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lv Solid Particles in Suspension INTRODUCTION In addition to dissolved substances, drinking water typically contains small amounts of very finely divided solid particles of several kinds. These particles, ranging in size from colloidal dimensions to about 100 ,um, are composed of inorganic and organic materials that are derived from soils and rocks and from the debris of human activity with which the raw water has come in contact. They include clays, acicular or fibrous particles of asbestos minerals, and organic particles resulting from the decomposition of plant and animal debris in the soil. Little is known about the effects that these suspended solids may have on the health of those who drink water that contains them. However, there is widespread concern over the biological effects of the asbestos mineral fibers that occur in water, since similar fibers are known to be carcinogenic when air heavily laden with them is inhaled for many years. In view of this concern that such fibers as occur in water may be injurious to health, their occurrence, characterization, analysis, and biological ejects are reviewed in some detail. No evidence has yet been discovered that either of the other classes of common particulate contaminants of drinking water-clays and organic colloids has any direct effect on health. Nevertheless, it is possible that both may indirectly affect the quality of drinking water because they can adsorb a variety of toxic substances, bacteria, and viruses from solution or suspension and bind them more or less strongly. By such means these 135

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136 DRINKING WATER AND HEALTH materials may serve to concentrate and transport some water pollutants and protect them from removal by water treatment. For this reason, the properties of clays and organic particulates are also discussed, together with the tendency of chemicals, bacteria, and viruses to become concentrated at the surfaces of such particles. Removal of suspended particles from water is briefly reviewed, together with the significance of measurements of turbidity as an index of water quality. CLAY PARTICLES AND THEIR INTERACTIONS Clay is usually defined on a particle-size basis, the upper limit being 2,um diameter. Soils and sediments in nature will therefore have vaporing proportions of clay material containing clay-mineral components (usual- ly the phyllosilicates), as well as nonclay-mineral material that may include a variety of substances such as iron and aluminum oxides and hydroxides, quartz, amorphous silica, carbonates, and feldspar. The clay minerals themselves are classified in Table IV- 1 (Grim, 1968~. Clays are ubiquitous in soils and sediments derived from soils. They may be formed in soils during soil development through the weathering of various minerals, or they can be inherited essentially without change from the parent material upon which the soils are formed. Parent material, climate, topography, and vegetation determine the kinds of clays that are found. Hydrothermal activity may also lead to clay formation. As erosion acts on the landscape, clays may be suspended in water and carried until they are deposited by sedimentation. Most sedimentary rocks contain more or less clay as, for example, shales (almost exclusively clay), limestones, and sandstones. A number of scientific techniques are useful for studying clays, but the most useful for identification and indication of relative abundance is X- ray diffraction. The diffraction properties of the various clay minerals, as well as the methods of treatment and sample preparation, can be found in publications of Grim (1968), Brown (1961), and Whittig (1965~. Infrared spectroscopy is a valuable adjunct to X-ray diffraction in characterizing clays, and this subject has recently been reviewed by Farmer (1975~. Infrared spectroscopy is the most powerful method for study of organic- clay interactions (Mortland, 1970; Theng, 1975~. Other techniques useful in characterizing clays are electron microscopy (Gard, 1971), thermal methods (Mackenzie, 1957), and chemical analysis (Weaver and Pollard, 1973~. The layer-lattice clay minerals, in themselves, do not appear to have

