Chapter 1 - The Impacts of X-Ray Absorption Spectroscopy on ... · During the last two decades, XAS has devel-oped into a mature technique for obtaining the speciation (e.g., oxidation

Developments in Soil Science, Volume 34

C H A P T E R

1
The Impacts of X-Ray AbsorptionSpectroscopy on UnderstandingSoil Processes and Reaction
Mechanisms
Matthew Ginder-Vogel and Donald L. SparksDepartment of Plant and Soil Sciences, Delaware Environmental Institute,
University of Delaware, Newark, DE, USA

O U T L I N E

1. Introduction and Information Gainedfrom XAS 2

2. Complementary/Alternative Techniques 32.1. Synchrotron X-Ray Diffraction 42.2. X-Ray Photoelectron Spectroscopy 42.3. Vibrational Spectroscopy

(IR and Raman) 42.4. Mossbauer Spectroscopy 52.5. Nuclear Magnetic Resonance and

Electron Paramagnetic ResonanceSpectroscopies 5

2.6. Electron Microscopy 6

3. Historical Aspects of XASApplications to Mineral/WaterInterfacial Reactions 6

4. Applications of XAS to Elucidate SoilChemical Processes and Reactions 74.1. Adsorption of Metal(loids) on Soil

Components 74.2. Metal(loid) Surface Precipitation/

Dissolution 94.3. Redox Reactions on Soil Components 114.4. Speciation of Soil Contaminants 15

5. Future Needs 21

1
Copyright # 2010, Elsevier B.V.
All rights reserved.

2 1. IMPACT OF XAS ON UNDERSTANDING SOIL PROCESSES

1. INTRODUCTION ANDINFORMATION GAINED FROM XAS

The determination of element speciation anddistribution within heterogeneous environmen-tal systems, such as soils, is crucial to developinga thorough understanding of soil processes andreaction mechanisms that control not only thetransport of nutrients and contaminants, but alsobiomineralization and soil formation processes.The application of numerous spectroscopic andanalytical techniques is required to solve theseand other complex problems facing soil scientists.These techniques include X-ray absorption spec-troscopy (XAS) (Brown et al., 1997a,b; Schulzeand Bertsch, 1995; Sparks, 2005; Sparks andBorda, 2008), Fourier-transform infrared (FTIR)(Hind et al., 2001), Raman (Johnston and Wang,2002), nuclear magnetic resonance (NMR)(Bleam, 1991), electron paramagnetic resonance(EPR) (Flogeac et al., 2004), and Mossbauer (McCammon, 2004; Rancourt, 1998) spectroscopies.In addition, scanning and transmission electronmicroscopy (SEM and TEM), and atomic forcemicroscopy (AFM) (Cail and Hochella, 2005;Hillner et al., 1992; Weaver and Hochella, 2003)complement these spectroscopic techniques.

During the last two decades, XAS has devel-oped into a mature technique for obtaining thespeciation (e.g., oxidation state) and short-range structure of elements present in hetero-geneous soils and sediments (Fig. 1.1). XASencompasses both X-ray absorption near-edgestructure (XANES) spectroscopy and extendedX-ray absorption fine structure (EXAFS) spec-troscopy. XAS has a number of advantageousqualities for studying soils and sediments,which include elemental specificity, sensitivityto the local chemical and structural state of anelement, and the ability to analyze materialsin situ (Fig. 1.1). This information allows accu-rate determination of oxidation state, type ofnearest neighbors, coordination number, bonddistance, and orbital symmetries of the X-ray

absorbing element. Fendorf (1999) and Kellyet al. (2008) provide a detailed discussion of appli-cations of XAS to soil, clay, and environmental sci-ence and include summaries of the physical basisof XAS, experimental considerations, and dataanalysis techniques. Additionally, soil scientistshave recently begun employing variations on tra-ditional bulk XAS, including bulk and microfo-cused synchrotron X-ray diffraction (S-XRD andm-XRD, respectively), microfocused-XANES andEXAFS spectroscopies (m-XANES and m-EXAFS,respectively) (Brown and Sturchio, 2002) in addi-tion to transmission or fluorescence tomography(Chapter 3) in the analysis of heterogeneous sys-tems. X-ray fluorescence (XRF) mapping helpsto determine the areas of interest to be analyzedin combination with m-XRD, m-XANES, andm-EXAFS (Fig. 1.1). Another variation is quick-scanning XAS, which allows the collection of acomplete XANES or EXAFS spectrum in as littleas 300 ms (Dent, 2002; Ginder-Vogel et al., 2009).

The previous decade has seen a rapid devel-opment in the number and variety of synchro-tron user facilities at least partially dedicatedto the soil and environmental sciences. In theUnited States, these facilities include, GeoSoi-lEnviroCARS (GeoSoilEnviroCARS, 2008), withbeamlines located at both the Advanced PhotonSource (APS) and the National SynchrotronLight Source (NSLS); Envirosuite (EnviroSuite,2008), with beamlines located at the NSLS;and the Molecular Environmental Interface Sci-ence (MEIS) group (MEIS, 2008), located at theStanford Synchrotron Radiation Laboratory(SSRL). Recently, several excellent reviews onthe use of synchrotron techniques in environ-mental sciences have been published. Thesereviews focused on the use of synchrotrontechniques in low-temperature geochemistryand environmental science (Fenter et al.,2002), clay and soil sciences (Schulze et al.,1999), interpretation of data collected fromenvironmental systems (Kelly et al., 2008), andthe study of heavy metals in the environment(Sparks, 2005). Additionally, recent journal

Spatial

resolution Ele

men

tal

spec

ifici

ty

Sensitivity to local atomic

structure

In situanalysis

XAS

Surfaceprecipitation

Mineralogicaltransformation

Sorption/ transformation

Defects

Cr

0

Cou

nts

Fe

FeNi

Ni

1 2 3 4 5 6Energy (keV)

7 8 9 10 11

Mn

100 μm

Cr Fe

FIGURE 1.1 Advantageous characteristics of X-ray absorption spectroscopy (XAS) for analysis of heterogeneous soilsand sediments.

32. COMPLEMENTARY/ALTERNATIVE TECHNIQUES

issues on the impact of user facilities inenvironmental sciences have highlighted the crit-ical importance of synchrotron research in soilscience and related disciplines (Sutton, 2006).In this chapter, we will examine the applicationof awide variety of synchrotron X-ray techniquesto fundamental issues in environmental soilchemistry. However, first we provide a brief dis-cussion of complementary spectroscopic andmicroscopic techniques that are oftenused in con-junction with XAS investigations.


Although XAS is a powerful analytical tool,it is best applied as one of several analyticaltechniques in order to obtain the clearest pic-ture of the processes controlling elementalmobility in soil environments. In addition tothe traditional analytical techniques used insoil chemistry, other advanced spectroscopic


techniques are often used to complement theelementally specific data obtained from XAS.

2.1. Synchrotron X-Ray Diffraction

X-ray diffraction (XRD) is a fundamentaltechnique used in the characterization andidentification of crystalline materials in soilsand sediments. Conventional XRD uses mono-chromatic X-rays generated when high voltageis applied to an anode, generally Cu. Conven-tional XRD is generally limited to crystallinematerials; additionally, it is time consumingand often requires a relatively large sample vol-ume. The low flux and relatively low energy oflaboratory-based X-ray sources generally limitsthe geometry of the diffractometer, whichrequires data collection of diffracted X-rays,rather than transmitted X-rays, making analysisof hydrated or oxygen samples quite difficult.However, S-XRD offers exceptional resolutionand sensitivity, even on very small samples.This permits the identification and quantifica-tion of trace phases, which is not possibleusing conventional X-ray sources. Additionally,amorphous materials, thin films, and hydratedsamples can be analyzed due to the high fluxof synchrotron X-ray sources and specializedinstrumental configuration. With the use ofarea detectors or CCD cameras, similar to pro-tein crystallography beamlines, it is possibleto collect a single XRD pattern in a matter ofseconds, allowing for in situ, real-time monitor-ing of aqueous reactions.

