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Particle—particle interactions in aquatic systems
Colloids and Surfaces, 39 (1989) 255-271 Elsevier Science Publishers B.V., Amsterdam
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Particle-Particle Interactions in Aquatic Systems CHARLES
R. O’MELIA*
Swiss Federal Institute for Water Resources and Water Pollution Control, 8600 Diibendorf (Switzerland) (Received 20 September
1988; accepted
16 January
1989)
ABSTRACT The significance of particles in the transport and fate of substances in aquatic environments is illustrated and present knowledge of the speciation and structure of interfaces in these environments is assessed. The importance of particle deposition and aggregation in ground water aquifers and in lakes is illustrated. Attention is focussed on the adsorption of natural organic macromolecules and polyelectrolytes at interfaces in natural systems and the impacts of these substances on colloidal stability and deposition in natural systems. Finally, the status of present understandings in this area is assessed and directions
for future study are proposed.
INTRODUCTION
This paper is written with three objectives. First, the significance of interfacial processes and particle-particle interactions in the transport and fate of pollutants in the environment is illustrated. Second, some approaches used in modelling interfacial processes and particle-particle interactions in natural waters are considered. Third, some modelling applications in natural systems are presented. Modelling particle-particle interactions requires knowledge or assumptions about the stoichiometry, speciation, structure, and streamlines in the interfacial regions between phases. In considering this physicochemical problem, the physical aspects of the system (e.g., the streamlines in interfacial regions) are drawn from fluid mechanics and the chemical properties of interfaces (stoichiometry, speciation and structure) are derived from chemistry. This separation neglects some aspects of interfacial processes that may derive from the coupling of these disciplines, such as electrodynamic effects on the viscous resistance when two charged particles approach each other in water [ 11. The chemicals used in modelling natural systems may be classified into three *Permanent address: Department of Geography and Environmental Hopkins University, Baltimore, MD 21218, U.S.A.
0166-6622/89/$03.50
0 1989 Elsevier Science Publishers B.V.
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types. First, there is the vast array of chemical substances that are present in aquatic systems, derived from natural or anthropogenic sources. Examples are humic substances, heterogeneous solid precipitates, weathering products, asbestos fibers and polychlorinated biphenyl compounds (PCBs ) . These are often described by collective or operational parameters such as dissolved organic carbon, filterable manganese, extractable ion, and reactive phosphorus. Actual chemical speciation in natural systems is often and perhaps normally unknown. Second, there are chemicals used in model systems, both conceptual and experimental. Examples include silver halide sols used extensively in studies of colloidal stability, metal oxide surfaces and simple metals and ligands used in adsorption studies, and the spherical latex particles used in studies of aggregation and deposition. These substances have well defined physical and chemical properties, but they are models of natural systems, not replicas of them. Third, there are mathematical or imaginary chemicals. Examples include the point charges in the Debye-Hiickel and Gouy-Chapman theories, and the linear, flexible macromolecules assumed in theories for polymer adsorption [ 2,3]. These chemicals exist only in the mind, but their use has contributed extensive insights into the colloid chemistry of natural and other systems. Effective modelling of the chemistry of particle-particle interactions in natural aquatic systems can require the use of all of these types of chemicals, with the appropriate mix depending on the problem to be addressed. PARTICLES, POLLUTANTS AND COLLOIDAL STABILITY
Solid particles are ubiquitous in natural waters, and these solid particles have a role in the transport and fate of pollutants affecting human health and aquatic ecosystems. Most pollutants in water are associated with solid particles or are particles themselves. Examples include, (i) trace metals adsorbed on clays, metal oxides, and metal carbonates, (ii) synthetic organic pollutants adsorbed on organic detritus, (iii) trace metals and organic substances associated with living biomass, (iv) pathogenic microorganisms, (v) asbestos fibers, and (vi) humic precursors of trihalomethanes. The thermodynamics and the kinetics of the transport and degradation of many pollutants are controlled by properties of the interfacial regions that surround phases (solids, liquids and gases) in aquatic, atmospheric and terrestrial environments. Particle-pollutant
reactions
Karickhoff [4] has reviewed the thermodynamics and kinetics of the sorption of organic pollutants in aquatic systems. The sorption of uncharged organic chemicals to natural particles is dominated by hydrophobic interactions and depends primarily on a chemical’s affinity for water, typically described by an octanol-water partition coefficient, and on the organic carbon content
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of the solid sorbent phase. The contribution of inorganic particles to the uptake of uncharged organic pollutants is negligible in most cases. An estimate of the sorption of an uncharged organic pollutant in natural systems can be made on the basis of the chemical’s octanol-water partition coefficient, the organic carbon content of the solid phase, and the concentration of solid sorbent in the water, the soil, or the sediment of interest. For hydrophilic organic pollutants, nonhydrophobic contributions to sorption can be important and inorganic surfaces such as clays and metal oxides can be significant adsorbents. The sorption of organic pollutants to suspended particles, soil and sediments can generally be viewed as rapid, although true equilibrium may take weeks or more to achieve and the process is often not completely reversible. Several investigators have studied the transport and fate of nonpolar organic compounds in natural waters. Representative studies are given by Schwarzenbath and co-workers [ 5,6]. Laboratory studies [ 51 using natural sorbents and nonpolar organic solutes indicated that the sorption of many organic compounds is proportional to the organic carbon content of the natural sorbent for solids with an organic carbon content greater than 0.1% and that sorption could be predicted from the octanol-water partition coefficient of the nonpolar organic solute and the organic carbon content of the natural solid sorbent. Modelling studies of Swiss lakes [6] indicate that particulate organic carbon from photosynthetic processes in the water column can be important in the transport and deposition of a nonpolar hydrophobic substance such as hexachlorobenzene in lake sediments. The adsorption of inorganic pollutants in aquatic systems has been reviewed by Dzombak and Morel [ 71. In contrast to the nonspecific adsorption observed for hydrophobic organic pollutants, the adsorption of inorganic solutes is viewed as a site-specific process in which ions bind chemically at functional groups on solid surfaces. The total energy of interaction includes an electrostatic contribution and a chemical contribution. Surface complex formation models of varying complexity are available to model pollutant adsorption. The reaction is dependent upon pollutant type and concentration, pH, ionic strength and solid or surface concentration. Although not stressed by these authors, organic surfaces and humic substances can be important in the speciation and adsorption of inorganic pollutants in natural waters. Sigg [8] has reported results from Lakes Constance and Zurich indicating that metals such as Cu, Zn, Cd, and Pb are continuously removed from lake waters by becoming attached to settling particles. The partition of these metals between the settling particles and the water determines the relative residence times of the metals in the lake and their residual concentrations in the water column. Biological surfaces are indicated as better scavengers for heavy metals than inorganic ones. A large part of the fresh sedimenting particles is of biogenie origin, and a productive lake with a high sedimentation rate represents a situation that is favorable for an efficient removal of metals. Xue and co-
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workers [ 91 have conducted laboratory experiments on the binding of metal ions to algal surfaces and indicate that the surfaces of algal cells have a high affinity for Cu (II) and Cd (II). These authors interpret the adsorption of metals on algal cells in terms of surface complex formation equilibria and extract average conditional equilibrium constants for the adsorption reactions from laboratory titration curves using voltametric methods to determine free metal concentrations. Humic substances and colloidal stability There is extensive evidence from fresh waters, estuaries, and the oceans that the surface properties and colloidal stability of particles in natural waters are affected by naturally occurring organic substances dissolved in these waters. These effects of natural organic matter in establishing colloidal stability in aquatic systems also are anticipated to occur in subsurface environments and to affect particle and pollutant passage/retention in subsurface systems. Organic substances affecting colloidal stability are expected to include not only naturally occurring materials but also related materials produced by biological processes in wastewater treatment plants and at hazardous waste sites, industrial surfactants and solvents and some other synthetic organic compounds that enter the environment. Humic substances are anionic polyelectrolytes of low to moderate molecular weight; their charge is due primarily to carboxyl and phenolic groups; they have both aromatic and aliphatic components and can be surface active; they are refractive and can persist for centuries or longer; they are produced by biological processes and are ubiquitous in natural waters; they have analogies to organic substances produced in large quantities in biological active landfills, hazardous waste sites and wastewater treatment facilities; they have been reported to stabilize particles in water. Some investigators have noted that major divalent metal ions in natural waters (in particular, Ca2+ ) can exert destabilizing effects on natural and synthetic particles at low metal ion concentrations. Some examples relating natural organic matter (NOM) and divalent metals to the surface properties and colloidal stability of particles in natural waters follow. Niehof and Loeb [lo] determined the electrophoretic mobilities of several inorganic solids after exposure to natural seawater and to seawater pretreated with ultraviolet radiation to photooxidize organic matter. All solids became moderately electronegative in natural seawater and all showed charges consistent with their chemical composition when immersed in inorganic-free seawater. The authors concluded that the surface charge exhibited by solids in seawater is due to adsorbed organic matter. Edzwald et al. [ll] studied the kinetics of particle-particle interactions in the coagulation of clays and estuarine sediments by solutions of NaCl and of synthetic (inorganic) seawater. All suspensions coagulated more rapidly in artificial seawater than in NaCl
solutions of equal ionic strength. The authors concluded that calcium and magnesium ions in seawater exert specific destabilizing effects on suspended particles. Sholkovitz and co-workers [ 12,131 studied the extent of flocculation of iron, aluminium and humic substances in estuaries. They observed that iron colloids in rivers can be stabilized by dissolved organic matter. These colloids are destabilized in estuaries by seawater cations (magnesium and especially calcium ions) through specific and electrostatic interactions with humic materials. Hunter and Liss [ 14 ] determined the electrophoretic mobilities of particles collected from four estuaries and evaluated them in terms of the salinity in the estuary at the site of collection. All particles at all locations were negatively charged. The authors concluded that dissolved organic matter formed surface coatings that gave the particles a consistent electronegative charge. They also showed that in calcareous rivers containing high calcium concentrations, the electrophoretic mobilities of the particles were smaller (less negative) than in rivers low in Ca’+. They concluded that some of the charge on the particles can be neutralized by incorporation of Ca2+ “into the fixed part of the double layer.” In studies of the kinetics of coagulation and deposition, it has become customary to introduce a sticking coefficient, stability ratio, attachment probability, or attachment factor. Two interchangeable approaches [ 15,161 are used. In considering the kinetics of slow or unfavorable coagulation, Smoluchowski [ 151 introduced a sticking coefficient (a) which can be defined as the rate at which colliding particles adhere to form aggregates divided by the rate at which they collide. For a completely unstable suspension (x = 1; for a perfectly stable suspension, a=O. Smoluchowski did not relate attachment to surface or solution chemistry and did not include some physical processes such as hydrodynamic retardation in considering interparticle contact rates. Nevertheless, his characterization of unfavorable particle-particle interactions by the stability factor c11 has been used extensively. Fuchs [ 16] first combined particle transport and chemistry. He considered Brownian diffusion in a force field and defined the stability ratio Was the ratio of the collision rate to the aggregation rate ( W= a-‘). W has been expressed in terms of the maximum or peak net repulsive interaction between two colliding charged particles and, in turn, to certain solution and surface characteristics. For a completely unstable suspension W= 1; for a perfectly stable suspension W= GO.The Fuchs’ stability factor, W, like Smoluchowski’s a!!,has been used extensively by investigators reporting experimental or theoretical studies of the kinetics of coagulation or deposition. Tipping and co-workers [ 17-201 studied the effects of humic substances and divalent metal ions on the surface properties of iron and manganese oxides. Measurements included electrophoretic mobility, adsorption of humic substances, and coagulation kinetics. Values of a were estimated to range from 0.01 to 1 ( W ranged from 100 to 1) for hematite coagulated in the presence of
