Basic Soil Chemistry

Soils are composed roughly equally of solid particles, about 90% of which are inorganic in nature and the rest organic matter, and of pore space, about half of which is air and half water. The inorganic particles are residues of weathered rock; chemically they are mainly silicate minerals. At the atomic level, these minerals consist of polymeric inorganic structures in which the fundamental unit is a silicon atom surrounded tetrahedrally by four oxygen atoms. Since these oxygen atoms are in turn each bonded to another silicon, etc., the resulting structure is an extended network. There are many variations on the silicate structural theme. Some networks have exactly twice as many oxygens (formally, 02~) as silicons (formally, Si4+) and correspond to electrically neutral Si02 polymers. In others, some of the tetrahedral sites are occupied by aluminum ions, Al3+, instead of Si4+; the extra negative charge in these networks is neutralized by the presence of other cations such as H+, Na+, K", Mg2+, Ca2_r, and Fe2+. Some common silicon-oxygen structural units are illustrated in Figure 16-7.

Over time, the weathering of the silicate minerals from rocks can involve chemical reactions with water and acids in which ion substitution occurs. Eventually, these reactions yield substances that are important examples of the class of soil materials known as clay minerals. A mineral having a particle size less than about 2 fxm is by definition a component of the clay fraction of soil.

In addition to clay, there are several other soil types; the definition of each type depends on particle size, as indicated in Figure 16-8. Notice the factor-of-10 increase in size with each transition in type: The upper boundary

Natalie Darrin

FIGURE 16-7 The common structural units in silicate minerals. Dark circles represent silicon atoms; open circles represent oxygens. [Source: R. W. Raiswell, P. Brimblecombe, D. L. Dent, and P. S. Liss, Environmental Chemistry (London: Edward Arnold Publishers, I960).]






20 fj,m


FIGURE 16-8 The soil particle size classification system of the International Society of Soil Science. [Source: G. W. vanloon and S. J. Duffy, Environmental Chemiitry (Oxford: Oxford University Press, 2000).]

for silt is 10 times that for clay, that for fine sand is 10 times that for silt, etc. Because the particle size of sand is large, it has a relatively low density and water runs through it easily. In contrast, soils composed of clay are dense and have poor drainage and aeration, since the clay particles form a sticky mass when wet, in contrast to sand and silt particles, which do not stick to each other. The best agricultural soils consist of a combination of soil types. The range of element content for major and minor elements in the mineral component of soil is given in Table 16-2.

Clay particles act as colloids in water. Because the clay particles are much smaller than those of sand or silt, their total surface area per gram is thousands of times larger. Consequently, most important processes in soil occur on the surface of colloidal clay particles.

Particles of clay possess an outer layer of cations that are bound electrostatically to an electrically charged inner layer, as illustrated in Figure 16-9. The most common cations in soil are FP, Na~, K+, Mg2+, and Ca2+. Depending on the concentration of cations in the water surrounding the clay

TABLE 16-2

Element Content of the Mineral Components of Soils

Major Elements (%)

Minor Elements (ppm)

































Source: G. W. vanLoon and S. J. Duffy, Environmental Chemistry (Oxford: Oxford University Press, 2000).

Source: G. W. vanLoon and S. J. Duffy, Environmental Chemistry (Oxford: Oxford University Press, 2000).

+ Ca2+, Mg2"1", H Na+, and CI"

FIGURE 16-9 Ion-exchange equilibria on the surface of a clay particle. The addition of K+ ions to the soil water displaces the exchange equilibria to the right, whereas removal of Kions from solution displaces it to the left. [Source: R.VV. Raiswell, P. Brimblecombe, D. L. Dent, and P. S. Liss, Environmental Chemistry (London: Edward Arnold Publishers, 1980).t

+ Ca2+, Mg2"1", H Na+, and CI"

FIGURE 16-9 Ion-exchange equilibria on the surface of a clay particle. The addition of K+ ions to the soil water displaces the exchange equilibria to the right, whereas removal of Kions from solution displaces it to the left. [Source: R.VV. Raiswell, P. Brimblecombe, D. L. Dent, and P. S. Liss, Environmental Chemistry (London: Edward Arnold Publishers, 1980).t particle, the cations on the particle are capable of being exchanged for them. For example, in water rich in potassium ions but poor in other ions, K+ ions will displace the ions bound to the surface of the clay particle (see Figure 16-9). If, on the other hand, the soil is acidic—i.e., rich in H+ ions—the metal ions on the surface will be displaced by H+ ions and the previously bound metal ions will enter the aqueous phase. Generally, the greater the positive charge on a cation, the more strongly it binds to the particle. Heavy metals dissolved in soil water are often bound to the surface of clay particles.

In addition to minerals, the other important components of soil are organic matter, water, and air. The proportion of each component varies greatly from one soil type to another The organic matter (1-6%), which gives soil its dark color, is primarily a material called humus. Humus is derived principally from phoiosynthetic plants, some components of which (such as cellulose and hemicellulose) have previously been decomposed by organisms that live in the soil. The undecomposed plant material in humus is mainly protein and lignin, both polymeric substances that are largely insoluble in water. A significant amount of the carbon in lignin exists in the form of six-membered aromatic benzene rings connected by chains of carbon and oxygen atoms (see Figure 11-4). Much of the organic matter in soil also consists of colloidal particles.

