Colloid Deposition and Release in Soils and Their...

Critical Reviews in Environmental Science and Technology, 41:336–372, 2011Copyright © Taylor & Francis Group, LLCISSN: 1064-3389 print / 1547-6537 onlineDOI: 10.1080/10643380902871464

Colloid Deposition and Release in Soilsand Their Association With Heavy Metals

GAO BIN,1 XINDE CAO,2,3 YAN DONG,3 YONGMING LUO,4

and LENA Q. MA3

1Agricultural and Biological Engineering, University of Florida, Gainesville, FL, USA2School of Environmental and Engineering, Shanghai Jiao Tong University, Shanghai, China

3Soil and Water Science Department, University of Florida, Gainesville, FL, USA4Institute of Soil Science, Chinese Academy of Sciences, Nanjing, P.R. China

Colloid transport in subsurface has received considerable atten-tion recently because mobile colloids can facilitate the transport ofheavy metals in soils to contaminate groundwater. Many studieson colloid mobility in the subsurface consider soils as well-definedporous media. Though similar in many aspects, soils are differ-ent from well-defined porous media. The authors emphasize theimpacts of soil properties on soil-colloid deposition, release, and as-sociation with heavy metals to provide an overview of colloidaldynamics in natural soils. The electrical double layer andDerjaguin-Landau-Verwey-Overbeek (DVLO) theories are summa-rized in Section II as theoretical bases for further discussions ofcolloid dynamics in soils, and their interactions with heavy metals.After discussions of theory developments and experimental resultson the characteristics of soil colloids in Section III and soil porousmedia in Section IV, the authors compares the deposition of col-loidal particles in well-defined porous media with that in naturalsoils in Section V. In Section VI, processes that affect colloid releasein soils are summarized and theories of ion transfer processes insoils during colloid release are reviewed and discussed. Finally,the authors give a brief overview of the adsorption and precipita-tion of heavy metals to soil colloidal particles and their influenceson colloid surface charge development in Section VII. The authorsconclude with remarks on the importance of colloid deposition andrelease in soils and their association with heavy metals.

Address correspondence to Lena Q. Ma, Soil and Water Science Department, Universityof Florida, Gainesville, FL, USA 32611-0290. E-mail: [email protected]

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Colloid Deposition and Release in Soils 337

KEYWORDS: colloid, contamination, DLVO theory, groundwater,heavy metal, soil

I. INTRODUCTION

It is well known that colloids, which are ubiquitous in soils, can be mobilizedand transported by water flow to a distance in soil environments (Ryan andElimelech, 1996; Sen and Khilar, 2006). The movement of colloidal particlesin soil pores can be faster than that of water flows due to the size exclusioneffect (Bradford et al., 2005; Kersting et al., 2004). As a result, the mobilityof sparingly soluble heavy metals associated with colloids in soils can beenhanced (DiCarlo et al., 2006; Ma et al., 2005; Metreveli et al., 2005; Zhangand Selim, 2007). Colloid-facilitated transport has been considered one ofthe most important mechanisms making reactive heavy metals mobile in soils(Barton and Karathanasis, 2003; Grolimund and Borkovec, 2005; Kretzschmarand Schafer, 2005).

The term colloid generally applies to suspended particles between 1nm and 10 µm (Stumm, 1977), which include abiotic colloids (e.g., clay,metal oxides, and humic substances) and biocolloids (i.e., virus, bacteria,and protozoa). Interactions among colloidal particles bigger than 1 µm (non-Brownian) are mainly controlled by physical forces (e.g., gravity and fluiddrag). The interactions among submicron particles (1 nm–1 µm, Brownian),however, are mainly controlled by the interfacial characteristics of particle-solution (Stumm, 1977). Submicron colloids are of particular interest to usbecause of their large surface area to mass ratios and their high mobility insoils (Ryan and Elimelech, 1996). Although biocolloids are also very commonin soils, in the present paper we focus only on abiotic soil colloids excludingcolloidal-sized microbes.

Several processes responsible for colloid deposition and release in well-defined porous media have been established (Bradford et al., 2006; Johnsonet al., 2007b). In principle, the Derjaguin-Landau-Verwey-Overbeek (DVLO)theory is widely accepted in describing the interactions between colloids andsoil grains (Muller, 1994). The theory states that the net interaction energybetween two particles is the sum of the interactions of electrical double layers(EDLs) and van der Waals-London (WL) forces, which vary with separationdistance between particles. As particles approach each other, the net interac-tion energy experiences a secondary minimum first, then a primary minimum(φmin) after a maximum energy barrier (φmax; Figure 1). Particles may be at-tached or deposited when the net attractive forces are close to either theprimary or secondary energy minimum (Figure 1). Both the magnitude ofthe energy barrier and the depth of the primary and secondary minimum,however, are affected by solution chemistry. In order for a deposited particleto be released, repulsive forces have to be generated between the surfaces


338 G. Bin et al.

FIGURE 1. General potential energy profile of the interaction of surfaces with inclusion ofvan der Walls attraction, electrical repulsion, and Born repulsion, which shows both primary(φmin) and secondary minimums and an energy barrier (φmax) as well as the zones in whichrelease and deposition take place (qualitative schematic).

of particles and stationary grains, perhaps as a result of changes in solutionchemistry or flow velocity.

Understanding of colloid release and deposition is still far from com-plete. Colloid deposition under favorable conditions, in the presence ofattractive interactions, can be predicted reasonably well using the DLVOtheory. But colloid deposition under unfavorable conditions, in the pres-ence of repulsive interactions, cannot be predicted by the theory (Grassoet al., 2002; Tufenkji and Elimelech, 2004b). Experimentally observed col-loid deposition rates are many orders of magnitude greater than thosepredicted and are independent of the size of colloid particles. Elim-elech (1994) presented an extensive discussion on possible explanationfor these discrepancies. Although most of the studies on colloid deposi-tion are developed for well-defined porous media, they are applicable tosoils.

In this review we discuss only two characteristics of colloid deposition:transient phenomenon and straining. Though they are important to colloiddeposition in natural soils, they received relatively less attention in the liter-ature. Compared to colloid deposition, colloid release is even more poorlyunderstood, which we emphasize in this review.

Investigations of colloid transport in the soil vadose zone have demon-strated that colloid particles tend to adsorb onto the water-air interface andthe sorption process is almost irreversible (Sirivithayapakorn and Keller,2003). These findings can trace their theoretical and experimental basisto the flotation discipline (Williams and Berg, 1992). Recent studies have



also revealed other governing mechanisms of colloid retention and releasewithin soil pores (i.e., pore straining; Bradford et al., 2006), thin-water films(i.e., film straining; Wan and Tokunaga, 1997), air-water-solid interfaces (i.e.,air-water interface capture; Gao et al., 2008), and immobile waters (i.e.,immobile-water trapping; Gao et al., 2006). To maintain the simplicity, wefocuses only on colloid dynamics in saturated soils excluding the vadosezone mechanisms.

Association of soil colloids with heavy metals has been an importantissue in recent years (Hu et al., 2008; Zhang and Selim, 2007), but relativelyless attention has been paid to the influences of colloid-metal associations oncolloid mobility (Kretzschmar and Sticher, 1997). In the flotation processeswidely used in mining industry, it is well documented that aqueous heavymetals can be potential determining ions (i.e., ions have the ability to prefer-entially dissolve over their counterions) or specific adsorbing ions (i.e., ionshave the ability to be adsorbed to charged surfaces) to charged surfaces ofminerals (Fuerstanau and Palmer, 1976), therefore the association of heavymetal with colloid surfaces may alter the surface charge significantly. Thepotential impact of the association on colloid mobility is emphasized in thisreview.

Most of the existing reviews on colloid transport in soils have focusedon well-defined porous system (Bradford et al., 2006; Keller and Auset, 2007;Tufenkji et al., 2006). There are apparently differences between well-definedporous media and natural soils. In soils, ionic strength and pH of soil solutionand redox potential of grain surface are influenced by rainfall, bioactivity,and changes in land use. The physiochemical heterogeneities in soils are notnegligible, and various interfaces among grain, water, and air need to betaken into account to understand colloid mobility in soils. Colloid depositionand release in natural soils and their association with heavy metals are moredynamic and complicated than those in well-defined porous media. Thoughwe do not cover all the differences between well-defined porous media andsoils, we emphasize the unique electrochemical features of soils affectingcolloid dynamics.