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Solid Particles in Suspension 137 TABLE IV-1 Classification of the Clay Minerals (Grim, 1968) I. Amorphous Allophane group II. Crystalline A. Two-layer type (sheet structures composed of units of one layer of silica tetra- hedrons and one layer of alumina octahedrons) 1. Equidimensional Kaolinite group Kaolinite, nacr~te 2. Elongate Halloysite group B. Three-layer types (sheet structures composed of two layers of silica tetrahedrons and one central dioctahedral or tr~octahedral layer) 1. Expanding lattice a. Equidimensional Montmor~llonite group (smectite) Vermiculite Montmor~llonite, sauconite b. Elongate Montmor~llonite group Nontronite, saponite, hector~te 2. Nonexpanding lattice Illite group Regular mixed-layer types (ordered stacking of alternate layers of different types) Chlorite group D. Chain-structure types (hornblende-like chains of silica tetrahedrons linked to- "ether by octahedral groups of oxygens and hydroxyls containing Al and Mg atoms) Attapulgite Sepiolite Pal ygors kite deleterious effects when ingested by humans. Some of them are, in fact, constituents of pharmaceuticals such as kaopectate (kaolinite). Other indications (some from folklore) suggest beneficial results from ingestion of clays. The effect of ingestion of fibrous clay minerals of the chain- structure types (e.g., attapulgite, palygorskite, sepiolite, sometimes called "asbestos"), is still open to question and is the subject of extensive study at the present time. If layer-lattice clay minerals have deleterious e~ects on human health, they are probably indirect, through adsorption, transport, and release of inorganic and organic toxicants, bacteria, and viruses. Several reports have shown that concentrations of many pollutants are much higher in sediments of streams and lakes than in the waters with which they are associated. Clays and organic particulates are the

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138 DRINKING WATER AND HEALTH materials chiefly responsible for such concentrations. Since clays are ubiquitous in many waters used as sources for human consumption, it is to be expected they will appear as particulate matter in some drinking waters and thus it is of interest to consider the kinds of interactions they have with dissolved materials. Considerable knowledge exists regarding the surface chemistry and adsorptive properties of clays, and thus, with information on the nature of a solute, it is possible to have some idea of their interaction. Clays are very adsorptive substances. The possibility exists that clays could act as vehicles for transport of toxic compounds through adsorption in one environment, followed by release of the toxic material when the clay entered a different environment. It has been well established that some pesticides applied to watersheds can be adsorbed by soil components and subsequently removed into water by erosional processes (Bailey et al., 1974; Nicholson, 1969; Nicholson and Hill, 19701. Inorganic Pollutants This classification of pollutants would include metal cations and some anions. Among the metal cations that have been found to be polluting some water and soils are Pb, Cr. Cu. Zn, Co, Mn, Ni, Hg, and Cd, while radioactive isotopes of Pu, Cs, and Sr, among others, over potential threats as pollutants. On the other hand, anionic species such as phosphates, arsenate, borate, and nitrates are considered pollutants in some situations. The interactions of metal cations with clays include adsorption by ion exchange, precipitation as hydroxides or hydrous oxides on clay surfaces, and adsorption as complex species. ObviouslY. OH and Eh are critical 1 J ' AL factors In determining the nature of the interactions between clays and some transition and heavy metal ions. Hodgson's review (1963) includes some reference to earlier work on clay interactions with some of the transition ions and heavy metals. Jenne (1968) has electively described the various factors controlling the concentrations of transition cations in waters and soils, while Jenne and Wahlberg (1968) and Tamura (1962), among others, have considered the interaction of radionuclides with clays. Holdridge (1966) has reported adsorption studies of heavy metal cations on ball clay. With regard to phosphate, it is likely that its interactions with calcium ion and amorphous hydroxides of Fe3+ and A13+ and with allophane are more important than adsorption by clay minerals in affecting its concentration in natural waters. In addition to adsorption by simple ion exchange, much work indicates the retention of transition and heavy metals at clay mineral surfaces via

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Solid Particles in Suspension 139 precipitation of insoluble compounds, notably hydroxy and oxidehy- droxy polymers. The incorporation of A13+, Mg2+, and Fe3+ hydroxy polymers within the interlamellar space of swelling clays to form chlorite- like species is a well-known pedogenic process. It has also been shown that these brucite-gibbsite-like materials may often be withdrawn if the mineral is subjected to a different environment, usually one involving a change in pH. Gupta and Malik (1969) have reported the incorporation of Ni2+ in smectite to form a nickel-chlorite, while Blatter (1973) found similar reactions of smectite with Hg2+ . Thus it seems that many of these kinds of metal cations have the ability to form interlayer complexes in swelling clays. It would seem likely that in