1

EPA/540/S-92/018

October 1992

United States

Environmental Protection

Agency

Office of Solid Waste

and Emergency

Response

Office of

Research and

Development

Ground Water Issue

Behavior of Metals in Soils

Robert S. Kerr Environmental

Research Laboratory

Ada, Oklahoma

Superfund Technology Support Center for

Ground Water

Joan E. McLean* and Bert E. Bledsoe**

Technology Innovation Office

Office of Solid Waste and Emergency

Response, US EPA, Washington, DC

Walter W. Kovalick, Jr., Ph.D.

Director

The Regional Superfund Ground-Water Forum is a group of

scientists, representing EPA's Regional Superfund Offices,

organized to exchange up-to-date information related to

ground-water remediation at Superfund sites. One of the

major issues of concern to the Forum is the mobility of metals

in soils as related to subsurface remediation.

For the purposes of this Issue Paper, those metals most

commonly found at Superfund sites will be discussed in terms

of the processes affecting their behavior in soils as well as

laboratory methods available to evaluate this behavior. The

retention capacity of soil will also be discussed in terms of the

movement of metals between the other environmental

compartments including ground water, surface water, or the

atmosphere. Long-term changes in soil environmental

conditions, due to the effects of remediation systems or to

natural weathering processes, are also discussed with respect

to the enhanced mobility of metals in soils.

For further information contact Bert Bledsoe (405) 332-8800 or

FTS 700-743-2324 at RSKERL-Ada.

Introduction

The purpose of this document is to introduce to the reader the

fundamental processes that control the mobility of metals in

the soil environment. This discussion will emphasize the basic

chemistry of metals in soils and will provide information on

laboratory methods used to evaluate the behavior of metals in

soils. The metals selected for discussion in this document are

the metals most commonly found at Superfund sites and will

be limited to lead (Pb), chromium (Cr), arsenic (As), cadmium

(Cd), nickel (Ni), zinc (Zn), copper (Cu), mercury (Hg), silver

(Ag), and selenium (Se).

Metals are defined as any element that has a silvery luster

and is a good conductor of heat and electricity. There are

many terms used to describe and categorize metals, including

trace metals, transition metals, micronutrients, toxic metals,

heavy metals. Many of these definitions are arbitrary and

these terms have been used loosely in the literature to include

elements that do not strictly meet the definition of the term.

Strictly speaking arsenic and selenium are not metals but are

metalloids, displaying both metallic and non-metallic

properties. For this paper, the term metal will be used to

include all the elements under discussion.

The average concentration of select metals in soils is listed in

Table 1. All soils naturally contain trace levels of metals. The

presence of metals in soil is, therefore, not indicative of contamination. The concentration of metals in uncontaminated

soil is primarily related to the geology of the parent material

from which the soil was formed. Depending on the local

geology, the concentration of metals in a soil may exceed the

ranges listed in Table 1. For example, Se concentration in

non-seleniferous soils in the U.S. range from 0.1 to 2 mg/Kg.

In seleniferous soils, Se ranges from 1 to 80 mg/Kg, with

reports of up to 1200 mg/Kg Se (McNeal and Balistrier, 1989).

Use of common ranges or average concentration of trace

metals in soils as an indicator of whether a soil is

contaminated is not appropriate since the native concentration

of metals in a specific soil may fall out of the listed ranges.

Only by direct analysis of uncontaminated soils can

background levels of metals be determined.

* Utah Water Research Laboratory,Utah State University, Logan,

UT 84322-8200

** Environmental Scientist, Robert S. Kerr Environmental

Research Laboratory, Ada, OK 74820
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Table 1. Content of Various Elements in Soils (Lindsay, 1979)

Metal Selected Average Common Range

for Soils for Soils

mg/kg mg/kg

Al 71,000 10,000-300,000

Fe 38,000 7,000-550,000

Mn 600 20-3,000

Cu 30 2-100

Cr 100 1-1000

Cd 0.06 0.01-0.70

Zn 50 10-300

As 5 1.0-50

Se 0.3 0.1-2

Ni 40 5-500

Ag 0.05 0.01-5

Pb 10 2-200

Hg 0.03 0.01-0.3

importance when considering the migration potential of metals

associated with soils.

Multiphase equilibria must be considered when defining metal

behavior in soils (Figure 1). Metals in the soil solution are

subject to mass transfer out of the system by leaching to

ground water, plant uptake, or volatilization, a potentially

important mechanism for Hg, Se, and As. At the same time

metals participate in chemical reactions with the soil solid

phase. The concentration of metals in the soil solution, at any

given time, is governed by a number of interrelated processes,

including inorganic and organic complexation, oxidationreduction reactions, precipitation/dissolution reactions, and

adsorption/desorption reactions. The ability to predict the

concentration of a given metal in the soil solution depends on

the accuracy with which the multiphase equilibria can be

determined or calculated.

Most studies of the behavior of metals in soils have been

carried out under equilibrium conditions. Equilibrium data

indicate which reactions are likely to occur under prescribed

conditions, but do not indicate the time period involved. The

kinetic aspect of oxidation/reduction, precipitation/dissolution,

and adsorption/desorption reactions involving metals in soil

matrix suffers from a lack of published data. Thus the kinetic

component, which in many cases is critical to predict the

behavior of metals in soils, cannot be assessed easily.

Without the kinetic component, the current accepted approach

is to assume that local equilibrium occurs in the soil profile.

Equilibrium thermodynamic data can then be applied not only

to predict which precipitation/dissolution, adsorption/

desorption, and/or oxidation/reduction reactions are likely to

occur under a given set of conditions, but also to estimate the

solution composition, i.e., metal concentration in solution, at

equilibrium. This approach relies heavily on the accuracy of

thermodynamic data that can be found in the literature.

Soil Solution Chemistry

Metals exist in the soil solution as either free (uncomplexed)

metal ions (e.g., Cd2+, Zn2+, Cr3+), in various soluble

complexes with inorganic or organic ligands (e.g., CdSO4

o,

ZnCl+, CdCl3

-

), or associated with mobile inorganic and

organic colloidal material. A complex is defined as an unit in

which a central metal ion is bonded by a number of associated

atoms or molecules in a defined geometric pattern, e.g

ZnSO4

o, CdHCO3+, Cr(OH)4-

. The associated atoms or

molecules are termed ligands. In the above examples, SO4

2-,

HCO3

-

, and OH- are ligand. The total concentration of a

metal, MeT, in the soil solution is the sum of the free ion

concentration [Mez+], the concentration of soluble organic and

inorganic metal complexes, and the concentration of metals

associated with mobile colloidal material.

Metals will form soluble complexes with inorganic and organic

ligands. Common inorganic ligands are SO4

2-, Cl-

, OH-, PO4

3-,

NO3

-

 and CO3

2-. Soil organic ligands include low molecular

weight aliphatic, aromatic, and amino acids and soluble

constituents of fulvic acids. Formation constants for various

metal complexes are available in the literature (e.g., see

Nordstrom and Munoz, 1985; Lindsay, 1979; Martell and

Smith, 1974 -1982). Organic complexation of metals in soil is

not as well defined as inorganic complexation because of the

difficultly of identifying the large number of organic ligands that

Metals associated with the aqueous phase of soils are subject

to movement with soil water, and may be transported through

the vadose zone to ground water. Metals, unlike the

hazardous organics, cannot be degraded. Some metals, such

as Cr, As, Se, and Hg, can be transformed to other oxidation

states in soil, reducing their mobility and toxicity.

Immobilization of metals, by mechanisms of adsorption and

precipitation, will prevent movement of the metals to ground

water. Metal-soil interaction is such that when metals are

introduced at the soil surface, downward transportation does

not occur to any great extent unless the metal retention

capacity of the soil is overloaded, or metal interaction with the

associated waste matrix enhances mobility. Changes in soil

environmental conditions over time, such as the degradation

of the organic waste matrix, changes in pH, redox potential, or

soil solution composition, due to various remediation schemes

or to natural weathering processes, also may enhance metal

mobility. The extent of vertical contamination is intimately

related to the soil solution and surface chemistry of the soil

matrix with reference to the metal and waste matrix in

question.

Fate of Metals in the Soil Environment

In soil, metals are found in one or more of several "pools" of

the soil, as described by Shuman (1991):

1) dissolved in the soil solution;

2) occupying exchange sites on inorganic soil constituents;

3) specifically adsorbed on inorganic soil constituents;

4) associated with insoluble soil organic matter;

5) precipitated as pure or mixed solids;

6) present in the structure of secondary minerals; and/or

7) present in the structure of primary minerals.

In situations where metals have been introduced into the

environment through human activities, metals are associated

with the first five pools. Native metals may be associated with

any of the pools depending on the geological history of the

area. The aqueous fraction, and those fractions in equilibrium

with this fraction, i.e., the exchange fraction, are of primary
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Figure 1. Principal controls on free trace metal concentrations in

soils solution (Mattigod, et al., 1981).

Complex

Formation

Mass Transfer

Free Metal

Concentration

in Soil Soltuion

Redox

Reaction

Acid-Base

Reaction

Ion

Exchange

and

Adsorption

Precipitation

and

Dissolution

of Solids

Atomic absorption spectrophotometers (AA) and inductively

coupled plasma emission spectrometers (ICP) are commonly

used to determine the metal concentration in soil solutions.