II. DLVO THEORY

The classic DLVO theory has been extensively used to describe colloid inter-actions and stability (Verwey and Overbeek, 1948). According to the DLVOtheory, the total interaction energy (φTotal) between colloidal particles (ora particle and a stationary grain) is the sum of electrostatic repulsion en-ergy (φEL, arising from the overlap of electrical double layers) and attractiveenergy (φWL, due to WL force), which is a function of particle separation dis-tance. The generalized curve of interaction energy versus particle separation


340 G. Bin et al.

distance shows one energy maximum (φmax) and two energy minimums (pri-mary and secondary minimums; Figure 1). Aggregation or coagulation occurswhen particles collide with sufficient kinetic energy to overcome the energybarrier (φmax) so they reach a distance within the φmin. On the other hand,if coagulated particles gain enough energy under perturbation to overcomethe energy well (φmax − φmin) with increasing separation distance, they arereleased. The shape, intensities, and positions of the maximum and mini-mums are determined by the interactions between particle surfaces, whichchange with surface properties of particles and solution, and interfacial char-acteristics.

Significant discrepancies have been found between experimental obser-vations and the prediction based on the DLVO theory (Elimelech and Omelia,1990; Tufenkji and Elimelech, 2004b). Various updated theories of colloidstability have been proposed in recent years (Grasso et al., 2002; Tufenkjiand Elimelech, 2005). The accuracy of prediction is somewhat increasedusing more sophisticated models in some systems at the expense of sim-plicity. Nevertheless, stability of clay colloids in solution can be qualitativelydescribed by the DLVO theory in some systems (Verwey and Overbeek,1948). Although measured collision efficiency between colloids and porousmedia surfaces is much greater than the theoretical prediction (Elimelechand Omelia, 1990), colloid deposition and release can still be qualitativelydescribed by the theory (Roy and Dzombak, 1996). It has been generallyagreed that the concept of DLVO theory is correct even though it is incom-plete (Swanton, 1995).

To reasonably predict colloid interactions in a natural system, the clas-sic DLVO theory needs some modification. Direct measurements of colloidinteractions indicate that additional forces exist between particles (Cail andHochella, 2005). Interactions between polar electron acceptors (Lewis acids)in solution and polar electron donors (Lewis bases) from colloid particles aretermed as AB force and are usually dominant in aqueous systems in additionto the electrostatic and WL forces (Grasso et al., 2002). It can account forup to 85% of interactions between soil mineral particles in aqueous solution;therefore, the DLVO forces combined with the AB force adequately describessome anomalous stability of mineral colloids (Wu, 2001). The approach tocombine classic DLVO forces and non-DLVO forces to calculate total col-loid interaction energy is termed as the extended DLVO theory. Besides theAB force, other non-DLVO forces between particles (e.g., osmotic and stericinteractions) may also be important to colloid interactions in soils (Grassoet al., 2002). The extended DLVO theory should also be corrected for colloidsurface roughness and heterogeneity of colloid compositions, charges andsizes (Johnson et al., 2007a; Tufenkji and Elimelech, 2005). Unfortunately,none of the theories or models has successfully taken all these factors intoconsideration. Nevertheless, the DLVO theory has provided a guideline to at



least qualitatively describe colloid interactions in soils (Elimelech and Omelia,1990). A simple approach (Vanoss et al., 1988, 1990) of combining the ABforce into the extended DLVO theory can present much better estimationsof colloid interaction energies in soils:

φTotal = φE L + φWL + φAB (1)

The extended DLVO theory is more successful than the classic one in de-scribing colloid behaviors (Vanoss et al., 1990). Detailed description of Eq.(1) can be found in Vanoss et al (1990).

As shown in Figure 1, in the presence of an energy barrier (φmax) thedeposition of a particle from solution onto a surface is mainly affected by themagnitude and shape of this barrier extending from its maximum into thesolution phase. These are primarily determined by long-range forces suchas electrostatic, WL, and AB forces. In contrast, particle release depends oninteraction at separations between the energy minimum and the barrier max-imum, which is greatly influenced by additional repulsion at short separation(Born repulsion; Ruckenstein and Prieve, 1976). Over this region the energyprofile is very sensitive to the specific characteristics of the interacting sur-faces and intervening liquid layer. This may explain why colloid release ispoorly understood at the present time.

III. CHARACTERISTICS OF SOIL COLLOIDS

A. Charge DevelopmentA1. INTRODUCTION

Solid particles present in water are often charged. The mechanisms of chargedevelopment have been extensively investigated (Lopez-Garcia et al., 2007;Ottewill, 1994). Based on the nature of solid particles, charge developmentcan be divided into three categories: (a) lattice substitution is the most com-mon in clay particles and referred to as permanent charge; (b) specificchemical interactions between surfaces and solution, including hydrolysisof surface functional group (e.g., hydroxyl and carbonyl) and chemical ad-sorption; and (c) preferential dissolution resulting from preferential hydrationof surface atoms.

In addition, based on the contribution from electrolytes in solution tothe charges of particles, ions can be classified into three categories: poten-tial determining ions, specific adsorbed ions, and indifferent ions. Potentialdetermining ions are the constituent ions of solid particles (e.g., H+ andOH− for Fe oxides, and Ca2+ and CO2−

3 for CaCO3) and their concentrationsprimarily determine the surface potential of particles (Parks, 1967). Specificadsorbed ions (SAIs) can change the magnitude and sign of the surface


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charge (e.g., Ca2+ and Pb2+) by adsorbing onto the surfaces and forminga stern plane in the EDL (Parks, 1975). Indifferent ions adsorb physicallyto surfaces and change the magnitude of the surface charge (e.g., Na+ andCl−), and form the diffusive layer in EDL resulting from a balance of theirelectrostatic interaction with the surface and osmotic interaction with bulksolution (Parks, 1975).

Soil colloids are heterogeneous in composition and consist of inorganicand organic constituents, or a mixture of the two. They are combinationsof various inorganic minerals and organic substances. Heterogeneity in thecomposition and structures of colloidal particles makes their charge develop-ment a complex process. This may be further complicated by the dynamicnature of soil solution chemistry, which may cause the interfacial compo-sition and structure of colloidal particles to vary temporally. Nevertheless,numerous studies have shown that their charge developing process canstill be described using the EDL theory (Grasso et al., 2002; Vanoss et al.,1990).

A2. ELECTRICAL DOUBLE LAYER BETWEEN METAL OXIDE-WATER INTERFACES

According to the EDL theory, a charged particle in solution is surrounded bytwo thin liquid shells: an inner region (Stern layer) where the counterions arestrongly bound and an outer region (diffuse layer) where they are less firmlyassociated (Haydon and Ottewill, 1965). In this review we mainly discussthe EDL of metal oxides in water because metal oxides are important naturalcolloids and have been relatively well understood among various surfacesof colloid particles (Dzombak and Morel, 1990; Hofmann and Liang, 2007).When specific absorbing ions other than H+ and OH− are absent in a system,the charge developing process on the surfaces can be described as follows(Schindler and Stumm, 1987):

SOH+2 = SOH + H+ Kint

a1 (2)

SOH = SO− + H+ Kinta2 (3)

Where SOH denotes a surface site and K inta1 and K int

a2 are the intrinsic sur-face acidity constants. Schindler and Stumm (1987) summarized the intrin-sic acidity constants (K int

a1 and K inta2 ) of various metal oxides and found

that the constants are generally correlated to those in solution. The affinityof protons to the surfaces of metal oxides, however, is much more com-plicated and influenced by changes in surface composition and structure(Koopal, 1996a; Preocanin and Kallay, 2006). The values of these constantsare difficult to determine and different electrostatic models may give con-tradictive predictions. A simplified approach is often used by combining



Eqs. (2) and (3):

SOH+2 = SO− + 2H+ Kint

a (4)

Kinta is simply related to the point of zero charge (PZC or pH0), which is

experimentally measurable and independent of electrostatic models for thesolid-water interface (Bourikas et al., 2005). When the concentration of pos-itively charged surface species is equal to that of the negatively chargedspecies (i.e., a zero-charge surface), the pH0 is related to the intrinsic equi-librium constant:

pH0 = 0.5 log Kinta = 0.5 log Ka (5)

This means that intrinsic acidity constant (K inta ) is equal to the apparent

acidity constant (Ka) at a surface potential of zero. Eq. (5) not only in-dicates that pH0 is a measure of K int

a , it also shows that the bonding ofprotons at metal oxide surfaces is more analogous to the bonding of similarcomplexes in a aqueous phase (Schindler and Stumm, 1987) or to that inbulk crystal structure (Sverjensky, 1993). The correlation between the sur-face acidity constants of metal oxides and those in solution supports the firstscenario.