Ca2+ and humic substances [ 191. Tipping and co-workers concluded that adsorbed humic substances produced negative charges on the particle studied. This charge was reduced by the presence of divalent metal ions, especially Ca2+. Humic substances slowed coagulation rates and calcium ions enhanced them, possibly by specific interactions with the humic substances. In addition to enhancing coagulation rates, calcium also substantially increased the adsorption of humic substances on hematite. Davis [42] studied the adsorption on alumina and kaolinite of natural organic matter extracted from lake sediments. Measurements included electrophoretic mobility, charge titration and adsorption of NOM. The author concluded that NOM was readily adsorbed on these surfaces, giving them a negative charge. Calcium enhanced the adsorption of organic matter on alumina at high pH and reduced its adsorption at pH less than 7. Davis proposed that adsorption of NOM occurs by complex formation between surface hydroxyl groups and acidic functional groups on the organic matter. Ali et al. [43] reported the results of laboratory measurements of a! for the coagulation of natural particles from the sediments of the Loch Raven Reservoir (Maryland) in suspensions at pH 8.1 and containing various combinations of calcium ions and aquatic fulvic acid. In the absence of Ca2+, ar decreased from 0.035 to 0.006 ( W increased from 30 to 170) as the concentration of fulvic acid added to the suspensions increased from 0 to 20 mg 1-l expressed as dissolved organic carbon (DOC ) . Most of this decrease in cv (increase in stability) occurred in the DOC range from 0 to 5 mg 1-l. Calcium acted as a destabilizing agent, a! increased from 0.035 to 0.2 ( W decreased from 30 to 5) as calcium was increased from 0 to 10e3 M. Concurrent additions of fulvic acid and Ca2+ yielded intermediate effects. The ranges of Ca2+ and DOC studied in this work are representative of most fresh water systems and indicate that wide variation in a! can be expected among natural waters, including both surface and ground waters. All measured values of cv (0.006-0.23 ), however, were greater than expected from theoretical predictions, for which a+0 and W-co Ali [ 441 reported extensive measurements of cyfor the surface waters of the Loch Raven Reservoir and limited measurements for other lakes, Statistical analysis of the data suggest that variations of cyin these natural environments were related to variations in DOC. Weilenmann [45 ] reported measurements and modelling of coagulation and sedimentation in Swiss hardwater lakes. Measured stability factors (a values) indicated substantial variation in particle stability among the Swiss lakes studied. These differences were interpreted qualitatively in terms of surface and solution chemistry. Divalent metal ions decreased particle stability, probably by specific adsorption. Humic substances stabilized particles, probably by electrostatic and possibly also by steric effects.
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Interfacial structure and particle stability The effects of humic substances and divalent metal ions on the colloidal stability of natural particles are common observations without clear origins. Early studies showed that particles carry a negative charge in the presence of natural organic matter, regardless of the composition of the solid phase. This was taken as indirect evidence that humic substances stabilize particles electrostatically. Calcium ions, in turn, might contribute electrostatically to particle destabilization. More recently, there has been speculation that nonelectrostatic (steric) effects are involved in the stabilization of particles by humic materials and that specific chemical interactions (complex formation) are involved in colloid destabilization by divalent metal ions. There are, however, no conclusive evidence and no consistent theory for these hypotheses. Some speculations follow. A consideration of the interaction between two suspended particles (e.g., in a lake) or between a suspended particle and a stationary collector (e.g., in a ground water aquifer) can begin with a determination of the structure of a single solid-solution interface. With this established, the interaction between two solid bodies (as characterized by force, free energy or disjoining pressure) can be addressed. The magnitudes of the interaction forces that determine stability in particle-particle interactions depend on the properties of each solid-solution interface. Electrostatic forces are almost universal in water since few particles are uncharged in aquatic environments, at least over a substantial pH range. Electrolytes and pH strongly affect the magnitude of these electrostatic forces. Specific chemical interactions between solutes and surfaces can be decisive in establishing surface properties and structure. Steric effects are potentially most significant in systems containing uncharged particles with adsorbed nonionic macromolecules. In this case, the amount of adsorbed polymer and solvent quality are important factors. The relative importance of electrostatic and steric effects is not clear in systems containing charged surfaces and adsorbed macromolecules, especially if the macromolecules are ionizable. Two distinct approaches to the structure of a solid/solution interface are considered here: (1) an electric double layer (termed EDL) and (2) a macromolecular adsorbed layer (denoted here as MAL) . Separately these two structures are understood fairly well, at least in simplified cases. However, less is known about the effects of charge in the EDL on the conformation of adsorbed macromolecules in the MAL and, reciprocally, about the effects of charged and uncharged macromolecules comprising the MAL on EDL properties. These two distinct approaches result in two distinct mechanisms for colloidal stability termed electrostatic and steric stabilization. Electrostatic stabilization results from energetically unfavorable overlap of the diffuse ion atmospheres surrounding all charged particles in water. The well-known Derjaguin-
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Landau-Verwey-Overbeek (DLVO) theory [21] of the electrical double layer and colloidal stability provides a model for evaluation of the combined effects of electrostatic repulsion and van der Waals attraction between surfaces in the absence of adsorbed macromolecules and when all solute chemicals can be treated as point charges. When electrostatic repulsion dominates over van der Waals attraction, the results are termed “slow” coagulation and “unfavorable” deposition and filtration. A good didactic summary has been given by Lyklema [ 221, The DLVO model has successfully described the effects of counterion charge on the concentration of simple electrolytes required to coagulate many suspensions (the Schulze-Hardy rule) but it has been unsuccessful in describing the effects of particle or solution properties on the kinetics of unfavorable coagulation or deposition processes. This severely restricts its use in natural systems where rate processes are often the principal interest. Steric stabilization can result from the adsorption of polymers at solid-water interfaces. Large polymers can form adsorbed segments on a solid surface with loops and tails extending into solution [ 2,3,22]. Steric stabilization results from energetically unfavorable intermolecular interactions and entropically unfavorable compression in overlapping macromolecular adsorbed layers. Steric effects are likely to be important when the thickness of the macromolecular adsorbed layer is about the same or larger than the diffuse layer thickness. In the absence of charge effects, the energy of interaction between the adsorbed layers on two particles can be calculated by summing the various contributions to the free energy of the system as a function of the distance separating the surfaces. Under conditions of unfavorable intermolecular interactions, a repulsive interaction energy develops when the particles are close enough so that dangling tails of adsorbed macromolecules interact. On close approach, conformational entropy losses can give rise to strong repulsion if desorption of the macromolecules does not occur. When surface coverage is incomplete so that open sites remain available for adsorption, dangling tails can adsorb to the opposite surface, leading to aggregation. In natural aquatic systems it is useful and realistic to consider an interfacial region as comprised of small ions and larger organic macromolecules, i.e., as a system with properties of.both the EDL and an MAL. A definitive theoretical description of this combination remains to be developed. Uncharged macromolecules have significant effects on the properties of the electric double layer. For example, potentiometric titration of a silver iodide-polyvinyl alcohol (PVA) system showed a decrease in the slope of the curve of surface charge versus pAg with increasing amount of PVA adsorbed, as well as a shift in the point of zero charge (p.z.c. ) to smaller pAg [ 231. The reduction in slope indicated a decrease in the differential capacitance and the shift in the p.z.c. indicated change in the surface potential. In interpreting these observations, a useful approximation is to consider train segments of the polymer to be located in the Stern layer, and loop and tail segments to be in the diffuse layer [24]. The
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dimension of a polymer segment is of the order of 0.5 nm, and train segments displace specifically adsorbing ions and water molecules, Loop and tail segments displace ions and water molecules in the diffuse layer. Possible consequences include: (1) Reduction of the differential capacitance in the Stern Layer due to displacement of specifically adsorbing ions or to a reduction in the dielectric constant. (2) Change in the surface potential due to replacement of water dipoles by nonpolar polymer segments. (3) Change in local activity coefficients. (4) Outward shift of the plane of electrokinetic shear. (5 ) Expansion of the diffuse layer. (6) Binding of water molecules in the diffuse layer by hydrophilic polymer groups, reducing the local dissolving power of water [ 251. Charged macromolecular species have even more complex effects on EDL properties than do uncharged polymers. Ionized macromolecular segments contribute fixed charge in the diffuse layer as well as in the Stern layer, so that the local distribution of mobile indifferent electrolytes is affected. The conformation of the adsorbed polyelectrolytes (the distribution of charged segments in the EDL) is also dependent on the local electrolyte concentration. Finally, specific chemical interactions between small ions (e.g., Cazt ) and macromolecular functional groups (e.g., carboxyl and phenolic groups in humic substances) can be expected to alter the charge and conformation of the macromolecular groups in the interfacial region. PARTICLE PASSAGE/RETENTION
IN GROUND WATER AQUIFERS
When colloidal particles are stable, i.e., when they do not aggregate readily or adhere to surfaces, they may be expected to travel long distances in ground water systems and to carry particle-reactive pollutants with them. When they are unstable, i.e, when they aggregate rapidly or attach to solid surfaces, they can be expected to deposit extensively in porous media and to reduce aquifer permeability. Both types of behavior have been observed. One important key in this problem is the chemistry of the solid-solution interface and, in turn, the composition of the aqueous phase in natural systems and at hazardous waste sites. There have been several excellent studies of the transport of particles and associated pollutants in subsurface environments. A comprehensive summary of theories and observations has been prepared by McDowell-Boyer et al. [ 261. Here a few observations are cited and then attention is focussed on colloid chemical aspects of the subject. Gerba and co-workers report the transport of bacteria and viruses for distances of several hundred meters in aquifers [ 27-29 1. In contrast, Jang et al.