2.2. X-Ray Photoelectron Spectroscopy

X-ray photoelectron spectroscopy (XPS) usesX-rays to excite electrons from molecular orbi-tals into the continuum. However, rather thanmeasuring X-ray absorption as a function ofenergy, XPS is conducted by using a fixedenergy source to excite electrons from the sam-ple and then measuring their kinetic energy.Since their kinetic energy is dependent on their

binding energy, different chemical speciescan be identified. As with conventional XRD,these experiments are commonly carried outin laboratories, using Al or Mg anodes to pro-duce monochromatic radiation. Since XPS mea-sures electrons, it is highly surface sensitive,generally interrogating only the first few 100sof angstroms of a surface. Additionally, recentadvances in synchrotron-based, ambient-pres-sure XPS measurements allow the measurementof hydrated samples (Ghosal et al., 2005).

2.3. Vibrational Spectroscopy(IR and Raman)

The first reported study of soil systems usingvibrational spectroscopy is the study of watersorption to montmorillonite in 1937 (Buswellet al., 1937). Consequent to this study, the useof vibrational spectroscopies, such as infrared(IR) and Raman, have greatly contributed toour understanding of the structure, bonding,and reactivity of soil particles (Johnston andWang, 2002). IR and Raman modes are verysensitive to small changes occurring in the localenvironment of the constituent atoms; there-fore, these spectral perturbations may be usedto gain insight into the interaction betweenabsorbates and the particle surface. The recentdevelopment of attenuated total reflectionFourier-transform infrared (ATR-FTIR) meth-ods has allowed the investigation of the particlesolid-water interface under environmentallyrelevant (i.e., hydrated) conditions and hasbecome one of the tools of choice to investigatethe solid-liquid interface (Hind et al., 2001).Additionally, time-resolved ATR-FTIR can beused to follow the surface chemistry and deter-mine kinetics of reactions occurring at mineralsurfaces in real time (Parikh et al., 2008).The development of Raman microscopeshas allowed Raman spectroscopy to be usedto analyze spatially heterogeneous samplesand is particularly useful in the analysis oforganics and mineral phases (Johnston and


Wang, 2002). The fate and transport of pollu-tants, behavior of exchangeable cations insoils, bonding mechanism of pesticides, specificadsorption of anions, and the behavior of waternear mineral surfaces are examples of researchareas that have benefited from vibrational spec-troscopy (Johnston and Wang, 2002). Vibra-tional spectroscopies are important tools usedto investigate the mineral-water interface; how-ever, they suffer from interferences from water,impurities, and the bulk solid itself (Johnstonand Wang, 2002).

2.4. Mossbauer Spectroscopy

Mossbauer spectroscopy is a nuclear spec-troscopy that has energy resolution sufficientto resolve the hyperfine structures of nuclearlevels (Murad and Cashion, 2003; Rancourt,1998). In a given experiment, it always operateson a single g-ray transition, the Mossbauertransition, between the ground state and anexcited state of one isotope in the sample.Mossbauer spectroscopy probes the local elec-tronic structure and, by extension, the localcrystallographic, magnetic, and chemical envir-onments. One of the primary limitations of itsapplication to environmental systems is thatthe isotope of interest has a large enough recoil-less fraction to exhibit measurable Mossbauereffects. To date, only 45 isotopes, including57Fe, have a large enough Mossbauer effect tobe measurable. Other isotopes of interest toenvironmental scientists that may be measuredvia Mossbauer spectroscopy include 119Sn,151Eu, 129I, and 121Sb. Another factor that maylimit the application of Mossbauer spectros-copy to environmental systems is the complex-ity of interpreting the parameters controllingthe relevant hyperfine interactions; spectralanalysis often combines advanced data analysisand statistical methods with quantum mechan-ical and physical constraints in order to obtaina stable solution (Murad and Cashion, 2003).Nevertheless, Mossbauer spectroscopy has

been extensively used in the study of Fe-bearing minerals in environmental settings.Recently, Williams and Scherer (2004) usedMossbauer spectroscopy to determine the reac-tion mechanism of Fe(II) sorption and oxida-tion at Fe(III) (hydr)oxide mineral surfaces.Many additional examples of the applicationsof Mossbauer spectroscopy are available inMurad and Cashion (2003).

2.5. Nuclear Magnetic Resonance andElectron Paramagnetic ResonanceSpectroscopies

NMR and EPR spectroscopies have beenimportant analytical tools in chemistry, mate-rial sciences, and organic geochemistry for sev-eral decades (Nanny et al., 1997); however, theyhave only recently become recognized as usefultools in the wider field of environmental sci-ence. NMR is a noninvasive probe that can beused to identify individual compounds andaids in determining structures of large macro-molecules and examination of certain reactionkinetics. NMR spectroscopy measures the mag-netic properties of NMR active nuclei (i.e., 1H,13C, 15N, 23Al, 31P, and 133Cs) that are influ-enced not only by their chemical environments,but also by their physical interactions with theirenvironments. Examination of environmentallyrelevant systems is significantly hampered bythe low prevalence of most NMR active nuclei;however, with further developments in instru-mentation, the sensitivity and utility of thistechnique is increasing. NMR has most fre-quently been applied to the study of soilorganic matter; however, NMR has the poten-tial to examine the pore size distribution(Chen et al., 2002b) as well as water flow andtransport processes in soil (Chen et al., 2002a;Kinchesh et al., 2002).

EPR is a technique for studying chemical spe-cies that have one or more unpaired electrons. Insoil science, it is generally applied to transitionmetals. Although the basic physical principles of


EPR are similar to those of NMR, the techniquemeasures the relaxation of excited electron spinsrather than excited nuclear spins. While EPR isless widely used than NMR, its ability to measureonly paramagnetic species gives it tremendouselemental specificity, and EPR spectra can berecorded in just a few milliseconds, making it anideal technique for measuring transition metalreaction kinetics. For example, Fendorf et al.(1993) studied Mn(II) sorption to birnessite on atimescale ofmillisecondsusing stop-flowelectronparamagnetic resonance (SF-EPR) spectroscopy.

2.6. Electron Microscopy

Scanning electron microscopy (SEM) andtransmission electron microscopy (TEM) pro-vide unique techniques to visually examineenvironmental samples. SEM is often used toexamine the morphology of environmental par-ticles >100 nm in size. Additionally, when cou-pled with energy dispersive spectroscopy(EDS), it allows mapping of major and minorelements on a scale of �1 mm. TEM analysiscan be used to examine much smaller featuresof soil particles, with high-resolution TEM(HR-TEM) providing a point-to-point resolutionbetter than 1 A. TEM is often coupled with EDSto provide elemental makeup. Selected areaelectron diffraction (SAED) can be used to iden-tify crystalline materials on the nano-scale, whileelectron energy loss spectroscopy (EELS) isbeing used to identify elemental oxidation stateon the nanometer scale. For example, Livi et al.(2009) used high-resolution TEM to confirm theamorphous nature of Ni layered double hydrox-ide (LDH) precipitates on pyrophyllite.

3. HISTORICAL ASPECTS OF XASAPPLICATIONS TO MINERAL/

WATER INTERFACIAL REACTIONS

The employment of in situ (under naturalconditions where water is present) molecular-

scale techniques, such as synchrotron-basedEXAFS and XANES spectroscopies, XRF, XRD,and microtomography, have been widelyused over the past 20 years to study metal(loid)and radionuclide sorption phenomena atthe mineral/water interface and to determinethe chemical form (speciation) and distributionof environmental contaminants in heteroge-neous natural materials, such as mineralsand soils. Arguably, they have revolutionizedmany scientific fields, including soil science,geochemistry, and environmental sciences(Fig. 1.1).

