mobilize Hg (Hakansson et al. 2008). Similar to soil washing, this process creates liquid wastes which require subsequent treatment (Hakansson et al. 2008). One of the limitations of this technique is that it is effective only when the moisture
17 content of the soil is between 14 and 18% (Hamby 1996). Another disadvantage is that reduced immobile forms of Hg present in soils can beco me mobile, and if not removed properly, they can lead to toxicity in contact with living organisms. Also, electro -remediation gives poor results in soils with low permeability (e.g., high clay content), which limits the transfer of Hg from solid to aqueous phases (Suer an d Lifvergren 2003). Methods for removal of elemental Hg from soil include amalgamation by adding aluminum, copper or other metals to soils. This ability of Hg to form amalgams with certain metals allows its removal from the bioavailable Hg pool, and therefore, from the methylation/demethylation cycle (DItri 1972). Volatilization or thermal desorption is a nother method for removal of Hg from soils (Chang et al. 1998; Kunkel et al. 2006). This process requires speci al equipment at the contaminated site to collect Hg vapor so that it does not contaminate the ambient air (Takeuchi et al. 2001; US-EPA 1997a; US-EPA 2007). The encapsulation process involves either mi xing of the soil with cementing agents or glassification of the soil in or der to encapsulate the contam inated area (DItri 1972; Hamby 1996). One of the key disadvantages of encapsulation is that it pe rmanently restricts future soil use. Phytoextraction, which makes use of plants to remove Hg and other metals from contaminated soils, has economic advantages over traditional engineer ing-based remediation methods (Barocsi et al. 2003; Garbisu and Alkorta 2001; LeDuc and Terry 2005; Meagher and Heaton 2005). However, to be efficient, hyper-a ccumulator plant species are needed and they should posses certain traits such as ability to tolerate metals, high biomass production, rapid growth, high bioaccumulation of the targeted me tal (Bismarck et al. 2000; Weatherford et al.
18 1997). Another disadvantage of this method is that although many heavy metal cations can be taken up and accumulated in plant roots, transloc ation of these cations to the shoots is often limited (Huang et al. 1997; Wang and Greger 2006) To enhance the bioavailability of metals and hence the uptake by plants, chelating agents su ch as EDTA may be added to the soil (Barocsi et al. 2003; Huang et al. 1997; St anhope et al. 2000). Only a few studies have proposed the use of phyto-extraction for Hg (Meagher and Heat on 2005; Moreno et al. 2005). In a study by Moreno et al. (2005), Hg uptake by Brassica juncea (Indian mustard) was facilitated by the application of sodium thiosulphate to the soil However, it was found that Hg volatilized from above ground portion of plants, ra ising questions on the implications of Hg emission (Moreno et al. 2005). An alternative to the conventional engineeri ng-based methods is immobilization of Hg in soil via adsorption to render it harmless. Ad sorption processes are of great importance in determining the fate and transport of Hg and othe r metals in soils (Sarkar et al. 2000). The main approach of Hg immobilization is to stabilize Hg thus reducing its bioavailability and consequently the environmental risks. The proce ss is inexpensive depending on what adsorbents are used to stabilize Hg. A number of adsorbents such as activat ed carbon, zeolites, silica gels, clays, and ion-exchange resins (Chen et al. 2004; Chiarle et al. 2000; Huang and Blankenship 1984; Mercier and Detellier 1995; Pr ado et al. 2002; Zamzow et al 1990) have been proposed to remove Hg from solutions, while only a few adsorb ents, such as used ti re rubber, have been proposed for Hg adsorption in soils (Meng et al. 1998). Even though used tire rubber was reported to have a relatively high adsorption capacity for Hg (14.6 mg/g), it is not known how its application would affect the overa ll soil quality. In the search fo r an ideal adsorbent for Hg
19 immobilization, drinking water treatment residuals (WTRs) are considered in this doctoral research. Drinking Water Treatment Residuals (WTRs) Alum inum and iron based WTRs are the mo st common by-products of drinking water treatment facilities. WTRs are formed as a result of addition of aluminum or iron salts and other coagulants (such as polymers) to raw water to re move colloids, silt and clay-size particles, and color (Dayton and Basta 2001; Makris and Harris 2006). WTRs are produced daily during drinking water treatment processes and are either disposed in landfills, stored in onsite lagoons, or simply discharged into the river systems (N ovak and Watts 2004). The disposal of this waste product is expensive and may incr ease the overall cost of wate r purification (Novak and Watts 2004). Extensive research has shown that as a re sult of their chemical composition and high reactivity, WTRs can potentially be used as am endments to increase so il phosphorus (P) sorption capacity and reduce the impacts of runoff and infi ltration on water quality (Codling et al. 2000; Dayton et al. 2003; Silveira et al. 2006). Based on laboratory studies, it ha s been suggested that the mechanism of P sorption onto WTRs matrices tends to be biphasic, with an initial fast sorption reaction on the surface due to electrosta tic interactions between P and WTRs, followed by a slower sorption step associat ed with the intraparticle diffusi on of P into micropores (Makris et al. 2004b). With regard to P sorption kinetics, reports on reaction order ar e not consistent. In a study by Novak and Watts (2005), P sorption data on WT Rs were best fit to a first-order reaction model, while P kinetic data by Makris et al. (2005) fitted best to a second-order reaction model. This discrepancy can be attributed to factors such as the lack of homogeneity and the difference in the composition of WTRs used in these two studies. However, these authors concur on the strong ability of WTRs to sorb P.
20 Overall, previous research has now establis hed the strong ability of WTRs to immobilize anions such as phosphate (Dayton et al. 2003), fluoride (Sujana et al. 1998), and perchlorate (Makris et al. 2006). Unlike the above anionic sp ecies, Hg and most metals occur in soil solutions primarily as cations. Reports on the e fficiency of WTRs in binding heavy metals are still lacking. Some preliminary efforts, such as the work by Brown et al. (2001) found that WTR amendments in soil reduced the NH4NO3 extractable Cd, Pb, and Zn, suggesting that WTRs might have a strong ability for immobilization of these metals of environmental concern. Although WTRs are found abundantly as Feor Al-bas ed materials, it is believed that the use of Fe-WTRs in the remediation of Hg contaminated soils is less attractive due to the highly redox sensitive nature of iron, which could result in iron dissolution under anoxic conditions and release of previously sorbed metal cations. Th erefore, in this study we will focus only on aluminum-based WTRs (Al-WTRs). Dayton and Basta (2001) investigated the prospe ct of seventeen WTRs (fourteen of which were Al-WTRs) to be used as soil substitu tes by comparing physicochemical properties and nutrient composition of the WTRs with typical so ils. They found that most of the studied AlWTRs had soil-like qualities with the potential to be used as so il substitutes for land reclamation purposes (Table 1-2). The physicochemical propert ies of WTRs were found adequate for plant growth (Table 1-3). If used solely for crop growth, the drawback of WTR application would be the reduction of P bioavailabilit y, which could lead to P deficiency in plants if applied at excessively high rates (>10% WTRs) (Dayton and Basta 2001). However, this shortcoming can be easily corrected by applying a P-fertilizer to the treated soil whenever necessary (Hyde and Morris 2000).
21 Dayton and Basta (2001) also evaluated the potential toxi city of Al-WTRs using the toxicity characteristic leaching procedure (TCLP), a US-EPA standard procedure used to profile waste materials as hazardous or non-hazardous for the purpose of landfilling. The heavy metal content of all studied Al-WTR s was significantly below the re gulatory levels for TCLP and therefore categorized as non-haza rdous waste. Indeed, metal con centrations of WTRs are not regulated by the US-EPA 40 CFR Part 503 for sewage sludge dis posal (US-EPA 1997b). Studies have shown that metal concentrations of studied WTRs were lower than the U.S. EPA regulatory limits for metals of environmental concern (Table 1-4) (Hyde and Morris 2000), and the dissolved Al concentrations in Al-WTRs extracts did not produce toxicity symptoms in soybeans and corn (Dayton and Basta 2001). Studie s also demonstrated that Al-WTRs did not cause acute toxicity to the aquatic invertebrate Daphnia similis in a 48 hour toxicity assay (Sotero-Santos et al. 2005). In addition, no evid ence of aluminum toxicity was found when alumbased sludge was used as a growth medium (Babatunde and Zhao 2007). Adsorption capacity, non-toxic nature, ease of us e, low cost, and availability are the factors that should be taken into account when choosing an adsorbent for soil Hg remediation (Stoodley et al. 2002). The results of our re search demonstrate that Al-WTRs, as Hg adsorbents, seem to fit the above mentioned criteria. An added advantage to this adsorbent is th at it exhibits soil-like qualities. Research Objectives and Dissertation Outline The purpose of this doctoral research is to ev alu ate the potential of using Al-WTRs in the remediation of Hg-contaminated sites. While nu merous studies have ev aluated the use of AlWTRs in controlling water pollution from agricultural-P runoffs, the potential of this material for the immobilization of metals in contaminated so ils has not been adequately studied. Based on a
22 thorough review of the peer-reviewed literature no data on the interaction between Hg and WTRs exist. Experiments consisted primarily of laboratory assays to determine Hg sorption capacities of Al-WTRs, the ideal Al-WTRs rate for Hg imm obilization in soils, redu ction of Hg toxicity and bioavailability, and the potential for the stability of formed Hg-complexes. The overarching hypothesis of this study is that both chemical and physical compositions of Al-WTRs favor the sorption a nd therefore the immobilization of Hg by formation of Hg-[AlWTRs] complexes that are high ly stable under environmental conditions. Therefore, Al-WTRs can be used to immobilize Hg in soils where Hg levels have been increased above natural background by anthropoge nic activities. The study was designed to meet the following main objectives. Objective 1: Evaluate the potential of Al-WTRs to sorb Hg from aqueous solutions through determination of the Hg sorption capacities of Al-WTRs and identify the effect of contact time and pH on Hg adsorption. Objective 2: Assess the potential of Al -WTRs to sorb and immobilize Hg in contaminated soils using column studies mimicking the effect of acid rain. Objective 3: Assess the ability of Al-WTRs to re duce/eliminate the bioavailability and toxicity of Hg from treated soils using microbial Hg methyl ation as a surrogate for its bioavailability and MetPLATETM, a microbial test that is spec ific for heavy metal toxicity. Objective 4: Initiate an investigation of the me chanisms of Hg immobilization on AlWTRs and the implications of such mechanisms on the long-term stabil ity of formed Hg-[AlWTRs] complexes.
23 Table 1-1. Selected Hg standards for different environmental compartments Florida Soil Clean Up Target Level Direct Exposure (mg/kg)c Maximum Contaminant Level (mg/L) a Hazardous Waste Disposal (mg/L) b Commercial/ Industrial Residential Florida SCTL Leachability based on groundwater criteriac (mg/kg) 0.002 0.2 17 3 2.1a a Established under the Safe Drinking Water Act (US-EPA 2003a); b Based on TCLP leachate concentration, if greater th an the limit, land disposal is prohibited (US-EPA 2003b); c Soil Clean Up target levels (SCTL) vary by states (FDEP 2005)
24Table 1-2. Comparison of chemical and physic al properties of WTRs with typical so il levels (from Dayton and Basta, 2001) WTR pH Electrical conductivity (dS/m) Cationexchange capacity (cmol/kg) Total N (g/kg) Organic carbon (g/kg) Bulk density (g/cm3) Plant available water (g/kg) 1 7.1 0.63 56.5 10.1 80 0.58 134 2 7.7 0.54 46.7 7.1 75 0.74 301 3 7.0 1.09 18.8 18.4 128 N/A 416 4 7.8 0.60 51.0 8.2 69 0.81 142 5 7.8 1.08 44.2 12. 65 N/A 144 6 7.6 0.37 13.6 1.3 17 N/A 172 7 7.7 0.44 34.8 3.9 28 0.95 71.8 8 6.6 0.28 29.6 7.6 32 0.91 130 9 7.0 0.27 20.3 5.6 46 0.79 27.3 10 6.9 0.40 29.5 4.8 62 0.82 26 11 7.7 0.59 29.9 2.3 48 1.17 206 12 5.3 0.43 31.7 5.9 78 0.63 16.3 13 7.5 1.03 29.7 14.6 149 0.56 194 14 7.2 0.67 30.5 7.9 86 0.93 139 15 7.0 0.22 17.8 7.3 60 0.96 100 16 7.0 0.80 31.9 7.9 63 0.97 77 17 6.6 0.22 16.4 2.8 23 1.3 66.8 WTR RANGE 5.3-7.8 0.22-1.1 13.6-56.5 1.3-18.4 17-149 0.56-1.3 26-416 WTR MEDIAN 7.1 0.5 30.0 7.0 63 0.9 139 SOIL TYPICAL 5.0-8.0 <4.0 3.5-35.6 0.2-5.0 <30 1.0-1.55 63-300
25Table 1-3. Comparison of WTR nutrient levels with soil nutrient levels adequate for gr owth of most crops (from Dayton and Basta, 2001) WTR Soluble P, G/L Olsen P, mg/k g M III P, mg/kg NH4N, mg/kg NO3-N, mg/kg K, mg/kg Mg, mg/kg Fe, mg/kg Zn, mg/kg Ca, g/kg 1 180.5 17.04 6.81 71.1 18.7 197 56.2 8 0.55 2.63 2 96.1 6.35 3.57 68.2 66 79 341.2 50.1 5.2 21 3 240.4 8.56 5.89 55.8 5.3 42.7 102.9 85.9 70 0.6 4 227.7 24.74 12.2 69.8 7.3 76 252.7 61.5 1.5 2.81 5 274.2 23.24 9.93 101.5 56.1 186 438 110.3 3 7.23 6 48.55 4.37 6.53 22 5.7 216.6 750 56.2 1.1 4.93 10 34 4.9 6.53 47.4 3.5 94.4 40.2 60.4 2.3 2.2 11 66 7.7 2.3 25.2 35.3 161.4 788 23.4 17:9 15.3 12 79 4.0 1.6 117.8 11.9 58 8.0 34.8 0.12 0.18 13 116.6 47.1 29.7 140 123 68.3 56 58.8 1.3 9.1 14 576 49 54.4 40.6 50 206.2 278.6 101 19.8 3.5 15 49.6 8.9 24.3 41.5 41.3 109 474 34.3 2.9 1.8 16 53.2 17.7 23.3 63.1 16 102.4 89 22 23.4 1.1 17 69.2 15 16.8 26.9 13.9 19 117 89.8 4 1.1 WTR RANGE 34-576 4-49 1.654.4 22-140 3.5-123 19-278 81231 8-231 0.12-70 0.1821 WTR MEDIAN 98 13.1 6.8 51 17 109 117 60.4 3 2.6 ADEQUATE SOIL LEVEL 50-200 >12 32.5 TOTAL INORGANIC N 50-200 125 50 4.5 0.8 >0.38
26 Table 1-4. Average concentrations of 15 metals in Al-WTRs determined using the EPA method 3050. Samples were collected ev ery 3 weeks over a period of 1 year (from Hyde and Morris 2000) Analyzed Metal Concentration Range (mg/kg) Mean (mg/kg) Median (mg/kg) SD Regulation Limits a (mg/kg) Ag 0.25-2.7 1.08 0.65 0.79-As 1.9-10.3 5.8 4.3 3.0 41 Ba 66.2-353 175.5 185.8 77.5-Be 0.03-2.18 0.81 0.78 0.60-Cd 0.03-9.31 1.56 0.26 3.0439 Cr 8.1-253 105.8 63.1 103 1200 Cu 120-475 247.5 232.2 113 1500 Hg 0.01-0.12 0.04 0.03 0.0317 Ni 8.1-62 29.2 18.8 21.3420 Pb 1.5-13.5 6.6 6.6 3.3 300 Sb 0.03-0.80 0.13 0.03 0.24-Se 0.05-5.9 2.0 1.7 1.7 36 TI 0.03-0.25 0.08 0.05 0.07-V 14.0-130 57.3 34.0 45.12800 Zn 44.4-206 90.4 70.9 50.3 a Regulation limits are the US-EPA 40 CFR Part 503 concentration limits for land application of sewage sludge
27 CHAPTER 2 ALUMINUMBASED DRINKING WATER TREA TME NT RESIDUALS (AL-WTRs) AS A SORBENT FOR MERCURY: MATRIX CH ARACTERIZATION AN D DETERMINATION OF SORPTION CAPACITIES Introduction Mercury (Hg) is one of the 13 m etals in th e US-EPA priority pollutant list with known negative impacts on both human health and ecological functions (Adriano, 2001). Its strong ability to bioaccumulate and biomagnify in food chains leads to toxicity in living organisms. The major sources of anthropogenic Hg in soils includ e chlor-alkali plants, mine wastes, and Hg from several diffuse sources such as coal-fired pow er plants (DItri 1972; Drasch et al. 2004). Although much is now known on the biogeochemistry of Hg in sedimentary environments, research on the development of cost-effective and environment-friendly remediation techniques remains challenging. So far, only a few remediati on techniques (e.g. stabilization/solidification and soil washing/acid extraction) have been test ed commercially. However, their widespread use remains limited due to several factors including their prohibitive costs (US-EPA 2007). In fact, the remediation of metal-contaminated soils rema ins one of the most in tractable problems of environmental restoration (Mulliga n et al. 2001), and this is both a national and international issue. It requires the developmen t of cost-effective and efficient re medial approaches that render Hg and other metal contaminants harmless, while avoiding adverse effects on the treated systems. Drinking-water treatment resi duals (WTRs) are waste by-pr oducts of the drinking water treatment processes and are produce d daily in large quantities in most municipalities worldwide. They are formed as a result of the addition of al uminum or iron salts to raw water in order to remove colloids, silt and clay-size particles, co lor, and pathogenic microorganisms (Dayton and Basta 2001; Makris and Harris 2006 ; Ward et al. 2005). Accordingly, WTRs consist of particles
28 that settle as a result of coagulation and flocculation processes. In the United States and in Europe, several million tons of WTRs are produ ced yearly (Babatunde and Zhao 2007; Prakash and Sengupta 2003). Previous studies have focu sed primarily on the ability of WTRs to immobilize negatively charged ions such as phosphate (Dayton et al. 2003), fluoride (Sujana et al. 1998), and perchlorate (Makri s et al. 2006). Studies on the e fficiency of WTRs in binding cations are still lacking. Some pr eliminary efforts, such as th e work by Brown et al. (2005) found that WTR amendments in soil would reduce the NH4NO3-extractable Cd, Pb, and Zn, suggesting that WTRs might have the capacity to immobilize these metals in soils. In this study, the potential of the inexpensive and readily available waste Al-WTRs material to sorb and immobilize Hg from aqueous solutions was assessed in batch experiments. Although WTRs are found abundantly as either Alor Fe-based materials, the limited choice to Al-WTRs in this study stems from the fact that Fe-based WTRs could be prone to redox driven changes. This could potentially lead to the di ssolution of previously immobilized metals, and therefore negate the potential for WTRs use in soil remediation. In this study, the maximum sorption capacities, sorption isotherms, as well as the effect of pH on Hg sorption on Al-WTRs were determined. Kinetic and intr aparticle diffusion models were used to gain insight in the potential sorption mechanisms of Hg on tested Al-WTRs particles. Materials and Methods Collection and Characterization of Al-WTRs Al-WTRs samples were collected from th e Manatee County Drinking W ater Treatment Plant in Bradenton, Florida, USA. In this water treatment plant, Al-WTRs is produced following the addition of alum and a small amount of copolymers of sodium acrylate and acrylamide (Makris et al. 2005). The Al-WTRs samples used in this study were collected from the Treatment Plant disposal site (open dry land in the vicinity of the water treatment plant) using a shovel.