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[30] reported that aggregating bacteria clogged porous media and reduced permeability. Gschwend and Reynolds [ 311 report extensive movement of inorganic ion phosphate colloids at a wastewater infiltration site and Champlin and Eichholz [ 321 observed extensive mobility of kaolinite clay stabilized with surfactants in laboratory columns. Nevertheless, clays are used commercially in petroleum extraction to reduce permeability and seal shafts [33]. The effects of colloidal stability on ground water quality are seen in these illustrative examples to range from extensive transport and passage for stable colloids and particle-reactive pollutants to extensive clogging and retention for other, often unstable, colloids and associated pollutants. Humic substances and other natural organic macromolecules are present in ground waters and in leachates from hazardous waste sites. These substances can bind synthetic organic substances and have been reported to enhance their transport in porous media (e.g., see Ref. [34] ). It is also plausible that the surfactant nature of naturally occurring organic matter, invoked to explain the fouling of membranes, marine surfaces, and conduits, could also lead to reduced permeabilities in porous media. The present state of modelling particle deposition in porous media may be summarized as follows (see also the paper by Tobiason [ 35 ] ) . ( 1) For favorable chemical conditions (i.e., with no repulsive EDL or MAL interactions), the initial deposition of Brownian or submicron particles in clean beds can be well described by considering convective diffusion and Happel’s [ 361 cell model for flow through porous media [ 371. Physical transport models that neglect repulsive chemical interactions agree with experimental laboratory observations made under chemical conditions in which repulsive interactions are absent. WZ 1 z a. (2 ) For favorable chemical conditions, the initial deposition of non-Brownian particles can be well described by a trajectory approach including hydrodynamic retardation and using Happel’s cell model for fluid flow. Physical transport models and experimental measurements agree; WZ 1 cs a. (3) For unfavorable deposition caused by electrostatic (EDL) repulsive interactions, increasingly unfavorable chemical conditions result in systematic and significant reductions in the deposition of both Brownian and non-Brownian particles. The effects of steric repulsion on particle deposition in porous media have not yet been studied, either theoretically or experimentally. (4) Although particle deposition is retarded under unfavorable chemical conditions caused by EDL interactions, it is still substantially greater than predictions by models incorporating present double-layer structures and interaction mechanisms. Experimental stability factors (a values) are all 2 10M3 while predictions are many orders of magnitude smaller. (5) Particle deposition in porous media is an unsteady state process. The removal of particles from a flowing suspension can in many cases reduce the capacity of the porous media to retain additional particles so that particle re-
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moval deteriorates over time. This is typically the case when chemical conditions produce unfavorable deposition. It is also possible for retained particles to act as additional collectors or filter media. In this case deposition improves over some time as collectors accumulate within the porous media. This occurs when the surface chemistry of the particles in suspension permits attachment to previously retained particles. In such cases, clogging of porous media can occur. Models for the transport and attachment of suspended particles over time are preliminary and limit predictions of particle concentrations and clogging rates. Modelling studies McDowell-Boyer et al. [ 261 have presented some valuable estimates of particle transport through porous media obtained by applying filtration theory to ground water conditions. The results presented here are taken from a theoretical and experimental study by Tobiason [ 381 and agree rather well with those published by McDowell-Boyer et al. Application of physical theories for particle-particle interactions to a selected ground water condition is illustrated in Fig. 1. The predicted travel distance required for removal of 99% of particles is presented as a function of the size of the particles in suspension at a flow velocity of 0.1 m/day. When attachment is effective (a = 1, favorable interactions), the predicted travel dis-
Log Particle Radius (Fm) Fig. 1. Model simulation of the effect of colloidal stability (LY)on particle transport in a ground water aquifer. The travel distance at which particle removal is 99% is plotted as a function of particle size for a! = 1 and (Y= 0.001. Calculations are made for a flow velocity of 0.1 m/day, an aquifer porosity of 0.4, aquifer media with a diameter of 0.5 mm, a suspended particle density of 1.05 g crne3 and a temperature of 25 ’ C. After Tobiason [ 381.
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tance does not exceed more than a few mm for any suspended particle size. These results are based on rather rigorous theoretical modelling that has been confirmed in extensive controlled laboratory experiments and are consistent with filtration practice. Small particles are transported to the surfaces of media grains by diffusion and larger ones are transported by gravity under ground water conditions. Particles with a size in the order of a micron have the greatest ability to.travel through the aquifer; this size corresponds to bacteria and some clays. Nevertheless, when attachment is effective, particles are predicted to deposit close to their point of introduction into the aquifer. Clogging is possible, but passage is not. When particle attachment is unfavorable, represented in Fig. 1 by the dashed line calculated for cy= 0.001, the required travel distance for 99% removal increases considerably, but does not exceed about 20 m. This assumed value of cy represents the lower limit of stability factors reported for natural systems and laboratory experiments. Reports of extensive transport of microorganisms and other colloidal particles noted previously in this paper [29,31] are not consistent with the calculations presented in Fig. 1. The origin of this discrepancy does not originate in the modelling estimates of particle transport in porous media used in Fig. 1; these have been tested extensively in laboratory systems and confirmed in filtration practice. Possible explanations include (1) particle transport through fissures and cracks in nonhomogeneous media, (2) saturation of the “filtration” capacity of the media, (3) growth of organisms and formation of solid precipitates within the aquifer, (4) sampling errors leading to dislodgement of particles during sample withdrawal and formation of particles in sample wells and containers, and (5) chemical conditions leading to the establishment of interfacial conditions that render the suspended colloidal particles more stable than available observations of surface waters would indicate but more consistent with available theories for colloidal stability. Possibly the factors listed as 1 to 4 have all occurred in some instance. With the exception of aquifer heterogeneity, however, they are not considered definitive here. Saturation of the media should be a rare occurrence; extensive deposition should normally lead to clogging of aquifer media, not extensive passage of suspended particles. Sampling errors have been observed and procedures developed recently to control them. The emphasis here is on the last factor listed above, the interactions among colloidal particles under the chemical and hydrodynamic conditions in subsurface environments and the interactions of suspended colloidal particles with aquifer media under these conditions. Favorable interactions are expected to produce clogging. Very unfavorable interactions (a << 10e3 or WB 1000) could lead to passage of colloidal particles through aquifers. Experimental measurements of deposition in laboratory systems dominated by electrostatic particle-
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particle interactions yield values of cy2 low3 ( WI 1000). Both experimental and conceptual studies are required to evaluate and predict the passage of colloidal materials in ground water aquifers, and these studies should be focussed on the chemical aspects of the particle-particle interactions in these systems. COAGULATION IN LAKES