Both techniques measure the total metal concentration in the

solution without distinguishing metal speciation or oxidation

state. Free metal, complexed metal ion concentrations and

concentration of metals in different oxidation states can be

determined using ion selective electrodes, polarography,

colorimetric procedures, gas chromatography-AA, and high

performance liquid chromatography-AA (see Kramer and

Allen, 1988). While these specific methods are necessary for

accurate measurements of metal speciation and oxidation

state, these methods are not routinely performed by

commercial laboratories nor are these procedure standard

EPA methods.

Metal concentrations determined by AA or ICP are often used

as inputs into a thermodynamic computer program, such as

MINTEQA2 (USEPA, 1987). This program can be used to

calculate the speciation and oxidation state of metals in soil

solution of known composition. Formation constants are

known for many metal complexes. There is, however, only

limited information for metal-organic complexes, including

formation constants for many naturally occurring ligands and

those in waste disposal systems. The required input data for

these models include: the concentration of the metal of

interest, the inorganic and organic ligands, and the major

cations and other metal ions, and pH. In specific cases the

redox potential and pCO2 also may be required. Output

consists of an estimation of the concentration of free metals

and complexed metals at equilibrium for the specified

conditions.

Many predictive methods, based on solution and solid phase

chemistry, do not adequately describe transport of metals

under field conditions. Solution chemistry considers the

interaction between dissolved species, dissolved being

defined as substances that will pass a 0.45μm filter. However,

in addition to dissolved metal complexes, metals also may

associate with mobile colloidal particles. Colloidal size

particles are particles with a diameter ranging from 0.01 and

10μm (Sposito, 1989). Gschwend and Reynolds (1987)

reported that colloidal particles of intermediate diameter,

0.1μm to 1μm, were the most mobile particles in a sandy

medium. Colloidal particles include iron and manganese

oxides, clay minerals, and organic matter. These surfaces

have a high capacity for metal sorption. Puls et al. (1991)

reported a 21 times increase in arsenate transport in the

presence of colloidal material compared with dissolved

arsenate. This increased transport of contaminants

associated with mobile colloidal material has been termed

facilitated transport.

Solid Phase Formation

Metals may precipitate to form a three dimensional solid

phase in soils. These precipitates may be pure solids (e.g.,

CdCO3, Pb(OH)2, ZnS2) or mixed solids (e.g., (FexCr1-x)(OH)3,

Ba(CrO4,SO4)). Mixed solids are formed when various

elements co-precipitate. There are several types of coprecipitation, inclusion, adsorption and solid solution

formation, distinguished by the type of association between

the trace element and the host mineral (Sposito, 1989). Solid

solution formation occurs when the trace metal is compatible

may be present in soils. Most of the metal-organic complex

species identified in the literature were generated from metal

interaction with fulvic acids extracted from sewage sludges

(Baham, et al., 1978; Baham and Sposito, 1986; Behel, et al.,

1983; Boyd et al., 1979; Boyd et al., 1983; Dudley, et al.,

1987; Lake et al., 1984; Sposito et al., 1979; Sposito et al.,

1981; Sposito et al., 1982). The soluble metal organic

complexes that may form in other waste systems, however,

have not been identified.

The presence of complex species in the soil solution can

significantly affect the transport of metals through the soil

matrix relative to the free metal ion. With complexation, the

resulting metal species may be positively or negatively

charged or be electrically neutral (e.g., CdCl3

+, CdCl-

, CdCl2

o).

The metal complex may be only weakly adsorbed or more

strongly adsorbed to soil surfaces relative to the free metal

ion. A more detailed discussion on the effect complex

formation has on metal mobility is given in the section: Effect

of anions on adsorption and precipitation. Speciation not only

affects mobility of metals but also the bioavailability and

toxicity of the metal. The free metal ion is, in general, the

most bioavailable and toxic form of the metal.

Several metals of environmental concern exist in soils in more

than one oxidation state: arsenic, As(V) and As(III), selenium,

Se(VI) and Se(IV), chromium, Cr(VI) and Cr(III), and mercury,

Hg(II) and Hg(I). The oxidation state of these metals

determines their relative mobility, bioavailability, and toxicity.

For example, hexavalent Cr is relatively mobile in soils, being

only weakly sorbed by soils. Hexavalent Cr is also extremely

toxic and a known carcinogen. Trivalent Cr, on the other

hand, is relatively immobile in soil, being strongly sorbed by

soils and readily forming insoluble precipitates, and it is of low

toxicity.
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with the element of the host mineral and thus can uniformly

replace the host element throughout the mineral. An example

of solid solution formation is the substitution of Cd for Ca in

calcium carbonate. Cadmium and Ca have almost identical

ionic radii so that Cd can readily substitute of Ca in this

carbonate mineral. Mechanisms of retention, whether surface

adsorption, surface precipitation, co-precipitation, and pure

solid formation are often difficult to distinguish experimentally.

Retention involves a progression of these processes. The

term sorption is used when the actual mechanism of metal

removal from the soil solution is not known.

Stability diagrams are used as a convenient technique for

illustrating how the solubility of metal compounds varies with

soil pH and with metal concentration (or activity). The

diagrams also allow some prediction of which solid phase

regulates metal activity in the soil solution. Methods for

constructing such diagrams is given in Sposito (1989) and

Lindsay (1979). Santillan-Medrano and Jurinak (1975) used

stability diagrams for predicting the formation of precipitates of

Pb and Cd in a calcareous soil. The stability diagrams

(Figures 2 and 3) illustrate the decrease in Pb and Cd

solubility with increasing pH, which is the usual trend with

cationic metals. Solution activity of Cd is consistently higher

than that for Pb indicating that Cd may be more mobile in the

environment. Lead phosphate compounds at lower pH and a

mixed Pb compound at pH>7.5 could be the solid phases

regulating Pb in solution. The authors concluded that

cadmium solution activity is regulated by the formation of

CdCO3 and Cd(PO4)2 or a mixed Cd solid at pH<7.5. At

higher pH, the system is undersaturated with respect to the Cd

compounds considered.

The formation of a solid phase may not be an important

mechanism compared to adsorption in native soils because of

the low concentration of trace metals in these systems

(Lindsay, 1979). Precipitation reactions may be of much

greater importance in waste systems where the concentration

of metals may be exceedingly high. McBride (1980)

Figure 2. The solubility diagram for Pb in Nibley clay loam soil

 (Santillan-Medrano and Jurinak, 1975).

0

2

4

6

8

10

6.0 6.5 7.0 7.5 8.0 8.5 9.0

PbO

PbHPO4

PbCO3 Pb(OH)2

Pb5(PO4)

3OH

Pb5(PO4)

3Cl

Pb3(PO4)2

pH

p

α Pb

p

α Cd

0

2

4

6

8

10

6.0 6.5 7.0 7.5 8.0 8.5 9.0

CdCO3

pH

Cd(OH)2

Cd3(PO4)2

Figure 3. The solubility diagram for Cd in Nibley clay loam soil

 (Santillan-Medrano and Jurinak, 1975).

concluded that calcite (CaCO3) serves as a site for adsorption

of Cd2+ at low concentrations of Cd, while CdCO3

precipitation, possibly as a coating on the calcite, occurs only

at higher Cd concentrations.

Surface Reactions

Adsorption is defined as the accumulation of ions at the

interface between a solid phase and an aqueous phase.

Adsorption differs from precipitation in that the metal does not

form a new three dimensional solid phase but is instead

associated with the surfaces of existing soil particles. The soil

matrix often includes organic matter, clay minerals, iron and

manganese oxides and hydroxides, carbonates, and

amorphous aluminosilicates.

Soil organic matter consists of 1) living organisms, 2) soluble

biochemicals (amino acids, proteins, carbohydrates, organic

acids, polysaccharides, lignin, etc.), and 3) insoluble humic

substances. The biochemicals and humic substances provide

sites (acid functional groups, such as such as carboxylic,

phenolics, alcoholic, enolic-OH and amino groups) for metal

sorption. A discussion of the nature of soil organic matter and

its role in the retention of metals in soil is given by Stevenson

(1991) and Stevenson and Fitch (1990). The biochemicals

form water soluble complexes with metals, increasing metal

mobility, as discussed in a previous section. The humic

substances consists of insoluble polymers of aliphatic and

aromatic substances produced through microbial action.

Humic substances contain a highly complex mixture of

functional groups. Binding of metals to organic matter

involves a continuum of reactive sites, ranging from weak

forces of attraction to formation of strong chemical bonds. Soil

organic matter can be the main source of soil cation exchange

capacity, contributing >200meq/100 g of organic matter in

surface mineral soils. Organic matter content, however,

decreases with depth, so that the mineral constituents of soil

will become a more important surface for sorption as the

organic matter content of the soil diminishes.

There have been numerous studies of the adsorptive

properties of clay minerals, in particular montmorillonite and
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kaolinite, and iron and manganese oxides. Jenne (1968)

concluded that Fe and Mn oxides are the principal soil surface

that control the mobility of metals in soils and natural water. In

arid soils, carbonate minerals may immobilize metals by

providing an adsorbing and nucleating surface (SantillanMedrano and Jurinak, 1975; Cavallaro and McBride, 1978;

McBride, 1980; Jurinak and Bauer, 1956; McBride and

Bouldin, 1984; Dudley et al., 1988; Dudley et al., 1991).

Soil surfaces carry either a net negative or positive charge

depending on the nature of the surface and the soil pH. The

permanent net negative charge on surfaces is due to charge

imbalance resulting from the isomorphous substitution of Al3+

for Si4+ in the tetrahedral layers and/or substitution of Mg2+,

Fe2+, etc. for Al3+ in the octahedral layers of aluminosilicate

clays. The charge on the surface is not affected by changes

in soil pH and hence it is termed a permanent charged

surface. pH dependent charged surfaces are associated with

the edges of clay minerals, with the surfaces of oxides,

hydroxides and carbonates, and with organic matter (acid

functional groups). The charge arises from the association

and dissociation of protons from surface functional groups.