Blesa et al. (1990) pointed out that there is a difference in Gibbs freeenergy between forming aqueous and surface ion complexes, and inferredthat the difference may result from the dehydration of the adsorbing ions. Tofavor adsorption, adsorbed species should occupy kinks, edges, or adatompositions where water molecules from the first coordination sphere must beremoved. For observed difference between the hydrolysis in solution and theformation of the surface complex, the easy dehydration of anionic speciesshould account, at least partially. Blesa et al. (1990) suggested that the sol-vation of metal ions must be considered. Following this point, combiningcrystal chemistry (Parks, 1965) and solvation theory (Sverjensky, 1993), Sver-jensky (1994) showed that surface protonation on a wide variety of mineralscan be accurately calculated from the dielectric constant of a mineral andthe ratio of the Pauling bond strength to the cation-hydroxyl bond length ofthe solid particles. This suggests that much more emphasis should be placedon the analogy between the bonding of the surface protonated species andthe bonding in the underlying crystal structure. In the approach of Sverjen-sky (1994), the property of crystals has been numeralized with its electricconstant instead of qualitative terms such as kinks, edge position, and atompositions. The details on modeling or calculating pH0 can be found in Kallayet al. (2007) and Preocanin and Kallay (2006).

Fokkink et al. (1989) demonstrated that pH and temperature congru-encies exist in the surface charge development on various metal oxides.They measured the surface charge σ 0 in a number of oxides (TiO2-rutile,RuO2, and α-Fe2O3-haematite) in KNO3 solution as a function of pH, ionic


344 G. Bin et al.

strength, and some other variables. When σ 0 is plotted as a function of(pH-pH0), where pH0 is the PZC, the curves coincide for the three oxideswithin the experimental error at three levels of ionic strength even thoughthey are different in pH0. In other words, the individual identity of metaloxides makes no difference to the S-shaped plot of σ 0 versus pH-pH0. In themeasurements of temperature dependence of surface charge development,changes in temperature only affect the positions of PZC, not the trend of thesurface charge development. They concluded that the EDLs in metal oxidescan be functionally divided to a specific and a generic part, with respect tothe natures of oxide and electrolyte. The specific part, which is determinedby the specific interactions of a surface with protons on the surface, deter-mines the PZC. The general part, which is controlled by the solution sideof the double layer, determines the surface charge development once thesurface charge deviates from pH0. It can be well described by the Nernstequation (ψ0 = 2.303 RT/F (pH-pH0)).

In the presence of specific adsorbing ions (SAIs) such as Ca2+ and Mg2+

in solution, the charge developing process of a solid can be more compli-cated because its PZC will be shifted (Duval et al., 2002; Koopal, 1996b).In the presence of SAIs, observations have also shown the shifting of com-mon intersection point (CIP) of titration curves under various ionic strengths.The CIP is the equal compensation point (Lyklema, 1987), at which charge-compensating cations and anions have an equal affinity to the surfaces. Fromthis point of view, they explained CIP �= PZC in the case of specific adsorp-tion as follows: at the PZC, adsorption of metal cations is favored over thatof anions if cations have a higher affinity to the surface. To reach a situationwhere the intrinsic adsorbability of cations and anions is identical (equalcompensation point), a positive charge on the surface has to be developed.This charge should be more positive with higher specific affinity of cationsto the surface. Once the equal compensation point has been reached, furtherincrease in concentrations of cations and anions do not lead to further shift(i.e., all successive curves at increasing concentrations will pass through thepoint). Similar reasoning applies when anions have a higher affinity overcations. This principle is important to understand the impact of soil solutionchemistry on soil colloid charge development. Following this line, it is easyto understand the effects of changes in species of cations and anions insolution on PZC and CIP.

The previous discussion only applies to ideal crystal surfaces, wherethe dissolution is negligible using the surface complex approach. How-ever, solubility of metal oxides be taken into account for surface chargedetermination, especially in the case of amorphous (hydr)oxides. Adsorp-tion and desorption of hydrolyzed metal complex ions then may be-come an alternative to determine surface charge other than the surfacecomplexation of H+ and OH−. Based on the minimum solubility the-ory (Bourikas et al., 2005; Reymond and Kolenda, 1999), PZC should be



identical to the isoelectric point (IEP) of an aqueous solution suspend-ing the particles. The IEP is defined as the pH when the concentrationsof positively and negatively charged complexes are equal in an aqueousphase (Parks, 1965), which is often found at the pH of minimum solubil-ity of a solid. Blesa et al. (1990) found that the PZC and IEP for metaloxides do not coincide. However, the deviation from IEP can be well ex-plained if considering the fact that the complex cations and anions requiredifferent dehydration energy when they are transferred from solution tosurfaces. They explained that monomeric cations hydrate more stronglythan anions in aqueous solution, so the shift should be small and nega-tive when the charge is determined by monomeric surface complexes andthe minimum solubility theory is generally valid. On the other hand, forpolymeric species more specific behaviors are expected. Either relativelylarge positive or negative shifts may occur if the charge surface complexesare polymeric. The underlying assumption of these approaches is that thesurface active site is more analogous to their solution species. This maybe more suitable to soil environment where various amorphous miner-als are abundant and aqueous species are better understood than surfacespecies.

A3. CHARGE DEVELOPMENT OF SOIL PARTICLES

Surface charge of soil particles can be classified into two types: permanentand pH-dependent charges. Detailed characterization of soil particle chargeshas been developed (Karak et al., 2005; Moulik et al., 2005). Charge devel-opment on soil particles is more complicated than that on the metal oxidesbecause their surface composition and structure vary greatly and potentialdetermining ions are not limited to just H+ and OH−. However, similar phe-nomena such as CIP have been observed, which is referred to as the pointof zero salt effect (PZSE). It is often found that true PZSE deviates from CIP(Chorover and Sposito, 1995).

Different from ideal crystallized metal oxides, soil is a mixture of vari-ous minerals and organic substances including significant amounts of amor-phous materials, such as amorphous Fe and Al. It is well documented thathydrolyzed species of Al and Fe on soil particles have significant impact ontheir pH-dependent charge (Itami and Fujitani, 2005; Marchi et al., 2006; Yuet al., 2005), and their complexation with organic matter further complicatesthe charge developing processes (Antelo et al., 2007). In addition, mineraldissolution may be significant for some soils and may affect the processesgreatly (Blesa et al., 1997). Combining all these factors with the dynamic na-ture of solution chemistry (Dong et al., 2000), it can be concluded that at thepresent time it is difficult to accurately predict surface charge developmentof soil particles.

Blesa et al.’s (1997) approach may be promising to provide a guidelineto estimate soil surface charge development. They emphasized the role of


346 G. Bin et al.

the partitioning of charged species between solid surfaces and solution dur-ing surface charge development, which is driven by the difference in Gibbsfree energy of a charged species between in solution and on surface. Theunderlying assumption is that the charged species in solution and on sur-faces are similar. This approach has a relatively sound theoretical basis andexperimental evidence, it is worth further examination.

A4. CHARGE DEVELOPMENT OF MOBILE COLLOIDS

Much research has been conducted to investigate charge development ofwater dispersive and mobile colloids (Itami and Fujitani, 2005; Kretzschmaret al., 1997). It is commonly observed that the stability of soil colloidal claysis many times greater than that of comparable reference clays. Clays iso-lated from surface soils were more dispersive than clays from the subsurfacehorizons of kaolinitic soils (Kretzschmar et al., 1997). Factors contributing tothe high dispersibility of soil clays include the presence of adsorbed organicmatter, minor quantity of smectite in kaolinitic clay, and larger surface rough-ness of soil clays compared with unweathered reference clays (Kretzschmaret al., 1995).