Synchrotron light sources were first avail-able to general users in 1974. The first pub-lished study that employed XAS to directlydelineate the type of surface complexesformed at mineral/water interfaces was theseminal research of Hayes et al. (1987). Study-ing selenite and selenate sorption at the goe-thite/water interface, selenite was observedto primarily form an inner-sphere surfacecomplex while selenate formed predomi-nantly an outer-sphere complex. This wasthe first direct confirmation that inner- andouter-sphere surface complexes, first pro-posed by Werner Stumm and his co-workers,form on natural mineral surfaces. This studycatalyzed a plethora of other studies overthe next two decades, resulting in the activeinvolvement by soil scientists in using syn-chrotron radiation techniques to study reac-tion mechanisms at mineral surfaces andspeciation of inorganic contaminants in soilsand biosolids. The growing use of these tech-niques, and their importance in addressinga multitude of environmental challengesinvolving soil, water, and air quality haveresulted in the creation of a new multidisci-plinary field referred to as molecular environ-mental science, which is the study of thechemical and physical forms and distributionof contaminants in soils, sediments, wastematerials, natural waters, and the atmosphereat the molecular level.

TABLE 1.1 Predominant sorption mechanisms formetals and metalloids on mineral surfaces

Predominant Sorption Mechanism

Metals

Cadmium Inner-sphere

Cobalt Inner-sphere (low loading), Co-Al hydroxideprecipitates (high loading)

Chromium Inner-sphere (low loading), Cr hydroxideprecipitates (high loading)

Copper Inner-sphere

Lead Inner-sphere (low loading), Surface polymers(high loading)

Nickel Inner-sphere (low loading), Ni-Al hydroxideprecipitates (high loading)

Strontium Outer-sphere

Zinc Inner-sphere (low loading), Zn-Al hydroxideprecipitates (high loading)

Metal(loids)

Arsenite Inner-sphere

Arsenate Inner-sphere

Borate Inner-sphere

Carbonate Inner-sphere

Chromate Inner-sphere

Phosphate Inner-sphere

Selenate Outer-sphere

Selenite Inner-sphere

Sulfate Outer-sphere

Note: These are the predominant sorption mechanisms thatform on mineral surfaces based on surface spectroscopy(e.g., XAS and FTIR) studies. However, environmental fac-tors such as pH, surface loading, time, and ionic strengthcan affect the type of surface complex and there may bemultiple surface products that occur on a particular min-eral. References for each metal(loid) can be found in Sparks(2002).

74. APPLICATIONS OF XAS TO ELUCIDATE SOIL CHEMICAL PROCESSES AND REACTIONS

4. APPLICATIONS OF XAS TOELUCIDATE SOIL CHEMICALPROCESSES AND REACTIONS

4.1. Adsorption of Metal(loids) on SoilComponents

Since 1987, hundreds of studies using bulkXAS (Brown and Parks, 2001; Brown andSturchio, 2002; Brown et al., 2006; Sparks, 1995,2004c) have provided extensive information onmetal and oxyanion sorption on metal hydr(oxi-des), phyllosilicates, humic substances, and othernatural materials. This information includesstructure, stoichiometry, attachment geometry(inner vs outer sphere, monodentate vs bidentateor tridentate), the presence of multinuclear com-plexes and precipitate phases, and the presenceof ternary surface complexes, when complexingligands are present in solution (Brown andSturchio, 2002; Sparks, 2004c) (Table 1.1). The typeof surface complexes onphyllosilicates andmetal-(oxyhydr)oxides that occurwith low atomic num-ber metals, such as Al, B, Ca,Mg, S and Si, are noteasy to ascertain using XAS under in situ condi-tions. However, major advances are being madein the area of soft X-ray spectroscopy that willenable direct determination of the types ofsurface complexes which form with these metals.

Based on molecular scale studies, includingmany using XAS, one can predict that alkalineearth cations, Mg2þ, Ca2þ, Sr2þ and Ba2þ, pri-marily form outer-sphere complexes while thedivalent first-row transition metal cations,Mn2þ, Fe2þ, Co2þ, Ni2þ, Cu2þ and Zn2þ, andthe divalent heavy metal cations such as Cd2þ,Hg2þ and Pb2þ primarily form inner-spherecomplexes. At higher metal loadings and pHs,sorption of metals such as Co, Cr, Ni and Znon phyllosilicates and metal-(hydr)oxides canresult in the formation of surface precipitates(Fig. 1.1 and Table 1.1). The formation of thesemultinuclear and precipitate phases will be dis-cussed in more detail later.

Although there are currently experimentallimitations to using in situ molecular scale tech-niques to directly determine the type of surfacecomplexes that the anions NO3

�, Cl� and ClO�4


form on mineral surfaces, one can propose thatthey are sorbed as outer-sphere complexes andsorbed on surfaces that exhibit a positivecharge (Zhang and Sparks, 1990). Someresearchers have also concluded that SO2�

4

(Zhang and Sparks, 1990) can be sorbed as anouter-sphere complex; however, there is otherevidence that SO2�

4 may also sorb as an inner-sphere complex (Paul et al., 2007). There isdirect spectroscopic evidence to show that sele-nate can sorb as both an outer-sphere and aninner-sphere complex depending on environ-mental factors (Hayes et al., 1987; Manceauand Charlet, 1994; Sparks, 2002, 2004a).

Most other anions such as molybdate, arse-nate, arsenite, selenite, phosphate, and silicateappear to be strongly sorbed as inner-spherecomplexes, and sorption occurs through aligand exchange mechanism (Table 1.1). Thesorption maximum is often not sensitive toionic strength changes. Sorption of anions vialigand exchange results in a shift in the pointof zero charge (pzc) of the sorbent to a moreacid value (Sparks, 2002, 2004c).

Bulk XAS probes an area of several squaremillimeters and provides information on thelocal chemical environment of a surface. Thus,where more than one type of surface speciesis present, bulk XAS may detect only the pri-mary (or average) type of surface product/

TABLE 1.2 Effect of I and pH on type of P

I (M) pH Removal from solution (%) Adso

0.1 6.77 86.7

0.1 6.31 71.2

0.006 6.76 99.0

0.006 6.40 98.5

0.006 5.83 98.0

0.006 4.48 96.8

Source: Strawn and Sparks (1999), with permission.aBased on results from XAFS data analysis.

species in the bulk sample (i.e., sums over allgeometric configurations of the target atom).Consequently, while it may be concluded thatthe primary surface complex is inner-sphere,this does not necessarily mean that outer-sphere complexation is not occurring. Recently,through the use of X-ray scattering measure-ments to study metal(loid) binding on singlecrystal surfaces, Catalano et al. (2008) showedthat arsenate surface complexation wasbimodal, with adsorption occurring simulta-neously as inner- and outer-sphere species.Environmental factors such as pH, surfaceloading, ionic strength, type of sorbent, andtime also affect the type of sorption complexor product. An example of this is shown forPb sorption on montmorillonite over an ionicstrength (I) of 0.006-0.1 and a pH range of4.48-6.77 (Table 1.2). Employing XAS analysis,at a pH of 4.48 and an I of 0.006, outer-spherecomplexation on basal planes in the interlayerregions of the montmorillonite predominated.At a pH of 6.77 and I of 0.1, inner-sphere com-plexation on edge sites of montmorillonite wasmost prominent, and at pH of 6.76, I of 0.006and pH of 6.31, I of 0.1, both inner- and outer-sphere complexation occurred. These data areconsistent with other findings that inner-spherecomplexation is favored at higher pH and ionicstrength. Clearly, there is a continuum of

b adsorption complexes on montmorillonite

rbed Pb(II) (mmol/kg) Primary adsorption complexa

171 Inner sphere

140 Mixed

201 Mixed

200 Outer sphere

199 Outer sphere

197 Outer sphere


adsorption complexes that can exist in soils(Sparks, 2002, 2004c).