As a result of the partial oxidation of some of the lignin, many of the resulting polymeric strands contain carboxylic acid groups, —COOH. This dark-colored portion of humus consists of humic and fulvic acids and is soluble in alkaline solutions due to the presence of the acid groups. By definition, humic acid is insoluble in acid solution, whereas fulvic acid is soluble. The humic acid is less soluble in acid than the fulvic not only because its molecular weight is much greater (100 to 1000 times greater), but also because its oxygen content is lower, so there are fewer — OH groups per carbon to form hydrogen bonds with the water. The acid groups are often adsorbed onto the surfaces of the clay minerals, to an extent dependent on the distribution of surface charge on the particles. Humic and fulvic acids form colloids that are hydrophilic, whereas those of clay are hydrophobic. Generally speaking, nonpolar organic molecules are more strongly attracted to the organic matter in soil than to the surfaces of particles derived from minerals.

Since some of the carbon in the original plant material has been transformed to carbon dioxide and thus lost as a gas to the surroundings, humus has relatively more nitrogen than the original plants; its other main components are carbon, oxygen, and hydrogen. Due to decomposition processes occurring in the organic component of soil, the 02 content of soil air is often only 5-10% rather than 20%, and its CO> concentration is often several hundred times that in the atmosphere.


The percentage composition of a typical fulvic acid is 50.7% carbon by mass, 45.1% oxygen, and 4-22% hydrogen. Derive the (simplest) empirical formula for the substance.

The Acidity and Cation-Exchange Capacity of Soil

If the soil at the surface contains minerals with elements in a reduced state, their oxidation by atmospheric oxygen can produce acid. An example is the oxidation of sulfur in pyrite, discussed as acid mine drainage in Chapter 13. Acid rain, of course, provides another source of acidity in certain areas (Chapter 4).

Quantitatively, the ability to exchange cations is expressed as the soil's cation-exchange capacity, CEC, which is defined as the quantity of cations that are reversibly adsorbed per unit mass of the (dry) material. The quantity of cations is given as the number of moles of positive charge (usually expressed as centimoles or millimoles), and the soil mass is usually taken to be 100 g or 1.00 kg. Typical values of the CEC for common clay minerals range from 1 to 150 centimoles per kilogram (cmol/kg). The CEC values are determined in large part by the surface area per gram of the mineral. CEC values for the organic component of soil are high, due to the large number of —COOH groups that can bind to and exchange cations; e.g., the CEC of peat can be as high as 400 cmol/kg.


The CEC for a soil sample is found to be 20 cmol/kg. What is the CEC value for this sample in units of millimoles per 100 grams?

Biologically, the exchange of cations by soils is the mechanism by which the roots of plants take up metal ions such as potassium, calcium, and magnesium. Although the roots release hydrogen ions to the soil in exchange for the metal ions, this is not the main reason why soil in which plants grow is often somewhat acidic. Most of the acidity is due to metabolic processes involving the roots and microorganisms in the soil, which result in the production of carbonic acid, H2CQ3, and of weak organic acids.

Rainwater that is acidic releases base cations from soil particles by exchanging them for H+ ions. The acidity of water flowing through the soil stays low for this reason. However, once the base cations have been exhausted, aluminum ions are released, as discussed in Chapter 4. It is now known that in the past, base cations from dust particles, especially those containing calcium and magnesium carbonates, neutralized some of the acidity in precipitation; but along with sulfur dioxide, their emission from industrial sources has been curtailed in recent decades.

The pH of soil can vary over a significant range for a variety of reasons. For example, soils in areas of low rainfall but high concentrations of the soluble salt sodium carbonate, Na2C03, become alkaline due to the (hydrolysis) reaction of the carbonate ion, CO,2". with water, as discussed in Chapter 13.

Soils that are too alkaline for agricultural purposes can be remediated either by the addition of elemental sulfur, which releases hydrogen ions as it is oxidized by bacteria to sulfate ion, or by the addition of the soluble sulfate salt of a metal, e.g., iron(III) or aluminum, which reacts with the soil's water to extract hydroxide ions and thereby release hydrogen ions:

2 S(s) + 3 02 + 2 HzO-> 4 H+ + 2 S042~

Fe3+ + 3 H;20-> Fe(OH)3(s) + 3 H+

The pH of water present in the soil is determined by the concentration of hydrogen and hydroxide ions. However, soil has reserve acidity due to the large number of hydrogen atoms in the —COOH and —OH groups in the organic fraction and on the cation-exchange sites on minerals that are occupied by H+ ions. In other words, soils act as weak acids, retaining their H+ ions in a bound condition until acted upon by bases. Thus the pH of soil tends to be buffered against large increases in pH, since this bound hydrogen ion can be slowly released into the aqueous phase. In the process of liming, which is the addition of salts such as calcium carbonate to soil, carbonate ions neutralize acids present in the uppermost soil zones, producing carbon dioxide and water. Once this process has occurred, calcium ions can replace hydrogen ions in the organic matter or clays. The additional carbonate ions that enter the aqueous phase combine with the newly released H+ ions, again to produce weakly acidic carbonic acid, which dissociates into carbon dioxide gas and water. Thus liming is a procedure by which the pH of a soil can be raised somewhat, and it is the practical method by which acidic soils can be remediated.