natural systems where the polluting species might be present in very low concentrations compared with other interlayer-forming species such as A13+, they might be incorporated within the gibbsite-like layer as it forms, in essentially isomorphous substitution for A13+, although reports of this phenomenon were not found. It would also appear that incorporation of a polluting metal cation within the intergrade clay is no guarantee that it might not again be released to the natural system when the clay, through erosion and deposition, is placed in a different environment where the interlayer material may be removed. This phenomenon has been shown for vermiculite-chlorite intergrades that, upon erosion from an acidic soil, are deposited in a calcareous freshwater lake or floodplain to form discrete vermiculite, within relatively short periods of time (Frink, 1969; Lietzke and Mortland, 1973~. The clay mineral vermiculite has a special affinity for K+ ion, the ion is initially adsorbed in the interlamellar regions of the mineral and then trapped by collapse of the layer structure. The ion is thus removed from direct interaction with the surrounding solution. This process is called potassium fixation, but will occur with other ions of similar diameter to more or less extent. Cations that might be considered pollutants that undergo this reaction with vermiculite are Ba2+ and radioactive Cs+ . in which The hydroxides and hydrous oxides of iron, manganese, and aluminum are often components of the clay fraction of sediments and have important e~ects on pollutant concentrations in natural waters. They often exist as coatings on the surfaces of other minerals and thus may exert chemical activity far out of proportion to their total concentrations. Jenne (1968) suggests that they furnish the principal control on the concentrations of heavy metals such as Co, Ni, Cu. and Zn through adsorption processes. The principal factors a~ecting adsorption and desorption of heavy metals from these kinds of particulates are pH, Eh, concentration of the metal in question, concentration of competing

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140 DRINKING WATER AND HEALTH metals, and the eRects of other adsorbents such as organic matter and clay minerals. Organic Pollutants These materials encompass a wide range of compounds, including pesticides, polychlorinated biphenyls, aromatic species of various kinds arising from industrial activity, and fluorine compounds in aerosols. Whether or not organic species adsorb or interact with clays depends upon the structure and properties of the compound and the nature of the clay and its exchangeable cations. Several mechanisms of interaction are possible and have been described in a number of recent reviews (Mortland, 1970; Bailey and White, 1970; Theng, 1974; and Rausell- Colom and Serratosa, 1975~. Organic cations adsorb on clays by ordinary ion exchange and are usually preferred over the inorganic ions by the exchange complex because of their large size and high molecular weights. Examples of organic compounds that are cationic and could be considered pollutants if transported outside their areas of application are the herbicides Paraquat and Diquat. These compounds are strong bases and are completely ionized in water. Other organic compounds, while neutral at the ambient pH of the solution phase, may become protonated after adsorption at the clay surface. The surface acidity of clays has been shown to be a considerably stronger proton donor system than pH measurements of the water-clay system would indicate. Thus, organic compounds containing basic nitrogen or carbony} groups may become protonated, and therefore cationic, after adsorption at clay surfaces. Another kind of organic-clay interaction is the coordination or ion- dipole type. Compounds with nitrogen, oxygen, sulfur, or olefinic groups have electron pairs that may be donated to electrophilic exchange cations to form complexes on the clay surface. In natural systems, an important consideration is the competitive eject of water for these adsorption sites. That is, the energy of ligand formation of an organic molecule with an exchange cation must be greater than the salvation energy of the cation in order to displace water molecules and obtain direct organic-cation coordination. In the laboratory these interactions are easily obtained by dehydration; however, in natural systems the competition of water is a major factor in determining whether or not these complexations occur. Nevertheless, it is likely that this kind of interaction does occur with some highly polar, electron-donating organic compounds. Another important factor is the nature of the exchange cation. Thus, for example, transition metal cations on the exchange complex, that have unfilled a'orbitals, will interact strongly with electron-supplying groups of organic molecules.