29 Samples were transferred into high density po lyethylene (HDPE) contai ners and transported back to the University of Florida in Gainesvi lle. In the laboratory, Al-WTRs samples were first air-dried at room temperature for a period of 4 weeks. Relatively homogeneous material was then obtained by passing dried Al-WTRs through a 2 mm screen. The pH and the electrical conductivity (EC) of Al-WTRs were measured in 1:1 (mass/volume (m/v)) and 1:2 (m/v) Nanopure water suspensions respectively, after a 4 hr equilibration period using a pH-meter (m odel 240, Corning) and EC meter (model 1054, Markson), respectively. The organic carbon conten t was measured according to the WalkleyBlack method in which organic carbon is oxidized by potassium dichromate (Walkley and Black 1934). This method is one of the most widely used procedures for obtaining a good estimate of oxidizable orgaic matter in soil samples (Hesse 1972). The effective cati on exchange capacity (CECe) which is a measure of the quantity of sites on soil surfaces that can retain cations by electrostatic forces, was determined as desc ribed by Sumner and Mill er (1996). Briefly, CECe was determined by extracting base cations (Ca2+, Mg2+, K+, and Na+) and aluminium (Al3+) with ammonium chloride, the concentrations of wh ich were then determined by the Inductively Coupled Plasma Atomic Emission Spectrometry (I CP-AES). The toxicity of Al-WTRs extracts was assayed with the MetPLATETM toxicity test, after equilibra tion by shaking of the mixture with Nanopure water in a 1:2.5 solid to liquid ratio for 2 hours (Boularbah et al. 1996). MetPLATETM is a metal specific toxicity test that is based on the inhibition of the enzyme galactosidase by metals at toxic levels in a mutant strain of E. coli, and does not respond to organic toxicants (Bitton et al. 1994). For total me tal analysis, about 1 g of dry Al-WTRs sample was digested overnight at 1100C with 30 ml of HNO3/H2SO4 mixture (7:3, v/v) in a closed Teflon vessel. The mixture upon cooling wa s diluted to 50 ml with Nanopure water. The
30 solution was then analyzed for total concentra tions of Al, Fe, Ca, Cu, Cr, Zn, and Pb by ICPAES and Hg by Cold Vapor Atomic Fluorescence Spectrometry (CV-AFS). An emphasis in the characterization of us ed Al-WTRs was on the measurement of the specific surface area (SSA), a key parameter that strongly influences the sorption capacity of solid surfaces (Goldberg et al 2001). As a pretreatment to SSA characterization, a known amount of Al-WTRs was introduced in a capillary glass tube and outgassed for 4 hours under helium flow at 700C (Makris et al. 2004a). The SSA was th en measured by a nitrogen adsorption (SSA-N2 at 77 K) and carbon dioxide (SSA-CO2 at 273 K) methods using Quantachrome Autosorb-1 (Quantachrome Corp.) apparatus. The SSA-N2 was calculated using the BrunauerEmmett-Teller or BET equation (Equation 2-1), where W is the weight of the gas adsorbed at a relative pressure P/P0, Wm is the monolayer capacity, and C is a BET constant. 0 0)( 1 1 )1( 1 P P CW C CW P P Wm m (Eq. 2-1) The SSA-CO2, which describes microporosity, was calculated using the DubininRadushkevich (DR) equation modified by Kaganer, also known as DRK equation (Equation 2-2) (Gregg and Sing 1982), and where W is the amount of the gas adso rbed at relative pressure P/P0, Wm is the monolayer capacity, D is a constant that character izes Gaussian distribution, P0 is the vapor saturation pressure of CO2 (26,140 mm Hg), and P is the equilibrium pressure (mm Hg). 2 0log loglog P P DWWm (Eq. 2-2) The BET-N2 method is applied in the regi on of relative pressures from P/P0=0.05 to P/P0=0.3, while the SSA-CO2 is carried out in the interval of relative pressures of P/P0=10-5 to P/P0 =0.0029) (Gregg and Sing 1982).
31 Finally, the zeta potential (ZP) of Al-WTRs as a function of pH was measured using the EKA electrokinetic analyzer (Anton Paar). The EKA is used to determine the ZP of larger (noncolloidal) particles and is based on the streaming potential method. The EKA automatically calculates the ZP by using the Helmholtz-Smoluchowski equation (AntonPaar 2003). The pH of the electrolyte solution (1mM KCl) was adjusted with 0.1 M HCl or 0.1M NaOH to the desired value in the pH range of 3 to 11. The electrolyte solution was pumped through a cylindrical cell that has two silver electrodes coated with AgCl. The ZP is then calculated based on the streaming potential, the flow of the solvent (pressure drop), the current, and the conductivity of the solution as the solvent flow s across the samples. More detailed information about this methodology can be found elsewher e (Bismarck et al. 2004). Determination of the Maximum Sorption Ca pacity of Al-WTRs and Sorption Isotherms The maximum sorption capacity of Al-WTRs fo r Hg was determined in batch adsorption experiments. A commercial Hg(NO3)2 standard solution obtained from Fisher Scientific was used to prepare Hg solutions in Nanopure water with final Hg concentr ations of 10, 20, 40, and 80 ppm. Obtained solutions were put in contact with dry Al-WTRs materials in a 3 to 5 ratio (m/v). All experiments were conducted at pH 6.5 and carried out in 50 ml capped and acid-cleaned polyethylene tubes which were continuously ro tated at about 30 rpm on a Roto-Shake Genie (Scientific Industries, Inc). After 96 hours of e quilibration, Al-WTRs slurries were centrifuged and the supernatant filtered (0.45m) and anal yzed for total Hg. Total Hg concentration in solution was measured by CV-AFS, following sample digestion according to EPA method 1631 (US-EPA 2001). Briefly, filtered samples were subjected to cold digestion using bromine monochloride (BrCl) and analyzed us ing the stannous chloride (SnCl2) reduction method (USEPA 2001) and a Tekran Series 2600 system (Tek ran, Ontrario, Canada). QA/QC criteria were met by running reagent blanks and standard soluti ons. The concentration of 10 analytical blanks
32 averaged 0.72 ng/L (n=10), and the detection lim it determined as 1 standard deviation was 0.5 ng/L. All containers used in this study were acid washed and only Teflon containers were used for storage of Hg containing solutions. The Hg loading capacities of Al-WTRs (mg/ g) were calculated us ing equation 2-3, and where Ci and Ce are initial and equilibrium Hg concentrations in mg/L, M is the weight of used Al-WTRs in grams, and V is the volume of solution in liters. M V) C(C (mg/g) Capacity Hg Loadingei (Eq.2-3) The percentage of Hg sorbed was calculated from the difference be tween the initial Hg concentration and Hg concentration remaining in the solution at equilib rium. The sorption data was then fit to Langmuir and Freundlich models and the best fit was used to calculate the maximum sorption capacity. The solid Al-WTRs material used in these sorption experiments was recovered and examined for sorbed Hg via scanning electron microscopy (SEM), carried out in the JSM-6330F field emission scanning electron microscope unit equipped with an X-ray energy dispersive spectrometer (SEM-EDS). To acco unt for Hg mass balance, the solid Al-WTRs material was also analyzed for total Hg by CV-AFS. Mercury Desorption from Al-WTRs For desorption study, Hg-spiked Al-WTRs we re prepared by bringing Al-WTRs into contact with Hg solution to achieve a Hg con centration of 30,000 mg/kg and equilibrated for 7 days (no shaking was applied). The amount of Hg sorbed was determined by analyzing the dried Hg-spiked Al-WTRs for total Hg concentration. Desorption of Hg was studied by first adding dried Hg-spiked Al-WTRs in a 1:20 (m/v) to an ex tracting solution with a pH of 4.22, prepared according to the US-EPA Synthetic Precipita tion Leaching Procedure (SPLP) (US-EPA 2003b)
33 and to simulate the leaching effect of acid rain. Next, the samples were rotated in capped centrifuge tubes at a bout 30 rpm for 18 hours on a Roto-Shake Genie, centrifuged, and the supernatant analyzed for total Hg as described in earlier in this chapter. The percentage of Hg desorbed was calculated using equation 2-4, where Cdes is the desorbed Hg concentration in SPLP extracting solution (mg/L), Cads is the Hg sorbed concentration on Al-WTRs (mg/kg), V is the volume of extracting solution (mL), and M is the weight of Al-WTRs (g). %100 Desorbed Hg % ads desCM V C (Eq. 2-4) Kinetics of Hg Sorption on Al-WTRs and Effect of pH Batch kinetic studies were performed to obtai n data on Hg sorption behavior versus time. Hg solution with an initial concentration of 40 ppm was used for this experiment. The Hg solution and Al-WTRs were mixed in a 2 to 5 ratio (m/v), pH adjusted to 6.5, and the mixture agitated by continuous stirring for the duration of the experiment. The contact time and sampling intervals were selected based on preliminary expe riments. Thus, at pre-decided time intervals, the supernatant was withdrawn, filtered (0.45m), and analyzed for total Hg. To investigate the possible mechanisms of sorption, ob tained data were fit in to di fferent kinetic (first and second order) and intraparticl e diffusion models. Finally, in addition to the above kinetic experiment at a fixe d pH, batch studies were also conducted to investigate the effect of changi ng pH on Hg sorption onto Al-WTRs. In these sorption experiments, 40 ppm Hg so lutions were used and the pH of Al-WTRs slurries (3:5 ratio; m/v) adjusted with either 0.1N HCl or 0.1N NaOH. The Hg solutions were then equilibrated with Al-WTRs for 96 hours. After the equilibration period, the supernatant was analyzed for total Hg.
34 Results and Discussion Characterization of Al-WTRs Images obtained from scanning electron micr oscopy showed that co llected Al-WTRs are heterogeneous mixtures of partic les with irregular shape and va riable sizes (Figure 2-1). The EDS elemental spectra analysis showed that Al-W TRs are predominantly composed of Al, Si, P, S, Ca, and Fe (Figure 2-2(A)). Other measured physicochemical characteristics of Al-WTRs are presented in Table 2-1. Overall, these parame ters are similar to Al-WTRs composition reported by others in previous studie s (Dayton and Basta 2001; Dayton et al. 2003; Hyde and Morris 2000; Makris et al. 2004b; Makr is et al. 2005). Although not re gulated by the US-EPA 40 CFR Part 503 for sewage sludge (US-EP A 1997a), it was confirmed that the concentration of Hg (0.02 mg Hg/kg Al-WTR) and that of ot her metals present in these Al-WTR samples were lower than the US-EPA regulatory limits for land applica tion of sewage sludge (T able 2-1) (US-EPA 40 CFR Part 503). With regard to high Al levels and its potential leachability and toxicity, the MetPLATETM bioassay showed no measurable toxic effect of Al-WTRs extracts. This observation is in line with prev ious studies where no aluminum toxicity was found in tomatoes (Dayton and Basta 2001) and broad beans (Skene et al. 1995) when grown in Al-WTRs used as a soil substitute. The SSA-N2 characterization based on the BET method (BET-N2) revealed a rather high surface area averaging 48 m2/g. The BET-N2 is applicable mostly to non-porous or mesoporous materials and tends to underestim ate the SSA of internal micropores when present (Lowell et al. 2004). The CO2 is a preferred adsorbent in micropore analysis even though CO2 (2.8 ) and N2 (3.0 ) have similar molecular dimensions (Low ell et al. 2004; Makris et al. 2004b). The reason being that CO2 analysis is performed at a hi gher temperature compared to N2 (T=273 K for CO2 and T=77 K for N2), as a result CO2 passage through micropores is less constricted (Gil and
35 Gandia 2003; Lowell et al. 2004). Our measured SSA-CO2 of 120 m2/g Al-WTR suggests that the Al-WTRs materials tested in this study have a large internal surface area that is not accounted by the BET-N2 (Table 2-1). The values for SSA-N2 and SSA-CO2 compare quite well with those reported for Al-WTRs collected from the same water treatment plant by Makris et al. (2004b; 2005). Determination of the Maximum Sorption Ca pacity of Al-WTRs and Sorption Isotherms Table 2-2 summarizes the results obtained from these experiments. The Hg loading capacities and the percentage of Hg adsorbed on to Al-WTRs calculated fo r each treatment show that Al-WTRs can efficiently remove dissolv ed Hg. Samples with 10 mg/L of initial Hg concentration had Hg levels reduced by almost 100% while solutions with initial Hg levels of 20 mg/L were reduced by about 90%. The maximum sorption capacity of 75.4 mg/g was obtained based on experiments with initia l Hg dissolved concentration of 80 mg/L. The EDS elemental spectra analysis of Al-WTR particles recovere d at the end of the Hg sorption experiments verified the presence of sorbed Hg on Al-WTR -particles (Figure 2-2(B)), and mass balance calculation on the distribution of Hg between the aqueous and solid phases accounted for 99 10% of the initial Hg concentrations. Additionall y, the shape of the obtained sorption isotherm (Figure 2-3) suggests that sorp tion sites with high affinity for Hg are present on Al-WTRs, leading to very low aqueous Hg equilibrium c oncentrations for samples with 10 and 20 mg/L initial Hg concentrations (Polo and Utrilla 2002). The sorption data was best fit to a Langmuir model (Figure 2-4, R2=0.98). Based on equation 2-5, a plot of Ce/q versus Ce allowed the determination of the maximum sorption capacity ( qm) and the binding constant ( Kads) from the slope and the intercept of the linear regression. In this equation, q is the adsorption density (mg/g), Ce is the equilibrium Hg
36 concentrations (mg/L), qm is the maximum sorption capacity (mg/g), Kads is the binding constant that measures the affinity of adsorbate for adsorbent (L/mg) (Sawyer et al. 2003). m e mads eq C qKq C 1 (Eq. 2-5) The maximum sorption capacity determined from the Langmuir model was 79.3 mg/g, which compares well with the maximum so rption capacity of 75.4 mg/g determined experimentally (Table 2-3). This sorption capacity is considerably la rger than that of some of the previously tested waste adsorbents such as us ed tire rubber with a maxi mum sorption capacity of 14.6 mg Hg/g (Meng et al. 1998). Additionally, the favorable natu re of sorption processes can be expressed in terms of a dimensionless constant separation factor (r) defined as: r=1/(1+KadsC0) where C0 is the initial Hg concentration (mg/L) and Kads is the binding consta nt from equation 2-5. For this study, the calculated (r) values were less than 1 and greater than 0, indicating a favorable sorption (McKay et al. 1982). It is worth noting that the Langmuir model assumes that the adsorption is limited to a single monolayer coverage and that all surface sites on the adsorben t have the same affinity for the adsorbate (Sawyer et al. 2003) Accordingly, this model app lies well to these short-term sorption experiments, dominated primarily by sorp tion of Hg to Al-WTRs external surface sites (Axe and Trivedi 2002). However, for microporous ad sorbents such as Al-WTRs, there is also a possibility for the intraparticle diffusion of Hg into the micropores. Therefore, it is likely that the true sorption capacity of Al-WTRs is even grea ter than the value obtained in these short-term sorption experiments. Intraparticle diffusion is a much slower sorption process and is observed by maintaining a constant boundary condition of a metal in the bulk aqueous phase (Xu et al.