Soil Salinity

In hot, dry climates, salts and alkalinity tend to accumulate in soil since there is little rainfall to leach ions from it. In contrast to other climates, the net movement of water in arid climates is upward rather than downward in the soil: Water evaporation and loss by transpiration of plants exceeds rainfall. The salts that accompany the upward migration of water remain at or near the surface when the water has escaped. Salt accumulation at the surface also occurs in semiarid regions due to the use of poor-quality irrigation water, whose salt content remains after the water has evaporated.

Ions are also liberated at the surface of soil in the weathering of otherwise insoluble minerals. A simple example is the reaction of olivine:

Mg2Si04 + 4 HzO-> Si(OH)4 + 2 Mg2+ + 4 OH"

Additional hydroxide ion is produced when the silicate ion, Si044-, produced by mineral weathering reacts as a strong base with water.

As a general rule, hydrolysis (reaction with water) of silicate minerals at the surface produces cations and hydroxide ions. In nonarid climates, the hydroxide is neutralized by acids that are naturally produced in the soil (see later section), but this does not occur in arid regions. There is very little organic matter in the soils of arid areas. The hydroxide reacts with atmospheric carbon dioxide that dissolves in water to produce bicarbonate and carbonate ions:

HC03~-> H+ + C032"

Consequently, bicarbonate and carbonate salts accumulate in arid soils. If the predominant cations in the soil are calcium and magnesium, most of the carbonate ion will be locked away as their insoluble carbonate salts. However, if the predominant cations are sodium and potassium, the soil when moist will have a high pH, since the carbonate salts of these ions are soluble and the free C032- will act as a base:

C032" + HzO ^^ HC03" + OH"

Increasing soil salinity is a major problem in Australia, especially in regions where wheat and other shallow-rooted crops have replaced natural, long-rooted vegetation. This replacement, plus irrigation of crops including rice and cotton, has resulted in a rise of the water table and, with it, the salt that was formerly deep in the soil.


Sediments are the layers of mineral and organic particles, often fine-grained, that are found at the bottom of natural water bodies such as lakes, rivers, and

FIGURE 16-10 U.S. watersheds where contaminated sediments may pose environmental risks. [Source: U.S. EPA, in B. Hileman, "EPA Finds 7% of Watersheds Have Polluted Sediments/' Chemical and Engineering News ■26 January 1998): 27.]

FIGURE 16-10 U.S. watersheds where contaminated sediments may pose environmental risks. [Source: U.S. EPA, in B. Hileman, "EPA Finds 7% of Watersheds Have Polluted Sediments/' Chemical and Engineering News ■26 January 1998): 27.]

Images Soil Chemistry

oceans. The ratio of minerals to organic matter in sediments varies considerably, depending on location. Sediments are of great environmental importance because they are sinks for many chemicals, especially heavy metals and organic compounds such as PAHs and pesticides, from which they can be transferred to organisms that inhabit this region. Thus the protection of sediment quality is a component of overall water management.

The map in Figure 16-10 shows watersheds in the continental United States where sediments are sufficiently contaminated to pose environmental risks. According to a U.S. EPA report, 7% of all watersheds pose a risk to people who eat fish from them and to the fish and wildlife themselves. The two pollutants found at high levels most frequently at contaminated sites are PCBs and mercury, although DDT (and its metabolites) and PAHs were also found at high concentrations at many sampling stations.

The transfer of hydrophobic organic pollutants to organisms may proceed by intermediate transfer to pore water, which is the water present in the microscopic pores that exist within the sediment material. Organic chemicals equilibrate between being adsorbed on the solid particles and being dissolved in the pore water. For this reason, pore water is often tested for toxicity in determining sediment contamination levels.

The Binding of Heavy Metals to Soils and Sediments

The ultimate sink for heavy metals, and for many toxic organic compounds as well, is deposition and burial in soils and sediments. Heavy metals often accumulate in the top layer of the soil and are therefore accessible for uptake by the roots of crops. For these reasons, it is important to know the nature of these systems and how they function.

Humic materials have a great affinity for heavy-metal cations and extract them from the water that passes through by the process of ion exchange. The binding of metal cations occurs largely by the formation of complexes with the metal ions by —COOH groups in the humic and fulvic acids. For example, for fulvic acids the most important interactions probably involve a —COOH group and an —OH group on adjacent carbons of a benzene ring in the polymer structure, where the heavy-metal Mz+ ion replaces two H+ ions:


I M = heavy metal

Humic acids normally yield water-insoluble complexes, whereas those of smaller fulvic acids are water-soluble.