Solid particles are ubiquitous in all lakes. Pollutants and nutrients are associated with these particles. Inter-particle collisions occur and, at times, aggregates are formed. Particles and their aggregates are both subject to sedimentation and removal from the waters of a lake. Although particle aggregation or coagulation occurs in all natural waters, its significance can vary widely. Weilenmann and co-workers [39,40] have applied models for coagulation and sedimentation to Swiss hardwater lakes. The kinetics of coagulation were described by Smoluchowski’s equations for heterodisperse suspensions and sedimentation was characterized by Stokes’ equation. Small particles collide with larger ones primarily by Brownian diffusion and differential settling. Larger particles and aggregates are removed by settling. One effect of coagulation in lakes is to increase settling velocities of particles by aggregation, thereby reducing their residence time and concentration in the water column and increasing the flux of particles and particle-reactive substances to the sediments. A second process in hardwater lakes is mutual coagulation of dense CaCO, with algae, leading to an increase in the sedimentation rates of organic biomass in aggregates with a bulk density greater than that of the free microorganisms. Experimentally measured particle stability coefficients (~1values) for fresh surface waters range over two orders of magnitude, from about 0.001 to more than 0.1, and support the view that solution chemistry can control the colloidal stability of natural particles in aquatic systems. Major divalent cations such as Ca2+ destabilize particles (increase cu) and dissolved NOM stabilizes them (decreases a). Since solution chemistry varies widely among lakes, it is expected and observed that stability or sticking factors vary accordingly. In turn, the importance of coagulation as a process in lacustrine systems is expected to depend on solution chemistry, with low aggregation rates in soft colored waters and rapid coagulation occurring in hard waters that are low in NOM. Field observations of Lakes Zurich and Sempach support this view [ 39,401. Because of the importance of the kinetics of slow or unfavorable coagulation in natural waters, it is useful to formulate models for coagulation rates in natural aquatic systems. Some observations about the present state of modelling the kinetics of aggregation reactions are pertinent in developing such an assessment. (1) For favorable chemical conditions (i.e., with no repulsive EDL or MAL interactions), the initial rates of aggregation of monodisperse suspensions by
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Brownian diffusion or laminar shear can be well described when hydrodynamic retardation is considered. Theories and experiments agree; WE 1 ZTa. (2) For favorable chemical conditions, the initial aggregation rates of heterodisperse suspensions by Brownian motion, laminar and turbulent shear flow, and by differential sedimentation have been modelled extensively, but experimental validation is limited. (3) For unfavorable or slow coagulation caused by electrostatic (EDL) repulsive interactions, increasingly unfavorable chemical conditions result in systematic reductions in aggregation rates. (4) Although particle aggregation is retarded under unfavorable chemical conditions caused by EDL interactions, it is still substantially faster than most predictions incorporating present double layer structures and interaction mechanisms. Most experimental stability factors (a) are 2 10 -3 while typical predicted values are orders of magnitude smaller. (5) Theories for the effects of charged and uncharged macromolecules on particle-particle interactions are presently being actively developed. A kinetic theory for the coagulation of sterically stabilized suspensions has not yet been developed, and experimental measurements under defined conditions are lacking. Since natural organic macromolecules are ubiquitous in natural waters and since they are readily adsorbed at interfaces, an understanding of natural coagulation rates should be derived from an assessment of the effects of macromolecules on interfacial structure and on chemically derived forces in particleparticle interactions. The influence of [H+ 1, ionic strength and especially the specific chemical interactions between adsorbed NOM and major divalent cations such as Ca2 + should be included in such an assesment. Present knowledge about interfacial structure under conditions in natural aquatic systems does not provide a sufficient base for formulating a quantitative model for aggregation kinetics in these systems. CONCLUDING REMARKS
In the absence of a definitive theory for the origins of the stability and the kinetics of aggregation of natural particles in natural waters, some speculation may be a permissible, although inadequate, substitute. Aquatic humic substances range in molecular weight from about 500 to over 100 000 daltons. Cornel et al. [41] report determinations of the hydrodynamic radii of humic substances in solution. These authors indicate that these anionic macromolecules resemble flexible, linear polyelectrolytes at low ionic strength or at high pH. At high ionic strength and at neutral and acidic pHs, the molecules assume coiled configurations. For humic substances with nominal molecular weights ranging from 50 000 to 100 000, the hydrodynamic diameter decreased from 15 to 1.7 nm as the ionic strength increased from 4. lop4 to 1 M. The configura-
tions of adsorbed polyelectrolytes differ substantially from those that exist in solution, so that the molecular dimensions reported by Cornel et al. do not apply directly to dimensions in the MAL. At low ionic strength, charged macromolecules can be extended along the surface of a solid, occupying a significant area but not projecting into the solution. As an extreme, at low ionic strength all polyelectrolyte segments could occupy positions in the Stern layer and form an MAL with a thickness of 1 nm or so. At high ionic strength, intramolecular electrostatic repulsion is reduced substantially, and an adsorbed macromolecule will have a coiled configuration more closely resembling its state in solution. The thickness of the MAL could increase with ionic strength even though the hydrodynamic size of the macromolecules in solution decreases as ionic strength increases. The thickness could approach an upper limit equal to the hydrodynamic diameter of a macromolecule in solution at high ionic strength. The ionic strength of fresh waters may vary from about 10e4 to lo-‘, with corresponding diffuse layer thicknesses ranging from 30 to 3 nm. Consideration of saline lakes and the ocean extends the ionic strength of natural waters up to almost 1 and reduces diffuse layer thicknesses to 0.3 nm. The dimensions of the EDL and the MAL can be therefore comparable in many natural systems. This indicates that the colloidal stability of particles in aquatic environments can be affected by the chemistry of both small, primarily inorganic solutes and larger natural macromolecules, and that the interactions between these groups of species can be important in determining particle stability in natural systems. At low ionic strength and in the absence of Ca2+, the charge on anionic humic substances could stretch these molecules and cause them to adsorb in a flat configuration, limiting both the amount of organic matter adsorbed and the distance to which it extends from the surface into and through the diffuse layer (which would be in the order of 30 nm in thickness). Under these circumstances the thicker EDL could dominate particle stability. At the high ionic strength of oceanic waters, humic substances could coil, adsorb more extensively, and extend through the diffuse layer which would have a thickness of about 0.3 nm. The thickness of the EDL would be less than the MAL, so that the charge and configuration of the adsorbed organic macromolecules could control particle stability. At intermediate ionic strengths characteristic of many aquatic systems, both electrostatic (EDL) and macromolecular (MALI effects could contribute to particle-particle interactions. When divalent metal ions such as Ca2+ are present, they could react specifically with functional groups on humic substances, altering macromolecular conformations in the MAL, increasing adsorption, and probably also reducing both the charge and the thickness of the MAL. The latter effects would reduce colloidal stability (increase (x) and lead to enhanced rates of coagulation and deposition. The transport of particles and particle-reactive pollutants in natural aquatic systems such as lakes and in ground water aquifers is affected by particle-
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particle interactions in these systems. These interactions depend on the physical and chemical characteristics of the interfacial regions surrounding solid phases in these environments and, in turn, on surface and solution chemistry. The development of conceptual and experimental bases for describing the kinetics of aggregation and deposition processes in aquatic systems should be founded on present understandings of these processes derived from model systems and established in the field of colloid chemistry. In building on this base, new knowledge about the conformations of natural organic macromolecules and polyelectrolytes at interfaces and about the effects of these substances on particle-particle interactions will be needed. ACKNOWLEDGEMENTS
The financial support of the U.S. Environmental Protection Agency through grant number R812760, the U.S. National Science Foundation through grant number CEEBl-21501 and the Lyonnaise des Eaux are gratefully acknowledged.
REFERENCES V.M. Muller, Langmuir, 3 (1987) 621. J.M.H.M. Scheutjens and G.J. Fleer, Phys. Chem., 83 (1979)
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L. Sigg, in W. Stumm 1985, pp. 283-310.
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H.-B. Xue, W. Stumm and L. Sigg, Water Res., 22 (1988) 917. R.A. Niehof and G.I. Loeb, Limnol. Oceanogr., 17 (1972) 7.