Using an iron oxide surface functional group as an example,

the association of protons with the functional group results in a

positive charge [-Fe-OH2

+] and dissociation of protons, under

more alkaline conditions, results in a negative charge [-Fe-O-

].

At the point of zero net proton charge (PZNPC) the functional

group is neutral [-Fe-OHo]. For all pH dependent charged

surfaces, whether organic or inorganic, as the pH decreases,

the number of negatively charged sites diminishes. Under

more acidic conditions, the majority of pH dependent surfaces

will be positively charged and under more alkaline conditions,

the majority of sites will be negatively charged. The pH

dependent charged surfaces in soils differ widely in their

PZNPC.

The structural charge developed on either a permanent

charged surface or a pH dependent charged surface must be

balanced by ions of opposite charge at or near the surface.

The cation exchange capacity is a measure of the negatively

charged sites for cation adsorption and anion exchange

capacity is a measure of the positively charged sites for anion

adsorption. The anion capacity is, however, very small

relative to the cation adsorption capacity of soils.

A surface complexation model is often used to describe

adsorption behavior (Sposito, 1989). Several types of surface

complexes can form between a metal and soil surface

functional groups and are defined by the extent of bonding

between the metal ion and the surface (Figure 4). Metals in a

diffuse ion association or in an outer sphere complex are

surrounded by waters of hydration and are not directly bonded

to the soil surface. These ions accumulate at the interface of

the charged surfaces in response to electrostatic forces.

These reactions are rapid and reversible with only a weak

dependence on the electron configuration of the surface group

and the adsorbed ion. These two metal-surface interactions

have also been termed exchange reactions because the

introduction of other cations into the system, in sufficient

concentration, causes the replacement or exchange of the

original cations. Metals associated with exchange sites may,

depending on the environment, be relatively mobile.

Exchangeable metals may be the most significant reserve of

potentially mobile metals in soil (Silveira and Sommers, 1977;

Latterell et al., 1978).

Outer-Sphere

Complex

Diffuse Ion

Inner-Sphere

Complex

Figure 4. The three mechanisms of cation adsorption on a

 siloxane surface (e.g., montmorillonite). (Spositi,

 1989).

With inner sphere complexation, the metal is bound directly to

the soil surface, no waters of hydration are involved. It is

distinguished from the exchangeable state by having ionic

and/or covalent character to the binding between the metal

and the surface. A much higher bonding energy is involved

than in exchange reactions, and the bonding depends on the

electron configuration of both the surface group and the metal.

This adsorption mechanism is often termed specific

adsorption. The term specific implies that there are

differences in the energy of adsorption among cations, such

that other ions, including major cations, Na, Ca, Mg, do not

effectively compete for specific surface sites. Specifically

adsorbed metal cations are relatively immobile and unaffected

by high concentrations of the major cations due to large

differences in their energies of adsorption.

At low concentrations, metals are adsorbed by the specific

adsorption sites. These adsorbed metals are not removed by

the input of major cations. With increasing concentration of

the metal, the specific sites become saturated and the

exchange sites are filled (Hendrickson and Corey, 1981;

Lehmann and Harter, 1984; Garcia-Miragaya et al., 1986;

O'Connor et al., 1984; O'Connor et al., 1983). Metals

associated with these nonspecific sites are exchangeable with

other metal cations and are thus potentially mobile. For

example, in an adsorption study using Cd, O'Connor et

al.(1984) showed two mechanisms were responsible for metal

retention by soil. The authors attributed the first mechanism,

active at low concentration (0.01-10mg/L added Cd), to

specific adsorption. At higher concentrations (100-1000mg/L

added Cd), adsorption was attributed to exchange reactions.

Desorption studies showed that the added Cd at low

concentration was not removed by 0.05M calcium solutions,

whereas at the higher loading rates, the calcium salt removed

significant amounts of the adsorbed Cd. These results

indicate that the observed affinity of a metal for soil surfaces is

concentration dependent. These results also emphasize the

importance of using literature or laboratory generated values

that cover the range of metal concentration of interest at a

specific location. Use of data generated in the wrong

concentration range may lead to misinterpretation of the metal

binding strength of the soil.
6

 Table 2. Relative affinity of metals for soils and soil constituents

Soil or Soil Relative Order of Sorption Reference

Constituent

goethite Cu>Pb>Zn>Co>Cd Forbes et al., 1976

Fe oxide Pb>Cu>Zn>Cd Benjamin and Leckie, 1981

montmorillonite Cd=Zn>Ni Puls and Bohn, 1988

kaolinite Cd>Zn>Ni Puls and Bohn, 1988

soils Pb>Cu>Zn>Cd>Ni Biddappa et al., 1981

soils Zn>Ni>Cd Tiller et al., 1984

mineral soils Pb>Cu>Zn>Cd Elliott et al., 1986

organic soils Pb>Cu>Cd>Zn Elliott et al., 1986

soil Pb>Cu>Zn>Ni Harter, 1983

The relative affinity of a soil surface for a free metal cation

increases with the tendency of the cation to form strong

bonds, i.e., inner sphere complexes, with the surface. The

general order of preference for monovalent cations by

montmorillonite is Cs > Rb > K = NH4 > Na > Li. For the

alkaline earth metals the order is Ba > Sr > Ca > Mg. The

preference series indicates a greater attraction of the surface

for the less hydrated cations that can fit closer to the clay

surface. For transition metals, the size of the hydrated cation

cannot be used as the only predictor of adsorption affinity

since the electron configuration of a metal plays an important

role in adsorption. Table 2 reports on results from various

researches on the relative sorption affinity of metals onto a

variety of soils and soil constituents. Although there is

consistently a higher affinity of these surfaces for Pb and Cu

compared with Zn or Cd, the specific order of sorption affinity

depends on the properties of the metals, surface type, and

experimental conditions.

Anions in the Soil Environment

Common anionic contaminants of concern include: arsenic

(AsO4

3- and AsO2-

), selenium (SeO3

2- and SeO42-), and

chromium in one of its oxidation states (CrO4

2-). Soil particles,

though predominantly negatively charged, also may carry

some positive charges. The oxide surfaces, notably iron,

manganese, and aluminum oxides, carbonate surfaces, and

insoluble organic matter can generate a significant number of

positive charges as the pH decreases. The edges of clay

minerals also carry pH dependent charge. These edge sites

may be important sites of retention of anions at pHs below the

point of zero charge (PZC).

Clay minerals, oxides, and organic matter exert a strong

preference for some anions in comparison to other anions,

indicating the existence of chemical bonds between the

surface and the specific anion. Phosphate has been the most

extensively studied anion that exhibits this specific adsorption

(inner sphere complex) phenomenon. Selenite (SeO3

2-) and

arsenate (AsO4

3-) are adsorbed to oxides and soils through

specific binding mechanisms (Rajan, 1979; Neal, et al.,

1987b). Selenite (SeO4

2-) and hexavalent chromium are only

weakly bound to soil surfaces and are thus easily displaced by

other anions. Balistrieri and Chao (1987) found the sequence

of adsorption of anions onto iron oxide to be: phosphate =

silicate = arsenate > bicarbonate/carbonate > citrate =

selenite > molybdate > oxalate > fluoride = selenate > sulfate.

The adsorption capacity for anions is, however, small relative

to cation adsorption capacity of soils.

Soil Properties Affecting Adsorption

The adsorption capacity (both exchange and specific

adsorption) of a soil is determined by the number and kind of

sites available. Adsorption of metal cations has been

correlated with such soil properties as pH, redox potential,

clay, soil organic matter, Fe and Mn oxides, and calcium

carbonate content. Anion adsorption has been correlated with

Fe and Mn oxide content, pH, and redox potential. Adsorption

processes are affected by these various soil factors, by the

form of the metal added to the soil, and by the solvent

introduced along with the metal. The results of these

interactions may increase or decrease the movement of

metals in the soil water.

Korte et al. (1976) qualitatively ranked the relative mobilities of

11 metals added to 10 soils (Table 3) to simulate movement of

metals under an anaerobic landfill situation. The leachate

used was generated in a septic tank, preserved under carbon

dioxide and adjusted to pH of 5. Of the cationic metals

studied lead and copper were the least mobile and mercury(II)

was the most mobile (Figure 5). The heavier textured soils

with higher pHs (Molokai, Nicholson, Mohaveca and Fanno)

were effective in attenuating the metals, while sandy soils and/

or soils with low pH did not retain the metals effectively. For

the anionic metals, clay soils containing oxides with low pH

were relatively effective in retaining the anions (Figure 6). As

with the cationic metals, the light textured soils were the least

effective in retaining the anions. Chromium (VI) was the most

mobile of the metals studied. Griffin and Shimp (1978) found

the relative mobility of nine metals through montmorillonite

and kaolinite to be: Cr(VI) > Se > As(III) > As(V) > Cd > Zn >

Pb > Cu > Cr(III).
7

Figure 6. Relative mobility of anions through soil. (From Korte,

 Skopp, Fuller, Niebla, and Alesii, 1976).

Se V As Cr

 LOW MOBILITY

Increasing

Attenuation

Capacity

Increasing Mobility

SOILS

MOLOKAI c.

NICHOLSON si.c

MOHAVE s.l.

FANNO c.

MOHAVE (Ca) c.l.

DAVIDSON c.

AVA si. c.

KALKASKA s.

ANTHONY s.l.