Organic matter is an important source of negative charges in soils. Well-decomposed humus may have cation exchange capacity (CEC) >300 cmolkg−1, which is considerably greater than that of clays such as kaolinite (3–15),illite (30–40) and montmorillite (80–150; McBride, 1994). It has been esti-mated that 20–70% of the CEC of many soils are attributed directly to the soilorganic matter alone (Vaughan and Ord, 1984). In soils, dissolved organiccarbon (DOC) tends to adsorb onto solid particles such as clays and metaloxides, driven by ligand exchange, multivalent ion bridging, WL force, andhydrophobic interactions between DOC and the solid minerals (Murphy andZachara, 1995). In particular, for Fe and Al oxides, when the pH of a systemis lower than their PZC, adsorption of DOC onto the minerals increases sig-nificantly because of the electrostatic attraction between the two, resultingin significant increases of negative charge on the surfaces. There are nu-merous studies looking into the effects of organic coating onto metal oxides(Kretzschmar and Sticher, 1997) and clay minerals (Breiner et al., 2006) onsurface charge. Heil and Sposito (1993a, 1993b) found flocculation of illiticsoil colloids with organic coatings increased with pH in Ca solution, whereasit decreased when the organic matter is removed by H2O2. They concludedthat competition between H+ and Ca2+ for the acidic functional groups oforganic molecules is essential to the charge development on the coated col-loids. A similar result was also obtained by Kretzschmar and Sticher (1997)in humic-coated kaolinic soil clay. On the other hand, it has been reportedthat the electrical properties of the quartz surface dispersed in river wateris determined essentially by their interactions with inorganic cations suchas Al and Ca instead of organic matter (Findlay et al., 1996). These resultsare consistent in that the surface electric properties of minerals are a result



of interactions in the interfaces of particle and solution, including particleorganic matter, organic–inorganic ions, and particle inorganic ions. Particu-larly, types of cations and their concentrations are important for the chargedevelopment on particles in addition to organic matter.

Our ability to model surface charge development on soil colloids withreasonable certainty has not been fully realized; however, it has been con-ceptualized that soil particles can be treated as assemblages of crystalline andamorphous minerals and organic matters. Sophisticated models, such as thetriple layer model, fail to describe the electrified interfaces of soil colloids be-cause particle surfaces are irregular and the Stern layer often moves inward,inside the physical boundary of the particles. As a simple approximation,it is more practical to use the diffuse double layer theory to describe theinterfaces of soil particles with a specific component (i.e., a charged surface)and a generic component (i.e., a diffuse layer). The surface charge develop-ment then can be described by the partitioning of charged species betweensolution and surfaces assuming the charged species are similar between thetwo phases.

B. Hydration

When water molecules orientate around immerged particles, they form ahydration shell (Clifford, 1975). When these particles with such a shell ap-proach each other, additional forces between the particles arise. The originof these forces is the interactions between polar electron acceptors (Lewisacids) in water and polar electron donors (Lewis bases) from colloidal parti-cles, and termed AB forces as we discussed previously (Vanoss et al., 1990).Surface electron donicity of colloid particles is crucial for water molecules tobe orientated along the surface. Strong surface electron donicity causes moreordering of water molecules along the surfaces, resulting in a repulsive forcebetween the particles, while weak surface electron donicity causes less or-dering of water molecules, resulting in an attractive force. Vanoss et al. (1990)demonstrated that between particles dispersed in water, the interaction freeenergy from AB, whether repulsive or attractive, is commonly as much as100 times greater than WL energy, and ≥10 times greater than electrostaticrepulsive energy at close range (1–5 nm). The AB energy can account for upto 85% of interactions between mineral particles in aqueous solution; there-fore, the DLVO theory combined with the AB force can adequately describesome anomalous colloid stability that cannot be explained by the originalDLVO theory (Vanoss et al., 1994a; Wu et al., 1994b). Wu et al. (1994a,1994b) experimentally evaluated the particle interactions of montmorilloniteand calcite by examining flocculation of their suspensions while increasingCa2+ concentration. They concluded that Ca2+ not only reduced particle zetapotential, but also drastically lowered the electron donicity of the polar sur-face of the particles, resulting in an attractive AB force, which is primarily


348 G. Bin et al.

responsible for the flocculation of the suspensions. It is suggested that sur-face electron donicity and ζ -potential of particles are related; therefore, SAIssuch as Ca2+ alter not only ζ potential when they adsorb onto a particle, butalso decrease the surface electron donicity (Wu et al., 1994a, 1994b).

C. Size Development of Mobile Colloids

The behaviors of colloids are size dependent. Individual colloidal particlesmay aggregate depending on the magnitude of the energy barrier of theirinteraction (Figure 1), which is similar to colloid deposition. If the energybarrier of an interaction is low, fast coagulation occurs and the colloid ag-gregation rate depends on particle diffusion rate; otherwise, the aggregationrate is controlled by the magnitude of the energy barriers. These two ag-gregation mechanisms result in two different fractal dimensions, which havebeen discussed by Riscovic et al. (1996). Direct measurements and theoreti-cal simulations of colloid size distribution in natural system have shown thatthe concentration of colloids <0.1 µm might be negligible because most ofthem coagulate instantly (Buffle and Leppard, 1995; Filella and Buffle, 1993).Kaplan et al. (1997) examined the possibility of mobile colloids’ aggregationfrom a reconstructed soil profile. They analyzed the size distribution of par-ticles in the suspensions with or without sonification (which is supposed tobreak down the aggregates of particles in the suspension). They found thatthe suspension without sonification exhibits practically the same bimodaldistribution of particle sizes as does the one with sonification. This holdsfor soil suspensions collected from two different soils: a loamy sand and asandy soil. Their results suggest that colloid aggregation in soil pore spaceis limited especially when soil colloid particles are highly charged (Kaplanet al., 1997). Similar bimodal size distribution of mobile colloids were re-ported by Ronen et al. (1992) who characterized the suspended particlescollected from groundwater in the coast plain phreatic aquifer of Israel. Inthe bimodal distribution, the colloids of smaller size (<0.4 µm) may resultfrom long-distance translocation to sampling points, while those of largersize (<1 µm) may come from local dispersion at the points.

IV. CHARACTERISTICS OF SOIL POROUS MEDIA

Interactions between mobile colloids and surfaces of soil porous mediamay be more important than those between colloids in limited pore space.Therefore, pore structures and physicochemical surface characteristics havetremendous influence on colloid mobility. Extensive studies have been donein systems consisting of clean surfaces of porous media (Auset et al., 2005;Smith et al., 2008; Tong and Johnson, 2007; Tufenkji and Elimelech, 2005).



Large discrepancies, however, have been found in colloid deposition rate un-der unfavorable conditions between theoretical prediction and experimentalobservation. This has been attributed to the hydrodynamic effect of flowand surface heterogeneity of the porous media, such as surface roughnessand local charge heterogeneity (Tufenkji and Elimelech, 2005). The latter hasbeen considered to be the most promising approach to understand the in-teraction between colloids and porous media in clean systems (Schumacheret al., 2005). Several exhaustive reviews on the interactions between colloidsand porous media are available (Elimelech, 1994; Ryan and Elimelech, 1996;Swanton, 1995).

Soil porous media are assemblages of a wide range of particles, whichvary greatly in mineralogy, surface composition, and dimensions. Unlike sys-tems with clean surfaces, where the surface properties of porous media maybe far different from those of mobile colloids, the surfaces of soil porousmedia may consist of potential mobile colloids. In sandy soil, clay-sized par-ticles of small amount may tend to pack closely along the surfaces of largesand grains driven by energy minimum. The arrangement between grainand colloidal particles has also been confirmed by fractal analysis (Bartoliet al., 1991). Based on this fact, the interactions between colloids and soilporous media can be treated, at a first approximation, as those between col-loidal particles, which have actually been conventional assumption. Swanton(1995) suggested that in an undisturbed soil profile, the active sites may becompletely occupied by mobile colloids. Ryan and Gschwend (1994a, 1994b)successfully related the observed colloid release rates from Fe oxide-coatedsand columns to energy barrier (EXP [φmax − φmin]/kBT , where kB is theBoltzmann constant and T is absolute temperature). Their estimation of theenergy barrier (φmax − φmin) is based on the interactions between colloidsrather than those between colloids and porous media surface. The under-lying assumption is that soil porous media surfaces are similar to colloidsurfaces.