4.2. Metal(loid) Surface Precipitation/Dissolution

As the amount of metal cation or anionsorbed on a surface (surface coverage or load-ing that is affected by the pH at which sorptionoccurs) increases, sorption can proceed frommononuclear adsorption to surface precipita-tion (a three-dimensional phase) (Fig. 1.1 andTable 1.1). There are several types of surfaceprecipitates that may arise via polymeric metalcomplexes (dimers, trimers, etc.) which formon mineral surfaces or via the sorption of aque-ous polymers (Chisholm-Brause et al., 1990).Homogeneous precipitates often form on a sur-face when the solution becomes saturated andthe surface acts as a nucleation site (Fig. 1.1).After adsorption attains monolayer coverage,it continues on the newly created sites, causinga precipitate on the surface (Chisholm-Brauseet al., 1990; Farley et al., 1985; McBride, 1991;O’Day et al., 1994; Sposito, 1986). When the pre-cipitate consists of chemical species derivedfrom both the aqueous solution and dissolutionof the mineral, it is referred to as a coprecipi-tate. The composition of the coprecipitate var-ies between that of the original solid and apure precipitate of the sorbing metal. The ionicradius of the sorbing metal and sorbent ionsmust be similar for coprecipitates to form.Thus, Co(II), Mn(II), Ni(II), and Zn(II) (Chapter8) form coprecipitates on sorbents containingFe (III), Al(III), and Si(IV) and possibly Pb(II),which is considerably larger (1.20 A). Copreci-pitate formation is generally limited by the rateof mineral dissolution, rather than the lack ofthermodynamic favorability (McBride, 1994;Scheidegger et al., 1998). If the formation of aprecipitate occurs under solution conditionsthat would, in the absence of a sorbent, beundersaturated with respect to any knownsolid phase, this is referred to as surface-

induced precipitation (Sparks, 2004c; Sparksand Borda, 2008; Towle et al., 1997).

Thus there is often a continuum betweensurface complexation (adsorption) and surfaceprecipitation. At low surface coverage, surfacecomplexation (e.g., outer- and inner-sphereadsorption) tends to dominate. As surface load-ings increase, nucleation occurs and results inthe formation of distinct entities or aggregateson the surface. As surface loadings increasefurther, surface precipitation becomes the dom-inant mechanism.

Using in situ bulk XAS, it has been shown by anumber of scientists that multinuclear metalhydroxide complexes and surface precipitatesCo(II), Cr(III), Cu(II), Ni(II) and Pb(II) can formon metal oxides, phyllosilicates, soil clays andsoils (Bargar et al., 1995; Charlet and Manceau,1992; Chisholm-Brause et al., 1990, 1994; Elzingaand Sparks, 1999; Fendorf et al., 1994; Ford andSparks, 2000; O’Day et al., 1994; Papelis andHayes, 1996; Roberts et al., 1999; Roe et al.,1991; Scheckel and Sparks, 2001; Scheideggeret al., 1996a,b, 1997, 1998; Thompson et al.,1999a,b; Towle et al., 1997). These metal hydrox-ide phases occur at metal loadings below a theo-retical monolayer coverage and in a pH rangewell below the pH where the formation of metalhydroxide precipitates would be expected,according to the thermodynamic solubilityproduct (Scheidegger and Sparks, 1996; Sparks,2002, 2004b; Sparks and Borda, 2008).

Scheidegger et al. (1997) were the first toshow that sorption of metals, such as Ni, onan array of phyllosilicates and Al oxide canresult in the formation of mixed metal-Alhydroxide surface precipitates, which appearto be coprecipitates. The precipitated phaseshares structural features common to thehydrotalcite group of minerals and the layereddouble hydroxides (LDH) observed in catalystsynthesis. The LDH structure is built of stackedsheets of edge-sharing metal octahedra, con-taining divalent and trivalent metal cationsseparated by anions between the interlayer

LDH

Ni+Al

Ni+Al

NO3, Cl, CO3, [SiO2]n

α-Ni(OH2)

Ni

Ni

NO3, Cl, CO3, [SiO2]n

FIGURE 1.2 Structure of Ni-Al layerdouble hydroxide (LDH) showing brucite-like octahedral layers in which Al3þ substi-tutes for Ni2þ creating a net positive chargethat is balanced by hydrated anions in theinterlayer space (Ford et al., 1999).


spaces (Fig. 1.2). The general structural formulacan be expressed as [Me2þ1�x Me3þx (OH)2]

xþ�(x/n)An-mH2O, where for example, Me2þcould be Mg(II), Ni(II), Co(II), Zn(II), or Mn(II), and Fe(II) andMe3þ could be Al(III), Fe(III), and Cr(III) (Towleet al., 1997). The LDH structure exhibits a net posi-tive charge per formula unit, which is balanced byan equal negative charge from interlayer anionsAn�, such as Cl�, Br�, I�, NO�

3 , OH�, ClO�4 , and

CO2�3 ; water molecules occupy the remaining

interlayer space (Allmann, 1970; Taylor, 1984).The minerals takovite, Ni6Al2(OH)16CO3�H2Oand hydrotalcite, Mg6Al2(OH)16CO3�H2O areamong the most common natural mixed-cationhydroxide compounds containing Al (Sparks,2004b; Sparks and Borda, 2008; Taylor, 1984).Recently, Livi et al. (2009), using an array ofmicro-scopic techniques including analytical electronmicroscopy (AEM), high-resolution transmissionelectron microscopy (HR-TEM), and powderXRD, conducted studies to elucidate the natureof Ni hydroxide precipitates, using the same envi-ronmental conditions employed by Scheideggeret al. (1997) and reaction times ranging from 1 hto 5 years. While the precipitate phase had abonding environment like a Ni-Al LDH, the pre-cipitate was amorphous.

Mixed Co-Al and Zn-Al hydroxide surfaceprecipitates also can form on aluminum-bearing metal oxides and phyllosilicates (Fordand Sparks, 2000; Thompson et al., 1999a,b;Towle et al., 1997). This is not surprising asCo2þ, Zn2þ, and Ni2þ all have ionic radii simi-lar to Al3þ, enhancing substitution in the

mineral structure and formation of a coprecipi-tate. However, surface precipitates have notbeen observed with Pb2þ, as Pb2þ is too largeto substitute for Al3þ in mineral structures(Sparks, 2002, 2004c; Sparks and Borda, 2008).

The mechanism for the formation of metalhydroxide surface precipitates is not clearlyunderstood. It is clear that the type of metal iondetermineswhethermetal hydroxide surfacepre-cipitates form, and the type of surface precipitateformed, that is, metal hydroxide or mixed metalhydroxide is dependent on the sorbent type.Additionally, the reaction pH, which controlsthe degree of metal loading, is a major factor indetermining the formation of metal hydroxideprecipitates: for example below pH 6.5, metalhydroxide precipitates generally do not form onmineral surfaces or in soils (Elzinga and Sparks,1999, 2001; Roberts et al., 1999, 2003).

The formation of metal hydroxide surfaceprecipitates is an important mechanism ofmetal sequestration. As the surface precipitatesage, metal release is greatly reduced (Fig. 1.3).Thus, the metals are less prone to leachingand less bioavailable to plants and microbes.Peltier et al. (2010) reacted three soils of vary-ing mineralogy and organic matter contentwith 3 mM Ni at two pHs, 6 and 7.5, and eval-uated Ni bioavailability using a biosensor. AtpH 6, surface precipitates did not form andmost of the Ni was bioavailable. However, atpH 7.5, where precipitates were observed toform from XAS analyses, Ni bioavailability ismarkedly reduced.