Draw the structure that would be expected if a dipositive metal ion M2+ were to be bound to two —COO~~ groups on adjacent carbons in a benzene ring.

Heavy metals (Chapter 15) are retained by soil in three ways:

• by adsorption onto the surfaces of mineral particles,

• by complexation by humic substances in organic particles, and

• by precipitation reactions.

The precipitation processes for mercury and cadmium ions involve the formation of the insoluble sulfides HgS and CdS when the free ions in solution encounter sulfide ion, S2~. Significant concentrations of aqueous sulfide ion occur near lake bottoms in summer months when the water is usually oxygen-depleted, as discussed in Chapter 13. However, the total concentration of mercury in soil water can exceed the limits set by the solubility product of HgS because some of the mercury will take the form of the moderately soluble molecular compound Hg(OH)2 and does not participate in the equilibrium with the sulfide.

In acidic soils the concentration of Cd2+ can be substantial, since this ion adsorbs only weakly onto clays and other particulate materials. Above a pH of 7, however, Cd2_r precipitates as the sulfide, carbonate, or phosphate, since the concentration of these ions increases with increasing hydroxide ion levels. Thus the liming of soil to increase its pH is an effective way of tying up cadmium ion and thereby preventing its uptake by plants.

Like many other chemicals, heavy-metal ions are often adsorbed onto the surfaces of particulates, especially organic ones that are suspended in water, rather than simply being dissolved in water as free ions or as complexes with soluble biomolecules such as fulvic acids. The particles eventually settle to the bottom of lakes and are buried when other sediments accumulate on top of them. This burial represents an important sink for many water pollutants and is a mechanism by which the water is cleansed. Before they are covered by subsequent layers of sediments, however, freshly deposited matter at the bottom of a body of water can recontaminate the water above it by desorption of the chemicals, since adsorption and desorption establish an equilibrium. Furthermore, the adsorbed pollutants can enter the food web if the particles are consumed by bottom-growing and -feeding organisms.

Just as the total concentration of organic material in sediments may not be a good measure of the amounts that are biologically available, the same is true for the levels of heavy-metal ions present. Different sediments with the same total concentration of the ions of a heavy metal can vary by a factor of at least 10 in terms of the toxicities to organisms arising from the metal. This variation occurs principally because sulfides in the sediments control the availability of the metals. If the concentration of sulfide ions exceeds the total of that of the metals, virtually all the metal ions will be tied up as insoluble sulfide salts such as HgS, CdS, etc. and will be unavailable biologically at normal pH values. However, if the sulfide concentration is less than that of the metals, the difference is, biologically available. The sulfide ion that is available to complex with metals is the amount that will dissolve in cold aqueous acid and is termed acid volatile sulfide, AVS. Industrially polluted sediments may have AVS concentrations of hundreds of micromoles of sulfur per gram, whereas uncontaminated sediments from oxidizing environments can have values as low as 0.01 jimol/g.

Although mercury in the form of 1 lg"+ is firmly bound to sediments and does not readily redissolve into water, environmental problems have arisen in several bodies of water due to the conversion of the metal into methyl-mercury and its subsequent release into the aquatic food web. The overall cycling of mercury species among air, water, and sediments was illustrated in Figure 15-1,

As previously discussed, anaerobic bacteria methylate the mercuric ion to form Hg{CH3)2 and CH3HgX, which then rapidly desorb from sediment particles and dissolve in water, thereby entering the food web. Although the levels of methylmercury dissolved in water can be extremely low {of the order of hundredths of a part per trillion), a biomagnification factor of 108 results in ppm-range concentrations in the flesh of some fish. The devastating

FIGURE 16-11 Lead and mercury concentrations in the sediments of Halifax Harbor versus depth (and therefore year of deposit). [Source: D. E. Buckley, "Environmental Geochemistry in Halifax Harbour," WAT on Earth (1992): 5.|

consequences of methylmercury poisoning have already been described (Chapter 15).

Excavation and analysis of the sediments at the bottom of a body of water can yield a historical record of contamination by various substances. For example, the curves in Figure 16-11 show the levels of mercury and of lead in the sediments of the harbor in Halifax, Nova Scotia, as a function of depth and therefore of year. For decades raw sewage has been dumped into this harbor; consequently, its sediments are a historical record of the levels of pollutants in sewage. The metal pollution peaked about 1970, having begun to increase dramatically about 1900. These trends are also typical of heavy-metal levels in other water bodies, such as the Great Lakes; the characteristic decreases of the past few decades are due to the imposition of pollution controls.

In summary, then, both soils and sediments act as vast sinks and reservoirs in the containment of heavy metals.

Mine Tailings

In modern times, many minerals (and in some cases fossil fuels) are extracted from much, much larger quantities of rock (or sand, etc.) than in the past, since the remaining supplies of the minerals occur in dilute form. This practice produces huge quantities of unwanted crushed rock in the

Lead (ppm)

Lead (ppm)




1960 1940 1920 1900


1840 1820

Mercury (ppm)

form of dry, coarse-grained waste that must be disposed of, usually in slag heaps or landfills close to the mine. Eventually these waste piles are covered with soil and vegetation.