WAGRAM l.s.

 HIGH MOBILITY

 MODERATE

 MOBILITY

Factors Affecting Adsorption and Precipitation

Reactions

Although the principles affecting sorption and precipitation are

similar for cationic and anionic metals, for clarity, the following

section will concentrate on a general discussion of factors

affecting the behavior of cationic metals in soils. Factors

affecting anion adsorption and precipitation will be discussed

for each individual metal anion in a later section.

Effect of competing cations

For specific adsorption sites, trace cationic metals are

preferentially adsorbed over the major cations (Na, Ca, Mg)

and trace anionic metals are preferentially adsorbed over

major anions (SO4

-

, NO3

-

, soluble ionized organic acids).

However, when the specific adsorption sites become

saturated, exchange reactions dominate and competition for

these sites with soil major ions becomes important. Cavallaro

and McBride (1978) found that adsorption of Cu and Cd

decreased in the presence of 0.01M CaCl2. They attributed

this decrease to competition with Ca for adsorption sites.

Cadmium adsorption was more affected by the presence of

Ca than Cu. The mobility of Cd may be greatly increased due

to such competition. Likewise, Harter (1979) indicated the Ca

in solution had a greater effect on Pb adsorption than on Cu.

In another study, Harter (1992) added Cu, Ni and Co to

calcium saturated soils. The presence of Ca, a common ion in

soils with pH>5.6, did not affect Cu sorption but did limit the

sorption of Co and Ni. The author emphasized the importance

of these results in that standard management practice for

metal contaminated soils is to raise the pH to 7, often using a

Ca buffered system. The addition of Ca, as low as 0.01M Ca,

may increase the mobility of some metals by competing for

sorption sites.

Trace metals also will compete with each other for adsorption

sites. Although there have been several studies on the

relative adsorption affinities of trace metals by soils and soil

constituents (see Table 2), these studies have compared how

much of each metal, added to the soils as individual

Figure 5. Relative mobility of cations through soil. (From Korte,

 Skopp, Fuller, Niebla, and Alesii, 1976).

SOILS

MOLOKAI c.

NICHOLSON si.c

MOHAVE Ca c.l.

FANNO c.

MOHAVE s.l.

DAVIDSON c.

AVA si. c.l.

KALKASKA s.

ANTHONY s.l.

WAGRAM l.s.

Cu Pb Be Zn Cd Ni Hg

 LOW MOBILITY

 MODERATE

 MOBILITY Increasing

Attenuation

Capacity

Increasing Mobility

 HIGH MOBILITY

Table 3. Characteristics of the soils (Korte et al., 1976)

Free

 Soil Order pH CEC Surface Area Fe oxides Clay Texture

meq/100g m2/g % %

Wagram Ultisol 4.2 2 8.0 0.6 4 loamy sand

Ava Alfisol 4.5 19 61.5 4 31 silty clay loam

Kalkaska Spodosol 4.7 10 8.9 1.8 5 sand

Davidson Ultisol 6.2 9 61.3 17 61 clay

Molokai Oxisol 6.2 14 67.3 23 52 clay silty clay

Chalmers Mollisol 6.6 26 125.6 3.1 35 loam

Nicholson Alfisol 6.7 37 120.5 5.6 49 silty day

Fanno Alfisol 7 33 122.1 3.7 46 clay

Mohave Aridisol 7.3 10 38.3 1.7 11 sandy loam

Mohave ca Aridisol 7.8 12 127.5 2.5 40 clay loam

Anthony Entisol 7.8 6 19.8 1.8 15 sandy loam
8

components, was adsorbed and not whether the adsorption of

one metal will interfere with that of another. Few studies have

looked directly at the competitive adsorption of metals. Kuo

and Baker (1980) reported that the presence of Cu interfered

with the adsorption of Zn and Cd. Adsorbed Cu was not

significantly affected by added Zn but the presence of Cu, at

concentrations as low as 15 μg/L, completely prevented Zn

adsorption in one soil with a low cation exchange capacity

(Kurdi and Doner, 1983). In contrast, McBride and Blasiak

(1979) found that Cu was ineffective in competing for Zn

adsorption sites over a pH range of 5-7. The inability of Cu to

block Zn adsorption in this study was taken as evidence that

Zn and Cu were preferentially adsorbed at different sites.

Simultaneous addition of Cd and Zn to Mn oxide lowered the

adsorption of both metals (Zasoski and Burau, 1988).

The presence of other cations, whether major or trace metals,

can significantly effect the mobility of the metal of interest.

Use of data from the literature, generation of laboratory data,

or use of computer models that do not reflect the complex

mixture of metals specific to a site may not be useful to

understand or accurately predict metal mobility.

Effect of complex formation

Metal cations form complexes with inorganic and organic

ligands. The resulting association has a lower positive charge

than the free metal ion, and may be uncharged or carry a net

negative charge. For example, the association of cadmium

with chloride results in the following series of charged and

uncharged cadmium species: Cd2+, CdCl+, CdCl2

o, CdCl3-

.

Benjamin and Leckie (1982) stated that the interaction

between metal ions and complexing ligands may result in

either a complex that is weakly adsorbed to the soil surface or

in a complex that is more strongly adsorbed relative to the free

metal ion. In general, the decrease in positive charge on the

complexed metal reduces adsorption to a negatively charged

surface. One noted exception is the preferential adsorption of

hydrolyzed metals (MeOH+) versus the free bivalent metal

(James and Healy, 1972). The actual effect of complex

formation on sorption depends on the properties of the metal

of interest, the type and amount of ligands present, soil

surface properties, soil solution composition, pH and redox

conditions, as is illustrated by the follow research results.

In the presence of the inorganic ligands Cl-

 and SO4

2-, the

adsorption of Cd on soil and soil constituents was inhibited

(O'Connor, et al., 1984; Hirsch et al., 1989; Egozy, 1980;

Garcia-Miragaya and Page, 1976; Benjamin and Leckie, 1982)

due to the formation of cadmium complexes that were not

strongly adsorbed by the soils. Using much higher

concentrations of salt than normally encountered in soil

solutions (0.1 to 0.5M NaCl), Doner (1978) concluded that the

increased mobility of Ni, Cu, and Cd through a soil column

was due to complex formation of the metals with Cl-

. The

mobility of Cd increased more than that of Ni and Cu, Ni being

the least mobile. These observed mobilities are in the same

order as that of the stability constants of the chloride

complexes of these metals. Within normal concentration of

electrolytes in soil solution, Elrashidi and O'Connor (1982)

found no measurable change in Zn adsorption by alkaline soils

due to complex formation of Zn with Cl-

, NO3

2-, or SO42- ions.

Under these conditions (anion concentration of 0.1M), anion

complex formation did not compete with the highly selective

adsorption sites for Zn. Shuman (1986), using acid soils,

observed a decreased adsorption of Zn in the presence of Clat the concentration of CaCl2 used by Elrashidi and O'Connor

(1982) but no effect at lower concentrations. McBride (1985),

using aluminum oxide, and Cavallaro (1982), using clays,

found that high levels of phosphate suppressed adsorption of

Cu and Zn. Phosphate did not form strong complexes with Cu

or Zn but it was strongly adsorbed to soil surfaces thus

physically blocking the specific adsorption sites of Cu and Zn.

Other researchers (Kuo and McNeal, 1984; Stanton and

Burger, 1970; Bolland et al., 1977), using lower concentrations

of added phosphate, demonstrated enhanced adsorption of

Zn and Cd on oxide surfaces. At the concentration of

phosphate used in these studies, the adsorption of phosphate

onto the oxide surfaces increased the negative charge on the

oxide surface, thus enhancing adsorption of the metal cations.

Complex formation between metals and organic ligands

affects metal adsorption and hence mobility. The extent of

complexation between a metal and soluble organic matter

depends on the competition between the metal-binding

surface sites and the soluble organic ligand for the metal.

Metals that readily form stable complexes with soluble organic

matter are likely to be mobile in soils. Overcash and Pal

(1979) reported that the order of metal-organic complex

stabilities, for the system they studied, was Hg > Cu > Ni > Pb

> Co > Zn > Cd. Khan et al. (1982) showed that the mobility

of metals through soil followed the order: Cu > Ni > Pb > Ag >

Cd. The high mobility of Cu and Ni was attributed to their high

complexing nature with soluble soil organic matter. Amrhein,

et al. (1992) also showed the increased mobility of Cu, Ni, and

Pb in the presence of dissolved organic matter. In this study,

the Cd leached from the columns was not associated with

dissolved organic carbon but was associated with Cl or

acetate anions. Metals, such as Cd and Zn, that do not form

highly stable complexes with organic matter are not as greatly

affected by the presence of dissolved organic matter in the soil

solution as metals that do form stable complexes, such as Cu,

Pb, or Hg. Dunnivant et al. (1992) and Neal and Sposito

(1986), however, demonstrated that dissolved organic matter

does reduce Cd sorption due to complexation formation under

their experimental conditions.

In systems where the organic ligand adsorbs to the soil

surface, metal adsorption may be enhanced by the

complexation of the metal to the surface-adsorbed ligand.

Haas and Horowitz (1986) found that, in some cases, the

presence of organic matter enhanced Cd adsorption by

kaolinite. They interpreted these findings to suggest that the

presence of an adsorbed layer of organic matter on the clay

surface served as a site for Cd retention. Davis and Leckie

(1978) found Cu adsorption to iron oxide increased in the

presence of glutamic acid and 2,3 pyrazinendicarboxylic acid

(2,3 PDCA) but decreased in the presence of picolinic acid.