V. DEPOSITION OF COLLOIDS IN SOILS

A. Theoretical Background in Ideal Porous Media

Colloid mobility in porous media is mainly controlled by the net rate of col-loid deposition and release. Deposition of colloids onto porous media canbe viewed as a two-step process: transport of colloids from bulk solutionto the proximity of the surfaces of porous media and then their attachmentto the surfaces, which depends upon the nature of particles-surfaces inter-action. If the energy barrier (φmax) is ≤0, the colloid flux approaching tothe surfaces is equal to that of the deposition and attachment efficiency ofcollectors (porous media) is 100% (Yao et al., 1971). The energy barriermust otherwise be overcome in order for a colloidal particle to attach to the


350 G. Bin et al.

surfaces. Therefore, deposition rate constant (kdep) is a function of energybarrier φmax with kdep ∝ EXP (–φmax/kBT ; Ryan and Gschwend, 1994b).

The main mechanisms of colloid contact to a collector are inertialimpaction, interception, sedimentation, electrostatic forces, Brownian dif-fusion, and straining (Tien, 1989). The overall single collector contactefficiency can be theoretically calculated based on the filtration theorythat considers the processes of interception, sedimentation, and Brown-ian diffusion (Tufenkji and Elimelech, 2004a). The single collector con-tact efficiency only represents the physical processes of colloid filtrationin porous media, which is independent of the chemistry of the system.In the classic filtration theory, a separate term, removal efficiency (α)is used to reflect the role of chemistry in colloid deposition in porousmedia.

The classic filtration theory has been widely used in predicting colloiddeposition and transport in porous media; however, a great discrepancybetween the prediction and experimental observation under unfavorabledeposition conditions has been observed. This has been attributed to thedistribution of surface and physical properties, surface charge heterogeneityof solids, surface roughness, interfacial electrodynamics, and colloid depo-sition in secondary minimum (Tong and Johnson, 2007; Tufenkji and Elim-elech, 2004a). In addition, as particles deposit on a collector, the depositionrate may change after initial stage of deposition depending on the natureof particle–particle interaction (Yao et al., 1971). If the net interaction is re-pulsive, the collector surfaces become progressively occluded as particlesaccumulate and colloid deposition rate should decline accordingly. This sur-face exclusion phenomenon is termed blocking. There are several excellentrecent reviews on this topic (Bradford and Bettahar, 2006; Ryan and Elim-elech, 1996; Swanton, 1995).

B. Colloid Deposition in SoilsB1. TRANSIENT PHENOMENON IN COLLOID DEPOSITION

Because there are no clean surfaces in soils, the results obtained in idealporous media cannot be directly extrapolated to describe colloid depositionin soils. At a given time, in most soils, the most favorable deposition sitesmay already be occupied. If there is an influx of mobile colloids in soil, thedeposition site may be limited during the early stages of exposure because allfavorable deposition sites are filled, which is different from the clean porousmedia. Over a prolonged period of exposure for a colloid flux, soil porousmedia may develop surfaces that are similar to those of mobile colloids(Swanton, 1995), which will also limit the deposition of colloids. Therefore,colloid deposition may just be a transient phenomenon in soil porous media.



Dispersion of soil clays by exhaustively shaking in water provides ameasure of the fraction of colloid that can potentially become mobilized.Miler and Baharuddin (1986) observed that the amount of colloidal particlesremaining in suspension after 36 hr of shaking was highly correlated to sur-face soil loss in southeast USA under high intensity rainfall. Kaplan et al.(1997) found that mobile colloids are similar in mineralogy to the water dis-persible clays in two soils and that they are many orders of magnitude lowerthan the amount of water dispersible clays. There is a significant amount ofpotentially mobile colloids attached to the soil porous media (de Jonge etal., 2004); however, during that period of time they functioned as a part ofstationary porous media. They may be detached from the porous media (mo-bilized) once they are exposed to a chemical or hydrological perturbation(Saiers and Lenhart, 2003; Zhuang et al., 2007). The mobilization of theseloosely attached colloids is referred as transient phenomena in colloid depo-sition, which is important for understanding the distinctive features of colloidmobility in soil porous media. Take surface soil, for example, a perturbationoccurs often from the top such as rainfall. This type of perturbation may re-sult in the release of deposited colloids (Ryan and Elimelech, 1996), and themobilized colloids move down over the surfaces of porous media consistingof the deposited colloids. Therefore, at the front of the downward-movingperturbation, there are two opposite processes operating simultaneously: thedeposition and release of soil colloids. The deposition is mainly influencedby the blocking effect, and the release is discussed in Section VI.

B2. PORE STRAINING

If the effective size of a colloid particle is larger than the pores through whichwater flows, the particle is retained in porous media, which is referred to asstraining. Straining occurs in granular bed filtration if the ratio of the sus-pended particle diameter to the grain diameter is greater than some criticalvalues (Bradford et al., 2004; Xu et al., 2006). It is reported that straining isone of the most important mechanisms of colloid deposition in well-definedporous media (Bradford and Bettahar, 2005; Bradford et al., 2006; Xu et al.,2006). Previous studies have also examined pore straining of colloids in soilporous media. Seta and Karathanasis (1997) examined the transportabilityof water dispersible clay through intact soil columns and found that porestraining is one of the most important factors in determining the concen-trations of colloids in leachates. They concluded that relatively low colloidtransportability in some soils they studied is attributed to larger sizes of soilcolloids (i.e., straining). Jacobsen et al. (1997) examined the transportabilityof illite through soil columns containing macropores and found that illiteconcentrations moving through the soil columns are related to the sizes ofactive macropores in the soil columns. They also found that there was nosignificant difference in mobility between the illite and that coated with hu-mic acid through the columns. These results suggest that size straining is a


352 G. Bin et al.

physical process that is not affected by changes in surface potentials resultingfrom the coatings of humic acid. This was further confirmed by the fact thatcolloid concentrations in leachates increased with the intensity of infiltrationbecause high intensity of infiltration depresses the effect of straining. Kret-zschmar et al. (1995), however, found that coating colloids with humic acidresults in much lower collection efficiency in columns packed with cleansaprolite particles. This contradiction demonstrates the difference of colloiddeposition between soil porous media and clean porous media.

VI. COLLOID RELEASE IN SOILS

Different from colloid deposition, which is essentially determined by theinteractions at larger separation distance, colloid release depends on inter-actions at separation between the surface and energy barrier (Figure 1) andis greatly influenced by additional repulsion at short separation. Colloid re-lease is generally expected from the repulsion between porous media andcolloids. The efficiency of colloid release depends on whether the process iscaused by diffusion alone or by an applied external force. Without externalforces the rate of particle release is a function of the diffusional escape prob-ability of the deposited colloids through the energy barrier (i.e., the energywell: φmax − φmin; Ryan and Elimelech, 1996). It can also be reflected bythe shape of interaction energy profile between the energy minimum andthe primary maximum (Figure 1): the smaller the slope (dφ/dh) of the curvethe less force it is needed for colloid release, where h is the separationdistance.

It is clear that the two opposite processes, colloid deposition and release,are controlled by different mechanisms and factors. For a given system,mobile colloid concentration is determined by the relative magnitude ofcolloid deposition and release rates (Sen and Khilar, 2006). The residencetime of colloid-carrying solution in the system also plays an important role.A longer resident time represents a condition closer to the equilibrium ofcolloid deposition and release, while a short one may favor deposition overrelease and vice versa. For example, colloid dispersion in a batch experimentis a result of balancing between colloid deposition and release because ofthe longer residence time in general, whereas colloid release from a shortcolumn may be free of redeposition effect (Smith et al., 2008). Extra cautionmust be taken when extrapolating colloid stability obtained from a batchexperiment to colloid transportability in a column experiment where kineticeffect is more pronounced.

A. Processes of Colloid Release

Similar to colloid deposition, colloid release rate is generally determined byboth colloid detachment from porous media and transport to bulk solution.



Ryan et al. (1994b) examined the release rates of colloidal hematite fromhematite-coated sand columns under different ionic strengths and flow rates,and found that colloid release rates decrease as ionic strength increases.In addition, colloid release rate also decreases with increasing flow rate,which is contrary to the expectation that greater mobilization would occur atgreater flow rate due to greater hydraulic stress on deposited colloids (Ryanand Elimelech, 1996). They suggested that the rate-limiting step in colloidrelease is the transport of detached colloids to the bulk fluid when rapid col-loid release corresponds to conditions where the energy barrier has vanishedfrom the potential energy profile. Coincidentally, a similar phenomenon wasobserved by Jacobsen et al. (1997) in intact soil-column experiments wherenatural particles are leached with tap water. Their results showed that particlemobilization is not influenced by increases in flow rate. The plot of accu-mulated amount of mobilized particles versus square root of time showed afairly linear relation, implying diffusion limited kinetics.