37 2006). Further long-term experiments are theref ore needed to study the potential diffusion mechanisms of Hg into Al-WTRs micropores. Mercury Desorption from Al-WTRs Finally, Hg desorption studied by leaching of formed Hg-[Al-WTRs] complexes with synthetic acid rainwater solution (SPLP) led to the release of just 1.5% of Hg previously sorbed onto Al-WTRs particles. Although preliminary, thes e results suggest that the leaching potential of Hg from Al-WTRS could be very low. Kinetics of Hg Sorption on Al-WTRs Figure 2-5 shows the effect of contact time on Hg adsorption on Al-WTRs in a series of sorption experiments with initial Hg concentrat ion of 40 mg/L and pH 6.5. Equilibrium was attained in about 32 hour s, with nearly all in itial dissolved Hg (~99% ) sorbed onto Al-WTRs particles. Based on these studies, kinetic and intraparticle diffusi on models were used to gain insight in the potential mechanisms of Hg sorption and fixation on Al-WTRs. Pseudo-first order Lagergren model: The Lagergrens pseudo-first order equation (Lagergren 1898) is widely used for modeling sorption of metals from solutions (Goel et al. 2005; Gupta and Sharma 2002; Manc hon-Vizuete et al. 2005). In lin ear form it is expressed as follows 303.2 log)(log1 10 10tk q qqe te (Eq. 2-6) where qe is Hg concentration at equilibrium per unit mass of Al-WTRs (mg/g), qt is the amount of Hg adsorbed at time (t) per unit mass of Al-WTRs (mg/g), and k1 is the sorption rate constant of the pseudo-first order (h-1). By plotting log(qe-qt) versus time ( t) a straight line with a correlation coefficient of R2=0.9944 was obtained (Figure 2-6). The parameters k1 and qe determined from the slope and the intercept of th e plot are presented in Table 2-4. Based on the
38 heterogeneous composition of Al-WTRs, it is likel y that several reaction types/mechanisms are involved in the fixation of Hg on Al-WTRs. However, the use of equation 2-6 tends to suggest that the rates of these reactions could be approximated by a pse udo-first order kinetic model. Pseudo-second order model: Figure 2-7 was obtained by plotting t/qt versus time ( t ), based on equation 2-7 (Ho and McKay 2000), where qe2 is the Hg concentra tion at equilibrium per unit mass of Al-WTRs (mg/g), qt is the amount of Hg adsorbed at time (t ) per unit mass of Al-WTRs (mg/g), and k2 is the sorption rate constant (g/mg h). 2 2 221ee tq t qkq t (Eq. 2-7) The parameters qe2 and k2 obtained from the intercept and the slope of the plot are listed in Table 2-4. The correlation coefficien t for this pseudo-second order model (R2=0.9901) is slightly lower when compared to the pseudo-first order model, suggesting that the adsorption of Hg by Al-WTRs under these experimental conditions can be described by a pseudo-first order kinetic model. Intraparticle diffusion: The potential role of intrapartricle diffusion on Hg sorption process was first explored by using the Weber-M orris intraparticle diffusion model shown in equation 2-8 (Ho and McKay 1998; We ber and Morris 1963), and where, qt is the amount of Hg adsorbed at time (t) per unit mass of Al-WTRs (mg/g), ki is the intraparticle diffusion rate constant (mg/g h), and C is the sorption constant, which describes the thickness of the boundary layer. Ctkqit (Eq. 2-8) Figure 2-8 was obtai ned by plotting qt versus t and the obtained relationship (R2=0.9915), seems to indicate that intraparticl e diffusion may play a role in the sorption
39 process. Indeed, Figure 2-8 can be divided in two distinct linear portions. The first portion could be attributed to either boundary layer diffusion effects or the mass transf er effects on external surfaces, while the second portion tends to sugge st an additional slower and gradual sorption stage during which intraparticle diffusion would likely dominate (Mane et al. 2007). The role of intraparticle diffusion as a rate-limiting step on sorption process was further investigated by use of the Banghams model shown in equation 2-9 (Aharoni et al. 1979; Tutem et al. 1998), and where C0 is the initial concentration of adsorbate in solution (mg/L), V is the volume of solution (mL), m is the weight of adsorbent used per liter of solution (g/L), qt is the amount of Hg adsorbed at time (t) per unit mass of Al-WTRs (mg/g), (<1) and kb are constants. )log( 303.2 log loglog0 0t V mk mqC Cb t (Eq.2-9) A plot of the Banghams model is presented in Figure 2-9. The linearity of the plot (R2=0.9948) confirms the applicability of Bangham s equation and indicate s that intraparticle diffusion is a likely rate-limiting step of Hg sorption on Al-WTRs (Jain et al. 2004). The mechanism of intraparticle diffus ion is also supported by the fact that Al-WTR particles have a large internal network of micropores as mentioned in earlier sections. In general, the amorphous structure of Al-WTR s is believed to be caused by the formation of amorphous Al oxides and hydroxides (Mak ris et al. 2005; Nova k and Watts 2005). The concentrations of amorphous aluminum (hydr)o xides in Al-WTRs have been found to range from 50 to 150 g/kg (Dayton and Basta 2001). Theore tically, the prevalence of such geochemical phases in Al-WTRs would make them behave like aluminum oxyhydroxides. Results from previous studies tend to suggest that sorption of heavy metals to amorphous metal oxides is usually a two-step process with a rapid sorption of the metal first to the external surface, followed by a slow intraparticle diffusion along the oxide micr opore walls (Trivedi and Axe
40 2000). For remediation purposes, it is the long-term slow proce ss which might play the most significant role in the sorption of the Hg fraction that become s non-exchangeable, and therefore fully immobilized. Further studies are necessary to validate these observations and to help determine surface diffusivities. Effect of pH on Hg sorption by Al-WTRs: In interfacial phenomena involving liquids, pH is a key parameter governing the surface ch arge of solid materials (Stumm and Morgan 1996). Although Al-WTR is a hetero geneous material, its reactivity has been found to be similar to that of amorphous aluminum (hydr)oxides. The reported point of zero charge (PZC) of aluminum (hydr)oxides is usually greater than 7.7 (Gayer et al. 1958; Goldberg et al. 2001). Theoretically, the affinity of Hg and other meta l cations for (hydr)oxides and therefore Al-WTRs should increase for pH greater than pHzpc and decrease at pH < pHzpc, based on electrostatic interactions. The effect of th e pH of the solution on the sorption of Hg on Al-WTRs is presented in Figure 2-10. These results show the highest Hg removal from aqueous phase at the highly acidic pH of 3, while the lowest Hg removal le vel was observed at pH 5. Overall, Hg sorption decreased first from pH 3 to 5, and then increased gradually with increasing pH from 5 to 8. This trend is not supported by the empi rical PZC concept described above. Zeta potential (ZP) of the Al-WTRs was measured across a wide range of pH (3 to 11) to tentatively help explain the unusual sorption trend shown in Figure 2-11. Usua lly, the ZP should become more positive with the decrease in pH because of the build up of positively charged protons. The experimental data show that as the pH decreases from about 5.5 to 3, the ZP of Al-WTRs becomes more negative. This unusual trend could explain the high Hg sorption at pH 3. Similar ZP trends have been reported for several organic polymers such as ce llulose and lignin, the most abundant organic polymers in nature (Bismarck et al. 2001; Bismarck et al. 2000). The organic carbon (OC)
41 content of used Al-WTRs averaged 13%, and besides the anthropogeni c inputs due to the addition of organo-compounds during the treatment process, OC originates from natural raw waters that undergo the treatme nt process. This high OC content as well as the chemical composition of these organic compounds may be in fluencing the ZP vs. pH trend of Al-WTRs. The fact that Hg removal was very efficient at low pH values is signif icant from remediation standpoint, because often times contaminated soils ma y have a very low pH and, if left untreated, may leach out the Hg. As the pH increases from about 5.5 to 11, th e ZP becomes more negative, because of the build up of negatively charged hydroxyl ions. Figur e 2-10 and Figure 2-11 show correlation of Hg sorption behavior vs. pH on one hand, and be tween the measured ZP vs. pH, indicating that electrostatic attractions play a role in Hg sorption onto Al-WTRs. However, as the pH increases above about pH 4, Hg shoul d occur predominantly as 0 )(2)(aqOHHg based on geochemical equilibrium predictions calculated using MINEQL+ (Schecher and McAvoy 2001). Therefore, it is likely that electrostatic attraction does not c onstitute the only force that governs the sorption of Hg onto Al-WTRs. These data suggest that additi onal forces are involved in Hg sorption. One could speculate on the role of steric and hydrophobic interactions in the Hg sorption process on Al-WTRs. However, further invest igations are needed to study the mechanism of Hg sorption on Al-WTRs. Conclusions Al-WTRs are readily availabl e and non-hazardous waste by-products of drinking water treatment processes, produced da ily and in large quant ities in most municipalities worldwide. They have been used successfu lly in experimental settings to sorb and immobilize phosphorus from agricultural impacted soils. However studies on retention of metal cations such as Hg by
42 Al-WTRs are lacking. In this study, the potential of Al-WTRs to sorb and immobilize Hg from aqueous solutions was evaluated. Sorption isotherms indicated a strong affinity of Hg for AlWTRs and a relatively high maximum sorption ca pacity of 79 mg Hg/g Al-WTRs. Also, AlWTRs effectively immobilized Hg in the pH range of 3 to 8. Sorption kinetic data was best fit to a pseudo-first order model, while the use of the Weber-Morris and Bangham models suggested that the intraparticle diffusion could be the rate-limiting step for Hg immobilization onto AlWTRs particles. Overall, the results from these short-term experiments demonstrate that AlWTRs can be effectively used to remove Hg from aqueous solutions. This ability points to the potential of Al-WTRs as sorbent in soil reme diation techniques ba sed on Hg-immobilization.
43 Table 2-1. Physicochemical pr operties of Al-WTRs collected from the Bradenton Drinking Water Treatment Facility (Florida, USA) Parameter Mean Value a Units Regulatory Limit f (mg/kg) pHc 5.6.01b n/ag ECd 0.36.01 dS/m n/a CECe 45.80.09 cmol/kg n/a SSA-N2 (BET) 48.3 m2/g n/a SSA-CO2 (micropore) 120.3 m2/g n/a Organic Carbon 12.7.08 % n/a Al 73.8.2 g/kg Ndh Fe 3.7.1 g/kg Nd Ca 2.263.5 g/kg Nd As 8.01.1 mg/kg 41 Cu 141.4 mg/kg 1500 Cr 81.1.3 mg/kg 1200 Hg 0.02.003 mg/kg 17 Pb 1.99.4 mg/kg 300 Zn 14.37.3 mg/kg 2800 a All the values are means of triplicates; b values represent standard errors of the mean; cAt a soil/water ratio of 1:1 (m/v); dAt a soil/water ratio of 1:2 (m/v); e Effective cation exchange capacity; f US-EPA 40 CFR Part 503, pollutant limits for meeting land exceptional quality criteria; g Not applicable. h Not defined
44 Table 2-2. Determination of Hg maximu m sorption capacity (mg/g) at pH 6.5 Initial Hg (mg/L) Mean Equilibrium conc. (mg/L) a Loading capacity (mg/g) % Hg adsorbed 10 0.03.01b 16.6 99.7 20 2.01.33 29.9 90.0 40 8.72.37 52.1 78.2 80 45.13 75.4 56.6 aAll the values are mean s of four replicates. b values represent standa rd errors of the mean Table 2-3. Sorption isotherm parameters for Hg sorption on Al-WTRs at pH 6.5 Table 2-4. Kinetic Parameters for Hg Sorp tion on Al-WTRs at pH 6.5 and initial Hg concentration of 40 mg/L Parameter Value Parameter Value Pseudo-first order Pseudo-second order k1 (h-1) 0.057 K2 (g/mg h) 9.12*10-4 qe (mg/g) 80.2 qe2 (mg/g) 103 R2 0.9944R2 0.9901 Intraparticle diffusion Weber and Morris Model Bangham Model ki (mg/g h) 0.0677 kb (mL/g/L) 297 C 0.1328 0.7375 R2 0.9915R2 0.9948 Parameter Value Kads (L/mg) 0.4 qm (Langmuir) 79.3 qm experimental (mg/g) 75.4 R2 (Langmuir) 0.9858
45 Figure 2-1. SEM micrographs of the original Al-WTRs collected from the Bradenton Drinking Water Treatment Facility (Florida, USA) showing irregular size and nonhomogeneity of particles.
46 A B Figure 2-2. Elemental spectra (EDS) of the (A) original Al-WTRs before use in Hg sorption experiments and of the (B) recovered Al-WTRs after the so rption experiments confirming Hg sorption on Al-WTRs.
47 Hg Equilibrium Concentr ation in Solution (mg/L) 0510152025303540Hg Sorbed Amount (mg/g) 10 20 30 40 50 60 70 80 90 Figure 2-3. Mercury sorpti on isotherm with initial Hg concentr ations of 10, 20, 40, and 80 mg/L and pH 6.5. Plotted values are averages (n=4 ) and error bars represent standard errors of the mean. Error bars for 10, 20, and 40 mg/L are very small and overlap with the point.
48 C e (mg/L) 01 02 03 04 0C e / q 0.0 0.1 0.2 0.3 0.4 0.5 R2=0.9858 Figure 2-4. Langmuir plot for Hg sorption on Al-WTRs at initial Hg concentrations of 10, 20, 40, and 80 mg/L and pH 6.5. Where q is the adsorption density (mg/g) and Ce is the equilibrium Hg concentration (mg/L).
49 time (hours) 020406080100120140160180Hg Sorbed Amount (mg/g) 0 20 40 60 80 100 120 Figure 2-5. Effect of cont act time on Hg sorption on Al-WTRs at 40 mg/L initial Hg concentration and pH 6.5. The values are averages (n=2) and error bars represent standard errors of the mean.
50 time (hours) 05101520253035log( q e q t ) 1.0 1.2 1.4 1.6 1.8 2.0 R2=0.9944 Figure 2-6. Lagergren, pseudo-firs t order kinetic plot for Hg so rption on Al-WTRs at initial Hg concentration of 40 mg/L and pH 6.5. Where qe and qt are Hg concentrations at equilibrium and at time (t) (mg/g).
51 time (hours) 05101520253035t/qt 0.0 0.1 0.2 0.3 0.4 0.5 R2=0.9901 Figure 2-7. Pseudo-second order kinetic plot for Hg sorpti on on Al-WTRs at initial Hg concentration of 40 mg/L and pH 6.5. Where qt is the amount of Hg adsorbed at time (t ) per unit mass of Al-WTRs (mg/g).
52 123456q t (mg/g) 0 10 20 30 40 50 60 70 80 90 R2=0.9915Portion 1 Portion 2 t Figure 2-8. Weber and Morris intr aparticle diffusion plot for Hg sorption on Al-WTRs at initial Hg concentration of 40 mg/L and pH 6.5. Where qt is the amount of Hg adsorbed at time (t ) per unit mass of Al-WTRs (mg/g).
53 log(t) 0.20.40.60.81.01.21.41.6log log( C o /( C o q t m ) -1.2 -1.0 -0.8 -0.6 -0.4 -0.2 0.0 R2=0.9948 Figure 2-9. Bangham intr aparticle diffusion plot for Hg so rption on Al-WTRs at initial Hg concentration of 40 mg/L and pH 6.5. Where C0 is the initial concentration of adsorbate in solution (mg/L), m is the weight of adsorbent used per liter of solution (g/L), and qt is the amount of Hg adsorbed at time (t) per unit mass of Al-WTRs (mg/g).
54 pH 23456789Hg (%) Remaining in Solution 0 10 20 30 40 50 Figure 2-10. Effect of pH on Hg sorption on Al -WTRs at initial Hg concentration of 40 mg/L. The values are averages (n=3) and error bars represent standard errors of the mean.
55 pH 1234567891011Zeta Potential (mV) of Al-WTRs -11 -10 -9 -8 -7 -6 -5 -4 -3 -2 Figure 2-11. Effect of pH on Zeta Potential of Al-WTRs.
56 CHAPTER 3 IMMOBILIZATION OF MERCURY IN CONTAMINATED SOILS BY ALUMINUMBASE D DRINKING WATER TREATMENT RESIDUALS (Al-WTRs) Introduction The average background concentration of mercur y (Hg) in pristine soils in the United States has been estimated to be about 0.11 mg/kg (Adriano 2001; Bringmark 1997). The introduction of Hg to the environment through diff erent anthropogenic activi ties increases levels of this toxic metal in the atmosphere and in bot h terrestrial and aquatic systems. Because of the well-established ability of Hg to bioaccumulate and biomagnify in food chains (Adriano 2001), its introduction into natural systems and the result ing environmental and health implications have been the subject of numerous investigations wo rldwide (Adriano 2001; Lacer da et al. 1995; Sigel and Sigel 1997). With regard to soils, anthropogenic Hg originates mainly from sources such as mine tailings, chlor-alkali plants, and diffuse sources such as emissions from coal-fired power plants and waste incinerators (DI tri 1972; Drasch et al 2004). For example, soils collected from gold-mining impacted sites were reported to cont ain total Hg (THg) concen trations ranging from 132 mg/kg to 635 mg/kg (Kim et al. 2003), while soils from areas surr ounding a chlor-alkali plant had THg concentrations ranging from 4. 3 mg/kg to 1150 mg/kg (B ernaus et al. 2006). Overall, Hg contamination of soils remains a worl dwide problem due to either historic or present anthropogenic activities (Biester et al. 2002; Gray et al. 2002; Nriagu and Wong 1997). The toxicity and bioavailability of Hg in soil depends greatly on Hg speciation (Bringmark 1997). Mercury present in soils occurs primarily in three different oxidation states: 0, +1, and +2 (Alloway 1995). In its elemental form (Hg0), Hg is poorly retained onto soil particles, and is therefore easily lost through volatilization, inf iltration, and horizontal transport mechanisms (Bringmark 1997). Although not abundant Hg is also present in soil in the +1 oxidation state in a number of Hg-minerals. While Hg2I2 is considered the most stable Hg(I)-mineral, several other
57 Hg(I)-minerals exist and most are very soluble. Examples include the following in decreasing order of solubility: Hg2SO4 > Hg2CO3 > Hg2(OH)2 > HgHPO4. In fact, Hg2SO4, Hg2CO3, and Hg2(OH)2 are so soluble that they ra rely precipitate in soils (Lindsay 1979). In contrast, Hg(II)compounds are not only more stab le but also the most common a nd the most abundant in soils (Alloway 1995; Martinez and McBride 1998). Besides the prevalen t occurrence of Hg compounds in the inorganic form, soil-Hg occurs also in alkylated form s. Unlike the inorganic Hg species, alkylated Hg species, namely methyl -Hg, are produced prim arily through microbial methylation of soluble and bi oavailable inorganic Hg-speci es, while the potential for nonbiological (abiotic) Hg-alkylation has been also reported under specific laboratory conditions (Adriano 2001; Alloway 1995; Bringmark 1997; Celo et al. 2006). Mercury speciation and mobility in soils are greatly affected by certain key environmental parameters such as pH and redox conditions (All oway 1995). The pH and pe of most soils range from 3.5 to 9 and -6 to +12; respectively (Bourg and Loch 1995) The pH is one of the most important parameters that can influence Hg adso rption to soil particles and thus its fractionation and mobility (Mulligan et al. 2001; Yin et al. 1996). In general, metals are more mobile and bioavailable under acidic conditions (Alloway 19 95). In soil solution with acidic pH, positively charged (e.g. HgCl+) and neutral (HgCl2) species are formed. As th e pH increases, hydroxo-Hg species such as Hg(OH)2 become more predominant (A driano 2001). Mercury chloride compounds (e.g. HgCl2, Hg2Cl2) and hydroxo-Hg species (e.g., Hg(OH)2 and HgO) are predominant under oxidizing conditions (Adria no 2001). Under moderately oxidizing to reducing conditions and for pH >5, inorganic Hg can also be found in the elemental (Hg0) form (Schuster 1991). In addition, redox processes can induce strong acidification of soils because oxidized components are more aci dic than their redu ced counterparts (Bourg and Loch 1995).