Picolinic acid complexed Cu and the resulting complex was

not adsorbed by the oxide surface. The glutamic acid and 2,3

PDCA were adsorbed to the oxide surface, then complexed

the added Cu. Using natural organic matter, Davis (1984)

demonstrated the adsorption of Cu but not Cd to an organic

coated aluminum oxide.

The effect of complexation formation on sorption is dependent

on the type and amount of metal present, the type and amount

of ligands present, soil surface properties, soil solution

composition, pH and redox. The presence of complexing

ligands may increase metal retention or greatly increase metal
9

mobility. Use of literature or laboratory data that do not

include the presence of complexing ligands, both organic and

inorganic, present at the particular site of interest, may lead to

significant overestimation or underestimation of metal mobility.

Effect of pH

The pH, either directly or indirectly, affects several mechanisms of metal retention by soils. Figure 7 shows the impact

of soil pH on the adsorption of Pb, Ni, Zn, and Cu by two soils

adjusted to various pHs ranging from approximately 4.3 to 8.3

(Harter, 1983). As is true for all cationic metals, adsorption

increased with pH. The author, however, points out that the

retention of the metals did not significantly increase until the

pH was greater than 7. Figure 8 illustrates the adsorption of

selenite, SeO3

2-, on five soils adjusted to various pHs. As is

true with all oxyanions, i.e., arsenic, selenium and hexavalent

chromium, sorption decreases with pH.

The pH dependence of adsorption reactions of cationic metals

is due, in part, to the preferential adsorption of the hydrolyzed

metal species in comparison to the free metal ion (McBride,

1977; McLauren and Crawford, 1973; Davis and Leckie, 1978;

Farrah and Pickering, 1976a,b; James and Healy, 1972;

McBride, 1982; Cavallaro and McBride, 1980; Harter, 1983).

The proportion of hydrolyzed metal species increases with pH.

Cavallaro and McBride (1980) found that copper adsorption by

soils showed a stronger pH dependence than Cd. This finding

is consistent with the hypothesis that hydrolysis of Cu at pH 6

increases its retention by soil, while cadmium does not

hydrolyze until pH 8. Zinc was shown to be retained in an

exchangeable form at low pH in four Fe and Mn oxide

dominated soils but became nonexchangeable as the pH was

increased above 5.5 (Stahl and James, 1991). The

researchers attributed this change in mechanism of sorption

as being due to the hydrolysis of Zn and the adsorption of the

hydrolysis species by the oxide surfaces.

Many adsorption sites in soils are pH dependent, i.e., Fe and

Mn oxides, organic matter, carbonates, and the edges of clay

minerals. As the pH decreases, the number of negative sites

for cation adsorption diminishes while the number of sites for

anion adsorption increases. Also as the pH becomes more

acidic, metal cations also face competition for available

permanent charged sites by Al3+ and H+.

All trace metal hydroxide, oxide, carbonate, and phosphate

precipitates form only under alkaline conditions (Lindsay,

1979). The dissolution of these metal precipitates is strongly

dependent on the pH of the system. Jenne (1968) stated that

hydrous oxides of Fe and Mn play a principal role in the

retention of metals in soils. Solubility of Fe and Mn oxides is

also pH-related. Below pH 6, the oxides of Fe and Mn

dissolve, releasing adsorbed metal ions to solution (Essen and

El Bassam, 1981).

Work by McBride and Blasiak (1979) showed increased

retention of Zn with increasing pH, as is usual for metal

cations. When the pH was increased above 7.5, however, the

solution concentration of Zn increased. This phenomena has

been observed in other studies when acid soils were adjusted

to pH>7 (Kuo and Baker, 1980) and it has been attributed to

Pb

Zn

Ni1 Cu

Ni2

Dekalb A Dekalb B

Cu

Pb

Ni1

Zn

Ni2

10

9

8

7

6

5

4

3

2

1

Hagerstown A Hagerstown B

Zn

Pb

Cu

Ni

Pb

Cu

Ni1

Zn

Ni

Metal Adsorption Maximum (c mol • kg

2

-1)

9

8

7

6

5

4

3

2

1

4 8 5 6 7 8 4 5 6 7

pH

Figure 7. Effect of soil pH level on maximum Pb, Cu, Zn, and Ni

 retention by Dekalb and Hagerstown A and B horizons.

 Ni1 and Ni2 refer to two apparent sorption maxima.

(Harter, 1983).

Figure 8. Selenite adsorption envelope for five alluvial soils. The

intitial total selenite concentration was approximately 2

μmol kg-1 (Neal, et al., 1987a).

100

80

60

40

20

4.0 5.0 6.0 7.0 8.0 9.0 10.0

% Selenite Sorbed

pH

Initial [Selenite] = μmol kg-1

Ionic Strength = 50 mol m-3 NaCl

= Altamont

= Ciervo

= Los Banos

= Panhill

= Panoche
10

the solubilization of organic complexing ligands which

effectively compete with the soil surfaces for the metal cation.

Most functional groups of complexing ligands are weak acids,

thus the stability of the metal complex is pH-dependent with

little association in acid media. The degree of association

increases with pH. Baham and Sposito (1986) and Inskeep

and Baham (1983) demonstrated that the adsorption of Cu to

montmorillonite, in the presence of water soluble ligands

extracted from sludges and various other organic materials,

decreased with increasing pH. This behavior is the opposite

of the typical relationship between metal adsorption and pH.

Figure 9, taken from Baham and Sposito (1986), illustrates

that nearly 100% of the Cu added to the clay in the absence of

the organic ligands was removed from solution at pH>7. In

the presence of the organic ligands, the maximum amount of

Cu removed from solution was at pH35.5. As the pH was

increased above 5.5, adsorption of Cu decreased. The

explanation for this phenomena is that at low pH, H+

competes with the Cu for complexation with the organic

matter. As the pH increases, more of the Cu can be

complexed with the organic matter and less is therefore

adsorbed by the clay. This phenomena has important

implications with regards to the practice of liming acid soils to

raise the pH increasing metal retention. In soils with

significant levels of dissolved organic matter, increasing soil

pH may actually mobilize metal due to complex formation.

The pH of the soil system is a very important parameter,

directly influencing sorption/desorption, precipitation/

dissolution, complex formation, and oxidation-reduction

reactions. In general, maximum retention of cationic metals

occurs at pH>7 and maximum retention of anionic metals

occurs at pH<7. Because of the complexity of the soil-waste

system, with its myriad of surface types and solution

composition, such a generalization may not hold true. For

example, cationic metal mobility has been observed to

increase with increasing pH due to the formation of metal

complexes with dissolved organic matter.

Figure 9. Adsorption of Cu [50 mmol m-3 (50 mM)] by Na-

 montmorillonite in the presence and absence of water

 soluble extract of sewage sludge (WSE). GEOCHEM

 simulations were constructed employing the "mixture

 model" (Baham and Sposito, 1986).

Effect of oxidation-reduction

Almost half of the metals under consideration have more than

one oxidation state in the soil environment and are directly

affected by changes in the oxidation-reduction (redox)

potential of the soil. The redox potential of a soil system is the

measure of the electrochemical potential or availability of

electrons within a system. A chemical reaction in which an

electron transfer takes place is called an oxidation-reduction

process. Metals or elements which gain electrons and lose in

valence are undergoing reduction, while those losing electrons

and gaining in valence are becoming oxidized. A measure of

the redox potential (electron availability) indicates whether the

metals are in an oxidize or reduced state.

In soils, reducing conditions are brought about by the absence

of oxygen (anaerobic). This is caused by the oxygen being

utilized or consumed at a greater rate that it can be

transported into the soil system. This can be caused by

water-logged soils or soils contaminated with oxygen

consuming compounds. The consumption could either be

chemical or biological. The biological consumption of oxygen

is the results of microbes utilizing the organic contaminant

which have entered the soil system. Oxidizing conditions

(aerobic) are normally found in well-drained soils as well as

soils that have not been subjected to contamination by spills

or leaks.

The degree of oxidation or reduction is indicated by the redox

potential measurement. The four general ranges of redox

conditions as suggested by Patrick and Mahapatra (1968)

which may be encountered in soils are at pH 7, oxidized soils

> +400 millivolts (mv); moderately reduced soils, from +400 to

+100 mv; reduced soils, from +100 to -100 mv; highly reduced

soils, -100 to -300 mv. The redox state of a soil, as discussed

above, usually is closely related to the microbial activity and

the type of substrate available to the organisms.

Redox reactions can greatly affect contaminant transport, in

slightly acidic to alkaline environments, Fe(III) precipitates as

a highly adsorptive solid phase (ferric hydroxide), while Fe(II)

is very soluble and does not retain other metals. The

reduction of Fe(III) to Fe(II) will bring about the release of

ferrous iron to the pore waters and also any metals that were

adsorbed to the ferric hydroxide surfaces. The behavior of

chromium and selenium also illustrates the importance of

redox conditions to metals movement in soils. Hexavalent

Cr(VI) is both toxic and a relatively mobile anion while trivalent

Cr(III) is far less toxic, relatively insoluble, and strongly

adsorbs to surfaces. Selenate (Se(VI) is mobile, but less toxic

than selenite (Se(IV) which is more toxic, but less mobile. In

general, oxidizing conditions favor retention of metals in soils,

while reducing conditions contribute to accelerated migration.