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Solid Particles in Suspension 141 Still another kind of organic-clay interaction is hydrogen bonding. These interactions can be classified into three types: 1. Hydrogen bonding between water molecules directly solvating exchangeable cations and polar functional groups, such as carbonyl, on organic molecules. The water molecules thus act as a "bridge" between the cation and organic species. 2. Hydrogen bonding between functional groups such as alcoholic and amino groups and oxygens of the silicate surfaces. Infrared spectroscopy has indicated that these are relatively weak bonds, being within the lower range of energies where hydrogen bonding is found. 3. Intermolecular hydrogen bonding between two organic species on the clay curface. Other factors involved in clay-organic interactions include physical forces and entropy effects. With this general description of the interactions of organic species with clays, it is now appropriate to mention some special clay-organic properties that have relevance to organic pollutants. Many organic compounds, including aromatics and particularly the halogenated types such as DDT, chlorinated and brominated phenyls and biphenyls, are adsorbed to little if any extent on clay surfaces from aqueous solution. In the natural environment they are more likely to be adsorbed in organic components of soils and sediments. These materials usually have limited solubility in water, since they are hydrophobic. It is thus not surprising that they are not attracted to the hydrophilic surfaces of clays. The above discussion, however, suggests that, in natural systems, clay-organic complexes may act as adsorbing media for some organic pollutants that are not adsorbed at all by pure inorganic clays. Another phenomenon that may take place when organic species are adsorbed at clay surfaces is that of catalytic alteration. This has particular relevance for organic pollutants since there is much interest in their fate in the environment. Much work has been reported on catalytic reactions on clays at high temperatures, but it is only recently that much attention has been paid to catalysis by clays in conditions resembling the natural environment. One mechanism by which clays can act as catalysts is via their Bronsted acidity. Examples of this are the hydrolysis of esters demonstrated by McAuli~e and Coleman (1955), the conversion of atrazine to hydroxyatrazine by Russell et al. (1968), the decomposition of alkylammonium ions by Chaussidon and Calvet (1965), and the hydrolysis of nitrites to amides by Sanchez et al. (1972~. In many decomposition reactions involving Bronsted acidity, carbonium ion formation is undoubtedly involved. On the other hand, Lewis acid sites

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142 DRINKING WATER AND HEALTH may exist in clays that also will catalyze many organic reactions. These sites (electron acceptors) may be part of the basic structure of the mineral itself as, for example, ferric iron within the octahedral layer or exposed aluminum on the edges of the minerals. In addition, some cations on exchange sites function in this capacity, particularly those of the transition metal group. Solomon et al. (1968) have demonstrated catalytic properties of Lewis sites located on edges of clay minerals. The activity of some transition metal cations on exchange sites has also been amply demonstrated as, for example, the decomposition of urea to ammonium ion when complexed with Cu2+, Mn2+, or Ni2+ smectite. No such reaction was noted for urea complexed with alkali metal or alkaline earth-saturated clay (Mortland, 1966~. Aromatic molecules such as benzene will complex via pi electrons with clay minerals saturated with Cu2+, under mildly desiccating conditions. Under more vigorous dehydrating conditions, a radical cation of benzene is formed that will react with molecular benzene to give polymers containing phenyl groups as well as fragmented benzene rings (Doner and Mortland, 1969~. Anisole (methoxybenzene) will also form radical cations that react with molecular anisole to give 4,4'-dimethoxybiphenyl (Fenn et al., 1973~. Other cationic species with oxidizing abilities as great as Cu2+, such