One necessity for diffusion kinetics being dominant is that the barriermaximum of mobile colloids vanishes, which may be very common in soilbecause of the transient phenomena we discussed previously. The lack offlow rate effect in Jacobsen et al. (1997) may be caused by the larger slopeof energy barrier for potential mobile colloids at a given condition, whichmay otherwise be overcome by increases in hydraulic stress. Subsequently,these colloids will not become mobile. This is based on the heterogeneityof the interactions between colloids (mobile and potential mobile colloids)and media in soil discussed previously. In addition to ionic strength andflow rate, other physicochemical effects (e.g., charge and steric effects andpH perturbations) also play an important role in governing colloid release insoils (Grolimund et al., 2001a, 2001b).

B. Mobile Colloids and Water-Dispersible Clay

Kaplan et al. (1997) examined the release of colloids from an Ultisol soilin the southeast United States. Mobile colloids were collected during andafter mild rain events from the reconstructed pedons. They found that themobile colloids are similar to the water-dispersible clay in mineralogy, whichis consistent with the results of Seta and Karathanasis (1997).

However, there were some differences between the two results. Ka-plan et al. (1997) found that the sizes and compositions of the mobilecolloids differed from those of the water-dispersive clays, with the formerbeing smaller and generally enriched with kaolinite, Fe oxides, gibbsite,and organic carbon. In addition, compared to the total clay fractions inthe reconstructed pedon, the mobile colloids were more dilute in quartzand hydroxy-interlayered-vermiculite. Based on the results from scanningelectron microscopy and photon-correlation spectroscopy, essentially all themobile colloids (>90%) had diameters of about 230 nm and moved through


354 G. Bin et al.

the soil as discrete, nonaggregated particles. They concluded that the par-ticles enriched in mobile colloids are not only readily dispersible but alsosmaller in size than water dispersive clay. Moreover, they proposed colloidmobilization in the pedons is a result of two consecutive processes: dis-persion of highly charged particles due to changes in soil chemistry and/orinduced water flow, which is determined by particle composition and canbe evaluated by water dispersibility of a soil, and their transport through apedon, which is size dependent.

C. Influences of Colloid Stability on Mobilization ofWater-Dispersible Soil Colloids

The dispersability of water-dispersable clay has been studied extensively(Kjaergaard et al., 2004; Koopmans et al., 2005; Shaw et al., 2003). Eventhough we do not fully understand dispersability of soil colloids at the presenttime, it is undoubtedly affected by factors such as pH, ionic strength, DOC,and exchangeable sodium percentage. In their column leaching study, Setaand Karathanasis (1997) found that colloid release in soils strongly dependson pore water pH and total exchangeable bases. The effect of pH on colloidrelease can be explained by its effect on colloid stability: the higher the pHabove the pH0, the greater the colloid stability. The strong correlation be-tween total exchangeable bases and colloid recovery is anticipated becauseof its high correlation with pH. Using similar soil leaching studies, Kaplanet al. (1997) found that colloid release is highly correlated to the exchange-able sodium percentage (ESP) of the soils. The positive correlation betweensoil colloid dispersability and ESP has also been found by other researchers(Redman et al., 1999; Suarez et al., 1999).

There are sufficient evidences to show that ESP is consistently correlatedto the dispersability of soil colloids in a soil. Rengasamy and Oades (1977),however, demonstrated that even Ca-saturated clay (i.e., ESP = 0) could bedispersed provided that a soil has been free of electrolyte by dialysis. Further,in a soil system, solution chemistry is dynamic and therefore concentrationsand types of cations in soil solution and exchangeable sites may vary greatly.Parameters such as total exchangeable bases, which take cations besides Na+

into account, may be more reasonable than ESP to describe colloid trans-portability. The effects of cation exchange reactions on colloid release havebeen demonstrated by many studies, which are discussed in the followingsection.

D. Ion Transfer Processes During Colloid ReleaseD1. ION TRANSFER DURING COLLOID RELEASE IN WELL-DEFINED SYSTEMS

Kallay and Zalac (2001) examined the effects of pH-neutral electrolytes onthe detachment of spherical colloidal particles of goethite from glass surfacesin basic media. They found that colloid release is enhanced with increasing



concentration of NaNO3, which is unexpected. This behavior corresponds toa double layer that undergoes relaxation during particle detachment. As thedistance between the particle and the surface increases, to keep the potentialconstant, the adsorption of the potential determining ion (OH−) must occurwith consequent equilibration of all ionic species in the Stern and diffuselayers. During particle detachment, ion transport from solution phase into theinterfacial layer is accelerated with increasing distance between the surfacesdue to solution influx from bulk. Kallay and Zalac (2001) summarized thatthis unusual behavior can be observed under the following conditions: col-loid redeposition is minimal (short column), particles and media have alikecharges, and surface potentials are constant during colloid detachment. Thisapparently unusual behavior can be understood by taking into considerationof the energy profile as a function of distance near the surfaces (Figure 1).The probability of colloid detachment depends on the depth of energy well(φmax − φmin), however, the probability of colloid deposition depends onthe height of the energy maximum (φmax). As the concentration of NaNO3 inthe system increases, the depth of energy well may be reduced, resulting inan increased rate of colloid detachment; at the same time the energy maxi-mum may decrease, resulting in an increased rate of deposition. The amountof colloidal particles one observed in effluents is determined by the net rateof these two opposite processes.

Kallay and Zalac’s (2001) rationale is that a change in electrical envi-ronment of the interfaces resulted from colloid detachment may induce ionflow between solution and interfaces (EDL), which may facilitate colloiddetachment. This has been known as so-called surface charge or potentialregulation, which has been discussed by Chan et al. (1975). They showedthat there may be significant ion flow between solution and interfaces dur-ing EDL interaction if the particles carry pH-dependent charge. This suggeststhat ion transfer may also be important during colloid release because pH-dependent charge is generally abundant in soil colloid particles.

D2. ION TRANSFER DURING SOIL COLLOID MOBILIZATION

Shainberg et al. (1981a, 1981b) examined the effects of electrolyte concen-trations on hydraulic conductivity of a sodic soil. They found that both claydispersibility and hydraulic conductivity (HC) of the soil were very sensitiveto the levels of exchangeable Na+ and to the salt concentrations of a per-colating solution. When salt concentration in the solution was 3.0 meq/L,clay dispersion increased and HC decreased with ESP only if ESP >12%.Conversely, when salt concentration was maintained at ∼0.5 meq/L, claydispersion increased and HC decreased with ESP even when ESP >1–2%.These results indicate that ESP itself cannot determine soil dispersibility,and salt concentrations in solution have to be taken into account. A simpleprinciple behind this is that colloids flocculate when the critical flocculationconcentration of a salt is reached. Shainberg et al. (1981a, 1981b) concluded


356 G. Bin et al.

that the response of soils to low ESP and leaching with low-electrolyte wa-ter depends on the concentration of electrolytes in the soil solution that thesolid phases of a soil maintain. Their results demonstrated that salt concentra-tion in soil solution was determined by the dissolution rate of soil minerals.Higher rate of mineral dissolution depresses the impact of exchangeable Naon soil dispersion. Based on these studies, it is recognized that the interac-tion between cations on exchangeable sites and bulk solution significantlyimpacts colloid mobilization.

There is sufficient evidence that exchangeable Na has significant in-fluence on colloid release from soil at low ionic strength. Cummins andKelley (1923) first demonstrated the hydrolysis of exchangeable Na+ in soil,or the replacement of exchangeable Na+ by H+ from the dissociation ofwater. They found that in the absence of CaCO3 and CO2, a Na-saturatedsoil leached with distilled water yields a NaOH solution. In an experimentwith Na-montmorillonite, where the hydrolysis products are removed con-tinuously, Baron and Shainberg (1970) found a yield of 0.1 M Na+ solution.Shainberg (1973) demonstrated that Na-montmorillonite releases Na+ evenwhen the reaction products are not removed and the specific conductanceof the clay suspension was proportional to the square root of time. Theseobservations are consistent with a hydrolysis mechanism consisting of twoconsecutive reactions: a rapid exchange between exchangeable Na+ and H+

in solution, which results in an acidic surface, and a slower, first-order trans-formation of H+ clay to Mg2+ or Al3+ clay, which increases the amount ofexchangeable Na+ release.