Effect of co-waste

Most soil-metal interaction studies have been performed using

a specific, well characterized background solution, such as an

inorganic salt solution (0.01M CaCl2, Na2SO4, etc.) or a water

soluble extraction of organic matter (leaf litter, sewage

sludges, etc.). These studies, as reported above, have led to

an understanding of the effects that metal type, metal

concentration, solution composition, and soil surface type

have on the retention of metals by soils. The behavior of

metals associated with various industrial or mining wastes in

0

50

100 34 5 6 7 8 9

EXPERIMENTAL DATA

Cu-ligands absent

 CTS 4.3 mM

 CuTS 50 μM

GEOCHEM MODELS

Cu and ligands

Cu and model ligands

Cu-ligands absent Cu(OH)2 (s)

CuTS in Solution (%)

pH
11

soil systems has not been extensively studied, however. In

such wastes the metal concentration may be much greater

than used in studies of native metals and metals associated

with the controlled application of fertilizers and sewage

sludges, and may be associated with a myriad of inorganic

and organic chemicals that have not been characterized but

may have a great effect on predicting metal mobility. Below

are examples in which investigators have used various waste

mixture for the background solution in sorption studies. In all

cases, the results were highly dependent on the waste type

used. These examples have been included to emphasize the

importance of performing laboratory studies or using literature

data that mimic the actual matrix of the waste or soils-waste

system being investigated.

The retention of Cd, Cu, and Zn by two calcareous soils using

a water extract of an acidic milling waste as the background

solution (pH=4.0, dominant major cation was Ca and anion

was sulfate) was studied by Dudley et al. (1988, 1991). The

presence of carbonate minerals is known to effectively

immobilize Cd and Cu by providing an adsorbing or nucleating

surface and by buffering pH (Santillan-Medrano and Jurinak,

1975; Cavallaro and McBride, 1978; McBride and Bouldin,

1984). For the soil with a lower carbonate content (0.2%

CaCO3), the sorption of Cd and Zn was slow to reach

equilibrium (114 hours) due to the complex set of reactions

that occurred when the soil (pH 8.6) and acid milling extract

(pH 4.0) were combined. The dissolution of carbonates in the

acid medium controlled the rate and extent of Cd and Zn

sorption. The authors concluded that Cd and Zn were

retained by an exchange mechanism only after the pH of the

system reached equilibrium (pH 5.5), allowing time for

significant transport of these metals. Copper sorption was

independent of calcite dissolution. The soil with the higher

carbonate content (30%) showed a significant drop in pH (pH

9.1 to 6.6) with the addition of the acid leachate but had

sufficient carbonates to buffer the system and sorbed all three

metals.

Kotuby-Amacher and Gambrell (1988) studied the retention of

Cd and Pb on subsurface soils using a synthetic municipal

waste leachate and a synthetic acid metal waste leachate,

compared with Ca(NO3)2 as the background solution.

Sorption of the two metals was diminished in the presence of

both synthetic leachates. The presence of competing cations

and complexing organic and inorganic ligands in the synthetic

wastes decreased the retention of Cd and Pb by the soils.

Boyle and Fuller (1987) used soil columns packed with five

different soils to evaluate the mobility of Zn in the presence of

simulated municipal solid waste leachate with various

amounts of total organic carbon (TOC) and total soluble salts

(TSS). Zinc transport was enhanced in the presence of higher

TOC and TSS. Soil properties considered important for

retaining Zn in this study were surface area, CEC, and percent

clay content. The authors, however, concluded that the

leachate composition was more important than soil properties

for determining the mobility of Zn.

Puls et al. (1991) studied the sorption of Pb and Cd on

kaolinite in the presence of three organic acids, 2,4-

dinitrophenol, p-hydroxybenzoic acid, and o-toluic acid. The

acids were selected based on their frequent occurance at

hazardous waste sites and their persistence in soils. Sorption

of Pb decreased in the presence of all the acids due to the

formation of 1:2 metal-organic complex resulting in an

uncharged form of Pb. Sorption of Cd decreased in the

presence of two of the acids but increased in the presence of

2,4-dinitrophenol. The authors attributed the increase in

sorption as being due to either direct sorption of the acid to the

clay with the subsequent sorption of Cd or to the enhanced

sorption of the 1:1 complex formed between Cd and the acid.

Sheets and Fuller (1986) studied the transport of Cd through

soil columns with 0 to 100% ethylene glycol or 2-propanol as

the leaching solution. Soils sorbed less Cd from the ethylene

glycol solutions than when the columns were leached with

water. The 2-propanol increased sorption in one of the soils

tested. The effect on Cd sorption was attributed to the change

in soil permeability and surface characterization due to the

presence of the solvents.

Metal mobility in soil-waste systems is determined by the type

and quantity of soil surfaces present, the concentration of

metal of interest, the concentration and type of competing ions

and complexing ligands, both organic and inorganic, pH, and

redox status. Generalization can only serve as rough guides

of the expected behavior of metals in such systems. Use of

literature or laboratory data that do not mimic the specific site

soil and waste system will not be adequate to describe or

predict the behavior of the metal. Data must be site specific.

Long term effects also must be considered. As organic

constituents of the waste matrix degrade, or as pH or redox

conditions change, either through natural processes of

weathering or human manipulation, the potential mobility of

the metal will change as soil conditions change. Few long

term studies have been reported.

Behavior of Specific Metals

Copper

Copper is retained in soils through exchange and specific

adsorption mechanisms. At concentrations typically found in

native soils, Cu precipitates are unstable. This may not be the

case in waste-soil systems and precipitation may be an

important mechanism of retention. Cavallaro and McBride

(1978) suggested that a clay mineral exchange phase may

serve as a sink for Cu in noncalcareous soils. In calcareous

soils, specific adsorption of Cu onto CaCO3 surfaces may

control Cu concentration in solution (Cavallaro and McBride,

1978; Dudley, et al., 1988; Dudley et al., 1991; McBride and

Bouldin, 1984). As reported in the adsorption sequence in

Table 2, Cu is adsorbed to a greater extent by soils and soil

constituents than the other metals studied, with the exception

of Pb. Copper, however, has a high affinity for soluble organic

ligands and the formation of these complexes may greatly

increase Cu mobility in soils.

Zinc

Zinc is readily adsorbed by clay minerals, carbonates, or

hydrous oxides. Hickey and Kittrick (1984), Kuo et al. (1983),

and Tessier et al. (1980) found that the greatest percent of the

total Zn in polluted soils and sediments was associated with

Fe and Mn oxides. Precipitation is not a major mechanism of

retention of Zn in soils because of the relatively high solubility

of Zn compounds. Precipitation may become a more

important mechanism of Zn retention in soil-waste systems.

As with all cationic metals, Zn adsorption increases with pH.
12

Zinc hydrolysizes at pH>7.7 and these hydrolyzed species are

strongly adsorbed to soil surfaces. Zinc forms complexes with

inorganic and organic ligands that will affect its adsorption

reactions with the soil surface.

Cadmium

Cadmium may be adsorbed by clay minerals, carbonates or

hydrous oxides of iron and manganese or may be precipitated

as cadmium carbonate, hydroxide, and phosphate. Evidence

suggests that adsorption mechanisms may be the primary

source of Cd removal from soils (Dudley et al., 1988, 1991).

In soils and sediments polluted with metal wastes, the greatest

percentage of the total Cd was associated with the

exchangeable fraction (Hickey and Kittrick, 1984; Tessier et

al., 1980; Kuo et al., 1983). Cadmium concentrations have

been shown to be limited by CdCO3 in neutral and alkaline

soils (Santillan-Medrano and Jurinak, 1975). As with all

cationic metals, the chemistry of Cd in the soil environment is,

to a great extent, controlled by pH. Under acidic conditions

Cd solubility increases and very little adsorption of Cd by soil

colloids, hydrous oxides, and organic matter takes place. At

pH values greater than 6, cadmium is adsorbed by the soil

solid phase or is precipitated, and the solution concentrations

of cadmium are greatly reduced. Cadmium forms soluble

complexes with inorganic and organic ligands, in particular Cl-.

The formation of these complexes will increase Cd mobility in

soils.

Lead

Soluble lead added to the soil reacts with clays, phosphates,

sulfates, carbonates, hydroxides, and organic matter such that

Pb solubility is greatly reduced. At pH values above 6, lead is

either adsorbed on clay surfaces or forms lead carbonate. Of

all the trace metals listed in Table 2, Pb is retained by soils

and soil constituents to the greatest extent under the

conditions of these studies. Most studies with Pb, however,

have been performed in well defined, simple matrices, i.e.,

0.01M CaCl2. Puls et al. (1991), and Kotuby-Amacher and

Gambrell (1988) have demonstrated decrease sorption of Pb

in the presence of complexing ligands and competing cations.

Lead has a strong affinity for organic ligands and the formation

of such complexes may greatly increase the mobility of Pb in

soil.

Nickel

Nickel does not form insoluble precipitates in unpolluted soils

and retention for Ni is, therefore, exclusively through

adsorption mechanisms. Nickel will adsorb to clays, iron and

manganese oxides, and organic matter and is thus removed

from the soil solution. The formation of complexes of Ni with

both inorganic and organic ligands will increase Ni mobility in

soils.

Silver

Published data concerning the interaction of silver with soil are

rare. As a cation it will participate in adsorption and

precipitation reactions. Silver is very strongly adsorbed by

clay and organic matter and precipitates of silver, AgCl,

Ag2SO4 and AgCO3, are highly insoluble (Lindsay, 1979).

Silver is highly immobile in the soil environment.