58 Under reducing conditions in soils with high sulfur content, Hg is precipitated as Hg sulfide, which has a very low solubility (Adriano 2001; Sc huster 1991). If sulfur is not available under such reducing conditions, the Hg fraction that is bound to Fe a nd Mn (hydr)oxides in soils may be mobilized and released to soil pore wate r (Bourg and Loch 1995). Studies suggest that reducing conditions are the most favorable for th e formation of methyl-H g by anaerobic bacteria (Adriano 2001; Chen et al. 1997; Compeau and Bartha 1984). The speciation of Hg in soils is rather complex, and besides the effects of pH and redox discussed above, it is also affected by the interacti ons of Hg with various ligands present in soils. In fact, concentrations of free metal ions in so il solutions are mostly determined by levels of binding to the solid ph ase (Weng et al. 2001), and would therefore be a function of soil physicochemical composition (Bernaus et al. 2006; Lestan et al. 2003). This specific aspect controls metal adsorption-deso rption reactions which occur mainly due to chemical bond formation, complex formation, and ion-exchange (Lestan et al. 2003). A ccordingly, adsorption and desorption processes are of a special importance in pollutant dynamics in soils as they tend to control the fate and transport of metals including Hg in soils (Sarkar et al. 2000; Yin et al. 1996). In addition to the interactions with the solid phase, metal ions can also form complexes with dissolved organic matter and several other dissolved ligands. Therefore, the quantity and quality of dissolved organic carbon (DOC) as well as the type and abundance of inorganic soluble ligands in soil pore water tend to enhanc e both the solubility and mobility of Hg (Adamo et al. 2002, Weng 2002). In contaminated soils DOC would likely act as vehicle for Hg infiltration from the surface and vadose zone soils to the groundwater. Overall, Hg in soils may occu r as: (i) dissolved (free ion or soluble complex), (ii) nonspecifically adsorbed (binding main ly due to electrostatic forces), (iii) specifically adsorbed
59 (strong binding due to covalent or coordinativ e forces), (iv) chelat ed (bound to organic substances), and (v) precipitated (e.g. sulfides carbonates, hydroxides, etc) species (Schuster 1991). With regard to the remediation of metal-co ntaminated soils, several methods have been proposed. The advantages and disadvantages of thes e methods have been discussed in chapter 1, and the in-situ immobilization of Hg in soil s through addition of sorbents is believed to be a costeffective alternative to most conventional remediation t echniques (Gray et al. 2006). Immobilization techniques do not remove Hg from soils; but they reduce Hg mobility and bioavailability in-situ through adsorption proces ses. Preliminary investigations discussed in chapter 2 point to the ability of Al-WTRs to efficiently immo bilize Hg that may leach from contaminated soils. In addition, Al-WTRs are waste by-products of water treatment facilities widely distributed nationwide. Ther efore, they can be considered as low low-cost sorbents with a very high Hg adsorption capacity of ~79 mg Hg/g Al-WTRs. In natural systems, rain is the dominant sour ce of water, and the decr ease in soil pH as a result of acid rain could increa se the mobility and bioavailabil ity of Hg (Goyer et al. 1985; Schwedt 2001). In this study, th e potential of Al-WTRs to so rb and immobilize Hg from contaminated soils was assessed using column st udies mimicking the effect of acid rain. The ability of Al-WTRs to reduce/eliminate the bioavailability and toxicity of Hg was evaluated through biogeochemical and toxicological approaches. Materials and Methods The Al-WTRs and Soil Sample Collection, Characteriz ation, and Site Description The Al-WTRs samples used in this study were collected from the Manatee County Drinking Water Treatment Plant in Bradenton, FL The material was first air dried at room temperature for 4 weeks, and then sieved through a 2 mm screen. Se lected physiochemical
60 properties such as pH, electrical conductivity (E C), effective cation exch ange capacitiy (CEC), and total metal concentrations were determined. A detailed description of different characterization methods used in this study has been described ear lier (chapter 2). Two types of Hg-contaminated so ils were used to investigat e the ability of Al-WTRs to immobilize Hg from contaminated soil matrices. Th e first soil was an initially non-contaminated but Hg-spiked soil used in studies evaluating th e efficiency of Al-WTRs for Hg immobilization in a freshly contaminated and low adsorption capacity soil. The noncontaminated soil was collected from the top 3 feet of the E horiz on at the McCarty Woods of the University of Florida campus. This sandy soil was chosen because it was expected to have a high Hg leaching potential. The second type of soil used in this stu dy came from a historically contaminated site in the Carson River watershed in Nevada. Mercury in this soil was introdu ced over a century ago during the 30 years of intense gold and silver mining activities of late 1800s (Gustin et al. 1994). Amalgamation with Hg was used to extract gold and silver from the ores. Unfortunately, the negative effects of Hg introduced to the Ca rson River watershed by mining activities became known as a serious environmental issue only severa l decades after the end of the intense use of Hg in gold and silver extraction by amalgamati on processes. Thus, Hg accumulated in mine tailings has been redistributed throughout the Carson River watershed resulting in some of the highest environmental Hg levels reported in North America (Gustin et al. 1994). As a result, soils, surface waters, sediments, fish and wildlife in the Carson River basin are contaminated or have been affected by Hg contamination (Bonzongo et al. 1996; Lechler et al. 1997). Because of these elevated Hg levels, the Carson River watershed was placed on the National Priorities List by the US-EPA in 1990 (Gustin et al. 1994). This Hg -contaminated site offers the opportunity to obtain soil samples in which Hg has been in contact with soil particles for several decades and to
61 use such soil in comparison with newly cont aminated materials through Hg-spiking. Soil samples from the Carson River Basin were collec ted with a shovel from the top 1 foot layer and placed into plastic bags. The soil samples were then shipped overnight in coolers to the University of Florida in Gainesville. In the laboratory, all soil samples used in this study were first air-dried at room temperature and then sieved through a 2-mm screen to remove coarse debris. The noncontaminated sandy soil was spiked with a HgCl2 solution added in excess of the soil saturation level, and the mixture of soil and Hg solution was left to equilibr ate for 7 days in capped plastic containers before being air dried again. At this point, the Hg contaminated soils were well homogenized and aliquot samples taken for aci d digestion and determ ination of total-Hg concentration as described later. The pH and the electrical conductivity (EC) of Al-WTRs and soils were measured in 1:1 (mass/volume (m/v)) and 1:2 (m/v) Nanopure water suspensions respectively, after a 4 hr equilibration period using a pH-meter (model 240, Corning) and an EC meter (model 1054, Markson). The organic carbon content was meas ured according to the Walkley-Black method (Walkley and Black 1934). The effective cation ex change capacity was determined as described by Sumner and Miller (Sumner and Miller 1996). Particle si ze distribution of soils was determined according to the USDA Soil Survey Lab Method (USDA 1992). For total metal analys is, about 1 g of dry soil sample was digested overnight at 1100C with 30 ml of HNO3/H2SO4 mixture (7:3, v/v) in a closed Teflon vessel. Upon cooling, the mixture was diluted to 50 ml with Nanopure water and analyzed for total metal concentrations by inductively coupled plasma atomic emission spectrometry (ICP-AES) and Hg by cold vapor atomic fluorescence spec trometry (CV-AFS).
62 The analysis of crystalline mineral phases of clay fraction of the Nevada soil was conducted by X-ray diffraction (XRD) in the soil mineralogy labor atory in the Department of Soil and Water Sciences, University of Florida on a computer-controlled X-ray diffractometer equipped with a stepping motor and a graphite crystal monochromator. Samples were scanned from 2-600 2 at a rate of 2 degrees per minute using CuK radiation (Yong et al. 2001). Column Leaching Studies Experimental Design: Soil columns used in this study were custom-made out of clear PVC (3.8 cm internal diameter, 30 cm length), equipped with a 2 cm drainage hole at the base and covered with netting and a la yer of glass wool to prevent soil loss and to minimize the dead end volume. The columns were packed with 300 g of contaminated soil mixed with Al-WTRs to reach a final concentration of 2.5%, 5%, and 10% of Al-WTRs (mass Al-WTRs/mass soil). In columns with 2.5% and 5% treatments, Al-WTR s was uniformly mixed with the entire soil column. The 10% treatment included two incorp oration schemes: (i) uniform mixing of AlWTRs with the entire soil mass placed in the co lumn, and (ii) Al-WTRs added as a bottom layer at the base of the column. The latter will be refe rred to from now on as the liner treatment. The experimental design resulted in four Al-WTRs treatments and a control (i.e. no Al-WTRs addition) and is summarized in Table 3-1. Soil columns were vertically held on a wooden rack placed on a lab bench. Studies using the Floridas sandy soil we re run in triplicates, while experiments with the Nevada soil were based on duplicate due to limite d availability of soil samples. Leaching Protocol: The US-EPA Synthetic Precipitat ion Leaching Procedure (SPLP) solution (EPA method 1312) obtai ned by dissolving 40 g of HNO3 and 60 g of H2SO4 into 2 liters of Nanopure water with a final pH of 4.22.05, was used as leaching solution to mimic the effect of acid rain (Townsend et al. 2005). The pore volume of packed columns was
63 determined by first saturating the soil within a bottom plugged column with the leaching solution and then draining out the column. The pore vol ume was then calculated as the difference between the volume of leaching solution added initially to the column and the volume of leachate drained out of the colu mn. For the two soils used in this study, the determined pore volumes averaged 75 and 85 ml for the Nevada and Florida soil, respectively. Using the SPLP solution, each column was leached with 1 pore volume once a week to mimic the wet and dry cycles for rainy versus non rainy days. Following each leaching event, soil leachates were collected in acid pre-cleaned Teflon containers, immediately filtered through 0.45 m filters, and aliquots of the filtrate analyzed for (i) tota l-Hg concentrations by CV-AFS and (ii) toxicity using the MetPLATETM bioassay (Bitton et al. 1994). The Al concentrations and pH of the leachate were also measured to make sure th at Al-WTRs was not a major source of soluble aluminum due to the acidic pH of the SPLP solution used. Differences in Hg concentrations in leachat es obtained from different treatments were evaluated by use of the Independent-Sample t-test with SPSS statistical software (vs. 11.0). Tests were performed at a confidence level of 95%. At the end of the leaching experiments, soil columns were taken apart and the column contents air dried in plastic cont ainers for post-leaching analyses of the solid phase. One gram of the material from each column was digested as described earlier and an alyzed for total Hg by CV-AFS. For the columns where Al-WTRs was used as liner, soil materials from the column and sample Al-WTRs from the liner were remove d, air-dried and analyzed separately.
64 Toxicity of leachates: The metal toxicity of soil leachates was assessed only on leachates collected from leaching with the first and last pore volumes by using MetPLATETM toxicity assay (Bitton et al. 1994). MetPLATETM is a toxicity test specific to heavy metals and is based on the inhibition of the -galactosidase enzyme by metals at toxic levels in a mutant strain of Escherichia coli (Bitton et al. 1994). The a ssay was performed in triplicate with four dilutions which allowed the determination of the EC50. First, 0.1 mL of bacterial reagent suspension was added to 0.9 mL of soil leachates and to negati ve control in culture tubes which were then incubated at 35oC for 90 minutes. Moderately hard water (60 mg/L Ca, 60 mg/L Mg, pH = 7.47.8) served as the negative control and it was al so used to prepare the four dilutions of soil leachates. After the incubation, 0.2 mL from each tube was pipetted to a 96-well microplate and 0.1 mL of rehydrated MetPLATETM chromogenic substrate was added to each well. The microplate was then incubated at 35oC until the purple color develope d in the negative control. The -galactosidase activity was evaluated by the conversion of chlorophenol red galactopyranoside to chlorophenol red, which was determined by measuring absorbance with a microplate spectrophotometer (Maxline Micropl ate Readers, Molecu lar Devices, Sunnyvale, CA) at 570 nm. The EC50 values of soil leachates were then calculated based on four dilutions using a regression analysis. The toxicity unit (TU) (Bitton 1998 ) was used to express the me tal toxicity. The TU relates to EC50 as shown in equation 3-1, and the hi gher the TU, the higher the toxicity. 50EC 100 TU (Eq. 3-1) This approach allows the expressi on of experimental data in te rms of percent toxicity removal calculated as follows:
65 100 TU TU TU Removal Toxicity %control treatment control (Eq. 3-2) and where TUcontrol is the toxicity unit for leachate of control soils (no Al-WTRs added) and TUtreatment is the toxicity unit for leachate of soil with the Al-WTRs treatment. Assessing the Bioavailability of Mercury in Al-WTRs Treated Soils At the end of the leaching experiments, aliquot soil samples were taken from the columns and used in laboratory incubations to assess the po tential of the Al-WTRs treatment to limit the methylation potential of Hg. The premise behind this approach is that the formation of methylHg could be used as a surrogate for bioavailability. To promote the production of methyl-Hg, thick soils slurries were prepared in precleaned 50 ml glass serum bottles using Nanopure water in a 3:1 ratio (mass/volume). The prepared slurries were first bubbled with ultra high purity nitrogen to expedite the development of anoxic conditions that favor Hg methylation (Adriano 2001), and then capped with rubber sto ppers and sealed with aluminum caps. The slurries were next incubated in the dark for 10 days at room temperature. At the end of the incubation process, Hg methylation reactions were stopped by freezing (-18oC) the samples pending methyl-Hg analysis. Methyl-Hg was first separated from the soil matrix by extraction with 15 ml of an acidic solution containing a mixture of 18% (w/v) KBr and 5% (v/v) H2SO4, and 3 ml of a 1M CuSO4 solution. After shaking for 1 hour at 200 rpm and r oom temperature, the sl urry was centrifuged at 10,000 rpm for 30 minutes and the supernatant was transferred to a different clean container. Then, 10 ml of methylene chloride (CH2Cl2) were added to the supernat ant, the container tightly capped and the mixture shaken for another hour at 300 rpm at room temperature. The separation of the two liquid phases was accelerated by centr ifugation. Only 2 ml out of the 10ml of CH2Cl2 added initially to the sediment extract were tran sferred to a centrifugati on tube containing 10 ml
66 of Nanopure water. The mixture was next placed in a water bath at 650 C for a back extraction of methyl-Hg into the aqueous pha se as the organic solvent (i.e. CH2Cl2) evaporated over time. Prior to the analysis of the aqueous sample, 8 ml of the extracted methyl-Hg solution were transferred into headspace vials. Next, 0.5 ml of 2 M acetate buffer (mixture of sodium acetate and glacial acetic acid in Nanopure water, pH 5) was added, followed by the addition of 0.5 ml of a 1 wt % of sodium tetra-ethylborate (NaB(C2H5)4) solution for deriva tization, sealed and allowed to react for 30 minutes (Leermakers et al. 2005). This overall process was adapted from Bloom et al. (1997), in which Hg species are converted to highly volatile alkyl-species except for Hgo (Equations 3-3 and 3-4). H3C-Hg+ + NaB(C2H5)4 H3C-Hg-C2H5 + B(C2H5)3 + Na+ (Eq. 3-3) Hg2+ + 2NaB(C2H5)4 (C2H5)2Hg + 2B(C2H5)3 + 2Na+ (Eq. 3-4) Methyl-Hg was then determined using h eadspace gas chromatography separation of different volatile Hg species (Hg0, H3C-Hg-C2H5, and (C2H5)2Hg), followed by thermodecomposition to produce elemental Hg from the separated Hg species, and then detection by atomic fluorescence spectrometry at 253.7 nm. All tests were run in trip licates to allow the determination of statistical differences betw een treatments. The differences in methyl-Hg concentrations in treatments we re evaluated by use of the Inde pendent-Sample t-test with SPSS statistical software (vs. 11.0 ) at a confidence level of 95%. Chemical Fractionation of Mercury in Used Soils. The chemical fractionation of Hg was conduc ted only on soil colle cted from Nevada. Samples (~1g) of air-dried soil were extracte d sequentially using a procedure adapted from Tessier et al. (1979). The following fractions were targeted an d all extractions were carried out in triplicate.