0

100

75

50

25

05 10 15

+ + + + + + + + + + +

+

X

X X X X X X X X X X X X X X

HNO3 pH 6.0

2 Year1 Year 6 Months 3 Months 1 Month 24 Hours 12 Hours 1 Hour

Number of replenishments

Rel

ativ

e N

i rem

aini

ng o

n so

lid (

%)

FIGURE 1.3 Dissolution of Ni from surface precipitatesformed on pyrophyllite at residence times of 1 h to 2 years.The figure shows the relative amount of Ni2þ remaining onthe pyrophyllite surface following extraction for 24 h peri-ods (each replenishment represents a 24 h extraction) withHNO3 at pH 6.0.

Inner-sphere outer-sphere

Precipitation Solid phase

transformation Adsorption

Hydrotalcite-like precipitates

Recrystallization Ostwald ripening

sec min hours days months years

FIGURE 1.4 Changes in sorption processes with timeshowing a continuum from adsorption to precipitation tosolid phase transformation (Sparks, 2002).


The decrease in metal release and bioavail-ability is linked to the increasing silicationof the interlayer of the LDH phases withincreased residence time, resulting in a mineraltransformation from a LDH phase to a precur-sor phyllosilicate surface precipitate (Ford andSparks, 2000; Ford et al., 1999). The mechanismfor this transformation is thought to be due todiffusion of Si, originating from weathering ofthe sorbent, into the interlayer space of theLDH, replacing the anions such as NO�

3 . Poly-merization and condensation of the interlayerSi slowly transforms the LDH into a precursormetal-Al phyllosilicate. The metal stabilizationthat occurs in surface precipitates on Al-freesorbents (e.g., talc) may be due to Ostwald rip-ening, resulting in increased crystallization;however, as noted earlier, microscopic analyseshave shown that Ni-Al LDH phases appear tobe amorphous (Livi et al., 2009). Using acid-solution calorimetry and results from previouscalorimetry studies, Peltier et al. (2006) demon-strated that the enthalpy of formation of LDHphases is more exothermic, indicating greaterstability, in order of Cl<NO3<SO4<CO3<Si of

interlayer anionic composition, and that LDHphases were much more stable than a Ni(OH)2 phase.

In short,with time, one sees thatmetal sorptionon soil minerals can often result in a continuumofprocesses, from adsorption to precipitation tosolid-phase transformation (Fig. 1.4), particularlyin the case of certain metals, such as Co, Ni, andZn. The formation of metal surface precipitatescould be an important mechanism for sequester-ing metals in soils, such that they are less mobileand bioavailable. Such products must be consid-ered when modeling the fate and mobility ofmetals like Co2þ, Mn2þ, Ni2þ and Zn2þ in soiland water environments (Sparks, 2002, 2004c;Sparks and Borda, 2008).

4.3. Redox Reactions on SoilComponents

The mobility and environmental threat ofmany contaminants and nutrients is deter-mined in large part by their redox state andthe redox conditions of the soil environmentsin which they exist. XAS is an ideal tool fordetermining the oxidation states of elementsin complex environmental media such as soilsand sediments. Changes in oxidation state ofan element result in a change in the energyrequired to excite core electron into the contin-uum, which manifests itself by a shift in thehalf-height position of the absorption edge foran element and can be determined using

Multiple scattering1.50

1.25

1.00

0.75

0.50

0.25

0.0017,140 17,160 17,180

Energy (eV)

17,200 17,220 17,240

Nor

mal

ized

fluo

resc

ence

FIGURE 1.5 UraniumLIII-edgeXANESspectra of uraninite (dotted) and uranylnitrate hexahydrate (solid). The arrow indi-cates the contribution of multiple scatteringfrom the uranyl moiety.


XANES spectroscopy. In addition, a change inthe element’s local coordination environmentoccurs and is manifested in EXAFS spectra ofthe element. This is readily exemplified by thereduction of soluble U(VI)O2

2þ (and its variouscomplexes) to U(IV)O2, which results in theabsorption edge shifting �3 eV lower in energy(Fig. 1.5). Additionally, the transition from thehighly linear uranyl ion results in the loss of alarge shoulder, due to multiple scattering, thatimmediately follows the white line (Fig. 1.5)(Kelly et al., 2002, 2007, 2008). Further applica-tions of XAS to environmental uranium (U)chemistry are described in detail by Kelly inChapter 14 in this book.

In addition to U, XAS has been used toinvestigate metal and oxyanion cycling inmany redox active systems (Bank et al., 2007;Bostick et al., 2002; Fendorf et al., 1992, 2002;Ginder-Vogel and Fendorf, 2008; Hansel et al.,2003a, 2005; Manning et al., 2002; Webb et al.,2005a,b;). The primary elements of interestinclude environmentally predominant transi-tion metals, such as Fe and Mn, in addition toelements that exist in the environment primar-ily as oxyanions, such as As, Cr, and Se. Many

of these redox reactions are driven by bacterialmetabolism and result in the formation ofhydrated mineral phases, making in situ (i.e.,wet) analysis critical to determining the opera-tive redox pathways.

4.3.1. Transition Metal Oxides

The environmental prevalence of iron andmanganese oxides, hydroxides, and oxyhydr-oxides [(hydr)oxides], coupled with the largequantity of their highly reactive surface area,make them an important pool of reactivity, withregard to contaminant and nutrient absorptionand redox transformation. However, their reac-tivity, in terms of both sorption and redox chem-istry, is radically altered by changes in redoxconditions, especially as they impact the miner-alogy and surface chemistry of Fe and Mn(hydr)oxide phases.

Ferrihydrite, a poorly crystalline Fe(III)(hydr)oxide, is often found in soils and sedi-ments that oscillate between reducing and oxi-dizing conditions. Due to its high surface areaand intrinsic reactivity, ferrihydrite serves as acritical sink for contaminants (e.g., As and U)and nutrients (e.g., P); however, until recently,


the impact of reducing conditions on the evolu-tion of Fe(III) mineral mineralogy and itssubsequent ability to retain and interact withcontaminants and nutrients have been poorlyunderstood. XAS, coupled with other advancedanalytical techniques, has played a key role infurthering our understanding of this importantenvironmental process. Ferrihydrite is consid-ered the most bioavailable Fe(III) (hydr)oxidefor dissimilatory metal reducing bacteria,which couple the oxidation of organic matterand H2 to the reduction of Fe(III) and are ubiq-uitous in saturated soils and sediments. Upondissimilatory metal reduction, the productionof Fe(II) may result either in Fe(III) mineral dis-solution or the rapid transformation of poorlycrystalline Fe(III) (hydr)oxides into more crys-talline and/or mixed Fe(II)/Fe(III) (hydr)oxi-des, depending on the intensity of Fe(III)reduction. The application of XAS to the studyof Fe oxides is discussed in more detail inChapter 8.