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J.M.H.M. Scheutjens and G.J. Fleer, Phys. Chem., 84 (1980) 178. S.M. Karickhoff, Hydraul. Eng., 110 (1984) 707. R.P. Schwarzenbach and J. Westall, Environ. Sci. Technol., 15 (1981) 1360. D.M. Imboden and R.P. Schwarzenbach, in W. Stumm (Ed.), Chemical Processes in Lakes, Wiley-Interscience, New York, 1985, pp. l-30. D.A. Dzombak and F.M.M. Morel, Hydraul. Eng., 113 (1987) 430. (Ed.),
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J.K. Edzwald, J.B. Upchurch and C.R. O’Melia, Environ. Sci. Technol., 8 (1974) 58. J.M. Eckert and E.R. Sholkovitz, Geochim. Cosmochim. Acta, 40 (1976) 847. E.A. Boyle, J.M. Edmund and E.R. Sholkovitz, Geochim. Cosmochim. Acta, 41 (1977) 1313. K.A. Hunter and P.A. Liss, Nature, 282 (1979) 823. M. Smoluchowski, Z. Phys. Chem., 92 (1917) 129. M. Fuchs, Z. Phys., 89 (1934) 736. E. Tipping, Goechim. Cosmochim. Acta, 45 (1981) 191. E. Tipping and D. Cooke, Geochim. Cosmochim. Acta, 46 (1982) 75. E. Tipping and D.C. Higgins, Colloids Surfaces, 5 (1982) 85. E. Tipping and M.J. Heaton, Geochim. Cosmochim. Acta, 47 (1983) 1393. E.J.W. Verwey and J.Th.G. Overbeek, Theory of the Stability of Lyophobic Colloids, Elsevier, Amsterdam, 1948. J. Lyklema, in K.J. Ives (Ed.), The Scientific Basis of Flocculation, Sijthoff and Noordhoff, Alphen aan den Rhijn, 1978, pp. 3-36. L.K. Koopal and J. Lyklema, Discuss. Faraday Sot., 59 (1975)
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J. Lyklema, Pure Appl. Chem., 46 (1976) 149. B.V. Derjaguin, S.S. Dukhin and A.E. Yaroshchuk, J. Colloid Interface Sci., 115 (1987) 234. L.M. McDowell-Boyer, J.R. Hunt and N. Sitar, Water Resour. Res., 22 (1986) 1901. C.P. Gerba, C. Wallis and J.L. Melnick, Am. Sot. Civil Eng. J. Irrigation Drainage Div., 101 (IRl) (1975) 157. B.H. Keswick and C.P. Gerba, Environ. Sci. Technol., 14 (1980) 1290. C.P. Gerba and S.M. Goyal, in T. Asano (Ed.), Artificial Recharge of Groundwater, Butterworth, Stoneham, 1985, pp. 283-317. L.-K. Jang, P.W. Chang, J.E. Findley and T.F. Yen, Appl. Environ. Microbial., 46 (1983) 1066. P.M. Gschwend and M.D. Reynolds, J. Contaminant Hydrol., (1987) 309. J.B.F. Champlin and G.C. Eichholz, Health Phys., 30 (1976) 215. H. van Olphen, An Introduction to Clay Colloid Chemistry, 2nd edn, Wiley-Interscience, New York, 1977,318 pp. E.M. Perdue, In Transport of Contaminants in the Subsurface: The Role of Organic and Inorganic Colloidal Particles, U.S. Department of Energy, Washington, DC, 1986, pp. 25-37. J.E. Tobiason, Colloids Surfaces, 39 (1989) 53. J. Happel, Am. Inst. Chem. Eng., J., 4 (1958) 197. C.R. O’Melia, in W. Stumm (Ed.), Aquatic Surface Chemistry: Chemical Processes at the Particle-Water Interface, Wiley-Interscience, New York, 1987, pp. 385-403. J.E. Tobiason, Ph.D. dissertation, The Johns Hopkins University, Baltimore 1987,280 pp. U. Weilenmann, Ph.D. dissertation No. 8018, Eidgenossischen Technischen Hochschulen, Zurich 1986,163 pp. U. Weilenmann, C.R. O’Melia and W. Stumm, Limnol. and Oceanogr., 34 (1989) 1. P.K. Cornel, R.S. Summers and P.V. Roberts, J. Colloid Interface Sci., 110 (1986) 149. J.A. Davis, Geochim. Cosmochim. Acta, 46 (1982) 2381. W. Ali, C.R. O’Melia and J.K. Edzwald, Wat. Sci. Technol., 17 (1985) 701. W. Ali, Ph.D. dissertation, The Johns Hopkins University, Baltimore, 1985,222 pp. U. Weilenmann, Ph.D. thesis, No. 8018, Eidg. Tech. Hoch., Zurich, 1986,163 pp.
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Printed
in The Netherlands
Particle-Particle Interactions in Aquatic Systems CHARLES
R. O’MELIA*
Swiss Federal Institute for Water Resources and Water Pollution Control, 8600 Diibendorf (Switzerland) (Received 20 September
1988; accepted
16 January
1989)
ABSTRACT The significance of particles in the transport and fate of substances in aquatic environments is illustrated and present knowledge of the speciation and structure of interfaces in these environments is assessed. The importance of particle deposition and aggregation in ground water aquifers and in lakes is illustrated. Attention is focussed on the adsorption of natural organic macromolecules and polyelectrolytes at interfaces in natural systems and the impacts of these substances on colloidal stability and deposition in natural systems. Finally, the status of present understandings in this area is assessed and directions
for future study are proposed.
INTRODUCTION
This paper is written with three objectives. First, the significance of interfacial processes and particle-particle interactions in the transport and fate of pollutants in the environment is illustrated. Second, some approaches used in modelling interfacial processes and particle-particle interactions in natural waters are considered. Third, some modelling applications in natural systems are presented. Modelling particle-particle interactions requires knowledge or assumptions about the stoichiometry, speciation, structure, and streamlines in the interfacial regions between phases. In considering this physicochemical problem, the physical aspects of the system (e.g., the streamlines in interfacial regions) are drawn from fluid mechanics and the chemical properties of interfaces (stoichiometry, speciation and structure) are derived from chemistry. This separation neglects some aspects of interfacial processes that may derive from the coupling of these disciplines, such as electrodynamic effects on the viscous resistance when two charged particles approach each other in water [ 11. The chemicals used in modelling natural systems may be classified into three *Permanent address: Department of Geography and Environmental Hopkins University, Baltimore, MD 21218, U.S.A.
0166-6622/89/$03.50
0 1989 Elsevier Science Publishers B.V.
Engineering,
The Johns
types. First, there is the vast array of chemical substances that are present in aquatic systems, derived from natural or anthropogenic sources. Examples are humic substances, heterogeneous solid precipitates, weathering products, asbestos fibers and polychlorinated biphenyl compounds (PCBs ) . These are often described by collective or operational parameters such as dissolved organic carbon, filterable manganese, extractable ion, and reactive phosphorus. Actual chemical speciation in natural systems is often and perhaps normally unknown. Second, there are chemicals used in model systems, both conceptual and experimental. Examples include silver halide sols used extensively in studies of colloidal stability, metal oxide surfaces and simple metals and ligands used in adsorption studies, and the spherical latex particles used in studies of aggregation and deposition. These substances have well defined physical and chemical properties, but they are models of natural systems, not replicas of them. Third, there are mathematical or imaginary chemicals. Examples include the point charges in the Debye-Hiickel and Gouy-Chapman theories, and the linear, flexible macromolecules assumed in theories for polymer adsorption [ 2,3]. These chemicals exist only in the mind, but their use has contributed extensive insights into the colloid chemistry of natural and other systems. Effective modelling of the chemistry of particle-particle interactions in natural aquatic systems can require the use of all of these types of chemicals, with the appropriate mix depending on the problem to be addressed. PARTICLES, POLLUTANTS AND COLLOIDAL STABILITY
Solid particles are ubiquitous in natural waters, and these solid particles have a role in the transport and fate of pollutants affecting human health and aquatic ecosystems. Most pollutants in water are associated with solid particles or are particles themselves. Examples include, (i) trace metals adsorbed on clays, metal oxides, and metal carbonates, (ii) synthetic organic pollutants adsorbed on organic detritus, (iii) trace metals and organic substances associated with living biomass, (iv) pathogenic microorganisms, (v) asbestos fibers, and (vi) humic precursors of trihalomethanes. The thermodynamics and the kinetics of the transport and degradation of many pollutants are controlled by properties of the interfacial regions that surround phases (solids, liquids and gases) in aquatic, atmospheric and terrestrial environments. Particle-pollutant
reactions
Karickhoff [4] has reviewed the thermodynamics and kinetics of the sorption of organic pollutants in aquatic systems. The sorption of uncharged organic chemicals to natural particles is dominated by hydrophobic interactions and depends primarily on a chemical’s affinity for water, typically described by an octanol-water partition coefficient, and on the organic carbon content
257
of the solid sorbent phase. The contribution of inorganic particles to the uptake of uncharged organic pollutants is negligible in most cases. An estimate of the sorption of an uncharged organic pollutant in natural systems can be made on the basis of the chemical’s octanol-water partition coefficient, the organic carbon content of the solid phase, and the concentration of solid sorbent in the water, the soil, or the sediment of interest. For hydrophilic organic pollutants, nonhydrophobic contributions to sorption can be important and inorganic surfaces such as clays and metal oxides can be significant adsorbents. The sorption of organic pollutants to suspended particles, soil and sediments can generally be viewed as rapid, although true equilibrium may take weeks or more to achieve and the process is often not completely reversible. Several investigators have studied the transport and fate of nonpolar organic compounds in natural waters. Representative studies are given by Schwarzenbath and co-workers [ 5,6]. Laboratory studies [ 51 using natural sorbents and nonpolar organic solutes indicated that the sorption of many organic compounds is proportional to the organic carbon content of the natural sorbent for solids with an organic carbon content greater than 0.1% and that sorption could be predicted from the octanol-water partition coefficient of the nonpolar organic solute and the organic carbon content of the natural solid sorbent. Modelling studies of Swiss lakes [6] indicate that particulate organic carbon from photosynthetic processes in the water column can be important in the transport and deposition of a nonpolar hydrophobic substance such as hexachlorobenzene in lake sediments. The adsorption of inorganic pollutants in aquatic systems has been reviewed by Dzombak and Morel [ 71. In contrast to the nonspecific adsorption observed for hydrophobic organic pollutants, the adsorption of inorganic solutes is viewed as a site-specific process in which ions bind chemically at functional groups on solid surfaces. The total energy of interaction includes an electrostatic contribution and a chemical contribution. Surface complex formation models of varying complexity are available to model pollutant adsorption. The reaction is dependent upon pollutant type and concentration, pH, ionic strength and solid or surface concentration. Although not stressed by these authors, organic surfaces and humic substances can be important in the speciation and adsorption of inorganic pollutants in natural waters. Sigg [8] has reported results from Lakes Constance and Zurich indicating that metals such as Cu, Zn, Cd, and Pb are continuously removed from lake waters by becoming attached to settling particles. The partition of these metals between the settling particles and the water determines the relative residence times of the metals in the lake and their residual concentrations in the water column. Biological surfaces are indicated as better scavengers for heavy metals than inorganic ones. A large part of the fresh sedimenting particles is of biogenie origin, and a productive lake with a high sedimentation rate represents a situation that is favorable for an efficient removal of metals. Xue and co-