67 Fraction 1 (F1), Water Soluble: One gram of soil was extracte d at room temperature with 8 ml of Nanopure water for 3 hours with continuous agit ation at a rate of 200 rpm. Following the agitation step, the mixture was centrif uged at 10,000 rpm for 30 min (Beckman J2-HS, Tritech Field Eng. Inc.) and the supernatant removed with a pipett e. Next, the residue was rinsed with 8 ml of Nanopure water for 5 minutes and the mixtur e centrifuged again at 10,000 rpm for 30 min. The supernatant was careful ly withdrawn and added to the first supernatant fraction. The combined supernatant was then filtered (0.45m) and used for the analysis of total-Hg. The solid residue was then used in the next extraction st ep. Besides the extraction step, the centrifugation, rinsing, and filtration steps were identical for all fractions (F1 to F5), except for F6. Fraction 2 (F2), Exchangeable: The residue from F1 was treated with 8 mL of 1M magnesium chloride (MgCl2, pH 7). The mixture was agitated on at 200 rpm at room temperature for 1 hour, prior to centrifugation, rinsing, and filtration as described in F1. Fraction 3 (F3), Bound to carbonates: The residue from F2 was extracted with 8 mL of 1M sodium acetate (NaOAc) soluti on with a pH adjusted to 5.0 with acetic acid (HOAc). The mixture was agitated on an orbital shaker at 200 rpm at room temperatur e for 5 hours, prior to centrifugation, rinsing, and filtration. Fraction 4 (F4), Bound to Fe-Mn oxides: The residue from F3 was extracted with 20 ml of 0.04 M hydroxylamine hydrochloride (NH2OH.HCl) in 25% (v/v) HOAc in containers placed in a water bath for 6 hours at 96 3C, with occasional agitation by hand. Upon cooling, the mixture was centrifuged, rinsed, and the s upernatant filtered as described in F1. Fraction 5 (F5), Bound to organic matter: The residue from F4 was extracted with 3 mL of 0.02 M nitric acid (HNO3) and 5 mL of 30% hydrogen peroxide (H2O2) that had pH adjusted to 2 with HNO3. The mixture was heated to 85 2 C in a water bath for 3 hours with
68 intermittent agitation by hand. An additional 3 mL of 30% H2O2 (pH adjusted to 2 with HNO3) was added and the mixture heated for an additional 3 hours at 85 2C. Upon cooling, and to prevent the adsorption of extracted Hg onto oxidized Al-WTRs components, 5 mL of 3.2 M ammonium acetate (NH4OAc) in 20% (v/v) HNO3 was added and the sample diluted to 20 mL and agitated continuously for 30 min on an orbita l shaker at 200 rpm. These steps were then followed by centrifugation and filtration as described in F1. Fraction 6 (F6), Residual: The residual fraction was qua ntified by subtracting the sum amount of Hg obtained in fractions F1-5 from the amount of Hg obtained from the total Hg analysis. The determination of total Hg in supe rnatants were conducted comparatively by inductively coupled plasma atomic emission spectrometry (ICP-AES) and cold vapor atomic fluorescence spectrometry (CV-AFS), following overn ight digestion with bromine monochloride (mixture of KBr and KBrO3 dissolved in concentrated HC l) at room temperature. Results and Discussion Column Leaching Studies: General Characteristics The main physicochemical characteristics of Al -WTRs were presented earlier in chapter 2 (Table 2-1), and were found to be rather similar to Al-WTRs composition reported in the literature by other researchers (Dayton and Basta 2001; Dayton et al. 2003; Hyde and Morris 2000; Makris et al. 2004b; Makris et al. 2005). The two soils used in this study differed in their physicochemical properties (Table 3-2). Based on the United States Departme nt of Agriculture (USDA) soil texture triangle, the Floridas soil can be characterized as sand and the Ne vada soil as sandy loam (Schwedt 2001). The average measured pH for Florida sandy soil was 6. 4, while the Nevada soil had an average pH of 7.2; falling within the 5-7 a nd 7-9 pH ranges characteristic of humid and arid regions,
69 respectively (Alloway 1995; Essing ton 2003). Thus, in arid regions rainfall is not heavy enough to leach out basic cations such as Ca2+ and Mg2+ from soils and these soils exhibit high CEC (Bourg and Loch 1995; Schwedt 2001). The high CEC of Nevada soil could be attributed to the presence of secondary clay minerals such as montmorillonite in the composition of the soil as determined by the X-ray diffraction analysis (dat a not shown) (Sparks 1995). In contrast, in humid regions, basic cations are mainly leached from the soils and the cation exchange sites are occupied by acidic ions such as hydrogen and aluminum (Bourg and Loch 1995; Schwedt 2001). Hence, pH in sandy soil of this study was lower than in Nevada soil. A total of 13 and 11 pore volumes of SPLP solution were used for the Florida and Nevada soils, respectively in leaching events that ex tended over a period of 13 weeks. The pH of leachate solutions obtained from Nevada soil varied from 7.46 to 8.39 and did not differ significantly among treatments. Thus the pH of collected leachate solutions were always >4.2 (pH of SPLP solution), indicati ng the high buffering capacity of the tested Nevada soil. The pH values of leachates obtained from Florida soil columns ranged from 5.12 to 6.48 and did not differ significantly among treatments. Used Floridas soil did not have a high buffering capacity and was characterized by a low CE C. In such soils, acid rain may cause an increase in acidity of the soil over time. Soil acidification enhances heavy metal solubility, promoting leaching into deeper soil horizons (Adriano 1992; Adriano 2001). Florida Sandy Soil Mercury leached from columns packed with Hg-spiked Florida sandy soil: Overall, Al-WTRs amended Floridas sandy soil showed consis tently low levels of Hg in leachates in comparison to control columns, i.e. columns w ith no Al-WTRs addition. In fact, throughout the whole study, the Al-WTRs treated columns leach ed out significantly lower amounts of Hg compared to the control columns (no Al-WTRs). With the first pore volume, leachates collected
70 from soil columns amended with Al-WTRs at a pplication rates of 5 and 10%, contained Hg levels 3 to 4 orders of magnitude lower than Hg levels measured in leachate obtained from control columns (no Al-WTRs) (e .g. 0.03 mg/L for the liner, 102 mg /L for control) (Figure 3-1). In fact, the first pore volume removed appr oximately 27% of Hg from control columns, corresponding to a Hg concentration about 4.5 times higher than the 2.1 mg/kg criteria for leachability of Hg to groundwater (FDEP 2005). S ubsequent leaching resulted in much lower Hg outputs, which decreased progressi vely over time with subsequent leaching step. For the first 5 pore volumes, the effect of the liner, 10% 5%, and 2.5% Al-WTRs on Hg immobilization differed significantly among each other, with the liner treatment having the lowest THg concentration in the leachate, followed by 10%, 5%, and 2.5% Al-WTRs application rates. From the time of the application of the 6th pore volume, there was no significant difference between the 2.5% and 5% treatments. However, the 10% and liner treatments performed slightly better than 2.5% and 5% treatments (Figure 3-2). Af ter the last leaching event with the 13th pore volume, the amount of Hg removed from all columns becam e lower than the action limit for Florida groundwater leachability criteria (2 .1 mg/kg) (Figure 3-3). However, even at this point, control columns still released significantly higher Hg le vels as compared to Al-WTRs treated columns (p<0.05) (Figure 3-2(A) and Figure 3-3). From all the treatments tested, the 5% appl ication rate was found i deal for Florida sandy soil. The 2.5 % may not be recommended for low adsorption capacity soils as it fails to significantly remove Hg from pore water with up to 4.28 mg Hg/L leached out by the first 85 ml (or 1st pore volume) as compared to 0.9 and 0.06 mg/L for columns with 5% and 10% Al-WTRs application rates, respectively. Even though the 10% and the liner treatments had leached out lower Hg amounts than the 5%, the actual difference between th e 10% and 5% treatments was
71 only ~0.3 mg/kg over the course of the entire experiment. This may not be significant from a pollution prevention viewpoint, and based on trends of Hg concentrations in collected leachates, one would anticipate that on a l ong-term basis, soil columns treated with Al-WTRs at 5% and 10% application rates would have a similar efficiency with regard to Hg immobilization. Although efficient, it would be costly to implement the liner approach in-situ as the excavation of the contaminated soil layer would be require d prior to the installati on of the Al-WTRs liner. Mercury was added in Florida sandy soil as HgCl2. In fact, chloride (Cl-) occurs in natural soil systems and may be regarded as one the mo st mobile and effective complexing agents for Hg (Adriano 2001; Schuster 1991). Chloride fo rms highly soluble Hg-complexes (e.g. HgCl2) resulting in limited sorption and high mobility in porous and saturated media (Adriano, 1986; Schluter 1997). This lim itation of Hg sorption by Clin soils is especially pronounced under acidic pH, where Hg complexes with chloride (e.g. HgCl+ and HgCl2) dominate, in contrast to alkaline pH conditi ons which are favorable to h ydroxo-Hg species (e.g. Hg(OH)2) (Adriano 2001). The influence of Clon Hg sorption by Al-(hydr)oxides has been investigated by several authors (Kim 1995; Meagher and Heaton 2005), and their findings support the fact that the formation of HgCl2 limits Hg sorption (Backstrom et al 2003; Kasprzyk-Hordern 2004). Because of the prevalence of amorphous aluminum (hydr)oxides in the structure of Al-WTRs, the reactivity of Al-WTRs has been found to be rather similar to reactivity of Al-(hydr)oxides. However, our results suggest that when Hg is applied to soil as HgCl2, Al-WTRs was able to immobilize a significant amount. These results point out the e ffectiveness of Al-WTRs and indicate that the behavior of Al-WTRs incorporated in soils may not be similar to that of Al(hydr)oxides. It is therefore hypot hesized that Al-WTRs would do ev en better with most particle reactive Hg species.
72 Finally, the analysis of the different leachate fr actions for dissolved aluminum resulted only in values below the detection limit (<10 g/L) of the ICP-AES used in this study. Toxicity of leachates obtained from Florida sandy soil columns: Another approach in the assessment of the efficiency of Al-WTRs to immobilize Hg from contaminated soils is to evaluate the toxicity of soil leachates. MetPLATETM has been used successfully to assess the potential toxicity of environmental samples such as raw water, wastewater, soils, sediments, and solid wastes (Bitton et al. 1994; Boular bah et al. 2006). The use of MetPLATETM on selected leachates (i.e. leachates correspond ing to the first and last pore volumes) obtained in this study helped accomplish this goal. Based on toxic ity tests performed on the first pore volume leachates, the control (no Al-WTR s added) showed the lowest EC50, which indicates the highest toxicity (Table 3-3). This result is in agreement with the results of the ch emical analyses, in that the first pore volume leachate in control treatme nts had the highest Hg concentrations. The EC50 for the control treatment was used to calculate the toxicity removal by Al-WTRs in treated soils (Eq. 3-1 and 3-2). The EC50 values for the liner treatment and 10% Al-WTRs were >100%, which implies no detectable toxicity effect. For a ll the treatments, the results indicated that the liner and 10% treatments resulted in the highest toxicity removal, followed in a decreasing order by soils with 5%, and 2.5% Al-WTR s application rates. In contra st, the toxicity of leachates resulting from the 13th (last pore volume) was too low (EC50>100%). Overall, the MetPLATETM results suggest that the application of Al-WTRs to Hg contaminated soils not only reduces the Hg concentration in leachates but also the toxicity and therefore the bioavailability of Hg in the soil leachates. Mercury bioavailability in Florida sa ndy soil treated by Al-WTRs: The effect of different Al-WTRs treatments on Hg methylation potentials in soil slurries is shown in Figure 3-
73 4. A significant difference in the potential for methyl-Hg production was observed between the control and the 3 Al-WTRs treatments (i.e. 2.5% (p=0.03), 5% (p=0.04), and 10% (p=0.03)). One average, produced methyl-Hg levels in cont rol samples were one order of magnitude higher than the Al-WTRs treated-soils (Figure 3-4). Additionally, methylated-Hg represented ~0.11% of the total-Hg concentration in control soils as compared to only 0.04 to 0.06% in the different Al-WTRs treated soils. In general, rates of Hg methylation by microbial populations in soils and sediments depend on a variety of factors in cluding pH, redox potential, temp erature, nutrient availability, and Hg speciation (Baldi 1997; Han et al. 2007). Also, the ability of a given soil to release Hg in pore waters would tend to control rates of methyl-Hg production as the concentratio ns of methyland total-Hg have been found to correlate very well in pore wa ters (Biester et al. 2000). The results of these rather short-term experiments tend to sugge st that the addition of Al-WTRs does significantly limit the production of methyl-Hg, by controlling Hg bioavailability to Hgmethylating bacteria. Nevada Soil Mercury leached from columns packed with the Nevada soil: In contrast to the Hgspiked Florida sandy soil where the amount of leached Hg decreased pr ogressively with the number of pore volumes used, the concentrati on of Hg in leachates obtained from columns packed with the Nevada soil fluctu ated within each treatments (Fi gure 3-5). On average, control columns leached out a higher Hg level than Al-WTRs treated soils. However the results were not statistically different, except for th e liner treatment (p = 0.04) (Fi gure 3-6). This could be due to the high variability of Hg leach ed out from control columns (Figure 3-6), which could be attributed to the non-uniform dist ribution of Hg in used soil. On ce again, the liner treatment was the most effective. Overall, the Hg fraction re moved from soil via leaching was very small as
74 compared to the total-Hg concentration. Theref ore, Hg downward transp ort to groundwater in this soil would be negligible as illustrated by leac hate Hg concentrations less than the 0.02 mg/L Hg maximum contaminant level for Hg under the Safe Drinking Water Act (US-EPA 2003a). Accordingly, the use of Al-WTRs in this type of soil would be recommended only for preventive purposes in case the stability of Hg-complexes pr esent in soils become disturbed by changes in key environmental parameters. The limited amount of leachable Hg indicates that Hg in Nevada soil is strongly associated with soil particles. Th e results of chemical fractionation studies carried out according to a procedure modified from Tessier et al. (1979), showed that the 91.6 mg Hg/kg of this used Nevada soil were distributed as follows: wate r soluble (0.08%), exchangeable (0.79%), carbonate (0.06%), iron/manganese bound fractions (0.16%), organic bound (2.74%), and with 96.2 % of total Hg is in the residual fraction. Hg in this Nevada soil was introduced primarily in the elemental form more than a century ago. Because of the limited rainfall amount in the arid climate characteristics of Nevada, Hg was not mo bilized immediately. Over time, a significant fraction of the initial Hg0 underwent oxidation producing reactive Hg species and resulting in the formation of stable Hg-compounds. Following the in teractions between Hg and soil constituents in combination with aging over time, the amount of leachable Hg diminished (Alexander 1995; Cruz-Guzman et al. 2003; Oorts et al. 2007). Mercury in tested Ne vada soil could be present as insoluble minerals (e.g. HgS ) or trapped in the pores of soil particles as shown in speciation studies using samples collected from sites impacted by 19th century gold mining in the Californian Sierra Mountains (Me ng et al. 1998; Yin et al. 1997). In this ca se, the authors found that 92 to 98% of total Hg was associated with fr actions representing Hg in mineral lattice and as mercury sulfides (Bloom et al. 2003; Kim 1995).
75 Finally, and similar to the Floridas soil data, the analysis of collect ed leachate fractions for dissolved aluminum resulted only in non detectable levels (detection limit <10 g/L). Toxicity of leachates obtained from Nevada soil columns: For the Nevada soil, the EC50 values could not be calculated because of the low toxicity (EC50>100%), suggesting that Hg in this soil was strongly associ ated with soil c onstituents and no t bioavailable. Mercury bioavailability in Nevada soil treated by Al-WTRs: The effect of different AlWTRs treatments on Hg methylatio n potential in tested Nevada soil is shown in Figure 3-7. Unlike the statistically signifi cant difference observed between Hg-methylation potentials in control and Al-WTRs treatments in Floridas soil, studies wi th the Nevada soil show no significant difference between Hg -methylation potentials in control and Al-WTRs treated samples. However, on average the control tr eatment had a much hi gher methyl-Hg production potential. The methylated Hg fraction in this so il versus the different types of treatments ranged from 0.008 to 0.017% of the soil to tal-Hg concentration. Overall, the strong a ssociation of Hg to Nevadas soil particles seems to control its fate w ith regard to methylation. This is illustrated by the lack of a significant differen ce in methyl-Hg levels produced in control and Al-WTRs treated soils. In addition to Hg trends in leachates and the toxicity results described earlier, these Hgmethylation data confirm the f act that Hg speciation in the Nevada soil would downplay the efficiency of the proposed Al-WTR in in-situ remediation approach. In fact, despite the high total-Hg concentration in this Nevada soil, the easily leachable fraction is very negligible, and therefore, not ideal for the investigation of th e efficiency of Al-WTRs to immobilize Hg, as stated in the objectives of this research.