XAS has played a vital role in the identifica-tion and quantification of Fe (hydr)oxide miner-alogy. In batch experiments using ferrihydrite,the solid phases formed during microbial ironreduction are dominated by the metabolicbyproducts of Fe reduction, including ferrousiron and bicarbonate; and generated mineralsare rich in Fe(II) and bicarbonate, including sid-erite, green rusts, and magnetite (Fredricksonet al., 1998; Liu et al., 2001). In general, this typeof concentrated batch reaction produced highenough concentrations of the products ofbiological Fe reduction to be detected using tra-ditional techniques, such as laboratory-basedXRD and Mossbauer spectroscopy. However,in order to investigate the transformation of Fe(hydr)oxide minerals under more natural condi-tions, Fendorf and co-workers (Hansel et al.,2003a, 2005; Kocar et al., 2006; Nico et al., 2010)transitioned to Fe oxide-coated sand, whichwas packed into columns. This allowed Fe bio-mineralization to be studied under conditionsthat transported microbial metabolites away

from the reaction site. Although a variety of ana-lytical methods were used to analyze the resul-tant Fe mineralogy, including SEM, HR-TEM,XRD, and Mossbauer spectroscopy, they foundthat the most accurate and quickest method thatallowed the quantification of the various ironmineral phases was EXAFS spectroscopy com-bined with linear combination analysis (Hanselet al., 2003a, 2005). Using this technique, Hanselet al. (2005) demonstrated that the mineralogicaltransformation of Fe(III) (hydr)oxides is depen-dent on the Fe(III) flux resulting from microbialiron reduction. At high Fe(II) concentrations,the dominant mineralogical products includemagnetite and goethite, while lepidocrocite andgoethite were observed at lower Fe(II) concen-trations. Interestingly, more reduced phases,such as green rusts and Fe(II)-bearing minerals,were not observed in any experiments whereflow conditions were used. This pioneeringwork has formed the basis for a series of studiesthat examine the effect of microbial iron reduc-tion on the incorporation of other metals andmetalloids, such as U and Cr, into iron oxidemineral structures during biological Fe reduc-tion and abiotic redox reactions.

Widespread U contamination is presentthroughout the world as a result of natural Udeposits, mining activities, and nuclear weaponsand fuel production. Environmental U specia-tion is dominated by two oxidation states withmarkedly different properties. U(VI), as the ura-nyl cation (UO2

2þ), is thermodynamically stableunder oxidizing conditions and is highly solublein the presence of dissolved bicarbonate.Although uranyl sorption to solids such as Fe(hydr)oxides may be appreciable, it is largelyreversible and is thus subject to changes in aque-ous conditions and Fe (hydr)oxide mineralogy.Dissolved U can be reduced through the trans-formation of U(VI) into U(IV), forming the soliduraninite (UO2), which is sparingly solubleunder most environmental conditions. How-ever, UO2 may be reoxidized, thereby releasingU, by many oxidants including O2, NO3

�, Mn


(IV) oxides, and Fe(III) (hydr)oxides (Ginder-Vogel and Fendorf, 2008). Given the prevalenceof Fe (hydr)oxide minerals in the environmentand their critical role in controlling U mobility,much research has focused on the interactionof U complexes with iron, particularly as(hydr)oxide mineral phases as discussed byKelly in Chapter 14 of the book. However, wehave only recently begun to understand the roleof iron (hydr)oxides in catalyzing uraninite oxi-dation and incorporation of U into the Fe(hydr)oxide mineral structure. Ginder-Vogelet al. (2006) and Ginder-Vogel and Fendorf(2008) used a combination of S-XRD and XASto identify and quantify Fe oxide minerals pro-duced during the oxidation of biologically preci-pitated uraninite by ferrihydrite. This workrevealed that ferrihydrite is capable of oxidizingbiogenic uraninite; however, the oxidation reac-tion is ultimately limited by the conversion offerrihydrite into more thermodynamically stableforms, such as lepidocrocite and goethite(Fig. 1.6). Interestingly, during Fe(II) catalyzedferrihydrite remineralization, U(VI) is substi-tuted for between 3% and 6% of iron sites intransformation products (Nico et al., 2010). Thesubstituted U is stabile even in the presence of

100

Ferrihydrite

Goeth

Lepido

80

60

Mol

e pe

rcen

t iro

n

40

20

00 10 20

Time (h)30 4

strongly complexing solutions (30 mM KHCO3)and during redox cycling (Stewart et al., 2009)and can be environmentally important as along-term U sink.

Similar to U, Cr predominantly exists in theenvironment in two oxidation states, Cr(III) orCr(VI). Chromium(VI), chromate, is highlytoxic to humans and is quite soluble and hencemobile under most common environmentalconditions. Conversely, Cr(III) is less toxic andgenerally immobile. There are many abioticand biotic pathways of Cr reduction in theenvironment; however, under neutrophilic con-ditions, kinetic considerations favor Cr(VI)reduction by either dissolved Fe(II) or Fe(II)-bearing mineral phases (Fendorf et al., 2002).The identification and quantification of Cr oxi-dation states, using XANES spectroscopy, isquite straightforward due to a prominent pre-edge peak at 5993 eV, when Cr is present asCr(VI) (Peterson et al., 1996, 1997), which iscaused by a bound-state 1s to 3d transition.This transition is forbidden for centrosymmet-ric, octahedral Cr(III)O6, but allowed for thenoncentrosymmetric, tetrahedrally coordinatedCr(VI)O4 molecule due to mixing of the Cr(3d)and O(2p) orbitals. The intensity of this pre-edge

ite

crocite

0 50

FIGURE 1.6 Transformation of ferri-hydrite by Fe(II) generated during urani-nite oxidation. Reactions were conductedwith 3.0 mM KHCO3, 30.7 m2/L UO2,and 23.5 m2/L mM-Fe as ferrihydrite in3 mM HCO3

� at pH 7.2. Percentages(�5%) were determined from linear com-bination fits of k3-weighted Fe EXAFSspectra (k ¼ 1-14) (Ginder-Vogel et al.,2010).


feature can be used to quantify the proportionof Cr(VI) in a sample, at Cr(VI) concentrations>1-5%. Using this technique, several studieshave examined Cr speciation in reducing, Cr-contaminated soils (Bank et al., 2007; Fendorfand Zasoski, 1992; Kendelewicz et al., 2000;Patterson et al., 1997;) and determined that lessmobile Cr(III) was the dominant species in thesesystems; however, this may not always be thecase in oxic systems. Using a combination ofbulk XAS, m-XRF, and m-XRD to analyzesediments from the arid, oxic, Cr-contaminatedHanford site in Eastern Washington state, inconjunction with flow-through column experi-ments, Ginder-Vogel et al. (2005) identifiedseveral Fe(II)-bearing phyllosilicate minerals asthe primary reductants of Cr(VI) in these sedi-ments. However, depending on geochemicalconditions, as much as 50% of the solid-phaseCr remained in the hexavalent form.

In large part, the solubility of Cr(III) that isabiotically reduced by Fe(II) is determined inlarge part by the ratio of Cr(III) to Fe(III) in the(hydr)oxide precipitate of the general formulaFexCr1-x(OH)3 (Sass and Rai, 1987). The solubil-ity of Cr(III)-hydroxide precipitates is propor-tional to the ratio of Cr(III) to Fe(III), withincreased quantities of Fe(III) stabilizing thesolid (Sass and Rai, 1987). Generally, the ratioof Cr to Fe in hydroxide solids is determinedusing ex situ extraction methods, which may becompromised by Fe contamination of environ-mental samples; however, EXAFS spectroscopypresents an alternative method for determiningthis molar ratio (Hansel et al., 2003b). The inten-sity of the corner-sharing peak at 3.48-3.56 A,normalized to the first Cr-Cr(Fe) shell, has a lin-ear relationship with the mole fraction of Cr(III)within a mixed Cr(III)-Fe(III) solid.