258
workers [ 91 have conducted laboratory experiments on the binding of metal ions to algal surfaces and indicate that the surfaces of algal cells have a high affinity for Cu (II) and Cd (II). These authors interpret the adsorption of metals on algal cells in terms of surface complex formation equilibria and extract average conditional equilibrium constants for the adsorption reactions from laboratory titration curves using voltametric methods to determine free metal concentrations. Humic substances and colloidal stability There is extensive evidence from fresh waters, estuaries, and the oceans that the surface properties and colloidal stability of particles in natural waters are affected by naturally occurring organic substances dissolved in these waters. These effects of natural organic matter in establishing colloidal stability in aquatic systems also are anticipated to occur in subsurface environments and to affect particle and pollutant passage/retention in subsurface systems. Organic substances affecting colloidal stability are expected to include not only naturally occurring materials but also related materials produced by biological processes in wastewater treatment plants and at hazardous waste sites, industrial surfactants and solvents and some other synthetic organic compounds that enter the environment. Humic substances are anionic polyelectrolytes of low to moderate molecular weight; their charge is due primarily to carboxyl and phenolic groups; they have both aromatic and aliphatic components and can be surface active; they are refractive and can persist for centuries or longer; they are produced by biological processes and are ubiquitous in natural waters; they have analogies to organic substances produced in large quantities in biological active landfills, hazardous waste sites and wastewater treatment facilities; they have been reported to stabilize particles in water. Some investigators have noted that major divalent metal ions in natural waters (in particular, Ca2+ ) can exert destabilizing effects on natural and synthetic particles at low metal ion concentrations. Some examples relating natural organic matter (NOM) and divalent metals to the surface properties and colloidal stability of particles in natural waters follow. Niehof and Loeb [lo] determined the electrophoretic mobilities of several inorganic solids after exposure to natural seawater and to seawater pretreated with ultraviolet radiation to photooxidize organic matter. All solids became moderately electronegative in natural seawater and all showed charges consistent with their chemical composition when immersed in inorganic-free seawater. The authors concluded that the surface charge exhibited by solids in seawater is due to adsorbed organic matter. Edzwald et al. [ll] studied the kinetics of particle-particle interactions in the coagulation of clays and estuarine sediments by solutions of NaCl and of synthetic (inorganic) seawater. All suspensions coagulated more rapidly in artificial seawater than in NaCl
solutions of equal ionic strength. The authors concluded that calcium and magnesium ions in seawater exert specific destabilizing effects on suspended particles. Sholkovitz and co-workers [ 12,131 studied the extent of flocculation of iron, aluminium and humic substances in estuaries. They observed that iron colloids in rivers can be stabilized by dissolved organic matter. These colloids are destabilized in estuaries by seawater cations (magnesium and especially calcium ions) through specific and electrostatic interactions with humic materials. Hunter and Liss [ 14 ] determined the electrophoretic mobilities of particles collected from four estuaries and evaluated them in terms of the salinity in the estuary at the site of collection. All particles at all locations were negatively charged. The authors concluded that dissolved organic matter formed surface coatings that gave the particles a consistent electronegative charge. They also showed that in calcareous rivers containing high calcium concentrations, the electrophoretic mobilities of the particles were smaller (less negative) than in rivers low in Ca’+. They concluded that some of the charge on the particles can be neutralized by incorporation of Ca2+ “into the fixed part of the double layer.” In studies of the kinetics of coagulation and deposition, it has become customary to introduce a sticking coefficient, stability ratio, attachment probability, or attachment factor. Two interchangeable approaches [ 15,161 are used. In considering the kinetics of slow or unfavorable coagulation, Smoluchowski [ 151 introduced a sticking coefficient (a) which can be defined as the rate at which colliding particles adhere to form aggregates divided by the rate at which they collide. For a completely unstable suspension (x = 1; for a perfectly stable suspension, a=O. Smoluchowski did not relate attachment to surface or solution chemistry and did not include some physical processes such as hydrodynamic retardation in considering interparticle contact rates. Nevertheless, his characterization of unfavorable particle-particle interactions by the stability factor c11 has been used extensively. Fuchs [ 16] first combined particle transport and chemistry. He considered Brownian diffusion in a force field and defined the stability ratio Was the ratio of the collision rate to the aggregation rate ( W= a-‘). W has been expressed in terms of the maximum or peak net repulsive interaction between two colliding charged particles and, in turn, to certain solution and surface characteristics. For a completely unstable suspension W= 1; for a perfectly stable suspension W= GO.The Fuchs’ stability factor, W, like Smoluchowski’s a!!,has been used extensively by investigators reporting experimental or theoretical studies of the kinetics of coagulation or deposition. Tipping and co-workers [ 17-201 studied the effects of humic substances and divalent metal ions on the surface properties of iron and manganese oxides. Measurements included electrophoretic mobility, adsorption of humic substances, and coagulation kinetics. Values of a were estimated to range from 0.01 to 1 ( W ranged from 100 to 1) for hematite coagulated in the presence of
Ca2+ and humic substances [ 191. Tipping and co-workers concluded that adsorbed humic substances produced negative charges on the particle studied. This charge was reduced by the presence of divalent metal ions, especially Ca2+. Humic substances slowed coagulation rates and calcium ions enhanced them, possibly by specific interactions with the humic substances. In addition to enhancing coagulation rates, calcium also substantially increased the adsorption of humic substances on hematite. Davis [42] studied the adsorption on alumina and kaolinite of natural organic matter extracted from lake sediments. Measurements included electrophoretic mobility, charge titration and adsorption of NOM. The author concluded that NOM was readily adsorbed on these surfaces, giving them a negative charge. Calcium enhanced the adsorption of organic matter on alumina at high pH and reduced its adsorption at pH less than 7. Davis proposed that adsorption of NOM occurs by complex formation between surface hydroxyl groups and acidic functional groups on the organic matter. Ali et al. [43] reported the results of laboratory measurements of a! for the coagulation of natural particles from the sediments of the Loch Raven Reservoir (Maryland) in suspensions at pH 8.1 and containing various combinations of calcium ions and aquatic fulvic acid. In the absence of Ca2+, ar decreased from 0.035 to 0.006 ( W increased from 30 to 170) as the concentration of fulvic acid added to the suspensions increased from 0 to 20 mg 1-l expressed as dissolved organic carbon (DOC ) . Most of this decrease in cv (increase in stability) occurred in the DOC range from 0 to 5 mg 1-l. Calcium acted as a destabilizing agent, a! increased from 0.035 to 0.2 ( W decreased from 30 to 5) as calcium was increased from 0 to 10e3 M. Concurrent additions of fulvic acid and Ca2+ yielded intermediate effects. The ranges of Ca2+ and DOC studied in this work are representative of most fresh water systems and indicate that wide variation in a! can be expected among natural waters, including both surface and ground waters. All measured values of cv (0.006-0.23 ), however, were greater than expected from theoretical predictions, for which a+0 and W-co Ali [ 441 reported extensive measurements of cyfor the surface waters of the Loch Raven Reservoir and limited measurements for other lakes, Statistical analysis of the data suggest that variations of cyin these natural environments were related to variations in DOC. Weilenmann [45 ] reported measurements and modelling of coagulation and sedimentation in Swiss hardwater lakes. Measured stability factors (a values) indicated substantial variation in particle stability among the Swiss lakes studied. These differences were interpreted qualitatively in terms of surface and solution chemistry. Divalent metal ions decreased particle stability, probably by specific adsorption. Humic substances stabilized particles, probably by electrostatic and possibly also by steric effects.