76 Conclusions The mass balance of Hg in the Florida and Neva da soils is presented in Tables 3-4 and 3-5 and data are expressed in mg of Hg per 300 g of soil (amount of soil in each column). For the Nevada soil, the distribution of Hg was not uniform as illustrated by the large standard deviation of total Hg data and low % recoveries. This ma y be representative of th e actual field conditions where Hg is not evenly distributed. In contra st, Florida soil was artificially contaminated, resulting in rather equally dist ributed amounts of Hg in soil. The biogeochemical cycling of Hg in the soil is very complex and depends on a variety of factors such as soil composition, microbial processes, climate, pH a nd redox conditions (Kim 1995; Lindqvist et al. 1991). The bioavailability, mobility, and chemical r eactivity of metals are often associated with their di stribution among soil fractions (Tu et al. 2001). In general, a decrease in soil pH improves the solubility and availability of metals. Also, metals are more available in sandy soils as de monstrated in our study. The results of this study demonstrated that l eachability of Hg in Al-WTRs treated soils is significantly reduced compared to non-treated soil, despite the fact that Hg was complexed with chloride and the soil leached with an acidic solution (pH 4.2) Mercury in Nevada soil was strongly associated with soil particles and was not highly mob ile. Therefore, the application of Al-WTRs to soil where Hg is present primarily in the residual/refractory fraction may not be an efficient remediation approach, unless used for a passive and long-term pr eventive measure. The biogeochemistry and chemical speci ation of Hg in the contaminated soil should be investigated prior to choosing Al-WTRs as remediation technique.
77 Table 3-1. Experimental design for column studies with soil weight in each column of 300 g Treatment Control 2.5% 5% 10% Liner Al-WTRs ratio to soila None 1:40 1:20 1:10 1:10 Incorporation Method N/Ab Mixed uniformly with an entire column Mixed uniformly with an entire column Mixed uniformly with an entire column Applied at the bottom of a column as a liner a Mass Al-WTRs/mass soil; b N/A=Not Applicable
78 Table 3-2. Physicochemical characteristics of soils used in column leaching studies Characteristics Floridas sandy soil Nevadas soil pHa 6.4.16 7.2 0.09 Electrical Conductivitya (dS/m) 0.02.003 0.073.001 CEC (cmolc/kg) 7.7 85.4 % Organic carbon 0.08 0.21 % Sand 95.4 61.6 % Silt 2.9 31.6 % Clay 1.7 6.8 Hg (mg/kg) a 121.8b 91.6 c a Mean of triplicates one standard deviation; b Total-Hg concentration after spiking with HgCl2 solution and air drying; c Nevada soil was not spiked
79 Table 3-3. Toxicity of Floridas sa ndy soil leachate collected from the 1st pore volume as determined by the MetPLATETM toxicity test. Al-WTRs Treatment EC50 of soil leachates (% )a Toxicity units (TU) of soil leachatesc Toxicity removal (%) Bottom liner >100b <1c 100 10% >100 <1 100 5% 7.92d 12.6 98.7 2.5% 3.8.4 28.7 97.2 0% (Control) 0.1.02 1005 n/ae a Mean of 3 replicates one standard deviation; b EC50>100% implies no toxicity; c TU=100/EC50; d Standard deviation not re ported. Only one replicat e exhibited toxicity; e not-applicable
80 Table 3-4. Mass balance of Hg in column studies using Floridas sandy soil. The initial total Hg (THg) mass was 36.5 5.8 mg in each colu mn, corresponding to a concentration of 121.8 mg/kg. Soil remaining in columns was analyzed after leaching with 13 pore volumes using SPLP solution a Mean of triplicates one standard deviation; b soil leachedTHg )THg THg( erycovReHg%leaching aftersoilcolumn100 ; c soil leachedTHg )THg THg THg( erycovReHg%lining WTRsAlin leaching aftersoilcolumn100 Control (0% AlWTRs) Soil +2.5% Al-WTRs Soil +5% Al-WTRs Soil +10% AlWTRs Soil +10% Al-WTRs as liner THg left in each soil column (mg) a 23.4.0 31.7.2 29.8.7 37.5.2 25.3 .9 THg in leachate (mg) a 10.21.6 0.43.03 0.123.0 08 0.026.0 01 0.020.00 9 THg in Al-WTRs liner (mg) a N/A N/A N/A N/A 7.2.4 % Recovery 92%b 88%b 81%b 103%b 89%c
81 Table 3-5. Mass balance of Hg in column studies using Nevadas soil. The initial total Hg (THg) mass was 36.5.8 mg in each column, correspon ding to a concentration of 91.6 mg/kg. Soil remaining in columns was analyzed after leac hing with 11 pore volumes using SPLP solution a Mean of duplicates one standard deviation; b soil leachedTHg )THg THg( erycovReHg%leaching aftersoilcolumn100 ; c soil leachedTHg )THg THg THg( erycovReHg%lining WTRsAlin leaching aftersoilcolumn100 Control (0% Al-WTRs) Soil +2.5% Al-WTRs Soil +5% Al-WTRs Soil +10% Al-WTRs Soil +10% Al-WTRs as a liner THg left in each soil column (mg) a 27.5 9.2 26.1.2 33.0 8.6 25.3 11.3 22.7.9 THg in leachate (mg) a 2328 799 357 324 56.3 THg in AlWTRs liner (mg) N/A N/A N/A N/A 37411 % Recovery 75%b 71%b 90%b 69%b 62%c
82 Pore Volume of Leachate 12345678910111213 Hg leached (mg/L) 2 4 6 8 10 15 100 Control 2.5% Al-WTRs 5% Al-WTRs 10% Al-WTRs Liner Figure 3-1. Mercury concentratio ns in Florida sandy soil leachate s of control (no Al-WTRs) and samples treated with Al-WTRs at applica tion rates of 2.5%, 5%, and 10%. The 10% application rate treatmen t had two Al-WTRs incorpor ation methods: (1) Al-WTRs mixed uniformly with the entire soil in the column and (2) incorporation as a liner at the bottom of the soil column (liner treatment).
83 Pore Volume of Leachate 5678910111213 [Hg] in leachate (mg/L) 0.0 0.2 0.4 0.6 Control 2.5% 5% 10% Liner Pore Volume of Leachate 5678910111213 [Hg] leached (mg/L) 0.00 0.02 0.04 0.06 0.08 0.10 2.5% 5% 10% Liner Figure 3-2. Mercury concentra tions in Florida sandy soil leac hates collected from the 5th to the 13th pore volume in control (no Al-WTRs) and samples treated with Al-WTRs at application rates of 2.5%, 5% and 10%. The 10% applica tion rate treatment had two Al-WTRs incorporation methods: (1) Al-WTR s mixed uniformly w ith the entire soil in the column and (2) incorporation as a liner at the bottom of the soil column (liner treatment). A B
84 Control and Al-WTRs treatments Control2.5%5%10%Liner Hg leached (mg/kg) 1 2 3 4 5 25 30 35 Figure 3-3. Cumulative mass of Hg leached per kg of soil from the control (no Al-WTRs) and Al-WTRs treated columns of Florida sa ndy soil after 13 pore volumes. Error bars represent one standard deviation of the mean.
85 Al-WTRs Treatment Control2.5% Al-WTRs5% Al-WTRs10% Al-WTRs Methyl Hg (mg/kg) 0.00 0.02 0.04 0.06 0.08 0.10 0.12 0.14 0.16 Figure 3-4. Concentrations of methyl-Hg produced in Florid a sandy soil after 10 days of incubation under water saturation and anoxi c conditions. On the x-axis, the control corresponds to soil without Al-WTRs additi on, while shown percentages of Al-WTRs correspond to the different application ra tes of used Al-WTRs as treatments in column leaching experiments.
86 1234567891011 0.000 0.001 0.002 0.003 0.004 0.005 0.006 Control 2.5% 5% 10% Liner [Hg] mg/LPore Volume 1234567891011 0.0000 0.0002 0.0004 0.0006 0.0008 0.0010 0.0012 0.0016 0.0018 2.5% 5% 10% Liner [Hg] mg/LPore Volume Figure 3-5. Mercury concentrations in Neva da soil leachates of control (no Al-WTRs) and samples treated with Al-WTRs at applica tion rates of 2.5%, 5%, and 10%. The 10% application rate treatmen t had two Al-WTRs incorpor ation methods: (1) Al-WTRs mixed uniformly with the entire soil in the column and (2) incorporation as a liner at the bottom of the soil column (liner treatment). A B
87 Control2.5%5%10%Liner 0.000 0.002 0.004 0.006 0.008 0.010 0.012 Figure 3-6. Cumulative mass of Hg leached per kg of soil from the control (no Al-WTRs) and Al-WTRs treated columns of Nevada soil af ter 11 pore volumes. Error bars represent one standard deviation of the mean.
88 Al-WTRs treatment Control2.5% Al-WTRs5% Al-WTRs10% Al-WTRs Methyl Hg (mg/kg) 0.000 0.005 0.010 0.015 0.020 0.025 0.030 Figure 3-7. Concentrations of methyl-Hg produced in Nevada so il after 10 days of incubation under water saturation and anoxic conditions. On the x-axis, the control corresponds to soil without Al-WTRs addition, while shown percentages of Al-WTRs correspond to the different application rates of used Al-WTRs as treatments in column leaching experiments
89 CHAPTER 4 DISTRIBUTION OF MERCURY SORBED ONTO ALUMINUM-BASED DRINKING WAT ER TREATMENT RESIDUALS (AL-WTRS) AND IMPLICATIONS FOR LONGTERM STABILITY OF FORMED MERCURY-AL-WTRS COMPLEXES General Introduction The ultimate objective of this research is to establish the potential of drinking water treatment residual (Al-WTRs) to efficiently immob ilize mercury (Hg) within its structure; and to determine the efficacy of Al-WTRs to form Hg -[Al-WTRs] complexes that are stable over long periods of time. If used in remediation of Hg-c ontaminated soils, Al-WTRs would be expected to adsorb Hg from soil pore water, and its effici ency for Hg removal from pore water and the stability of formed Hg complexes would depend on the types of Hg binding sites present in the Al-WTRs matrix. To meet the above objective, Hg distributio n among different mineral and organic fractions of Al-WTRs, as well as the sign ificance of surface versus internal sorption sites in Hg immobilization by Al-WTRs should be determined. This study is divided into tw o parts. The first section uses a selective and sequential chemical fractionation approach to assess the distribution of Hg within the Al-WTRs matrix. The second part focuses on changes in sorbents poros ity due to Hg incorpor ation into the Al-WTRs matrix, and on the determination of Hg diff usivities into the micropores of Al-WTRs. Part-1: Chemical Fractionation of Mercury in Al-WTRs by Sequential Extraction Introduction Soil contamination by Hg poses a threat to living organisms and the environment. However, the determination of total-Hg concentr ations in soils is insufficient to accurately predict the environmental risks associated with Hg pollution. This is because Hg behavior and impacts in soils depend largely on its speciati on (Adriano 2001; Ma and Rao 1997). The latter controls the degree of toxicity a nd mobility of Hg, and therefore, its bioavailability and leaching
90 potential from porous me dia (Bernaus et al. 2006; Bloom et al. 2003; Fernandez-Martinez and Rucandio 2005). In this research, it has been shown that Al-WTRs can efficiently immobilize Hg when applied to Hg-contaminated soils (chapter 3). With regard to the long-term fate of such immobilized Hg, chemical fractionation methods can be used to asse ss the distribution of sorbed Hg amongst the different mineral and organic fr actions of Al-WTRs, and help predict the potential for leachability and bioavailability. The distribution of Hg among the different frac tions of Al-WTRs can be assessed by use of selective sequential chemical extraction procedures (Tessier et al. 1979). Although these fractionation techniques cannot identify specific compounds, they can provide much-needed information on the biogeochemical behavior of various Hg associations to both mineral and organic phases under various environmental conditi ons (Bloom et al. 2003). Therefore, this approach can be used to evaluate the incorporation of Hg into various fractions of Al-WTRs, and allow for the assessment of its fate (Fer nandez-Martinez et al. 2006; Siegel 2002). Selective sequential extraction techniques generally follow a procedure in which the total metal content is subdivided into several operationally defined fractions that may include (1) mobile or easily soluble fraction (e.g. water so luble compounds), (2) metal easily exchangeable through cationic substitution (e.g exchangeable compounds), (3) metal bound to oxide minerals (e.g. associated with iron and manganese hyd r(oxides), (4) metal bound to organic matter, (5) metal present in residual fractions (e.g. mercuric sulfide) (Biester et al. 2002; Neculita et al. 2005; Tessier et al. 1979). The most bioavailable Hg is found in the easily soluble and easily exchangeable fractions. The least bioavailable Hg is generally in the residual fraction which is chemically stable over geologic time periods and re presents the least toxic Hg fraction (Han et al. 2003). Chemical reagents that are used as extrac ting agents for each Hg-fr action are believed to
91 be highly selective in solubili zing the targeted Hg-fraction. In this study, the procedure proposed by Tessier et al. (1979) one of the most widely used metal fractionation procedures (Ayyamperumal et al. 2006; Chlopecka et al 1996; Oygard et al. 2008) is used. The objective of this first portion of the study is therefore to inve stigate the chemical partitioning of Hg in aged Hg-spiked Al-WTR s, using an operationa lly defined chemical fractionation procedure. This appr oach should allow the assessment of the potential for the longterm stability of formed Hg-[Al -WTRs] complexes, and thus the determination of the efficacy of Al-WTRs application in Hg-contaminated soils as a technique for soil remediation. Material and Methods Preparation of aged mercury-spiked Al-WTRs A flooding approach was used in which Hg -spiked Al-WTRs were prepared by first bringing dry Al-WTRs into contact with a HgCl2 solution in 1:2 ratio (m/v). The concentration of total-Hg achieved in this spiking process and determined by ICP-AES was 24,050 mg Hg/kg Al-WTR (n=3). Details on sample digesti on and analysis are given in the total Hg analysis section below. Obtained slurries were th en equilibrated for 7 days in closed containers in a fume hood. Next, the mixtures were allowed to dry at room temperature by opening the lead of the containers for 7 days. Following this init ial step of Hg incorporation into Al-WTRs, wet and dry cycles were simulated by first saturati ng the Hg-[Al-WTRs] mixtures with deionized water for 5 days and then air-dry ing for another 5 days at room temperature. This experimental approach was designed with the intention to force Hg into th e micropores of the Al-WTRs as surface sites became progressively saturated over time. After a total of 12 wet-dry cycles (~4 months), aliquots of the Hg-spi ked Al-WTRs materials were used for the determination of Hg distribution among the different Al-WTRs mineral and organic fr actions. The control treatment was represented by the native Al-WTRs with no Hg addition.
92 Sequential extraction procedure Sub-samples (~1g) of air-dried Al-WTRs sa mples (e.g. Hg-spiked Al-WTRs and control Al-WTRs with no Hg added) were extracted sequ entially using a modified procedure originally proposed by Tessier et al. (1979). The following fractions were targ eted and all extractions were carried out in triplicate. Fraction 1 (F1), Water Soluble: One gram of soil was extracte d at room temperature with 8 ml of Nanopure water for 3 hours with continuous agit ation at a rate of 200 rpm. Following the agitation step, the mixture was centrif uged at 10,000 rpm for 30 min (Beckman J2-HS, Tritech Field Eng. Inc.) and the supernatant removed with a pipett e. Next, the residue was rinsed with 8 ml of Nanopure water for 5 minutes and the mixtur e centrifuged again at 10,000 rpm for 30 min. The supernatant was careful ly withdrawn and added to the first supernatant fraction. The combined supernatant was then filtered (0.45m) and used for the analysis of total-Hg. The solid residue was then used in the next extraction st ep. Besides the extraction step, the centrifugation, rinsing, and filtration steps were identical for all fractions (F1 to F5), except for F6. Fraction 2 (F2), Exchangeable: The residue from F1 was treated with 8 mL of 1M magnesium chloride (MgCl2, pH 7). The mixture was agitated on at 200 rpm at room temperature for 1 hour, prior to centrifugation, rinsing, and filtration as described in F1. Fraction 3 (F3), Bound to carbonates: The residue from F2 was extracted with 8 mL of 1M sodium acetate (NaOAc) soluti on with a pH adjusted to 5.0 with acetic acid (HOAc). The mixture was agitated on an orbital shaker at 200 rpm at room temperatur e for 5 hours, prior to centrifugation, rinsing, and filtration. Fraction 4 (F4), Bound to Fe-Mn oxides: The residue from F3 was extracted with 20 ml of 0.04 M hydroxylamine hydrochloride (NH2OH.HCl) in 25% (v/v) HOAc in containers placed
93 in a water bath for 6 hours at 96 3C, with occasional agitation by hand. Upon cooling, the mixture was centrifuged, rinsed, and the s upernatant filtered as described in F1. Fraction 5 (F5), Bound to organic matter: The residue from F4 was extracted with 3 mL of 0.02 M nitric acid (HNO3) and 5 mL of 30% hydrogen peroxide (H2O2) that had pH adjusted to 2 with HNO3. The mixture was heated to 85 2 C in a water bath for 3 hours with intermittent agitation by hand. An additional 3 mL of 30% H2O2 (pH adjusted to 2 with HNO3) was added and the mixture heated for an additional 3 hours at 85 2C. Upon cooling, and to prevent the adsorption of extracted Hg onto oxidized Al-WTRs components, 5 mL of 3.2 M ammonium acetate (NH4OAc) in 20% (v/v) HNO3 was added and the sample diluted to 20 mL and agitated continuously for 30 min on an orbita l shaker at 200 rpm. These steps were then followed by centrifugation and filtration as described in F1. Fraction 6 (F6), Residual: The residual fraction was qua ntified by subtracting the sum amount of Hg obtained in fractions F1-5 from the amount of Hg obtained from the total Hg analysis. Total Hg analysis: For solid Al-WTRs samples, 5ml of aqua regia and 5 ml of hydrofluoric acid were added to about 1 g of dr y Al-WTRs samples (i.e. Hg-spiked and control Al-WTRs) in a Teflon vessel, capped, and digested overnight at 1100C. The mixture upon cooling was diluted to 50 ml with a saturated so lution of boric acid to di ssolve formed fluorides. For the aqueous fractions, concentrations of Hg in supernatants were determined by inductively coupled plasma atomic emission spectrometry (ICP-AES), following overnight digestion with bromine monochloride (mixture of KBr and KBrO3 dissolved in concentrated HCl) at room temperature. This step allows the oxidation of soluble organic matter and
94 dissolution of colloidal Hg, th erefore dissolving Hg bound to or ganic ligands and colloidal materials In addition to mercury, Al, Fe, and Si were al so determined to help assess the relationship between solubilzed Hg and these mineral forming elements. Results and Discussion The distribution of Hg among the different fractions of Al-WTRs is presented in Table 4-1. When expressed as percent of the initial total-Hg concentration, Hg was distributed as follows: the water soluble fraction accounted for 0.6%; th e easily exchangeable fraction for 3.2%, the carbonate-bound fraction for 0.9%, the Fe and Mn oxide-bound fraction for 3.7%, the organicbound Hg for 9.0 % and the residua l fraction for 82.3 %. In cont rast, control Al-WTRs samples released no Hg in any of the chemical fractions extracted, suggesting that the background Hg concentration in used Al-WTRs was mostly in the residual form and averaged 0.02 mg/kg. In Hg-spiked Al-WTRs, the water soluble frac tion (F1) had the lowest percentage (0.6%) of total-Hg present in the solid phase, and this despite the fact that Al-WTRs was flooded with HgCl2 at a final concentration > 24,000 mg Hg/kg Al-WTRs. In comparison with other mercury species, HgCl2 is one of the most soluble forms and is generally stable in soil solution and less prone to sorption in the soil (Bac kstrom et al. 2003; Kasprzyk-Hor dern 2004). For example, in a study by Bloom et al. (20 03), almost 100% of HgCl2 was extracted in the water soluble fraction after HgCl2 in high concentrations was dispersed in ka olin clay. The water soluble fraction is the most mobile fraction and poses the highest envi ronmental risk. Since Al-WTRs in this study was spiked with Hg amount rarely encountered in th e environment, even in contaminated systems, the concentration of Hg leachable by water amounted to 146 mg/kg. Nonetheless, the relatively small value of Hg concentration in this fraction as compared to the initial total Hg level emphasizes the potential for Al-WTRs to limit Hg leaching if used in soil remediation.