as Vo2+ and Fe3+, were also found to produce radical cations from some aromatic species with subsequent polymer formation (Pinnavaia et al., 1974~. These reactions suggest the possibility that some pollutant species adsorbed on clay surfaces may undergo similar reactions to form radical cations and subsequently interact with themselves, or other organic compounds with formation of different chemical derivatives. Thus, pollutant degradation or alteration on clays by oxidation-reduction reactions involving ex- changeable transition-metal cations may be a real possibility in nature. In addition to the degradation of atrazine to hydroxyatrazine, mentioned above, a number of other clay-catalyzed pesticide reactions have been reported. For example, Fleck and Halter (1945) report the conversion of DDT to DDE by kaolinite and smectite samples, preheated to 400K. Also, degradation of heptachlor by palygorskite has been suggested by Malina et al. (1956~. The degradation of the organic phosphate insecticide, ronnel, by clays heated to various temperatures has been reported by Rosenfield and van Valkenburg (1965~. Organic phosphate pesticides have been observed by Mortland and Raman (1967) to be hydrolyzed in the presence of Cu2+-montmorillonite by a coordination mechanism. The much weaker catalytic e~ects of Cu2+- vermiculite, beidellite, and nontronite were attributed to reduced activity of the copper on these minerals, as compared with montmorillonite, due to charge location. While most of the degradation of pesticides in nature

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Solid Particles in Suspension 143 has been attributed to biological agencies, the above discussion would suggest that catalysis at mineral surfaces may also play a role. Natural organic material in soils forms complexes with clays that exert important influences on the physical, chemical, and biological properties of the soil (Greenland, 1965, 1971~. Since the exact chemical and physical nature of these organic materials is not known, the kinds of interaction they have with clays are less well known than those of well-defined organic compounds. However, some of the kinds of reactions described above are probably involved. It is obvious that clays eroded from soil surfaces into streams and lakes will probably be, to some degree, complexed with organic matter. Humic acids, a constituent of soil organic matter, may be strongly adsorbed by clays, presumably by interaction with positive sites on the edges of clay particles or with polyvalent cations on the cation exchange complex acting as "bridges." Schnitzer and Kodama (1967) have shown that fulvic acid (another constituent of soil organic matter) adsorption depends on pH, and is greater under acid than alkaline conditions. This is to be expected, since the fulvic acid would be relatively undissociated at low pH but considerably more anionic at alkaline pH. Schnitzer and Kodama (1972) showed that fulvic acid is very strongly bound to Cu2+ on the exchange sites of montmorillonite through a coordination type of reaction. In addition, they have shown such adsorption is typical for any electrophilic cation on the exchange complex, particularly for ions of the transition metal group. Summary Pollutant concentrations are higher in sediments than in the waters with which they are associated. It should be recognized that the consequences of pollutant adsorption by clays may be very important in natural systems and may affect drinking water quality. Clay-pollutant complexes may be mobilized by erosion from the landscape, or form when eroded clay enters a stream containing a polluting species. If the complex survives water treatment and enters the drinking water system, it would then be available for ingestion by humans. In the adsorbed state on the clay surface the pollutant is probably not toxic, but the possibility exists that the pollutant might be released from the clay in the environment of the alimentary tract and thus exert toxic effects. Whether or not such a process might take place would depend on the complex in question, so that no generalities are possible. Information is completely lacking in this area, and thus research should be encouraged and supported.