A more important environmental problem arises from disposal of the tailings of mining processing—fine-grained slurries that are more mineral-rich and often contain chemicals such as cyanide (Chapter 14) that were used to extract or process the ore. The toxic components of tailings are a potential source of pollution to local surface water, groundwater, and soil.

To prevent their dispersal into the environment, tailings are usually deposited as slurries in dams constructed on-site for the purpose. To prevent leakage into the soil, a clay or geomembranic liner is usually incorporated at the dam's base. Over time, the solid settles to the bottom of the dam and the water evaporates or is drained off, i.e., the tailings become dewatered, although this can take a long time to achieve. In some cases, toxic materials such as heavy metals, nitrates, or excess acidity is removed by treatment of the tailings. Eventually, to establish a vegetated cover, organic matter and fertilizer must be added, since the dried tailings themselves contain little or no organic matter and consequently are sterile as well as hostile to plants.

If the crushed rock or tailings contain iron pyrites, exposure to oxygen will produce sulfuric acid by the series of reactions discussed as acid mine drainage in Chapter 13. The acidity can be neutralized by the continuous addition of limestone.

The most important environmental problem associated with a tailings dam is its potential failure, resulting in a catastrophic discharge of the tailings into a waterway and/or onto land. The failure might be due to flood, earthquake, or simply the loss of stability of the dam to pressure over time. A number of such incidents have occurred within the last decade—e.g., in Spain—with devastating results to wildlife, fish, and, in some cases, agricultural land.

An alternative to the storage of tailings on land is disposal in the deep ocean by using pipelines reaching down 100 m or more. Lack of oxygen there will slow the process of oxidation, and within a few years the tailings will be covered by other debris. Some biologists are dubious about this plan, because the environment for bottom-dwelling organisms will suffer. Heavy-metal contamination of fish has occurred in areas where this means of disposal is practiced. Due to the fine-grained nature of the tailings, dispersal over wide areas of the ocean floor occurs.

The Remediation of Contaminated Soil

Even areas thought to be rather pristine can have localized areas of contaminated soil. For example, the large-scale forestry industry in New Zealand has resulted in the contamination of several hundred sites where lumber is reared with the preservative pentachloropheriol, PCP (Chapter 11). Furan and iioxin contaminants of the PCP are also found in the soils. All three con-aminant types have now leached into the groundwater at some of the sites md have begun to bioacccumulate in the food chain.

Contaminated soil is found most often not only near waste disposal sites md chemical plants, but also near pipelines and gasoline stations. The three nain types of technologies currently available for the remediation of con-:aminated sites are

* containment or immobilization,

* mobilization, and

* destruction.

In general, the technologies can be applied in situ, i.e., in the place of contamination, or ex situ, i.e., after removing the contaminated matter to another location. Owing to the costs and risks, e.g., air pollution arising from excavation, in situ processes are usually preferred.

Among the techniques associated with in situ containment (i.e., the isolation of wastes from the environment) are capping of the contaminated site, especially with clay, and/or the imposition of cut-off walls of low permeability that prevent the lateral spread of contaminants. Ex situ containment would consist of the placing of the excavated soil in a special landfill. Immobilization techniques, including solidification and stabilization, are especially useful for inorganic wastes, which tend to be difficult to treat by other methods. Stabilization can often be achieved by adding a substance to convert a heavy-metal ion into one of its insoluble salts, such as the sulfide in the case of mercury and lead, or the oxide in the case of chromium. A concentrated waste can be solidified by reaction with Portland cement, for example, or by entombing the wastes in molten glass in the process of vitrification. By these techniques, the solubility and mobility of the contaminants are reduced.

Mobilization techniques are mainly accomplished in situ and include soil washing and the extraction of contaminant vapor from soil for highly volatile, water-insoluble contaminants such as gasoline. Heating of the soil to increase the rate of evaporation and air injection wells are sometimes used in conjunction with soil vapor extraction, in which the contaminants are removed by drilling wells in the soil and applying vacuum extraction. As indicated in Table 16-3, this technique is the most frequently used innovative technology at Superfund sites (Box 16-1) in the United States. A related technology is thermal desorption, in which wastes are heated to cause volatile organic compounds to vaporize. Both soil vapor extraction and thermal desorption are useful to remediate both volatile and semivolatile organic compounds, the latter including many PAHs.

Common Innovative Remediation Technologies in Projects at U.S. Superfund Sites (as of 1996)

f ABLF 16-3

Common Innovative Remediation Technologies in Projects at U.S. Superfund Sites (as of 1996)

f ABLF 16-3

Sites in Design

Sites Operational


or Installation

or Completed

Total Number

Soil vapor extraction




Thermal desorption




Bioremediation (ex situ)




Bioremediation (in situ)

1 i


In situ flushing




Soil washing




Solvent extraction




Source of data: U.S. EPA, Innovative Treatment Technologies: Annual Status Report, 8th ed., 1996.