Similar phenomena have been observed in soils. Oster and Shainberg(1979) demonstrated that by washing three arid-zone soils with distilledwater the release of electrolytes can be related to the square root of timeexhibiting two linear rates. They concluded that the rate of the first, the morerapid of the two, which occurred right after the washing (<1∼2 hr), dependson exchangeable Na. Increasing ESP causes increases in the release rate inelectrolytes.

Ma and Dong (2004) examined how colloid mobility in a Pb-contaminated soil, collected from Montreal, Canada, is affected by water-flooding incubation. The soil packed in short columns was incubated ap-proximately for 3, 20, and 80 days, respectively. After the standing water onthe top of soil columns was removed, 0.01 mM CaCl2 solution was pumpedthrough the soil columns until the soil was saturated with CaCl2. This is de-signed to eliminate the artifacts resulting from column packing and to useCa as an index for ion transfer during the following colloid release. Theinfluent was then switched from CaCl2 to deionized distilled water (DDW)and effluent of ∼12 mL was continuously collected using a fraction collector.Total and dissolved metal concentrations were analyzed. The relation amongcolloidal Al, colloidal Fe, dissolved Ca, and pH with pore volumes of theeffluent is presented in Figure 2.



FIGURE 2. Relationship among colloidal Fe and Al, dissolved Ca, and pH in effluents fromPb-contaminated soil columns (collected from Montreal, Canada) after 3 days of incubation.Each point presents the mean of two replicates (unpublished data).

It is expected that, after saturating the soil columns with 0.01 mM CaCl2,Ca was the dominant cation in both bulk solution and EDL surroundingcolloidal particles or stationary solid phases in the soil. When the influentwas switched from 0.01 mM CaCl2 to DDW, the release of Ca from the soilmay be determined by various mechanisms, which came to play probably


358 G. Bin et al.

sequentially. At the beginning stage of switching influent to DDW, Ca re-lease was from the bulk solution, which was confirmed by the fact that Caconcentrations and pH in the effluent at pore volume = 1 were approxi-mately those of the influent ([Ca] = 0.01 mM, pH = 6.95; Figure 2). At thesecond state, Ca diffusion from EDL to the bulk solution driven by a con-centration gradient may occur. Once Ca in the bulk solution was depleted,Ca cations moved against electrostatic attraction away from the surfaces, ex-tending EDL surrounding colloids and solid phases to mobilize colloids. Anysignificant expending of EDL and resulted colloidal mobilization had to beassociated with significant amount of Ca released into bulk solution. Theexcess Ca2+ in the diluted bulk solution may be combined with hydroxylto release proton (Ca2+ + H2O ↔ [CaOH]+ + H+; Oster and Shainberg,1979), resulting in perturbations in pH, which is consistent with the datashown in Figure 2. Generally, Ca hydrolysis is weak in solution, it is thusnot likely for it to cause notable changes in pH; however, the heterogene-ity of the system may enhance Ca hydrolysis in bulk solution to cause theperturbations.

Although further validations are still needed, the results can be used asa mechanistic illustration for ion transfer between EDLs and bulk solutiondescribed by Eq. (5). Studies on ion transfer (Ca and OH−) during the col-loid release provides a very useful tool to evaluate colloid mobility in soilsbecause current understanding of colloid mobility in soils is still limited,especially in respect to heterogeneous systems.

VII. ASSOCIATION OF COLLOIDS WITH HEAVY METALS

There are two major issues involved here: sorption capacity of heavy metalsto soil colloidal particles, which has been well documented (Hayes and Bolt,1991; Kim and Tobiason, 2004), and influences of metal sorption on colloidmobility. The association of colloids with heavy metals can be classifiedinto adsorption, precipitation (and surface precipitation), and ion exchange.Their influences on surface charge are possibly by virtue of heavy metal ionsfunctioning as specific adsorbing, potential determining, and indifferent ions,respectively. Since most heavy metals are low in solubility in soil and theirconcentrations are much lower than those of electrolytes in soil solution, it isunlikely for them to function as indifferent ions. The same reasoning may alsoapply to the role of heavy metal ions to play as potential determining ions. Ithas been realized that adsorption is one of the most important mechanismsfor heavy metal association with colloids (Hayes and Bolt, 1991) and colloid-facilitated heavy metal transport (Grolimund and Borkovec, 2005; Sen andKhilar, 2006).

In soils, colloids are heterogeneous, including weathered minerals (e.g.,clay and metal oxides), CaCO3, silica, large organic molecules, and cellular



debris. However, it may be classified by surface functional groups: surfacehydroxyl, carbonyl, or organic complexation group. In this review, the sur-face functional group hydroxyl is used as an example to discuss the associ-ation of heavy metals with colloids.

A. Adsorption of Heavy Metals to Surface Hydroxyl Groupsof Colloids

Surface hydroxyl-bearing minerals are abundant in soils (metal oxides,quartz, and the edge of kaolinite). Similar to the expression for the deproto-nation processes as described in Section III, surface complexation reactionsof surface hydroxyl with heavy metals can be written as (Hayes and Bolt,1991):

SOH + Mz+ = SOMz−1 + H+ K app1 (6)

2SOH + Mz+ = (SO)2Mz−2 + 2H+ K app

2 (7)

Where SOH denotes a surface site and Mz+ represents a metal cation. Kapp1

and Kapp2 are the apparent surface equilibrium constants, and can be ex-

pressed as

K int1 = K app

1 exp((z − 1)F ψ/RT ) (8)

K int2 = K app

2 exp((z − 2)F ψ/RT ) (9)

where ψ is the potential difference between the binding site and bulk so-lution. The exponential term accounts for the coulombic contribution to theintrinsic equilibrium constants. Because the surface potential cannot be de-termined experimentally, it is generally formulated based on a variety ofmodels, such as constant capacitance, diffuse layer, and triple layer models.It has been found that stability constants of surface complexes correlate withthose of the hydroxyl complexes in aqueous phase (Schindler and Stumm,1987). This supports Eqs. (6–9), which assume that surface complex is anal-ogous to aqueous complex and that Gibbs free energy of sorption is the sumof intrinsic and coulombic terms. Blesa et al. (1997) pointed out the impor-tance of ion hydrolysis during adsorption, and Sverjensky (1993) divided theintrinsic term into two: a solvation contribution and a remaining term. Thusthe overall free energy of adsorption of an ion can be written as

�Gads = �Gii + �GS + �Gcoul (10)

In this equation, the coulombic term (�Gcoul.) represents its contributionto the overall free energy of adsorption owing to the interaction betweenthe ion and surface charge, the solvation term (�GS) accounts for the roleof solvation during sorption of the ion, and the remaining term, named as


360 G. Bin et al.

ion-intrinsic term (�Gii), is assumed to be a property of the ion alone. Withthis approach the experimental adsorption data for Pb2+, Cd2+, Cu2+, Mg2+,and Ca2+ can be closely reproduced (Sverjensky, 1993).

When there exist ligands other than H2O and OH− in solution (e.g.,Lewis acids or bases, which is often the case in soil), competition mayoccur among dissolved ligands, heavy metals, and surface adsorption sites.Formation of aqueous metal-DOC complexes stabilizes the metal in solutionand effectively decreases the amount of aqueous metal ions available foradsorption. Similarly, anions such as sulfate and phosphate compete withsurface sites of solids for metal cations, reducing the metal adsorption onsolids. Metal sorption may be enhanced by the formation of ternary surfacecomplexes. To our knowledge, no theory can describe these complicatedprocesses well. However, partition of any aqueous species may be describedunder the framework of Eqs. (6–10; Blesa et al., 1997; Sverjensky, 1993).

B. Effects of Heavy Metal Adsorption on Surface Chargeof Soil Colloids

Many studies indicated the importance of particle-solution interactions in de-termining the surface properties of colloids in natural waters. Electrokineticmeasurements of natural particles dispersed in seawater and natural freshwaters showed that their electrophoretic mobilities tend to fall within verylimited ranges of negative values. The apparent uniformity of the surfaceelectrical property has been attributed to the adsorption of organic materi-als, particularly humic compounds, at the particle surfaces, which has beensupported by a number of studies (Breiner et al., 2006). It is indicated thatthe effects of adsorbed organic compound on the surface properties of min-erals present in natural waters are modified to some extent by additionalion interactions (Citeau et al., 2006). The interactions of heavy metals withthe surfaces of soil colloids cannot be simply described by the interactionsbetween heavy metals and pure minerals because of the complexed surfacesof real soil colloids, which include the irregularity of colloid shapes andthree-dimensional interfaces between solution and colloids. Metal oxides areoften used as mimics of soil colloids because they possess not only similarcomposition with soil colloids but also three-dimensional interfaces (Breineret al., 2006; Day et al., 1994).