Mercury

The distribution of mercury species in soils, elemental mercury

(Hgo), mercurous ions (Hg2

2+) and mercuric ions (Hg2+), is

dependent on soil pH and redox potential. Both the

mercurous and mercuric mercury cations are adsorbed by clay

minerals, oxides, and organic matter. Adsorption is pH

dependent, increasing with increasing pH. Mercurous and

mercuric mercury are also immobilized by forming various

precipitates. Mercurous mercury precipitates with chloride,

phosphate, carbonate, and hydroxide. At concentrations of

Hg commonly found in soil, only the phosphate precipitate is

stable. In alkaline soils, mercuric mercury will precipitate with

carbonate and hydroxide to form a stable solid phase. At

lower pH and high chloride concentration, HgCl2 is formed.

Divalent mercury also will form complexes with soluble organic

matter, chlorides, and hydroxides that may contribute to its

mobility (Kinniburgh and Jackson, 1978).

Under mildly reducing conditions, both organically bound

mercury and inorganic mercury compounds may be degraded

to the elemental form of mercury, Hgo. Elemental mercury can

readily be converted to methyl or ethyl mercury by biotic and

abiotic processes (Roger, 1976, 1977). These are the most

toxic forms of mercury. Both methyl and ethyl mercury are

volatile and soluble in water. Griffin and Shimp (1978)

estimated that the removal of Hg from a leachate was not due

to adsorption by clays, but was due to volatilization and/or

precipitation. This removal of mercury increased with pH.

Rogers (1979) also found large amounts of mercury volatilized

from soils. Amounts of mercury volatilized appeared to be

affected by the solubility of the mercury compounds added to

soil. Volatilization was also found to be inversely related to

soil adsorption capacity. The form of Hg lost from the soil,

whether elemental Hg or methylmercury, was not determined

in this study.

Arsenic

In the soil environment arsenic exists as either arsenate,

As(V) (AsO4

3-), or as arsenite, As(III) (AsO2-

). Arsenite is the

more toxic form of arsenic.

The behavior of arsenate in soil is analogous to that of

phosphate, because of their chemical similarity. Like

phosphate, arsenate forms insoluble precipitates with iron,

aluminum, and calcium. Iron in soils is most effective in

controlling arsenate's mobility. Arsenite compounds are

reported to be 4-10 times more soluble than arsenate

compounds.

Griffin and Shimp (1978), in a study of arsenate adsorption by

kaolinite and montmorillonite, found maximum adsorption of

As(V) to occur at pH 5. Adsorption of arsenate by aluminum

and iron oxides has shown an adsorption maximum at pH 3-4

followed by a gradual decrease in adsorption with increasing

pH (Hingston et al., 1971; Anderson et al., 1976). The

mechanism of adsorption has been ascribed to inner sphere

complexation (specific adsorption), which is the same

mechanism controlling the adsorption of phosphate by oxide

surfaces (Hingston et al., 1971; Anderson et al., 1976;

Anderson and Malotky, 1979).

The adsorption of arsenite, As(III), is also strongly pHdependent. Griffin and Shimp (1978) observed an increase in
13

sorption of As (III) by kaolinite and montmorillonite over a pH

range of 3-9. Pierce and Moore (1980) found the maximum

adsorption of As(III) by iron oxide occurred at pH 7. Elkhatib

et al. (1984b) found adsorption of As(III) to be rapid and

irreversible on ten soils. They determined, in this study and

another study (Elkhatib et al., 1984a), that Fe oxide, redox,

and pH were the most important properties in controlling

arsenite adsorption by these soils.

Both pH and the redox are important in assessing the fate of

arsenic in soil. At high redox levels, As(V) predominates and

arsenic mobility is low. As the pH increases or the redox

decreases As (III) predominates. The reduced form of arsenic

is more subject to leaching because of its high solubility. The

reduction kinetics are, however, slow. Formation of As (III)

also may lead to the volatilization of arsine (AsH3) and methylarsines from soils (Woolson 1977a). Under soil conditions of

high organic matter, warm temperatures, adequate moisture,

and other conditions conducive to microbial activity, the

reaction sequence is driven towards methylation and volatilization (Woolson 1977a). Woolson's (1977b) study showed

that only 1 to 2 percent of the sodium arsenate applied at a

rate of 10 ppm was volatilized in 160 days. The loss of

organic arsenical compounds from the soil was far greater

than for the inorganic source of arsenic. Arsenite, As(III), can

be oxidized to As(V). Manganese oxides are the primary

electron acceptor in this oxidation (Oscarson et al., 1983).

Selenium

The behavior of selenium in soils has received great attention

in recent years. Studies were stimulated by the high incidence

of deformity and mortality of waterfowl at the Kesterson

National Wildlife Refuge in California that resulted from the

input of agricultural drainage water from the western San

Joaquin Valley that was high in Se. Such studies have led to

a better understanding of the distribution and movement of Se

in soils and ground water.

Selenium exists in the soil environment in four oxidation

states: selenide (Se2-), elemental selenium (Seo), selenite

(SeO3

2-), and selenate (SeO42-). The concentration and form

of Se in soil is governed by pH, redox, and soil composition.

Selenate, Se(VI), is the predominant form of selenium in

calcareous soils and selenite, Se(IV), is the predominant form

in acid soil.

Selenite, Se (IV) binds to sesquioxides, especially to Fe

oxides. Balistriera and Chao (1987) found the removal of

selenite by iron oxide to increase with decreasing pH. This

study not only demonstrates the effect of pH on selenite

adsorption but also the effect of concentration. The decrease

in the percentage of selenite adsorbed with increasing

concentration of selenite at a given pH indicated multiple sites

of selenite retention. At the two lower concentrations, high

energy specific adsorption sites were available. As the

concentration of selenite was increased these sites became

saturated and the lower energy sites were utilized. Griffin and

Shimp (1978) found maximum adsorption of selenite on

montmorillonite and kaolinite to occur at pH 2-3. Neal et al.

(1987a) used five soils from the San Joaquin Valley and found

that selenite adsorption by the soils decreased with increasing

pH in the range of 4-9. Selenite adsorption to oxides and soils

occurs through an inner sphere complexation (specific

adsorption) mechanism (Rajan, 1979; Neal et al., 1987b).

In studies of competitive adsorption using phosphate, sulfate,

and chloride (Neal, et al., 1987b) and phosphate and various

organic acids (Balistrieri and Chao, 1987), selenite adsorption

decreased dramatically in the presence of phosphate and the

organic acids but was not affected by the presence of sulfate

or chloride. Balistrieri and Chao (1987), using Fe oxide, found

the sequence of adsorption to be: phosphate = silicate =

arsenate > bicarbonate carbonate > citrate = selenite >

molybdate > oxalate > fluoride = selenate > sulfate.

Precipitation is not a major mechanism of retention of selenite

in soils. Manganese selenite may form, however, in strongly

acidic environments (Elrashidi et al., 1989).

Selenate dominates under alkaline conditions. In contrast to

selenite, selenate, Se(VI), is highly mobile in soils. Benjamin

(1983) found that selenate was adsorbed by amorphous iron

oxide as a function of pH. Maximum removal was at pH 4.5

and adsorption decreased with increasing pH. Bar-Yosef and

Meek (1987) found some indication of selenate adsorption by

kaolinite below pH 4. Selenate seems to be adsorbed by

weak exchange mechanisms similar to sulfate (Neal and

Sposito, 1989), in contrast to selenite that is specifically

adsorbed by soils and soil constituents. There has been some

evidence that selenate was adsorbed by alkaline soils (Singh

et al., 1981), but Goldberg and Glaubig (1988) found no

removal of selenate by calcareous montmorillonite. Neal and

Sposito (1989), using soils from the San Joaquin Valley,

showed no adsorption of added selenate over a pH range

from 5.5-9.0. Fio et al. (1991) also observed no sorption of

selenate by alkaline soil from the San Joaquin Valley, but did

observe the rapid sorption of selenite by this soil. No stable

precipitates of selenate are expected to form under the pH

and redox conditions of most soils (Elrashidi, et al., 1989).

Similar to other anionic species, selenium is more mobile at

higher pHs. Soil factors favoring selenium mobility, as

summarized by Balistrieri and Chao (1987) are; alkaline pH,

high selenium concentration, oxidizing conditions, and high

concentrations of additional anions that strongly adsorb to

soils, in particular phosphate.

Under reduced conditions, selenium is converted to the

elemental form. This conversion can provide an effective

mechanism for attenuation since mobile selenate occurs only

under well aerated, alkaline conditions.

Organic forms of selenium are analogous to those of sulfur,

including seleno amino acids and their derivatives. Like sulfur,

selenium undergoes biomethlyation forming volatile methyl

selenides.

Chromium

Chromium exists in two possible oxidation states in soils: the

trivalent chromium, Cr(III) and the hexavalent chromium,

Cr(VI). Forms of Cr(VI) in soils are as chromate ion, HCrO4

-

predominant at pH<6.5, or CrO4

2-, predominant at pH 6.5, and

as dichromate, Cr2O7

2- predominant at higher concentrations

(>10mM) and at pH 2-6. The dichromate ions pose a greater

health hazard than chromate ions. Both Cr(VI) ions are more

toxic than Cr(III) ions. Reviews of the processes that control

the fate of chromium in soil and the effect these processes

have on remediation are given in Bartlett (1991) and Palmer

and Wittbrodt (1991).
14

Because of the anionic nature of Cr(VI), its association with

soil surfaces is limited to positively charged exchange sites,

the number of which decreases with increasing soil pH. Iron

and aluminum oxide surfaces will adsorb CrO4

2- at acidic and

neutral pH (Davis and Leckie, 1980; Zachara et al., 1987;

Ainsworth et al., 1989). Stollenwerk and Grove (1985)

concluded that the adsorption of Cr(VI) by ground-water

alluvium was due to the iron oxides and hydroxides coating

the alluvial particles. The adsorbed Cr(VI) was, however,

easily desorbed with the input of uncontaminated ground

water, indicating nonspecific adsorption of Cr(VI). The

presence of chloride and nitrate had little effect on Cr(VI)

adsorption, whereas sulfate and phosphate inhibited

adsorption (Stollenwerk and Grove, 1985). Zachara et al.