Source of data: U.S. EPA, Innovative Treatment Technologies: Annual Status Report, 8th ed., 1996.

BOX 16-1

The Superfund Program

In 1980 the federal government of the United States established a program now known as Superfund to clean up abandoned and illegal toxic waste dumps, since dangerous chemicals from many such sites were polluting groundwater. The cleanup costs are shared by chemical companies, the current and past owners of the sites, and the government. Many billions of dollars have already been spent on remediation, and many billions more will eventually he required. Progress in the cleanups has been rather slow on account of the litigation involved and the huge amounts of money at stake. Many decades will pass before even the highest-priority sites are all cleaned up.

The Superfund program is administered by the Environmental Protection Agency, which has identified nearly 1,300 waste sites having such serious potential to cause harm to humans and the environment that they have been placed on a National Priorities List. New Jersey, Pennsylvania, and California have the greatest number of priority sites.

By the late 1990s, the EPA had finished cleanup work at 300 sites, begun work at more than 700 others, and conducted emergency removal of materials at more than 3000 additional locations. The rate at which cleanup work was completed at additional sites slowed appreciably in the mid-2000s. In all, over 30,000 sites have been identified as potentially in need of cleanup.

The most common contaminants at the Superfund sites are the heavy metals lead, cadmium, and mercury, and the organic compounds benzene, toluene, ethylbenzene, and trichloroethylene.

In situ soil washing is accomplished by injecting fluids through wells into subsurface soil and collecting them in other wells. The fluid can simply be water, which will remove water-soluble constituents, or an aqueous solution that is acidic or basic in order to remove basic and acidic contaminants, respectively. Other options in soil washing include the use of solutions containing chelating agents such as EDTA (see Chapter 15) to remove metals and of oxidizing agents to oxidize and thereby solubilize previously insoluble species. The solvents used to extract the metal-organic complex from the aqueous environment of the soil include hydrocarbons and supercritical carbon dioxide (a substance described later in this chapter). In order for the resulting complex to be electrically neutral and therefore preferentially soluble in the organic phase, a chelating agent with acidic hydrogens that are substituted by the metal is usually employed. If the metal exists as an oxyan-ion, e.g., chromium as Cr(VI), it may first have to be reduced in oxidation number before it will bind to a chelating agent.

Sometimes the washing solution uses surfactants, surface'active agents. These are substances such as detergents that possess both hydrophobic and hydrophilic components within the same molecule and therefore can increase the mobilization of hydrophobic contaminants into the aqueous phase. Biosurfac.tants produced by microbes have recently been discovered that can selectively remove certain heavy metals such as cadmium from soil. Currently, soil washing and flushing are the most common innovative technologies used to remove metals at Superfund sites.

The containment, mobilization, and immobilization techniques by themselves do not result in the elimination of the hazardous contaminants. Destruction techniques, principally incineration and bioremediation, do result in permanent elimination because they chemically or biochemically transform the contaminants. Organic contaminants in soil can be oxidized (mineralized) by feeding the excavated soil into the combustion chamber of an incinerator or by using incineration or one of the specialty oxidation techniques, to be discussed later, to treat the substances that have been extracted from the soil. Bioremediation uses the metabolic activities of microorganisms to destroy toxic contaminants and is also discussed in detail later.

Electrochemical techniques are sometimes used to remediate soil. Placing electrodes in the contaminated ground and applying a dc voltage between them results in ion transport within the soil: The ions travel within the groundwater electrolyte. If heavy-metal ions are dissolved in the electrolyte water, they will eventually move to the (negatively charged) cathode and be deposited on it. Indeed, other metal ions tend to he desorbed from their positions on negatively charged clay surfaces in the process, since hydrogen ion is released at the anode (as water is electrolyzed) and subsequently migrates in the groundwater toward the cathode (see Figure 16-12). Recall that heavy-metal ions are much more soluble in an acidic than in a neutral or alkaline environment.

dc nl



(with groundwater)


FIGURE 16-12 Electrochemical remediation of metal-contaminated soil.

In some applications of such electro^ kinetic methods, the process is stopped before the metals are deposited on the cathode but after they have migrated most of the distance toward it. The flow of hydroxide ion from the cathode (see Figure 16-12) precipitates many metals in any 2e- event. The metal-rich soil surrounding the electrode is then excavated and cleansed by washing it or by other techniques. By repeating this procedure and inserting the electrodes at different locations in the soil, more efficient extraction of the heavy metals is possible. The elec-trokinetic technique has been used successfully for copper, lead, cadmium, mercury, chromium, and some radioactive metals. The method is high in energy demand, and hence in cost, because so much of the applied electrical potential is lost by the electrolysis of soil water.