95 Fraction 2 represents Hg bound to metal oxid es, primarily through surface adsorption, and represents the fraction most easily substituted by ionic exchange processes (Ma and Rao 1997; Tessier et al. 1979). Although 3.2% of added Hg was recovere d by extraction with sodium acetate at pH 5, similar to F1, this fraction re mains relatively small as compared to the large amount of Hg added initially to the tested Al-WTRs. The fraction F3 represents Hg associated w ith carbonate minerals. Mercury bound to this fraction is mobilized when th e pH drops significantly, and as carbonate minerals undergo dissolution. Only 0.9% of the total-Hg added to the Al-WTR was recovered in this fraction. This rather low value could be due to either the low affinity of Hg for carbonate minerals or to the possible limited abundance of such minerals in used Al-WTRs (data not available). Mercury co-precipitated or ad sorbed onto iron and manganese (hydr)oxides (F4 fraction) is relatively stable under oxic conditions. However, such Hg can be mobilized into soil pore water if iron and manganese (hydr)oxi des undergo reduction during mi crobial oxidati on of organic matter and as these (hydr)oxides act as term inal electron acceptors under anoxic conditions (Bernaus et al. 2006; Gobeil and Cossa 1993; Tessier et al. 1979) Tested Al-WTRs contained a considerable amount of Fe, i.e. 3766 mg/kg. Fractionation results showed that most of Fe was released in the F4-fraction (Table 4-1, Figure 4-1), while Mn was not detectable (analytical detection limit of about 10 g Mn/L on used ICP-AES). However, the amount of Hg released in F4 fraction was only 3.7% of the initial total-Hg concentration, suggesting that Fe-(hydr)oxides may not play a dominant role as adsorption sites for Hg in these tested Al-WTRs. Mercury extracted in F5-fraction (9% of total Hg) represents Hg that is bound to various forms of organic matter (Tessier et al. 1979). Organic matter plays an important role in mercury speciation (Khwaja et al. 2006; Suer and Lifvergren 2003). In chapter 2, SEM-EDS analyses
96 showed that Al-WTRs particles have abundant S in their chemical composition (Figure 2-2). This sulfur could be part of the different organic compounds pres ent in Al-WTRs. In the organic fraction, Hg is probably associated with the S, N, and O containing functiona l groups, with the sorption affinity strongly decrea sing from sulfur to oxygen (Schlu ter 1997). In fact, Hg is a chalcophile and therefore, th e strong bonds between Hg and thiol (-SH) groups present in organic compounds could be highly specific, and when the capacity for bonding to SH is filled, Hg would then become attached to NH2 and COOH functional groups by somewhat weaker bonds (Bringmark 1997). Even thou gh most of organic carbon in Al-WTRs is associated with natural organic matter of raw water, Hg may also be complexed with organic polymers which are added to treated water in very low concentra tions (Makris et al. 2005) In this organic F5fraction, Hg may become mobile only following the decomposition and/or dissolution of organic matter. The majority of Hg was associated with the F6-residual fraction (ca. 82.3% of initial totalHg concentration). The residual fraction is the leas t bioavailable and the most chemically stable. It therefore represents th e least toxic Hg fraction (H an et al. 2003). This is the Hg fraction that is not expected to be released to the aqueous phase under conditions normally encountered in nature (Tessier et al. 1979). Me rcury sorption onto Al-WTRs in the residual fraction is very important because of the implications for th e long-term stability of formed Hg-[Al-WTR] complexes. Besides Hg, the residual fraction cons isted mostly of silica and aluminum (Figure 41), with mean concentrations of 51.6 mg/g and 67.7 mg/g for Al and Si, respectively. Also, the acid digestion procedure used fo r the determination of total-Hg concentration in Al-WTRs did not achieve a complete dissolution of the solid Al-WTRs samples as evidenced by the presence of a solid residue at the bottom of the container at the end of digestion process. Therefore, it is
97 possible that the concentrations of extracted el ements in the residual fraction could have been slightly underestimated. Finally, extracted Hg concentrations did not correlate with the concentrations of other elements (i.e. Al, Fe, and Si) measured in the supernatant solutions obtained from each of the ex traction steps (Figure 4-1). Previous studies have suggest ed that the oxidation of orga nic matter by acidified hydrogen peroxide in the organic fracti on (F5) may not be complete (T essier et al. 1979). However it represents a compromise because the use of stronger oxidizing agents may alter the material in the residual fraction (Tessier et al. 1979). To determine the pe rcentage of the organic matter remaining in the residual fraction after the top 5 extractions, total organic matter content was measured comparatively in the native Al-WTRs and in the remaining residual fraction of AlWTRs obtained after chemical fractionation. Loss on ignition (LOI) was used to estimate the total organic matter content. First, Al-WTR s samples were incubated for 48 hours at 1300C to dehydrate the samples. Next, the samples were combusted at 5500C for 2 hours in a muffler to ash the organic matter. Obtained results indicate that nearly 41% of total organic matter present in Al-WTRs was left in the residual fraction af ter the chemical fractiona tion that includes the oxidation of organic matter by use of 30% H2O2. The refractory organic ma tter likely consists of paraffin like material and highly resistant organic residues (Te ssier et al. 1979). According to Makris (2004), such organic matte r would likely be present in th e structure of solid phase AlWTRs. It is believed that during the formation of Al-WTRs, orga nic matter can become trapped in the micropores of Al-WTRs as the material coagulates and flocculates in the water. Amorphous Al-(hydr)oxides exhibit a high affinity for Hg as illustrated in several studies (Sarkar et al. 1999, Axe and Triv edi 2002). Therefore, it is possibl e that Al-(hydr)oxides play a dominant role in Hg immobilized in the residua l fraction. But to accurately determine the Hg
98 binding sites on Al-WTRs particle s, studies employing analytical techniques such as X-ray absorption spectroscopy (XAS) should be used. Part-2: Mercury Incorporation in Micropores of Al-WTRs and Modeling of Intraparticle Diffusion Introduction Results presented in earlier Chapter 2 suggest that Hg sorption onto Al-WTRs particles is a two step process, with the intraparticle diffusion being the likely rate-limiting step. The first step (rapid adsorption) could be attrib uted to either boundary layer diffusion effects or to the mass transfer effect on external surf aces, while the second step is a much slower and gradual sorption process during which intrapartic le diffusion would dominate. Al so, the chemical fractionation method used to study the speciation of Hg in aged Hg-spiked Al-WTRs reveal ed that most of the Hg immobilized onto Al-WTRs particles is associated with the residual fraction. These two findings are in agreement and demonstrate that Hg could be incorporated into the micropores of Al-WTRs, resulting in the formation of stable Hg complexes. The plausibility of the intraparticle diffusion mechanism is also supported by the fact that Al-WTRs particles have a large internal su rface area. The specific surface area (SSA-N2) and micropore surface area (SSA-CO2) are key parameters that st rongly influence the sorption capacity of solid surfaces (Goldberg et al. 2001). According to the International Union of Pure and Applied Chemistry (IUPAC) classification, pores are divided into three categories based on their width (w): (1) micropores for w< 20 (2) mesopores when 20
99 (chapter 2). This internal network of micropores represents potential site s for Hg incorporation into the Al-WTRs matrix. The objective of this portion of the study was to examine Hg incorporation into micropores of Al-WTRs by assessing the physicochemical pr operties of Al-WTRs before and after Hg sorption. Here, the specific surface area and mi cropore volume of Hg-free and Hg-spiked AlWTRs were measured. In addition, to predict th e diffusivity of Hg into the micropores and to assess the rates of Hg diffusion into the inner structure of Al-WTRs, site activation theory (Trivedi and Axe 1999) was used to predict Hg diffusivities into Al-WTRs particles. Materials and Methods Aged mercury-spiked Al-WTRs Mercury-spiked Al-WTRs used in this second portion of the study were sub-samples from the Hg-spiked Al-WTRs described earlier in this chapter. Determination of SSA-N2 and SSA-CO2 of Al-WTRs before and after mercury sorption The determination of specific surface area ( SSA) was performed on both aged Hg-spiked and Hg-free (control) Al-WTRs. First, as a pretreatment to SSA meas urement, a known amount of Al-WTRs was introduced into a capillary glass tube and outgassed for 4 hours under helium flow at 700C (Makris et al. 2004a). The SSA was th en measured by the Brunauer-Emmett-Teller nitrogen adsorption method at 77 K (BET-N2), while SSA applicable to the internal (micropore) surface area was measured using carbon dioxide at 273 K (SSA-CO2). Both BET-N2 and SSACO2 were measured using Quantachrome Auto sorb-1 (Quantachrome Corp.) apparatus at the Particle Research Engineering Center (PERC) at University of Florida. The BET-N2 is applicable mostly to non-porous or mesoporous materials and tends to underestimate the SSA of internal micropores when present (Lowell et al. 2004). The CO2 is a preferred adsorbent in micr opore analysis even though CO2 (2.8 ) and N2 (3.0 ) have similar
100 molecular dimensions (Lowell et al. 2004). The reason being that CO2 analysis is performed at a higher temperature (T=273 K) compared to N2 (T=77 K), as a result CO2 passage through micropores is less constricted (Gil and Gandia 2003; Lowell et al. 2004). The BET-N2 method is applied in the region of rela tive pressures ranging from P/P0=0.05 to P/P0=0.3, while the SSACO2 is carried out in the interval of relative pressures of P/P0=10-5 to P/P0 =0.0029) (Gregg and Sing 1982). The SSA-N2 was calculated using the Brunauer-Emmett-Teller or BET equation (Equation 4-1), where W is the weight of the gas adsorbed at a relative pressure P/P0, Wm is the monolayer capacity and C is a BET constant. 0 0)( 1 1 )1( 1 P P CW C CW P P Wm m (Eq. 4-1) The micropore volume was calculated from the CO2 adsorption isotherm using the Dubinin-Radushkevich (DR) me thod (Equation 4-2), by plotting 2 0log P Pversus logW, from which V0 the micropore volume can then be calculated (Lowe ll et al. 2004). 2 0 0log)log(log P P DVW (Eq. 4-2) Where W is the amount of the gas ad sorbed at relative pressure P/P0, is the liquid adsorbate density, V0 is the micropore volume, D is a constant that characterizes Gaussian distribution, P0 is the vapor saturation pressure of CO2 (26,140 mm Hg), and P is the equilibrium pressure (mm Hg). The SSA-CO2, which describes micropore specific surface area, was calculated using the Dubinin-RadushkevichKaganer (DRK) equation (Equation 4-3) (Gregg and Sing 1982). A plot
101 of log W versus 2 0log P Pis used to determine the monolayer capacity Wm, from which the surface area can be calculat ed using equation 4-4. 2 0log loglog P P DWWm (Eq. 4-3) M NAW Sm (Eq. 4-4) Where S is the specific surface area, A is the cross-sectional area, M is the molecular weight of the adsorbate, and N is the Avogadro number (Lowell et al. 2004). The Grand Canonical Monte Carl o simulation model was used to determine the pore size distribution of Al-WTRs. All the calculations were performed using the Quantachrome Autosorb-1 Software (vs. 1.54), which incorporates an advanced built-in database of parameters and functions necessary for data anal ysis and the Monte Carlo simulation. Data obtained from different experiments of Hg incorporation into the micropores of AlWTRs was evaluated by comparing the volumes of Hg free and Hg-spi ked Al-WTRs using the Independent-Sample t-test with SPSS statistical software (vs. 11.0) The tests were performed at a confidence level of 95%. Determination of the activation energy (Ea) of mercury sorption onto Al-WTRs particles Short term batch kinetic st udies were conducted at 110C, 250C, and 350C to obtain the activation energy value needed for the determina tion of Hg diffusivity using the intraparticle diffusion model. A commercial standard solu tion of Hg(NO3)2 obtained from Fisher Scientific was used to prepare solutions of 40 mg Hg /L in Nanopure water. For each the tested three temperatures, 20 grams of Al-WTRs were mixed with 1000 ml of the Hg solution and run in duplicate. The contact time and sampling intervals were selected based on preliminary
102 experiments. At pre-decided time intervals, th e supernatant was withdrawn, filtered (0.45m), and analyzed for total Hg concentrations. Total-Hg was measured by ICP-AES, following overnight cold digestion step using bromine monochloride (i.e. KBr and KBrO3 dissolved into concentrated ultra high purity HCl). Obtained data were then fit to different kinetic models to determine the best fit and to evaluate the effect of temperature on the rate c onstant. The relationship be tween the reaction rate constant and the temperature is described by the Arrhenius equation (E quation 4-5), which can be linearized as shown in equation 4-6 (Schnoor 1996): R T /EaAek (Eq. 4-5) T lR a E Alnkn1 (Eq. 4-6) and where k is the reaction rate constant, A is a constant that is characteristic of the reaction, Ea is the activation energy for adsorption (cal mol-1), T is the absolute temperature (Kelvin), and R is the universal gas cons tant (1.987 cal mol-1 K-1). Thus, the Ea can be determined from the slope of th e straight line obtained by plotting ln k versus 1/T plot (Equation 4-6). The Ea value was then used to predict the diffusivities of Hg sorption on Al-WTRs. Results and Discussion Determination of SSA-N2 and SSA-CO2 of Al-WTRs before and after mercury sorption The SSA-N2 characterization (BET method) for cont rol Al-WTRs revealed a rather high surface area averaging 48 m2/g. However, the measured SSA-CO2 (DRK method) was even higher (120 m2/g) suggesting that Al-WTRs have a large internal surface area that is not accounted by the BET-N2. The results of N2 and CO2 gas sorption analysis for control (no Hg
103 addition) and Hg-spiked Al-WTRs are illu strated on Figures 42 and 4-3. The SSA characterization of Hg-spiked Al-WTRs revealed a significantly lower SSA-N2 and SSA-CO2 as compared to those determined in control sample (p=0.006, p=0.00003) (Figure 4-4). A further interpretation of CO2 sorption data was done by use of the grand canonical Monte Carlo simulation model. Th is technique allows for the determination of the micropore volume distribution in Al-WTRs. The model pres ents the pore structur e as a collection of idealized slit-shaped pores with smooth grap hitic walls (Ravikovitch et al. 2000). The most decrease in micropore volume was in the 3.98-5. 01 pore width interval (Figure 4-5), and the diameter of the hydrated divalent Hg is about 4.74 (Trivedi and Axe 2001). Overall, Hg sorption caused significant reduc tion in micropore volume as calculated by the DR equation, suggesting that Hg occupied pores that would otherwise be accessible to CO2 (Figure 4-6). The porosity results support the hypoth esis that the predominant mechanism of Hg soption into AlWTRs is likely the intraparticle diffusion. As mentioned earlier, according to the IUPAC classification, pores are divided into micro-, meso-, and macro-pores. However, this division is somewhat arbitrary and is mainly based on the difference in the types of for ces that control the adsorption be havior within the different size ranges (Karger and Ruthven 1992). In mesopores, capillary forces are important, while in macropores, these capillary forces contribute very little to the adsorpti on capacity (Karger and Ruthven 1992). On the other hand, micropores exhibit unusually hi gh adsorption potentials due to the wall close proximity (Low ell et al. 2004). Thus, diffusi on in micropores is dominated by steric interactions between the diffusing molecule and the walls of the pores. Intraparticle diffusion of Hg into the micropores of Al-WTRs is an important aspect from the remediation viewpoint as it relates to the long -term stability of Hg. It has been suggested that the diffusing
104 molecules may never escape from the force fi eld of the micropore walls (Karger and Ruthven 1992). Modeling of intraparticle diffusion a nd prediction of mercury diffusivity The diffusion of molecules into micropores is an activated pr ocess proceeding by a sequence of jumps between regions of relati vely low potential energy (Karger and Ruthven 1992). Understanding the mechanism of Hg incorporation in Al-W TRs is very important from remediation standpoint, as it wi ll likely determine the long -term stability of Hg-AlWTRs complexes. Diffusion is generally defined as a random mo lecular motion that occurs as a result of concentration gradient (Axe and Anderson 1997 ; Axe and Anderson 1998; Karger and Ruthven 1992; Reeder et al. 1999). The quantitative descrip tion of diffusion was fi rst introduced by Adolf Fick, through an expression that is generally known as Ficks First Law of Diffusion (Karger and Ruthven 1992; Schnoor 1996) (Equation 4-7). z c DJ (Eq. 4-7) Where D is the transport diffusivity, c is the concentration, z is the distance, and J is the diffusive flux. The diffusivity (D) describes the time scale for diffusi on which may vary from seconds for gases to millennia for crystal line solids at ordinary temperat ures (Karger and Ruthven 1992). Therefore, the determination of the diffusivity during the intraparticle diffusion process in AlWTRs has a practical significance as it may give an estimation of the longevity of Hg sorption process in Al-WTRs. Several approaches have been proposed fo r the assessment of the intraparticle diffusion mechanism including the intrap article diffusion model proposed for phosphorus in WTR by