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Interfacial structure and particle stability The effects of humic substances and divalent metal ions on the colloidal stability of natural particles are common observations without clear origins. Early studies showed that particles carry a negative charge in the presence of natural organic matter, regardless of the composition of the solid phase. This was taken as indirect evidence that humic substances stabilize particles electrostatically. Calcium ions, in turn, might contribute electrostatically to particle destabilization. More recently, there has been speculation that nonelectrostatic (steric) effects are involved in the stabilization of particles by humic materials and that specific chemical interactions (complex formation) are involved in colloid destabilization by divalent metal ions. There are, however, no conclusive evidence and no consistent theory for these hypotheses. Some speculations follow. A consideration of the interaction between two suspended particles (e.g., in a lake) or between a suspended particle and a stationary collector (e.g., in a ground water aquifer) can begin with a determination of the structure of a single solid-solution interface. With this established, the interaction between two solid bodies (as characterized by force, free energy or disjoining pressure) can be addressed. The magnitudes of the interaction forces that determine stability in particle-particle interactions depend on the properties of each solid-solution interface. Electrostatic forces are almost universal in water since few particles are uncharged in aquatic environments, at least over a substantial pH range. Electrolytes and pH strongly affect the magnitude of these electrostatic forces. Specific chemical interactions between solutes and surfaces can be decisive in establishing surface properties and structure. Steric effects are potentially most significant in systems containing uncharged particles with adsorbed nonionic macromolecules. In this case, the amount of adsorbed polymer and solvent quality are important factors. The relative importance of electrostatic and steric effects is not clear in systems containing charged surfaces and adsorbed macromolecules, especially if the macromolecules are ionizable. Two distinct approaches to the structure of a solid/solution interface are considered here: (1) an electric double layer (termed EDL) and (2) a macromolecular adsorbed layer (denoted here as MAL) . Separately these two structures are understood fairly well, at least in simplified cases. However, less is known about the effects of charge in the EDL on the conformation of adsorbed macromolecules in the MAL and, reciprocally, about the effects of charged and uncharged macromolecules comprising the MAL on EDL properties. These two distinct approaches result in two distinct mechanisms for colloidal stability termed electrostatic and steric stabilization. Electrostatic stabilization results from energetically unfavorable overlap of the diffuse ion atmospheres surrounding all charged particles in water. The well-known Derjaguin-
262
Landau-Verwey-Overbeek (DLVO) theory [21] of the electrical double layer and colloidal stability provides a model for evaluation of the combined effects of electrostatic repulsion and van der Waals attraction between surfaces in the absence of adsorbed macromolecules and when all solute chemicals can be treated as point charges. When electrostatic repulsion dominates over van der Waals attraction, the results are termed “slow” coagulation and “unfavorable” deposition and filtration. A good didactic summary has been given by Lyklema [ 221, The DLVO model has successfully described the effects of counterion charge on the concentration of simple electrolytes required to coagulate many suspensions (the Schulze-Hardy rule) but it has been unsuccessful in describing the effects of particle or solution properties on the kinetics of unfavorable coagulation or deposition processes. This severely restricts its use in natural systems where rate processes are often the principal interest. Steric stabilization can result from the adsorption of polymers at solid-water interfaces. Large polymers can form adsorbed segments on a solid surface with loops and tails extending into solution [ 2,3,22]. Steric stabilization results from energetically unfavorable intermolecular interactions and entropically unfavorable compression in overlapping macromolecular adsorbed layers. Steric effects are likely to be important when the thickness of the macromolecular adsorbed layer is about the same or larger than the diffuse layer thickness. In the absence of charge effects, the energy of interaction between the adsorbed layers on two particles can be calculated by summing the various contributions to the free energy of the system as a function of the distance separating the surfaces. Under conditions of unfavorable intermolecular interactions, a repulsive interaction energy develops when the particles are close enough so that dangling tails of adsorbed macromolecules interact. On close approach, conformational entropy losses can give rise to strong repulsion if desorption of the macromolecules does not occur. When surface coverage is incomplete so that open sites remain available for adsorption, dangling tails can adsorb to the opposite surface, leading to aggregation. In natural aquatic systems it is useful and realistic to consider an interfacial region as comprised of small ions and larger organic macromolecules, i.e., as a system with properties of.both the EDL and an MAL. A definitive theoretical description of this combination remains to be developed. Uncharged macromolecules have significant effects on the properties of the electric double layer. For example, potentiometric titration of a silver iodide-polyvinyl alcohol (PVA) system showed a decrease in the slope of the curve of surface charge versus pAg with increasing amount of PVA adsorbed, as well as a shift in the point of zero charge (p.z.c. ) to smaller pAg [ 231. The reduction in slope indicated a decrease in the differential capacitance and the shift in the p.z.c. indicated change in the surface potential. In interpreting these observations, a useful approximation is to consider train segments of the polymer to be located in the Stern layer, and loop and tail segments to be in the diffuse layer [24]. The
263
dimension of a polymer segment is of the order of 0.5 nm, and train segments displace specifically adsorbing ions and water molecules, Loop and tail segments displace ions and water molecules in the diffuse layer. Possible consequences include: (1) Reduction of the differential capacitance in the Stern Layer due to displacement of specifically adsorbing ions or to a reduction in the dielectric constant. (2) Change in the surface potential due to replacement of water dipoles by nonpolar polymer segments. (3) Change in local activity coefficients. (4) Outward shift of the plane of electrokinetic shear. (5 ) Expansion of the diffuse layer. (6) Binding of water molecules in the diffuse layer by hydrophilic polymer groups, reducing the local dissolving power of water [ 251. Charged macromolecular species have even more complex effects on EDL properties than do uncharged polymers. Ionized macromolecular segments contribute fixed charge in the diffuse layer as well as in the Stern layer, so that the local distribution of mobile indifferent electrolytes is affected. The conformation of the adsorbed polyelectrolytes (the distribution of charged segments in the EDL) is also dependent on the local electrolyte concentration. Finally, specific chemical interactions between small ions (e.g., Cazt ) and macromolecular functional groups (e.g., carboxyl and phenolic groups in humic substances) can be expected to alter the charge and conformation of the macromolecular groups in the interfacial region. PARTICLE PASSAGE/RETENTION
IN GROUND WATER AQUIFERS
When colloidal particles are stable, i.e., when they do not aggregate readily or adhere to surfaces, they may be expected to travel long distances in ground water systems and to carry particle-reactive pollutants with them. When they are unstable, i.e, when they aggregate rapidly or attach to solid surfaces, they can be expected to deposit extensively in porous media and to reduce aquifer permeability. Both types of behavior have been observed. One important key in this problem is the chemistry of the solid-solution interface and, in turn, the composition of the aqueous phase in natural systems and at hazardous waste sites. There have been several excellent studies of the transport of particles and associated pollutants in subsurface environments. A comprehensive summary of theories and observations has been prepared by McDowell-Boyer et al. [ 261. Here a few observations are cited and then attention is focussed on colloid chemical aspects of the subject. Gerba and co-workers report the transport of bacteria and viruses for distances of several hundred meters in aquifers [ 27-29 1. In contrast, Jang et al.
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[30] reported that aggregating bacteria clogged porous media and reduced permeability. Gschwend and Reynolds [ 311 report extensive movement of inorganic ion phosphate colloids at a wastewater infiltration site and Champlin and Eichholz [ 321 observed extensive mobility of kaolinite clay stabilized with surfactants in laboratory columns. Nevertheless, clays are used commercially in petroleum extraction to reduce permeability and seal shafts [33]. The effects of colloidal stability on ground water quality are seen in these illustrative examples to range from extensive transport and passage for stable colloids and particle-reactive pollutants to extensive clogging and retention for other, often unstable, colloids and associated pollutants. Humic substances and other natural organic macromolecules are present in ground waters and in leachates from hazardous waste sites. These substances can bind synthetic organic substances and have been reported to enhance their transport in porous media (e.g., see Ref. [34] ). It is also plausible that the surfactant nature of naturally occurring organic matter, invoked to explain the fouling of membranes, marine surfaces, and conduits, could also lead to reduced permeabilities in porous media. The present state of modelling particle deposition in porous media may be summarized as follows (see also the paper by Tobiason [ 35 ] ) . ( 1) For favorable chemical conditions (i.e., with no repulsive EDL or MAL interactions), the initial deposition of Brownian or submicron particles in clean beds can be well described by considering convective diffusion and Happel’s [ 361 cell model for flow through porous media [ 371. Physical transport models that neglect repulsive chemical interactions agree with experimental laboratory observations made under chemical conditions in which repulsive interactions are absent. WZ 1 z a. (2 ) For favorable chemical conditions, the initial deposition of non-Brownian particles can be well described by a trajectory approach including hydrodynamic retardation and using Happel’s cell model for fluid flow. Physical transport models and experimental measurements agree; WZ 1 cs a. (3) For unfavorable deposition caused by electrostatic (EDL) repulsive interactions, increasingly unfavorable chemical conditions result in systematic and significant reductions in the deposition of both Brownian and non-Brownian particles. The effects of steric repulsion on particle deposition in porous media have not yet been studied, either theoretically or experimentally. (4) Although particle deposition is retarded under unfavorable chemical conditions caused by EDL interactions, it is still substantially greater than predictions by models incorporating present double-layer structures and interaction mechanisms. Experimental stability factors (a values) are all 2 10M3 while predictions are many orders of magnitude smaller. (5) Particle deposition in porous media is an unsteady state process. The removal of particles from a flowing suspension can in many cases reduce the capacity of the porous media to retain additional particles so that particle re-
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moval deteriorates over time. This is typically the case when chemical conditions produce unfavorable deposition. It is also possible for retained particles to act as additional collectors or filter media. In this case deposition improves over some time as collectors accumulate within the porous media. This occurs when the surface chemistry of the particles in suspension permits attachment to previously retained particles. In such cases, clogging of porous media can occur. Models for the transport and attachment of suspended particles over time are preliminary and limit predictions of particle concentrations and clogging rates. Modelling studies McDowell-Boyer et al. [ 261 have presented some valuable estimates of particle transport through porous media obtained by applying filtration theory to ground water conditions. The results presented here are taken from a theoretical and experimental study by Tobiason [ 381 and agree rather well with those published by McDowell-Boyer et al. Application of physical theories for particle-particle interactions to a selected ground water condition is illustrated in Fig. 1. The predicted travel distance required for removal of 99% of particles is presented as a function of the size of the particles in suspension at a flow velocity of 0.1 m/day. When attachment is effective (a = 1, favorable interactions), the predicted travel dis-
Log Particle Radius (Fm) Fig. 1. Model simulation of the effect of colloidal stability (LY)on particle transport in a ground water aquifer. The travel distance at which particle removal is 99% is plotted as a function of particle size for a! = 1 and (Y= 0.001. Calculations are made for a flow velocity of 0.1 m/day, an aquifer porosity of 0.4, aquifer media with a diameter of 0.5 mm, a suspended particle density of 1.05 g crne3 and a temperature of 25 ’ C. After Tobiason [ 381.