In situ chemical oxidation can often be used to remediate soils (and groundwater) contaminated with chlorinated solvents and/or with BTEX (Chapter 14). The oxidizing agent is injected directly by means of a well into the underground waste and may or may not be extracted at the other side of the contaminated zone. Typically, a salt of permanganate ion, Mn04~, is used for TCE, PCE, and MTBE deposits, whereas ozone or hydrogen peroxide is used for BTEX and PAHs or, in some cases, for the C2 chlorinated solvents as well. The Mn02 product of oxidation by permanganate is a natural constituent of soil. Hydrogen peroxide is often supplemented with ferrous salts (together called Fenton's Reagent) to create hydroxyl radical by a reaction described in Chapter 14.

The Analysis and Remediation of Contaminated Sediments

We now realize that many river and lake sediments are highly contaminated by heavy metals and/or toxic organic compounds and that such sediments act as sources to recontaminate the water that flows above them.

One way to determine the extent of contamination of a sediment is to analyze a sample of it for the total amounts of lead, mercury, and other heavy metals that are present. However, this technique fails to distinguish between toxic materials already present in either their toxic form or in a form that can be resolubilized into the water and those that are firmly bound to sediment particles and unlikely to become resolubilized. Thus a more meaningful test involves extracting from a sediment sample the substances that are soluble in water or in a weakly acidic solution and analyzing the resulting liquid. In this way, the permanently bound and therefore inactive toxic agents can be left out of the account. Finally, the effect of sediments on organisms that usually dwell in or on them can be determined by adding the organisms to a sediment sample and observing whether they survive and reproduce normally.

Several types of remediation have been used for highly contaminated sediments. The simplest solution is often to simply cover the contaminated sediments with clean soil or sediment, thereby placing a barrier between the contaminants and the water system. In other instances, the contaminated sediments are dredged from the bottom of the water body to a depth below which the contaminant concentration is acceptable. If the sediments are high in organic content and inorganic nutrients, they are often used to enhance soil used for nonagricultural purposes. In some cases, the sediment can be used for cropland provided that its heavy metals and other contaminants will not enter the growing food. Cadmium is usually the heavy metal of greatest concern in such sediments; if the pH of the resulting soil is 6.5 or greater, most of the cadmium will not be soluble, so a higher total concentration is often tolerated.

Several chemical and biological methods of decontaminating sediments are in use. For example, treatment with calcium carbonate or lime increases the pH of the sediments and thereby immobilizes the heavy metals. In some situations, contaminated sediments are simply covered with a chemically active solid, such as limestone (calcium carbonate), gypsum (calcium sulfate), iron(Ill) sulfate, or activated carbon, that gradually detoxifies the sediments. In other cases, the sediments are first dredged from the bottom of the water body and then treated. Heavy metals are often removed by acidifying the sediments or treating them with a chelating agent; in both cases, the heavy metals become water-soluble and leach from the solid. For organic contaminants, extraction of toxic substances using solvents and destruction by either heat treatment of the solid or the introduction of microorganisms that consume them are the main options. The cleaned sediments can then be returned to the water body or spread on land. These techniques for removing metals and organics from sediments are also often useful on contaminated soils.

Bioremediation of Wastes and Soil

Recall from Chapter 14 that bioremediation involves the use of living organisms, especially microorganisms, to degrade environmental wastes. It is a rapidly growing technology, especially in collaboration with genetic engineering, which is used to develop strains of microbes with the ability to deal with specific pollutants. Bioremediation is used particularly for the remediation of waste sites and soils contaminated with semivolatile organic compounds such as PAHs. It is a popular method for use at Superfund sites (see Table 16-3).

Bioremediation exploits the ability of microorganisms, especially bacteria and fungi, to degrade many types of wastes, usually to simpler and less toxic substances. Indeed, for many years it was thought that microorganisms could and would eventually biodegrade all organic substances, including all pollutants, that entered natural waters or the soil. The discovery that some compounds, chloroorganics especially, were resistant to rapid biodégradation was responsible for correcting that misconception. Substances resistant to biodégradation are termed recalcitrant or biorefractory. In addition, other substances, including many organic compounds, biodegrade only partially; they are transformed instead into other organic compounds, some of which may be biorefractory and/or even more toxic than the original substances. An example of the latter phenomenon is the potential conversion of the once widely used solvent 1,1,1 -trichloroethane (methyl chloroform, now banned as an ozone-depleting substance, as discussed in Chapter 2) into carcinogenic vinyl chloride, CHC1=CH2, by a combination of abiotic and microbial steps.

If a bioremediation technique is to operate effectively, several conditions must be fulfilled:

• The waste must be susceptible to biological degradation and in a physical form that is susceptible to microbes.

• The appropriate microbes must be available.

• The environmental conditions, such as pH, temperature, and oxygen level, must be appropriate.

An example of biodégradation is the degradation of aromatic hydrocarbons by soil microorganisms when land is contaminated by gasoline or oil. The largest bioremediation project in history was the treatment of some of the oil spilled by the Exxon Valdez tanker in Alaska in 1989. The bioremediation consisted of adding nitrogen-containing fertilizer to more than 100 km of the shoreline that had been contaminated, thereby stimulating the growth of indigenous microorganisms, including those that could degrade hydrocarbons. Both surface and subsurface oil was biodegraded in this operation. Some of the aromatic components in crude oil in marine spills become more susceptible to biodégradation once they are photooxidized by sunlight into more polar molecules.