105 Makris et al. (2004a), and the modeling approach for metal ca tions in amorphous oxides based on site activation theory descri bed by Trivedi and Axe (2000). Th e concentrations of amorphous aluminum (hydr)oxides in Al-WTRs have been found to range from 50 to 150 g/kg (Dayton and Basta 2001). Theoretically, the prevalence of such geochemical phases in Al-WTRs would make them behave like aluminum (hydr)oxides. Therefore, we chose the site activation theory model proposed by Trivedi and Axe (2000) to determine the diffusivity of Hg intraparticle diffusion into Al-WTRs (Equation 4-8). RT E aae m E D 2 (Eq. 4-8) Where D is the diffusivity, is the mean distance between sites, m is the molecular weight of the diffusing species, Ea is the activation energy required for a sorbed ion to jump to the neighboring site, and RT Eae is the Boltzmann factor. Since the molecular weight (m), R and T have known values, Ea and should be determined in order to calculate the D (Equation 4-8). In this study, Ea was determined using rate constants of reaction from batch sorption kinetic studies conducted at 11, 25, and 35o C. First, obtained kine tic data were fit to a first order kinetic-model from which reaction rate constants were determ ined. The relationship between the reaction rate (k) constant and temperature (T) is describe d by the Arrhenius equation (Equation 4-5). The Ea was then determined from the slope of the straight line of the lnk versus 1/T plot (Figure 4-7). The calculated EA value of 7.156 kcal/mol was then used to predict the intrapar ticle diffusion and determine theoretical diffusivities of Hg into Al-WTRs micropores (Equation 4-8). In contrast to Ea, is usually determined through long-term lab experiments which can last from months to years. Although the lite rature reports several values for different amorphous metal(hydr)oxides including values for Al-(hydr)oxides (Fan et al. 2005; Trivedi and Axe 2001), the
106 complex composition of Al-WTRs may result in values that are either similar to that of amorphous Al-(hydr)oxides due to its anticipa ted abundance in Al-WTRs matrices or far different from those reported for pure (hydr)o xide compounds. However, with the assumption that Hg binding sites on Al-WTR s are predominantly on Aland to some extent Fe-(hydr)oxides, theoretical diffusivities of Al-WTRs can be estimated by varying the value of in equation 4-8. Accordingly, Figure 4-8 shows the range of potential diffusivity values (D) of Hg into Al-WTRs with varying values. These predictions show that if Al-(hydr)oxides are the main sites of intraparticle diffusion ( values ranging from 2.0-8 to 3.5-8 cm), then the diffusivity (D) of Hg in Al-WTRs would range from about 9.55-11 to 1.67-10 cm2 s-1. These theoretical diffusivity values for Hg diffusion into Al-WTRs would then be within the range of published diffusivities for amorphous oxides of Al and Fe (10-10 to 10-14 cm2 s-1) (Axe and Trivedi 2002). Based on these predicted values of D for Hg in Al-WTRs and assuming no competition with other cations, it would take anywhere between 48 to 2.38 years for Hg to fully occupy the 120 m2 available in a gram of Al-WTRs. This tim e would likely be shortened in actual soil matrices due to competition for available binding sites between Hg and other ions. Nonetheless, these numbers suggest that if used in soil reme diation, Al-WTRs could behave as a long-term sink for Hg, and more importantly, with formation of very stable Hg-Al-WTRs complexes. General Conclusions Overall, the micro-porosity measurements with CO2 of Hg free and Hg-spiked Al-WTRs revealed the reduction of micropore volume in Hg -spiked Al-WTRs. This suggests that it is likely that Hg diffused into pores of Al-WTRs t hus blocking the pores th at otherwise would be accessible to CO2,
107 From the site activation theory the theoretical diffusivities of Hg were predicted with the assumption that amorphous aluminum (hydr)oxide s are the dominant geochemical phases that govern the intraparticle diffusion mechanism. To validate this interpre tation, experimental diffusivities need to be determined thr ough long term constant boundary condition (CBC) studies. In CBC studies Hg bulk adsorbate concen tration is maintained constant by adding Hg into solution for prolonged periods of time (Fan et al. 2005). This approach will help saturate the sites at the surface of the partic les and allow a gradual diffusion of Hg into micropores. With the obtained results, the approach proposed by Trivedi and Axe (2000) could be used to determine experimental diffusivities. If the values of theoretical diffusivities predicted in this study are comparable to the values of diffusivities obtaine d experimentally, then one could conclude that amorphous aluminum (hydr)oxides likely domin ate the intraparticle diffusion process. Overall, the combination of Hg accumulati on in the residual fraction (based on the chemical fractionation results), and the evidence of Hg incorporat ion into the micropores of AlWTRs suggest that Hg-[Al-WTRs] complexes that may form if Al-WTRs is used in remediation of Hg-contaminated soils would re sult in long-term immobilization of Hg in the solid phase, and therefore the elimination/reduction of Hg bi oavailability and its subsequent toxicity.
108 Table 4-1. Mercury, Al, Fe, and Si concentrations and distributi on (%) in different fractions of Hg-spiked Al-WTRs Element Fraction Concentration mg/ga % of total Hg Water Soluble 146 0.60 Exchangeable 766 3.18 Carbonate 230.1 0.95 Fe/Mn Oxide 889.5 3.69 Organic 2162.8 8.99 Residual 19855 82.3 Total 24050 -Al Water Soluble
109 Percent of Total 05101520253035404550556065707580859095100 F 6 F 5 F 4 F 3 F 2 F 1 Hg Al Fe Si Figure 4-1. Mercury, Al, Fe, a nd Si distribution (%) in the va rious fractions of Al-WTR (F1water soluble, F2-exchangeable, F3car bonate, F4-Fe/Mn bound, F5organic bound, F6-residual).
110 P/P0 0.000.050.100.188.8.131.520.35 N2 Volume (cm3/g) 6 8 10 12 14 16 No Hg Rep. 1 No Hg Rep. 2 No Hg Rep. 3 Hg-spiked Rep. 1 Hg-spiked Rep. 2 Hg-spiked Rep. 3 A B Figure 4-2. The N2 gas sorption on Al-WTRs with (A) no Hg and (B) Hg-spiked. The treatments are presented in three replicates. Replicate 1 and 2 for both (A) and (B) overlap on this graph.
111 P/P0 0.00000.00050.00100.00150.00200.00250.0030 CO2 Volume (cm3/g) 0 1 2 3 4 5 6 7 No Hg Rep. 1 No Hg Rep. 2 No Hg Rep. 3 Hg loaded Rep. 1 Hg loaded Rep. 2 Hg loaded Rep. 3 A B Figure 4-3. The CO2 gas sorption on Al-WTRs with (A ) no Hg and (B) Hg-spiked. The treatments are presented in three replicates. Replicate 2 and 3 for (A) and replicate 1 and 2 for (B) overlap on this graph.
112 Figure 4-4. Mean BET (N2) and micropore (CO2) surface area of Al-WTRs with no Hg (black bars) and Hg-spiked (white bars) Al-WTRs. Error bars represent one standard deviation of three replicat es. Different letters indicate a significant difference ( <0.05). Specific Surace Area (m2/g) 0 20 40 60 80 100 120 140 Al-WTRs (no Hg) Hg-spiked Al-WTRs BET CO2a b c d
113 Pore Width (Angstroms) Micropore Volume (cm3/g) 0.000 0.002 0.004 0.006 0.008 0.010 0.012 0.014 Al-WTRs with no Hg Hg-spiked Al-WTRs 6.30 7.94 7.94 10.00 10.00 12.58 3.98 5.01 5.01 6.30 Figure 4-5. Micropore volume based on pore size dist ribution of Al-WTRs with no Hg (black bars) and Hg-spiked Al-WTRs (grey bars).
114 Al-WTRs treatments No Hg Hg loaded Micropore Volume (cm3/g) 0.00 0.01 0.02 0.03 0.04 0.05 a b Figure 4-6. Mean micropore volum e of Al-WTRs with no Hg (cont rol) and Hg-spiked Al-WTRs. Error bars represent one standard deviati on of three replicates. Different letters indicate a significant difference ( <0.05).
115 0.003200.003250.003300.00335 0.003400.003450.003500.00355ln k -3.8 -3.6 -3.4 -3.2 -3.0 -2.8 -2.6 -2.4 R2=0.9954 y = -3601.8x + 9.071/T Figure 4-7. Arrhenius plot of Hg sorption on Al-WTRs at 110 C, 250 C, and 350 C, where T is the temperature expressed in Kelvin and k is the reaction rate constant.
116 mean distance between sites (cm) 2.0e-82.5e-82.9e-83.5e-84.0e-8Diffusivity (cm2 s-1) 8.59e-11 9.55e-11 1.07e-10 1.17e-10 1.27e-10 1.36e-10 1.46e-10 1.58e-10 1.67e-10 1.79e-10 1.89e-10 Figure 4-8. Diffusivities (D) of Hg in Al-WTRs as a function of varying values (mean distance between sorption sites). The -values for amorphous aluminum oxides range from 2.0-8 to 3.5-8 cm (Fan et al. 2005; Trived i and Axe 2001) and correspond to D values in the range from 9.55-11 to 1.67-10 cm2 s-1.
117 CHAPTER 5 CONCLUSIONS AND RECOMMENDATIONS Conclusions Water treatment residuals (WTRs) are non-haza rdous waste by-products of drinking water treatment processes. They are produced daily in large quantities worldwide and are either stored in onsite lagoons, discharged into river systems, or buried in landfills. The disposal of this waste product can be expensive as it increases the overa ll cost of the wa ter purification process. Since current remediation methods of Hg-polluted soil s are rather expensive and have several other disadvantages, the use of WTRs as soil amendm ents to immobilize Hg, and ultimately other toxic metals, would not only lowe r the cost of potable water by eliminating the need for landfilling of these waste materials but provide environmental benefits as a result of the reclamation of Hg-contaminated soils. Previous research has now established the st rong ability of WTRs to immobilize anions such as phosphate (e.g. Dayton et al. 2003), fl uoride (Sujana et al. 1998), and perchlorate (Makris et al. 2006). Unlike the above anionic species which are the subject of several ongoing investigations related to soil remediation through in-situ immobilization of pollutants; the use of WTRs in remediation of metal-contaminated soils are nonexistent. Therefore, this study emphasized the use of aluminum-based WTRs (A l-WTRs) as a sorbent for Hg and was designed primarily to assess the ability of Al-WTRs to: (1) sorb Hg from aqueous solutions, (2) to immobilize Hg in contaminated soils, and (3) initiate an investigation of the mechanisms responsible for Hg sorption and immobilization on to Al-WTRs, as well as the potential for longterm stability of formed mercury-WTRs complexes.
118 The major findings of this research can be summarized as follows: Batch sorption experiments showed that Al-W TRs materials have a great potential as sorbent for Hg with a determined average maximum sorption capacity of 79 mg Hg/g AlWTRs. Sorption kinetic data was best fit to a pseudo-first order model, while the use of the Weber-Morris and Bangham models suggested that Hg sorption to Al-WTRs is likely a two-step process with the intraparticle diffusion being the rate-limiting step. The role of intraparticle diffusion as ra te-limiting step on Hg sorption process was investigated using different approaches. Overall, we comp ared micropore volume of AlWTRs before and after Hg spiking, which demo nstrated significant decrease in micropore volume in aged Hg-spiked Al-WTRs. The sele ctive sequential extraction method used to study the speciation of Hg in aged Hg-spiked Al-WTRs revealed that most of the Hg is associated with the residual fraction of Al -WTRs. These findings demonstrate that it is likely that Hg was incorporated into micropor es of Al-WTRs as a result of which Hg is more stable and thus is found in the residual fraction of Al-WTRs. The stability of Hg in the micropores is explained by the fact that micropores exhibit great interaction potential with the diffusing molecule due to wall proximity as a result of which the diffusing molecule may never escape from the force field of the micropore walls. Results from the short-term experiments, s howed that Al-WTRs ef fectively immobilized Hg in the pH range of 3 to 8. The fact that Hg removal was very efficient at low pH values is significant from a remedi ation standpoint, because often times contaminated soils may have a very low pH and if left untreated, may leach out the Hg. The potential of Al-WTRs to sorb and i mmobilize Hg from contaminated soils was assessed using flow-through columns and mimicking the effect of acid rain. Two types of Hg contaminated soil were used, artificially contaminated with Hg chloride to mimic newly contaminated soils and well-aged soils from a contaminated site in Nevada. The addition of Al-WTRs to soil columns reduced Hg concentration in leachates, toxicity of leachates, as well as methylation potential of Hg compared to the control where no AlWTRs were added, however the effect was more pronounced in the newly contaminated soil. Overall, the results point out to the ab ility of Al-WTRs to immobilize Hg in soils. Recommendations The leaching potential and therefore the mobility of Hg in soil depends on its speciation. The results of sequential extr action procedures in severa l studies showed that Hg concentrations > 100 mg/kg can be found in mo bile fractions of so ils impacted by chloralkali plants or by Hg mini ng activities (Bernaus et al. 2006; Fernandez-Martinez et al. 2006). Such soils would be good candidates to be remediated through Al-WTRs additions. Hence, the chemical speciation of Hg in soil should be known before adopting this remediation approach. Future studies are needed to test the efficiency of Al-WTRs to immobilize Hg in a wide variety of soils that have di ffering physicochemical parameters. This could help develop a
119 predictive tool in which the ab ility of Al-WTRs to remediat e a given soil type can be determined based on routine geological survey data for a given soil. Obtained results showed that applying Al-W TRs as a liner in the soils was the most effective treatment. However, it could be more difficult to implement this approach in field situations versus just mixing the top layer of soil with Al-WTRs. Th e liner approach could have a potential use as bottom liner in landfill s. Heavy metals includ ing Hg are often found in landfill leachates (Yong et al. 2001), therefore, use of Al-WTRs as a liner in landfills would limit the leaching of Hg and potent ially other metals to the groundwater. Al-WTRs is a mixed adsorbent and it is possibl e that different constituents are responsible for sorption of Hg in different fractions. Ther efore, future studies should focus on using the X-ray absorption spectroscopy (XAS), which ha s species-specific dete ction capacities, to determine the speciation of Hg in Al-WTRs. In particular, one type of XAS, the XANES (X-ray Absorption Near Edge St ructure) has lower detection limits than XAS and has been suggested for determining the speciation of he terogeneous samples (Bernaus et al. 2006). In this study the diffusivitie s of Hg through intraparticle diffusion into Al-WTRs were theoretically predicted. Future studies ar e needed to determine the experimental diffusivities. Experimental diffusivities can be assessed through long-term experiments in which the constant boundary conditions ( CBC) for Hg are maintained. During the CBC studies Hg concentrations in the aqueous phase is monitored over time and brought back to initial levels through additions of small volume s of highly concentrated Hg solutions. This approach helps saturate the sites at the su rface of the particles and allows a gradual diffusion of Hg into micropores. With the obtained results, the approach proposed by Trivedi and Axe (2000) could be used to determine experimental diffusivities. If the experimental diffusivities are close to those predicted in th is study, then the amorphous aluminum (hydr)oxides are likely the predominant Hg sorbents that govern the intraparticle diffusion in Al-WTRs. This project focused primarily on sorption of Hg without taking into account the competitive adsorption of other metal cations. Fu ture studies are needed to validate this hypothesis by assessing the effect of comp etitive adsorption on Hg uptake by Al-WTRs. If all the above recommendations can be comple ted, then research on immobilization of Hg by Al-WTRs would be of greater significance because it woul d not only solve the problem of disposal of Al-WTRs but also will provide environmental benefits associated with the reclamation of Hg-contaminated soils. The us e of the developed method will help prevent further contamination of water bodies caused by Hg leaching from contaminated soils, thus protecting human health and the environmen t. Drinking water treatment facilities, however, use different water so urces and different additives along with the alum coagulant and therefore, the composition of Al-WTRs from different sources may vary. Accordingly, future studies are need ed to validate the ability of Al -WTRs from different facilities around the nation to sorb Hg.
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135 BIOGRAPHICAL SKETCH Anna Hovsepyan was born in Yerevan, Ar me nia in 1980. She received her bachelors degree in environmental engineer ing from the State Engineering University of Armenia in 2001. In 2003, she received her masters degree in en vironmental analysis and management from Troy State University, Alabama, USA. Currently, she is a Ph.D. candidate under the supervision of Dr. Jean-Claude J. Bonzongo at the Department of Environmental Engineer ing Sciences at the University of Florida, USA.