4.4. Speciation of Soil Contaminants

With the advent of third-generation synchro-tron light sources that have higher flux andhigher brightness X-rays, together with state-

of-the-art detectors and improved beamlineoptics, microfocused beams for spectromicro-scopy and imaging are available to study het-erogeneous natural systems such as soils,biosolids, and plants. With these techniques, itis possible to study areas of 1 mm2 and focusthe beam to spatial resolutions of <5 mm,enabling direct speciation of contaminants inheterogeneous materials. This has been a majoradvance in the past 10 years, and coupled withmicrofocused XRF, one can also determine thedistribution and association of metals andoxyanions in complex natural systems. Whilebulk XAS can and should be used to determinethe major species that are found in such sys-tems, one needs to couple these studies withm-XAS and XRF investigations. There are sev-eral disadvantages in using bulk XAS to speci-ate contaminants in heterogeneous systemssuch as soils. Soils are complex and contain anarray of inorganic and organic components,including humic substances, phyllosilicates,metal hydr(oxides), macro- and micropores,and microorganisms, all closely associated witheach other. In such systems, some microenvir-onments contain isolated phases in higherconcentrations relative to the average in thetotal (bulk) matrix. For example, the oxides,phyllosilicates, and microorganisms in the rhi-zosphere have quite a different chemical envi-ronment as compared to the bulk soil. Thesephases may be very reactive and important inthe partitioning of metals, but may be over-looked using other analytical techniques thatmeasure the average of all phases. In systemswhere more than one type of surface species ispresent, bulk XAS will detect only the primary(or average) type of surface product/species inthe bulk sample. However, themost reactive sitesin soils generally have particle sizes in themicrometer range, andmetal speciationmayvaryover regions of a few 100 mm2. The existence ofmultiple species in soils results in overlappingatomic shells and it is difficult to ascertainthe precise metal speciation with bulk XAS.


Minor metal-bearing phases, even though theymay constitute the most reactive or significantspecies, may not be successfully detected withbulk analyses (McNear et al., 2005).

Some excellent reviews have appeared onthe application of m-XAS and XRF to speciatecontaminants in heterogeneous systems(Bertsch and Hunter, 1998; Bertsch andSeaman, 1999; Hunter and Bertsch, 1998;Manceau et al., 2003). A number of investiga-tions have appeared over the past 10 years onthe speciation of metals and oxyanions, soils,plants, biosolids, and fly ash (see references inBrown and Sturchio, 2002).

4.4.1. Microscale Speciation of Metal(loids)

Some examples of studies that have usedm-XAS and XRD to speciate metal(loids) in soilsinclude those by Hunter and Bertsch (1998),Manceau et al. (2000), Isaure et al. (2002),Strawn et al. (2002), Roberts et al. (2005),Ginder-Vogel et al. (2005), Nachtegaal et al.(2005), Arai et al. (2006), McNear et al. (2007),and Grafe et al. (2008). A few examples of thesestudies will be summarized below. Additionaldetails on m-XAS applications and data ana-lyses methods for heterogeneous systems canbe found in Manceau et al. (2000), Bertsch andHunter (2001) and Chapter 2.

Nachtegaal et al. (2005) investigated the spe-ciation of Zn in smelter contaminated soilsfrom a large site in Belgium in which part ofthe site had been remediated by adding berin-gite, an aluminosilicate material, compost, andplanting metal-tolerant plants. The other por-tion of the site was not treated. The objectivesof the study were to determine how Zn specia-tion differed in the remediated and nonreme-diated soils. Specifically, the study sought todetermine if Zn-LDH phases were present inthe soils, and the stability of the Zn under dif-ferent environmental conditions.

In the study of Nachtegaal et al. (2005),both mineral (e.g., willemite, hemimorphite,spalerite) and sorbed (Zn-LDH) Zn species

predominated. The speciation differences inthe remediated and nonremediated soils wereslight, with the major difference being the pres-ence of kerolite, a Zn phyllosilicate phase,found in the remediated site (Fig. 1.7). Desorp-tion studies, using both CaCl2 and HNO3 at pH4 and 6, showed that the Zn in both remediatedand nonremediated soils was quite stable,reflecting again the role that metal surfaceprecipitates, that is, Zn-LDH phases, play insequestering metals such that mobility andbioavailability are diminished.

Grafe et al. (2008) used m-XRF and m-XAS tostudy the speciation of As in a chromated-copper-arsenate (CCA) contaminated soil todetermine the effects of co-contaminating metalcations (Cu, Zn, Cr) on As speciation. Data ana-lyses revealed that As occurred as a continuumof fully and poorly ordered copper-arsenateprecipitates (63-75% of As species) and surfaceadsorption complexes on goethite and gibbsitein the presence and absence of structural Zn.Other non-Cu-based precipitates such as scoro-dite, adamite, and ojuelaite were also found.These data point out that association of As withCu suggest that the speciation of As in acontaminated soil is not entirely controlled bysurface adsorption reactions, but influencedsignificantly by the co-contaminating metalcation fraction.

4.4.2. Speciation of Nutrients in Soils andBiosolids

A number of studies have used particularlyXANES to speciate nutrients such as P in soilsand biosolids (see references cited in Chapter11). Hesterberg et al. (1999) used XANES toinvestigate adsorbed and mineral phases ofphosphate on Al and Fe oxides. They alsoshowed that the predominant phase of phos-phate in a North Carolina soil sample wasdicalcium phosphate. Peak et al. (2002) investi-gated the speciation of P in alum-amended andnonamended poultry litter samples usingXANES. Alum is used in the poultry industry

Cu

Zn Fe

Region 3

Region 1 Region 2

Region 4

1

3

2

2

20 μmA 80 μm

20 μm 80 μm

1

2

FIGURE 1.7 (A) m-SXRF tricolor maps for the treated soil samples. The numbers indicate the spots where m-EXAFS spec-tra were collected.

Continued


to reduce the level of water soluble P in litterbefore it is land applied. In the unamended lit-ter, the predominant P species were a mixtureof aqueous phosphate, organic phosphate, anddicalcium phosphate. In the alum-amended lit-ter samples, the major species were P adsorbedon Al-oxides, organic phosphate, and aqueousphosphate. These results indicate that the addi-tion of alum reduces the mobility of P in the lit-ter material. There was no evidence of Alphosphate precipitates being present in thealum-amended samples. However, such phasescould have occurred immediately after alum

was added when the pH is low, ca. 5, butincreases with time to ca. 7. With the rise inpH, any initial Al phosphate precipitateswould undergo dissolution. XANES, alongwith more detailed speciation analyses, includ-ing linear combination fitting, was used in sev-eral studies (Beauchemin et al., 2003; Seiteret al., 2008; Shober et al., 2006; Toor et al.,2005) to more quantitatively speciate P in soils,turkey manures, and poultry litter. In order toget a better assessment of organic P species,which are known to be significant in poultrylitter, Seiter et al. (2008) used principal

Region 4 spot 2–90% Sphalerite

Region 4 spot 1–36% Willemite–59% Kerolite

Region 3 spot 1–35% Willemite–57% Zn-kaolinite-pH7

Region 2 spot 2–60% Hemimorphite–45% Kerolite

Region 2 spot 1–55% Zn-kaolinite-pH7–54% Kerolite

Region 1 spot 3–86% Sphalerite

Region 1 spot 2–100% Zn-kaolinite-pH7

Region 1 spot1–71% Willemite–32% Zn-kaolinite-pH7

3

c(k)

×k

3

4 5 6

Treated

NSS = 0.099

NSS = 0.214

NSS = 0.194

NSS = 0.221

NSS = 0.058

NSS = 0.073

NSS = 0.095

NSS = 0.159

k(Å–1)B7 8 9 10

FIGURE 1.7—Cont’d (B) m-EXAFS spectra from selected spots on thin sections from the treated soil.Continued


component analysis, target transformation, andlinear least squares fitting to assess P speciationin alum-amended and nonamended poultry lit-ter samples. They found that alum-amendedpoultry litter contained higher amounts of Al-bound P and phytic acid, whereas nona-mended samples contained Ca-P minerals andorganic P compounds.