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tance does not exceed more than a few mm for any suspended particle size. These results are based on rather rigorous theoretical modelling that has been confirmed in extensive controlled laboratory experiments and are consistent with filtration practice. Small particles are transported to the surfaces of media grains by diffusion and larger ones are transported by gravity under ground water conditions. Particles with a size in the order of a micron have the greatest ability to.travel through the aquifer; this size corresponds to bacteria and some clays. Nevertheless, when attachment is effective, particles are predicted to deposit close to their point of introduction into the aquifer. Clogging is possible, but passage is not. When particle attachment is unfavorable, represented in Fig. 1 by the dashed line calculated for cy= 0.001, the required travel distance for 99% removal increases considerably, but does not exceed about 20 m. This assumed value of cy represents the lower limit of stability factors reported for natural systems and laboratory experiments. Reports of extensive transport of microorganisms and other colloidal particles noted previously in this paper [29,31] are not consistent with the calculations presented in Fig. 1. The origin of this discrepancy does not originate in the modelling estimates of particle transport in porous media used in Fig. 1; these have been tested extensively in laboratory systems and confirmed in filtration practice. Possible explanations include (1) particle transport through fissures and cracks in nonhomogeneous media, (2) saturation of the “filtration” capacity of the media, (3) growth of organisms and formation of solid precipitates within the aquifer, (4) sampling errors leading to dislodgement of particles during sample withdrawal and formation of particles in sample wells and containers, and (5) chemical conditions leading to the establishment of interfacial conditions that render the suspended colloidal particles more stable than available observations of surface waters would indicate but more consistent with available theories for colloidal stability. Possibly the factors listed as 1 to 4 have all occurred in some instance. With the exception of aquifer heterogeneity, however, they are not considered definitive here. Saturation of the media should be a rare occurrence; extensive deposition should normally lead to clogging of aquifer media, not extensive passage of suspended particles. Sampling errors have been observed and procedures developed recently to control them. The emphasis here is on the last factor listed above, the interactions among colloidal particles under the chemical and hydrodynamic conditions in subsurface environments and the interactions of suspended colloidal particles with aquifer media under these conditions. Favorable interactions are expected to produce clogging. Very unfavorable interactions (a << 10e3 or WB 1000) could lead to passage of colloidal particles through aquifers. Experimental measurements of deposition in laboratory systems dominated by electrostatic particle-
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particle interactions yield values of cy2 low3 ( WI 1000). Both experimental and conceptual studies are required to evaluate and predict the passage of colloidal materials in ground water aquifers, and these studies should be focussed on the chemical aspects of the particle-particle interactions in these systems. COAGULATION IN LAKES
Solid particles are ubiquitous in all lakes. Pollutants and nutrients are associated with these particles. Inter-particle collisions occur and, at times, aggregates are formed. Particles and their aggregates are both subject to sedimentation and removal from the waters of a lake. Although particle aggregation or coagulation occurs in all natural waters, its significance can vary widely. Weilenmann and co-workers [39,40] have applied models for coagulation and sedimentation to Swiss hardwater lakes. The kinetics of coagulation were described by Smoluchowski’s equations for heterodisperse suspensions and sedimentation was characterized by Stokes’ equation. Small particles collide with larger ones primarily by Brownian diffusion and differential settling. Larger particles and aggregates are removed by settling. One effect of coagulation in lakes is to increase settling velocities of particles by aggregation, thereby reducing their residence time and concentration in the water column and increasing the flux of particles and particle-reactive substances to the sediments. A second process in hardwater lakes is mutual coagulation of dense CaCO, with algae, leading to an increase in the sedimentation rates of organic biomass in aggregates with a bulk density greater than that of the free microorganisms. Experimentally measured particle stability coefficients (~1values) for fresh surface waters range over two orders of magnitude, from about 0.001 to more than 0.1, and support the view that solution chemistry can control the colloidal stability of natural particles in aquatic systems. Major divalent cations such as Ca2+ destabilize particles (increase cu) and dissolved NOM stabilizes them (decreases a). Since solution chemistry varies widely among lakes, it is expected and observed that stability or sticking factors vary accordingly. In turn, the importance of coagulation as a process in lacustrine systems is expected to depend on solution chemistry, with low aggregation rates in soft colored waters and rapid coagulation occurring in hard waters that are low in NOM. Field observations of Lakes Zurich and Sempach support this view [ 39,401. Because of the importance of the kinetics of slow or unfavorable coagulation in natural waters, it is useful to formulate models for coagulation rates in natural aquatic systems. Some observations about the present state of modelling the kinetics of aggregation reactions are pertinent in developing such an assessment. (1) For favorable chemical conditions (i.e., with no repulsive EDL or MAL interactions), the initial rates of aggregation of monodisperse suspensions by
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Brownian diffusion or laminar shear can be well described when hydrodynamic retardation is considered. Theories and experiments agree; WE 1 ZTa. (2) For favorable chemical conditions, the initial aggregation rates of heterodisperse suspensions by Brownian motion, laminar and turbulent shear flow, and by differential sedimentation have been modelled extensively, but experimental validation is limited. (3) For unfavorable or slow coagulation caused by electrostatic (EDL) repulsive interactions, increasingly unfavorable chemical conditions result in systematic reductions in aggregation rates. (4) Although particle aggregation is retarded under unfavorable chemical conditions caused by EDL interactions, it is still substantially faster than most predictions incorporating present double layer structures and interaction mechanisms. Most experimental stability factors (a) are 2 10 -3 while typical predicted values are orders of magnitude smaller. (5) Theories for the effects of charged and uncharged macromolecules on particle-particle interactions are presently being actively developed. A kinetic theory for the coagulation of sterically stabilized suspensions has not yet been developed, and experimental measurements under defined conditions are lacking. Since natural organic macromolecules are ubiquitous in natural waters and since they are readily adsorbed at interfaces, an understanding of natural coagulation rates should be derived from an assessment of the effects of macromolecules on interfacial structure and on chemically derived forces in particleparticle interactions. The influence of [H+ 1, ionic strength and especially the specific chemical interactions between adsorbed NOM and major divalent cations such as Ca2 + should be included in such an assesment. Present knowledge about interfacial structure under conditions in natural aquatic systems does not provide a sufficient base for formulating a quantitative model for aggregation kinetics in these systems. CONCLUDING REMARKS
In the absence of a definitive theory for the origins of the stability and the kinetics of aggregation of natural particles in natural waters, some speculation may be a permissible, although inadequate, substitute. Aquatic humic substances range in molecular weight from about 500 to over 100 000 daltons. Cornel et al. [41] report determinations of the hydrodynamic radii of humic substances in solution. These authors indicate that these anionic macromolecules resemble flexible, linear polyelectrolytes at low ionic strength or at high pH. At high ionic strength and at neutral and acidic pHs, the molecules assume coiled configurations. For humic substances with nominal molecular weights ranging from 50 000 to 100 000, the hydrodynamic diameter decreased from 15 to 1.7 nm as the ionic strength increased from 4. lop4 to 1 M. The configura-
tions of adsorbed polyelectrolytes differ substantially from those that exist in solution, so that the molecular dimensions reported by Cornel et al. do not apply directly to dimensions in the MAL. At low ionic strength, charged macromolecules can be extended along the surface of a solid, occupying a significant area but not projecting into the solution. As an extreme, at low ionic strength all polyelectrolyte segments could occupy positions in the Stern layer and form an MAL with a thickness of 1 nm or so. At high ionic strength, intramolecular electrostatic repulsion is reduced substantially, and an adsorbed macromolecule will have a coiled configuration more closely resembling its state in solution. The thickness of the MAL could increase with ionic strength even though the hydrodynamic size of the macromolecules in solution decreases as ionic strength increases. The thickness could approach an upper limit equal to the hydrodynamic diameter of a macromolecule in solution at high ionic strength. The ionic strength of fresh waters may vary from about 10e4 to lo-‘, with corresponding diffuse layer thicknesses ranging from 30 to 3 nm. Consideration of saline lakes and the ocean extends the ionic strength of natural waters up to almost 1 and reduces diffuse layer thicknesses to 0.3 nm. The dimensions of the EDL and the MAL can be therefore comparable in many natural systems. This indicates that the colloidal stability of particles in aquatic environments can be affected by the chemistry of both small, primarily inorganic solutes and larger natural macromolecules, and that the interactions between these groups of species can be important in determining particle stability in natural systems. At low ionic strength and in the absence of Ca2+, the charge on anionic humic substances could stretch these molecules and cause them to adsorb in a flat configuration, limiting both the amount of organic matter adsorbed and the distance to which it extends from the surface into and through the diffuse layer (which would be in the order of 30 nm in thickness). Under these circumstances the thicker EDL could dominate particle stability. At the high ionic strength of oceanic waters, humic substances could coil, adsorb more extensively, and extend through the diffuse layer which would have a thickness of about 0.3 nm. The thickness of the EDL would be less than the MAL, so that the charge and configuration of the adsorbed organic macromolecules could control particle stability. At intermediate ionic strengths characteristic of many aquatic systems, both electrostatic (EDL) and macromolecular (MALI effects could contribute to particle-particle interactions. When divalent metal ions such as Ca2+ are present, they could react specifically with functional groups on humic substances, altering macromolecular conformations in the MAL, increasing adsorption, and probably also reducing both the charge and the thickness of the MAL. The latter effects would reduce colloidal stability (increase (x) and lead to enhanced rates of coagulation and deposition. The transport of particles and particle-reactive pollutants in natural aquatic systems such as lakes and in ground water aquifers is affected by particle-
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particle interactions in these systems. These interactions depend on the physical and chemical characteristics of the interfacial regions surrounding solid phases in these environments and, in turn, on surface and solution chemistry. The development of conceptual and experimental bases for describing the kinetics of aggregation and deposition processes in aquatic systems should be founded on present understandings of these processes derived from model systems and established in the field of colloid chemistry. In building on this base, new knowledge about the conformations of natural organic macromolecules and polyelectrolytes at interfaces and about the effects of these substances on particle-particle interactions will be needed. ACKNOWLEDGEMENTS
The financial support of the U.S. Environmental Protection Agency through grant number R812760, the U.S. National Science Foundation through grant number CEEBl-21501 and the Lyonnaise des Eaux are gratefully acknowledged.
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