There has been fair amount of research on how ionic strength or pHaffects the mobility of colloids and colloid-metal complexes in soils (Klitzkeet al., 2008; Zachara et al., 2007; Zhang and Selim, 2007). However, little isknown about the effects of heavy metals on colloidal charge developmentand stability of the complex colloids. Kretzschmar and Sticher (1997) explic-itly took into account the effects of Pb2+ and Cu2+ adsorption onto humic-coated-oxide colloids on colloid charge development and colloid transporta-bility, and found that replacement of Ca2+ with Pb2+ resulted in a slight



decrease in electrophoretic mobility of humic-coated hematite colloids whileholding the total concentration (Ca2+ + Pb2+) constant. When Ca2+ wascompletely replaced by Pb2+, the suspensions were destabilized and aggre-gated within 20 hr. In contrast, replacement of Ca2+ by Cu2+ had little effecton electrophoretic mobility and colloidal stability. Both Pb2+ and Cu2+ areknown to bind much more strongly to humic substances and Fe oxide sur-faces than Ca2+. At the pH of the suspensions investigated (pH 5.7), Cu2+

has a higher affinity for humic substances than Pb2+, whereas Pb2+ adsorbsmore strongly to hematite than Cu2+. It is expected that the effect of thedestabilizing of suspensions is more pronounced by replacing Ca2+ withPb2+ than with Cu2+, therefore, humic-coated oxide colloids can be stableand mobile in the presence of strongly adsorbing trace metals.

C. Partitioning of Heavy Metals in Soil Colloids

The ability of colloids in facilitating heavy metal transport depends on thedistribution of heavy metal among solution, mobile colloids, and stationarygrains. Several mathematical models successfully incorporated partition co-efficients to simulate colloid-facilitated metal transport in porous media (Kimand Kim, 2007; Li et al., 2004); however, there still exits a knowledge gapin understanding metal distribution among different phases. This is mainlydue to the complexities of the dynamic feature of soil colloid mobility (e.g.,introducing heavy metal to soil alters the distribution of colloids betweenwater and solid phases).

Amrhein et al. (1993) examined the potential of colloid-facilitated trans-port of heavy metals in roadside soils receiving deicing salts (i.e., NaCl andcalcium magnesium acetate-CMA). In their experiments, a series of 60 mLsyringes were packed with 30 g of soil. The columns were leached on acenturion vacuum extractor with three consecutive 30 mL aliquots of either0.1 ml L−1 NaCl or CMA. After initial leaching with either of the salts, thecolumns were leached with three consecutive 30 mL aliquots of deionizedwater. A portion of each leachate was saved without any further filtration andthe remainder of each solution was filtered through a 0.45 µm membranefilter. Then the filtrates were immediately placed into a stirred ultrafiltrationcell with a membrane of 1000 MWCO. This procedure separates the par-ticles into three fractions: >450 nm, between 1.0 and 450 nm, and <1.0nm. Their results showed that the cumulative leached metals through theentire procedure vary with the initial salt input, and NaCl tended to mobilizemore soil colloids compared with CMA (Amrhein et al., 1993). Because ofthe possibility of artifacts resulting from column packing in their study, it ismore reasonable to examine the cumulative leached metals excluding thefirst several pore volumes.

Figure 3 is a plot of leached cumulative metals after switching the in-fluent from salt solution to deionized water, showing mobilization of soil


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FIGURE 3. Cumulative mass of metals leached from Pb-contaminated soil columns afterswitching from salt solution to deionized water. The three size fractions were separated by450 nm membrane filter and 1 nm cellulose ultrafiltration membrane (based on Amrhein et al.,1993).

colloids upon reduction in ionic strength based on their data. The concen-trations of heavy metals (Cu, Pb, Fe, Ni, and Cr) are consistently higher inthe leachate from soil columns preleached with NaCl than those with CMAin all three size fractions. Among other fractions, the concentration of eachmetal is much higher for the columns preleached with NaCl than those withCMA. This is because Na promotes colloid mobility and facilitates the re-lease of colloid-associated metals in soil. Metals in fraction 1.0–450 nm aremuch higher than those in fraction >450 nm for the columns preleached withCMA, whereas they are not significantly different for the columns preleachedwith NaCl, except Fe and Cu. This suggests that CMA is more selective inmobilizing smaller colloidal particles than NaCl. Colloidal Cu does not cor-relate well with DOC (data not shown) between the two fractions for thecolumns preleached with NaCl, nor does colloidal Fe with other colloidalmetals preleached with CMA (Figure 3). This suggests that the mobility ofheavy-metal-bearing colloids varies with both size and solution chemistry.

Ma and Dong (2004) examined the effect of water flooding incubationon colloidal Pb mobility in soil columns. Detailed description about exper-iment procedures can be found in Section VI-D 2. As shown in Figure 4,colloidal Pb concentrations in leachates varied with incubation, with longerincubation resulting in lower concentrations (Figure 4). Interestingly, therewas no significant difference in the colloidal Fe and Al in the first severalpore volumes when the columns were incubated from 3 to 20 days (data notshown), while the concentrations of colloidal Pb decreased approximately 4



FIGURE 4. Changes in colloidal Pb concentrations in effluents from Pb-contaminated soilcolumns (collected from Montreal, Canada) under various incubation times. Each pointpresents the mean of two replicates (unpublished data).

times from 3 to 20 days (Figure 4). These results indicate that water floodedincubation reduces the releases of Pb from soil by decreasing either themobility of soil colloids or the associations between Pb and mobile colloids.

VIII. CONCLUDING REMARKS

Compared with well-defined porous media, soils have distinctive featuresinfluencing colloid mobility and colloid-metal complex. Our ability to accu-rately predict colloid mobility and colloid-facilitated transport of heavy metalin soils has not been fully realized.

Soil colloids are exposed to solutions with dynamic chemistry. It is dif-ficult to predict charge development of soil colloids because of the variousinteractions between surfaces and solution. Point zero charge of mineral par-ticles suspended in a simple solution can be theoretically predicted, but littlesuccess has been achieved in soils because soil particles are assemblages ofcrystalline or amorphous minerals, and organic matters. Sophisticated mod-els such as the triple layer model failed to describe the electrified interfacesof soil particles because their surfaces are irregular and the Stern layer oftenmoves inward inside the physical boundary of the particles. It is more prac-tical to model the particle-water interfaces as a diffusive double layer. Thesurface charge development may be described by partitioning of chargedspecies between solution and surfaces assuming the species are similar inthe two phases. Similarly, association of heavy metals with colloids may have


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significant influences on colloid mobility, which is also subject to changes indynamic solution chemistry.

Different from clean porous media, the surfaces of soil stationary phasesconsist of deposited colloids, which cause blocking effects and may enhancecolloid mobility in soils. In addition, these deposited colloids may be releasedback to the solution upon physicochemical perturbations, which is referred toas transient phenomena in deposition. There are two simultaneous, oppositeprocesses controlling colloid mobility in soils—deposition and release—theirrelative effects on colloid transport in soils rely on flow residence time in thesystem.

The DLVO theory provides a theoretical framework in describing colloiddispersion in soils, and colloid deposition and release under ideal conditions.However, modern colloid theories (including the DLVO theory) always fail todescribe colloid dynamics in soils because of the heterogeneities in soil andcolloid surfaces. Phenomenological and empirical approaches are commonlyused in present research of colloid deposition and release in soils. Experi-mental studies always found that colloid deposition and release in soils arecoupled with significant ion transfer between EDLs and bulk solution. Byevaluating the possibility of the ion transfer, it is possible to estimate colloidmobility in soils, at least qualitatively.

To better understand the mobility of heavy metals in soils, it is importantto consider the role of soil colloids. The complexity in the characteristicsof colloids and soils makes it difficulty to model the behaviors of colloid-facilitated transport of heavy metals in heterogeneous media like soils, whichmerits further study.

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