4.4.3. Coupling Kinetic Measurements withMolecular Scale Techniques in Real Time

In soil environments, chemical reactions atthe mineral/water interface occur over a rangeof temporal scales, ranging from microseconds

to years. Many important processes (e.g.,adsorption, oxidation-reduction, precipitation)occurring at mineral surfaces are characterizedby a rapid initial reaction on time scales ofmilliseconds to minutes (Scheidegger andSparks, 1996). Knowledge of these initial reac-tion rates is critical to determining chemicalkinetic rate constants and reaction mechanisms,both of which are necessary to understandenvironmental chemical processes. Kineticmeasurements, using traditional techniques,such as batch or stirred-flow techniques, typi-cally yield only a few data points during theinitial phases of the reaction and cannot

Cu

Zn Fe

Region 1 Region 2

Region 3 Region 4

1

1

1

1

2

3

80 μm 20 μm

80 μm 80 μmC

FIGURE 1.7—Cont’d (C) m-SXRF tricolor maps for the nontreated soil samples.Continued


capture important reaction rates on second andslower time scales. Chemical relaxation techni-ques such as pressure jump (p-jump) and con-centration jump (c-jump such as stopped-flow)allow rapid data collection on time scales ofmilliseconds. However, rate “constants” arecalculated from linearized rate equations thatinclude parameters that were determined fromequilibrium and modeling studies. Conse-quently, the rate “constants” are not directlydetermined.

Direct, in situ, molecular scale measurementof rapid reactions has, until recently, been quite

limited. Fendorf et al. (1993) used SF-EPR spec-troscopy to measure Mn(II) sorption to birnes-site (d-MnO2) on a time scale of milliseconds.More recently, Parikh et al. (2008) used in situ,ATR-FTIR spectroscopy to measure As(III) oxi-dation rates by hydrous manganese(IV) oxide(HMO) at a time scale of �2.5 s. However, bothof these techniques suffer from significant lim-itations. EPR can only be used to measureEPR-active nuclei, while FTIR requires bothIR-active functional groups and relatively highconcentrations of the reactants being examined(Parikh et al., 2008).

Region 4 spot 1− 95 % Zn-Al LDH

Region 3 spot 3− 99 % Hemimorphite

Region 3 spot 2− 64 % Sphalerite− 29 % Hemimorphite

Region 3 spot 1− 106 % Hemimorphite

Region 2 spot 1− 93 % Zn-Al LDH

Region 1 spot 1− 72 % Sphalerite− 24 % Zn-Al LDH

Non-treated

c(k

) ×

k3

3 4 5 6 7 8 9 10k(Å–1)D

NSS = 0.134

NSS = 0.089

NSS = 0.070

NSS = 0.079

NSS = 0.093

NSS = 0.067

FIGURE 1.7—Cont’d (D) m-EXAFS spectra from selected spots on thin sections from the nontreated soil. The solid lineindicates the raw k3 � chi(k) data and the dotted line indicates the best fits obtained with linear combination fitting(Nachtegaal et al., 2005).


Quick-scanning X-ray absorption spectros-copy (QXAS) overcomes both of these limita-tions. Depending on beamline instrumentationand flux, QXAS can be used to probe most ofthe atoms on the periodic table and to relativelylow concentrations (Khalid et al., 2010). Themajority of the quick-scanning beamlines inthe world collect a complete EXAFS scan in�1 min, by slewing the monochromator fromlow energy to high energy and repeating theprocess (Mitsunobu et al., 2006). An alternativemethod for rapidly collecting EXAFS datais to perform energy-dispersive measurements;however, this technique generally suffers frompoor sensitivity (Dent, 2002). However, usinga unique, cam-operated, continuous-scanningsetup at beamline X18B at the NSLS, it is

possible to collect XANES and EXAFS spectraas the monochromator travels both up anddown in energy (Khalid et al., 2010). Combinedwith electronics that collect 2000 data pointsper second, this unique setup allows the collec-tion of a single EXAFS scan in as little as 100ms. Ginder-Vogel et al. (2009) and Landrotet al. (2009) used this beamline to examine thekinetics of Cr(III) and As(III) oxidation byHMO using XANES spectroscopy at sub-second time scales. For examples, using Q-XAS,As(III) and As(V) concentrations were deter-mined every 0.98 s in batch reactions. Theinitial apparent As(III) depletion rate constants(t<30 s) measured with Q-XAS are nearlytwice as large as rate constants measured withtraditional analytical techniques (Ginder-Vogel

21REFERENCES

et al., 2009). Landrot et al. (2010) determined thechemical (i.e., independent of physical phenom-ena) kinetics of Cr(III) oxidation. These resultsdemonstrate the importance of developinganalytical techniques capable of analyzingenvironmental reactions on the same time scaleas they occur.

5. FUTURE NEEDS

As our knowledge of environmental systemscontinues to advance, it is critical that X-rayspectroscopic techniques continue advancingwith the field. Currently, a primary limitationof the application of X-ray techniques to envi-ronmental systems is the availability of ade-quate amounts of beamtime. This problem canbe resolved by advances in both user supportand instrumentation. Several improvementsin user administration and services, includingstandardized data collection software, state-of-the-art data analysis software, increasedbeamline support staff, and improvements inlaboratory support facilities, would allow forthe more efficient use of limited beamtime.The heterogeneity of environmental samplesrequires both the application of a wide arrayof techniques, including X-ray spectroscopicand traditional ones, and more time using eachindividual technique. The enhanced intensityand flux of third generation synchrotron lightsources allow for the analysis of both lowerconcentration elements and smaller samples;however, many environmental samples alsorequire a broad energy range (5-50 keV), high-energy resolution (0.1 eV), and a range ofspatial resolutions (bulk, micro, and nano).Additionally, unique end-station capabilities,including flow-through reaction cells andanaerobic gloveboxes, coupled with fast-scanning, high-energy resolution fluorescentdetection and simultaneous collection of XRDand XAS data, will take advantage of the

unique ability of X-ray absorption to analyzesamples in situ, allowing a new generation ofcomplex environmental problems to be solved.

ACKNOWLEDGMENTS

The authors would like to acknowledge fund-ing from the NSF EPSCoR Program, grant #EPS-0447610. A portion of this work was conductedat Stanford Synchrotron Radiation Laboratory, anational user facility operated by Stanford Uni-versity on behalf of theUSDepartment of Energy,Office of Basic Energy Sciences. The SSRL Struc-tural Molecular Biology Program is supportedby the Department of Energy, Office of Biologicaland Environmental Research, and by theNational Institutes of Health, National Centerfor Research Resources, Biomedical TechnologyProgram. A portion of this work was performedat GeoSoilEnviroCARS (Sector 13), AdvancedPhoton Source (APS), andArgonneNational Lab-oratory. GeoSoilEnviroCARS is supported by theNational Science Foundation—Earth Sciences(EAR-0622171) and Department of Energy—Geosciences (DE-FG02-94ER14466). Use of theAdvanced Photon Source was supported by theUS Department of Energy, Office of Science,Office of Basic Energy Sciences, under ContractNo. DE-AC02-06CH11357. Use of the NationalSynchrotron Light Source, Brookhaven NationalLaboratory, was supported by the US Depart-ment of Energy, Office of Science, Office of BasicEnergy Sciences, under Contract No. DE-AC02-98CH10886. The Advanced Light Source is sup-ported by the Director, Office of Science, Officeof Basic Energy Sciences, of the U.S. Departmentof Energy under Contract No. DE-AC02-05CH11231.

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Arai, Y., Lanzirotti, A., Sutton, S.R., Newville, M., Dyer, J.,Sparks, D.L., 2006. Spatial and temporal variability ofarsenic solid-state speciation in historically lead arsenatecontaminated soils. Environ. Sci. Technol. 40, 673–679.

Bank, T.L., Vishnivetskaya, T.A., Jardine, P.M., Ginder-Vogel, M.A., Fendorf, S., Baldwin, M.E., 2007. Elucidat-ing biogeochemical reduction of chromate via carbonamendments and soil sterilization. Geomicrobiol. J. 24,125–132.

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Chapter 1 - The Impacts of X-Ray Absorption Spectroscopy on ... · During the last two decades, XAS has devel-oped into a mature technique for obtaining the speciation (e.g., oxidation