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144 DRINKING WATER AND HEALTH ASBESTOS: NOMENCLATURE, OCCURRENCE AND REDISTRIBUTION IN WATER Structure and Nomenclature Asbestos is the name for a group of naturally occurring hydrated silicate minerals possessing fibrous morphology and commercial utility. This definition generally limits application of the term to the minerals chrysotile, some members of the cummingtonite-grunerite series, crocido- lite, anthophyllite, and some members of the tremolite-actinolite series. Amosite is commonly used to refer to a cummingtonite-grunerite asbestos mineral, but it is a discredited mineral name (Rabbit 1948; Committee on Mineral Names, 1949~. Mode of occurrence and fiber length are important determinants of commercial value. Of the commercially mined and processed asbestos minerals, chrysotile accounts for about 95%, the remainder being amosite and crocidolite (May and Lewis, 1970~. Crocidolite is the fibrous equivalent of riebeckite, and chrysotile belongs to the serpentine group of minerals, which contains other nonfibrous members (Deer et al., 1970~. Noncommercial deposits of asbestos minerals are also relatively com mon. The standard definitions of the Glossary of Geology (American Geological Institute, 1972; second printing, 1973) are given below. ASBESTOS: (a) a commercial term applied to a group of highly fibrous silicate minerals that readily separate into long, thin, strong fibers of sufficient flexibility to be woven, are heat resistant and chemically inert, and possess a high electric insulation, and therefore are suitable for uses (as in yarn, cloth, paper, paint, brake linings, tiles, insulation cement, fillers, and filters), where incombustible, nonconducting, or chemically resistant material is required. (b) a mineral of the asbestos group, principally chrysotile (best adapted for spinning) and certain fibrous varieties of amphibole (ex. tremolite, actinolite, and crocidolite). (c) a term strictly applied to the fibrous variety of actinolite. Syn: asbestos; amianthus; earth flax; mountain leather. ASBESTIFORM: Said of a mineral that is fibrous, i.e. that is like asbestos. ACICULAR (Cryst): Said of a crystal that is needlelike in form. Of: fascicular, sagenitic FIBROUS: Said of the habit of a mineral, and of the mineral itself (e.g. asbestos), that crystallizes in elongated thin, needle-like grains, or fibers. The nomenclature used in this report conforms generally to these definitions, subject only to the further qualifications that the term asbestos will not be used in its most restrictive sense (c, above);

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Solid Particles in Suspension 145 asbestiform will not be used; and the terms acicular and fibrous are to be understood as discussed below. The term asbestos has often been used in recent scientific literature to describe individual fibrous or acicular particles of microscopic and submicroscopic size. However, mineralogists and geologists have has- tened to point out that the term should be used only as defined above, in reference to the minerals in bulk. Ampian (1976) considers the terms asbestos and asbestiform to be synonymous and that they then may only be used to apply to the bulb fibrous forms occurring in nature. Asbestiform is often used to define the morphology of a mineral that is similar to asbestos, but does not necessarily occur in nature in a commercial deposit; to avoid ambiguity, the term will not be used here. The terms acicular and fibrous are used here to characterize any mineral particle that has apparent crystal continuity, a length-to-width aspect ratio of 3 or more and widths in the micrometer or submicrometer range. Although the two terms are not strictly synonymous, the use here of either one to describe a mineral particle should be taken to imply the other, unless otherwise qualified. Table IV-2 lists some of the naturally occurring minerals that can have, but do not always have, an acicular morphology. To this list could be added a number of synthetic fibers, although they are not naturally occurring minerals. Many of the minerals in Table IV-2 are common rock-forming minerals. Properties of Asbestos Minerals MINERALOGY The asbestos minerals belong