In contexts such as contaminated soils, the biodégradation of PAHs is slow since they are strongly adsorbed onto soil particles and are not readily released into the aqueous phase, where biodégradation could occur. PAH-contaminated soils are especially prevalent at gasworks sites used in the 1850-1950 period for the production of "town gas" from coal or oil. The pollution is mainly in the form of deposits of tars, which are waste products that are high-molecular-weight organic liquids denser than water that are mixed with the soil and contain high levels of both BTEX and PAHs. Unfortunately, groundwater that comes into contact with the tar can become contaminated if some of its more soluble constituents such as benzene and naphthalene dissolve, although most of the tar is insoluble in water. The other common soil contaminants at gasworks sites are phenols and cyanide.

Bioremediation processes can take place under either aerobic or anaerobic conditions. In the aerobic treatment of wastes, aerobic bacteria and fungi that utilize oxygen are employed; chemically, the processes are oxidations, as the microorganisms use the wastes as food sources. In some of the bioremediation procedures used for aerobic soil, oxygen-saturated water is pumped through the solid to ensure that 02 availability remains high. For example, about 85,000 tonnes of soil contaminated by gasoline, oil, and grease from a fuel plant in Toronto were decontaminated by first enclosing it in plastic and then pumping air, water, and fertilizer into it to encourage the population of aerobic bacteria to multiply and devour the hydrocarbon pollutants. The bioremediation process took only three months.

There are many examples of wastes that can be degraded by anaerobic microorganisms, although usually the most rapid and complete biodégradations are obtained with aerobic microorganisms. The anaerobic process usually works best when there are some oxygen atoms in the organic wastes themselves. In general, the process corresponds to the fermentation discussed in Chapter 13, in which biomass with an approximate empirical formula of CH20 decomposes ultimately to methane and carbon dioxide:

An advantage of anaerobic biodégradation is its production of hydrogen sulfide, which in situ precipitates heavy-metal ions as the corresponding sulfides.

Several strategies used in bioremediation are based on the fact that microorganisms evolve quickly—due to their short reproductive cycle—and they develop the ability to use the food source at hand, even if it is chemical wastes. One remediation strategy is to isolate the most efficient degrading biomicroorgansims flourishing at a contaminated site, grow a large population of them in the laboratory, and finally return the enhanced population to the site. Another strategy is to introduce microorganisms that were found to be useful at other sites in deg rading a particular type of waste, rather than waiting for them to evolve. Unfortunately, microorganisms adapted to one environment may not be capable of surviving in another if additional hostile contaminants are present. A third strategy is to encourage an increase in the population of indigenous microorganisms at the site by adding nutrients to the wastes and ensuring that the acidity and moisture levels are optimum.

Rather than waiting for microorganisms to evolve spontaneously, an alternative strategy is to use genetic engineering to develop microbes specifically designed to attack common organic pollutants. However, regulatory authorities have thus far been reluctant to allow genetically modified organisms to be released into the environment, since public opposition to such a move would be substantial.

In addition to bacteria, white-rot fungi can be used in biodégradation. This species protects itself from pollutants by degrading them outside its cell wall by secreting enzymes there that catalyze the production of hydroxyl



other acyclic products

other acyclic products

FIGURE 16-13 Example of the aerobic degradation of PCB molecules.

radicals and other reactive chemicals. Since hydroxyl radical in particular is quite nonspecific about which substances it oxidizes, the fungi are useful in degrading mixtures of waste, including various chlorinated substances such as DDT and 2,4,5-T, as well as PAHs.

Bioremediation of Organochlorine Contamination

It has been discovered that PCBs (Chapter 11) in sediments undergo some biodégradation. PCB molecules with relatively few chlorine atoms undergo oxidative aerobic biodégradation by a variety of microorganisms. For the reaction to begin, a pair of nonsubstituted carbons—one ortho to the point of connection between the rings and one meta site next to it—must be available on one of the benzene rings. After 2,3 hydroxylation at these two sites, the 1,2 carbon-carbon bonds to the ortho carbons split in sequence, thereby destroying the second ring and producing compounds that readily degrade (see Figure 16-13).


Deduce the chlorine substitution positions on the benzene ring in the benzoic acid that results from the aerobic degradation of 2,3 ' ,5-trichlorobiphenyl.

Although PCB molecules that are heavily substituted by chlorines will not undergo this process, since they are unlikely to have adjacent unsubsti-tuted carbons, they will instead undergo anaerobic degradation, as will per-chlorinated organic compounds such as TCE and HCB (Chapter 10). In the absence of oxygen, anaerobic microorganisms facilitate the removal of chlorine atoms and their replacement by hydrogen atoms, apparently by a reductive dechlorination mechanism that initially involves the addition of an electron to the molecule. In the case of PCBs, this reductive dechlorination occurs most readily with meta and para chlorines. Apparently, steric effects block the ortho position from being attacked in most anaerobic mechanisms.

Aroclor 1242

Aroclor 1242

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