to the serpentine and amphibole groups, and the amphiboles are further divided into those of the orthorhombic crystal system (orthoamphiboles) and amphiboles of the monoclinic crystal system (clinoamphiboles). Table IV-3 summarizes the basic properties of the asbestos minerals. Chrysotile is the asbestos mineral of the serpentine group. Its crystal structure is a double sheet, comprising a layer of silica tetrahedra and a layer of magnesia octahedra, arranged in a manner that is somewhat analogous to the alumina octahedra silica tetrahedra layering of kaolinite. The way in which the sheet structure is modified to develop a fibrous morphology is, in detail, very complex; but in essence the modification can be imagined as a buckling of the double sheet, due to misfits, to form a hollow tube (Deer et al., 1966~. This central tube may or may not be filled with electron-opaque material, and the appearance of its

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194 DRINKING WATER AND HEALTH Committee on Mineral Names. 1949. Am. Min. 34:339. Cotterell, K., and P.F. Halt. 1972. An examination of crocidolites from North West Cape and Transvaal Mines. Inst. Min. Metall. Trans. Sect. B. 81:169-171. Cralley, LJ., R.G. Keenan, J.R. Lynch, and W.S. Lainhart. 1968. Source and identification of respirable fibers. J. Am. Ind. Hyg. Assoc. 29:129-135. Cunningham, H.M. and R Pontefract. 1971. Asbestos fibers in beverages and drinking water. Nature 232:332-333. Davis, W.E. 1970. National inventory of sources and emissions of asbestos. NTIS PB 192252. Deer, W.A., R.A. Howie, and J. Zussman. 1966. An Introduction to the Rock-forming Minerals. Longmans, London. Durham, R.W., and T. Pang. 1976. Asbestiform fiber levels in Lakes Superior and Huron. Scientific Series no. 67. Inland Waters Directorate, Canada Centre for Inland Waters, Burlington, Ontario. Environment Canada. 1973. National inventory of sources and emissions of asbestos. Report APCD 73-4, Air Pollution Control Directorate, Environment Canada, Ottawa. EPA. 1975. Preliminary assessment of suspected carcinogens in drinking water. Interim Report to Congress. Environmental Protection Agency. Flentje, M.E., and RJ. Schweitzer. 1955. Further study of solution effects on concrete and cement in pipe. J. Am. Water Works Assoc. 49:1441. Hartman, P. 1963. Structure, growth and morphology of crystals. Z. Kristallog. 119:65-78. House, R.F. 1967. Dispersion of asbestos. U.S. Pat. Office no. 3,586, 639. Hutchinson, J.L., M.C. Irusteta, and E.J.W. Whittaker. 1975. High-resolution electron diffraction studies of fibrous amphiboles. Acta Crystallog. 31:794-801. Kay, G. 1973. Ontario intensifies search for asbestos in dunking water. J. Water Pollut. Control Fed:33-35. Kehieker, D.M. et al. 1967. Determination of elementary fiber size of chrysotile asbestos. Sov. Phys. Crystallog. 12:430-435. Kramer, J.R. 1976. Fibrous c~mmmgtonite in Lake Superior. Can. Min. 14:91-98. Kristiansen, H. 1974. The extraction of calcium by soft water from prestressed concrete pipes. Vatten, 1:70. May, T.C., and R.W. Lewis. 1970. Asbestos. In Mineral Facts and Problems. U.S. Bureau Mines Bulletin 650:851-865. NAS-NRC. 1971. Airborne asbestos. Committee on Biologic Effects of Atmospheric Pollutants. National Research Council. (See also Environment Canada, 1973.) Parks, G.A. 1967. Aqueous surface chemistry of oxides and complex oxide minerals. In W. Stuum, ed. Equilibrium Concepts in Natural Water Systems, Adv. Chem. no. 67, Am. Chem. Soc. Rabbit, J.C. 1948. Am. Min. 33:263-323. Rendall, R.E.G. 1970. The data sheets on the chemical and physical properties of the UICC standard reference samples. In H.A. Shapiro, ed. Pneumoconiosis. Oxford University Press. Ruud, C.O., C.S. Barrett, P.A. Russell and R.L. Clark. 1976. Selected area electron diffraction and energy dispersive X-ray analysis for the identification of asbestos fibers, a comparison. Micron 7: 1 15-132. Timbrell, V. 1970. Characteristics of the International Union Against Cancer Standard Reference Samples of Asbestos. In H.A. Shapiro, ed. Pneumoconiosis. Oxford University Press. Timbrell, V., F. Pooley, and J.C. Wagner. 1970. Characteristics of respirable asbestos fibers. In H.A. Shapiro, ed. Pneumoconiosis. Oxford University Press.

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Solid Particles in Suspension 195 Timbrell, V., and R.E.G. Rendall. 1972. Preparation of the UICC reference samples of asbestos. Powder Tec. 5:279-287. Whittaker, E.J.W. 1966. Diffraction contrast in electron microscopy of chrysotile. Acta. Crystallog. 21:461-466. Whittaker, E.J.W. and J. Zussman. 1971. The Serpentine Minerals. In J.A. Gard, ed. Electron Optical Investigations of Clays. The Mineralogical Society, London. Yada, K. 1964. Study of chrysotile asbestos by a high resolution electron microscope. Acta Crystallog. 23:704707. Zoltai, Tibor, and J.H. Stout. 1976. Comments on asbestiform and fibrous mineral fragments, relative to Reserve Mining Company taconite deposits. Report to Minn. Pollution Control Agency, Minneapolis. Wright, G.W. 1974. Does the use of asbestos-cement pipe for potable water systems cause a health hazard? J. Am. Water Works Assoc. 66:4-22. Asbestos Fiber Sampling and Analysis Beaman, D.R., and D.M. File. 1975. The quantitative determination of asbestos fiber concentrations. The Dow Chemical Company, unpublished report. Berkley, D., J. Churg, I.J. Selikoff, and W.E. Smith. 1965. The detection and localization of asbestos fibers in tissue. Ann. N.Y. Acad. Sci. 132:48-63. Berkley, C., A.M. Langer, and V. Baden. 1968. Instrumental analysis of inspired fibrous pulmonary particles. Trans. N.Y. Acad. Sci. 30:331-350. Birks, L.S., M. Fatemi, J.V. Gilfrich, and E.T. Johnson. 1975. Quantitative analysis of airborne asbestos by x-ray diffraction. Feasibility Study AD-A007530, Naval Res. Lab., Washington, D.C. Brown, A.L., Jr., W.F. Taylor, and R.E. Carter. 1976. The reliability of measures of amphibole fiber concentration in water. Environ. Res. 12:150-160. Clark, R.L., and C.O. Ruud. 1975. Transmission electron microscopy standards for asbestos. Micron 5:270. Cook, P.M., J.B. Rubin, C.J. Maggiore, and W.J. Nicholson. 1974. X-ray diffraction and electron beam analysis of asbestiform minerals in Lake Superior waters. In Trans. Inst. Electrical Electronic Eng. (in press). Crable, J.V., and M.J. Knott. 1966a. Application of x-ray diffraction to the determination of chrysotile in bulk and settled dust samples. Am. Ind. Hyg. J. 27:383-387. Crable, J.V., and M.J. Knott. 1966b. Quantitative x-ray diffraction analysis of crocidolite and amosite in bulk or settled dust samples. Am. Ind. Hyg. J. 27:449-453. Dement, J.M., R.D. Zumwalde, and K.M. Wallingford. 1975. Asbestos fiber exposures in a hard rock gold mine. In Proc. N.Y. Acad. Sci. Conf. Occup. Carcinogenesis. Ann. N.Y. Acad. Sci. 271:345-352 (1976). Ferrell, R.E., G.G. Paulson, and C.W. WaLlcer. 1975. Evaluation of an SEMEDS method for identification of chrysotile. Scanning Electron Microscopy:537-546. Julian, Y., and W.C. McCrone. 1970. Identification of asbestos fibers by microscopical dispersion staining. Microscope 18:1-10. Keenan, R.G., and J.R. Lynch. 1970. Techniques for the detection, identification and analysis of fibers. Am. Ind. Hyg. J. 31:587-597. Langer, A.M., A.D. Mackler, and F.D. Pooley. 1974. Electron microscopical investigation of asbestos fibers. Environ. Health Perspect. 9:63-80. Langer, A.M., I. Rubin, and I.J. Selikoff. 1975. Electron microprobe analysis of asbestos bodies. Histochem. Cytochem. J. 20:735-740.

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