(1987) and Zachara et al. (1989) found SO4

2- and dissolved

inorganic carbon inhibited Cr(VI) adsorption by amorphous

iron oxyhydroxide and subsurface soils. The presence of

sulfate, however, enhanced Cr(VI) adsorption to kaolinite

(Zachara et al., 1988). Rai et al. (1988) suggested that

BaCrO4 may form in soils at chromium contaminated waste

sites. No other precipitates of hexavalent compounds of

chromium have been observed in a pH range of 1.0 to 9.0

(Griffin and Shimp, 1978). Hexavalent chromium is highly

mobile in soils.

In a study of the relative mobilities of 11 different trace metals

for a wide range of soils, Korte et al. (1976) found that clay

soil, containing free iron and manganese oxides, significantly

retarded Cr(VI) migration (see Figure 6). Hexavalent

chromium was found to be the only metal studied that was

highly mobile in alkaline soils. The parameters that correlated

with Cr(VI) immobilization in the soils were free iron oxides,

total manganese, and soil pH, whereas the soil properties,

cation exchange capacity, surface area, and percent clay had

no significant influence on Cr(VI) mobility.

Rai et al. (1987) reported that Cr(III) forms hydroxy complexes

in natural water, including Cr(OH)2

+, Cr(OH)2+, Cr(OH)3o, and

Cr(OH)4

-

. Trivalent chromium is readily adsorbed by soils. In

a study of the relative mobility of metals in soils at pH 5, Cr(III)

was found to be the least mobile (Griffin and Shimp, 1978).

Hydroxy species of Cr(III) precipitate at pH 4.5 and complete

precipitation of the hydroxy species occurs at pH 5.5.

Hexavalent chromium can be reduced to Cr(III) under normal

soil pH and redox conditions. Soil organic matter has been

identified as the electron donor in this reaction (Bartlett and

Kimble, 1976; Bloomfield and Pruden, 1980). The reduction

reaction in the presence of organic matter proceeds at a slow

rate at environmental pH and temperatures (Bartlett and

Kimble, 1976; James and Bartlett, 1983a,b,c). Bartlett (1991)

reported that in natural soils the reduction reaction may be

extremely slow, requiring years. The rate of this reduction

reaction, however, increases with decreasing soil pH (Cary et

al., 1977; Bloomfield and Pruden, 1980). Soil organic matter

is probably the principal reducing agent in surface soils. In

subsurface soils, where organic matter occurs in low

concentration, Fe(II) containing minerals reduce Cr(VI) (Eary

and Rai, 1991). Eary and Rai (1991), however, observed that

this reaction only occurred in the subsurface soil with a pH<5.

The reduction of Cr(VI) occurred in all four subsurface soils

tested by decreasing the pH to 2.5.

Bartlett and James (1979), however, demonstrated that under

conditions prevalent in some soils, Cr(III) can be oxidized.

The presence of oxidized Mn, which serves as an electron

acceptor, was determined as an important factor in this

reaction.

Industrial use of chromium also includes organic complexed

Cr(III). Chromium (III) complexed with soluble organic ligands

will remain in the soil solution (James and Bartlett, 1983a). In

addition to decreased Cr(III) adsorption, added organic matter

also may facilitate oxidation of Cr(III) to Cr(VI).

Computer Models

Several equilibrium thermodynamic computer programs are

available for modeling soil solution and solid phase chemistry

by proving information on the thermodynamic possibility of

certain reaction to occur. In addition to calculating the

equilibrium speciation of chemical elements in the soil solution

and precipitate/dissolution reactions, models such as

GEOCHEM (Mattigod and Sposito, 1979) and MINTEQA2

(USEPA, 1987) provide information on cation exchange

reactions and metal ion adsorption. These models are used

to:

1) calculate the distribution of free metal ions and metalligand complexes in a soils solution,

2) predict the fate of metals added to soil by providing a

listing of which precipitation and adsorption reactions are

likely to be controlling the solution concentration of

metals, and

3) provide a method for evaluating the effect that changing

one or more soil solution parameters, such as pH, redox,

inorganic and organic ligand concentration, or metal

concentration, has on the adsorption/precipitation

behavior of the metal of interest.

These models are equilibrium models and as such do not

consider the kinetics of the reactions. These models are also

limited by the accuracy of the thermodynamic data base

available.

Analysis of Soil Samples

Total concentration of metals in soil

Measurement of the total concentration of metals in soils is

useful for determining the vertical and horizontal extent of

contamination and for measuring any net change (leaching to

ground water, surface runoff, erosion) in soil metal

concentration over time. The methods do not, however, give

an indication as to the chemical form of the metal in the soil

system.

The complete dissolution of all solid phase components in

soils requires a rigorous digestion using either a heated

mixture of nitric acid, sulfuric acid, hydrofluoric acid, and

perchloric acid (Page et al., 1982) or a fusion of the soil with

sodium carbonate (Page et al., 1982). Both methods require

special equipment and special safety considerations. A more

commonly used procedure is the hot nitric acid-hydrogen

peroxide procedure outlined in SW-846 Method 3050

(USEPA, 1986). This is a partial digestion of the soil solid

phase. The method probably releases metals associated
15

with a recent pollution source, i.e., exchangeable, specifically

adsorbed to clays, oxides or organic matter, and most

precipitates, but would not release metals associated within

solid phases that are not dissolved by the hot nitric acid and

oxidizing agent, i.e., within the structure of insoluble minerals.

Sequential extractions of metals in soils

Since the potential migration of metals in soil systems is

dependent on the chemical form of the metal, extraction

procedures have been developed to selectively remove

metals from these various geochemical forms. While these

procedures cannot be used to identify the actual form of a

given metal in a soil, they are useful in categorizing the metals

into several operationally defined geochemical fractions, such

as exchangeable, specifically adsorbed, and metals

associated with carbonates, organic matter, and/or iron and

manganese oxides.

Numerous extraction procedures have been developed for

metal cations (Sposito et al., 1984; Hickey and Kittrick, 1984;

Tessier et al., 1979; Grove and Ellis, 1980; Kuo et al., 1983)

and anions (Chao and Sanzolone, 1989; Gruebel et al., 1988).

Lake et al. (1984) reviewed a number of the procedures used

for cationic metal extraction. The extraction procedures

consist of reacting a soil sample with increasing strengths of

chemical solutions. Typically water or a salt solution (KNO3,

CaCl2, etc.) is the first extractant used. These are followed by

mild acids, bases, chelating agents, and oxidizing solutions.

Table 4 illustrates the wide variety of extractants that have

been used in the literature for metal cations.

The aqueous fraction and those fractions in equilibrium, i.e.,

the exchange fraction, with this fraction are of primary

importance when considering the migration potential of metals

in soils. In theory, mild extractants, such as salt solutions, are

more likely to extract metals that could be released to the soil

solution with input of water than metals associated with

stronger binding mechanisms, such as specifically adsorbed

or precipitated metals. Work by Silveira and Sommers (1977)

and Latterell et al. (1978) suggests that salt extractable metals

represent the potentially mobile portion of the total

concentration of metals in soils. Harrison et al. (1981) likewise

suggested that the mobility of metals decreases in the order of

the extraction sequence. Rigorous evaluation, however, of the

appropriateness of any extraction procedure for defining the

mobile fraction of metals in soils has not been reported in the

literature.

Hickey and Kittrick (1984) used a sequential extraction

procedure to separate Cd, Cu, Ni, and Zn in metal polluted

soils and sediments into five operationally defined

geochemical fractions: exchangeable (1.0M MgCl2), metals

associated with carbonates (acetate buffer, pH 5), metals

associated with Mn and Fe oxides (0.04M NH4OH•HCl),

metals associated with organic matter (0.02M HNO3+H2O2),

and residual metals (HF+HClO4). Figure 10 shows the

average distribution of the metals among the defined

geochemical fractions. Approximately 37% of the total Cd and

a significant portion of the Zn were in the exchange fraction

indicating the potential mobility of these two metals. Only a

small portion of the Cu and Ni were in the exchange fraction.

A significant portion of the Cu was associated with the organic

fraction, in agreement with the known affinity of Cu for organic

material. Nickel was mostly associated with the residual

Figure 10. Proportions of Cd, Cu, Ni, and Zn in each of the

 operationally defined geochemical fractions of the

 experimental samples (Hickey and Kittrick, 1984).

fraction and significant portions of Zn and Cu were associated

with the oxides. The authors concluded from this study that

the relative mobility of the metals followed: Cd > Zn > Cu =

Ni.

There has been recent criticism of these sequential extraction

procedures (Miller et al., 1986; Kheboian and Bauer, 1987;

Gruebel et al., 1988; Tipping et al., 1985; Rapin et al., 1986;

Calvet et al., 1990). The methods are not entirely specific for

a geochemical fraction of the soil and the extractant also may

remove metals associated with other fractions. Secondly,

readsorption of the extracted metals to the remaining solid

phase of the soil may occur leading to artificially low

concentrations of the metal being associated with that fraction.

Finally, no one extraction procedure would be universally

applicable for all metals and all soils. Perhaps the most

suitable extractant for defining the mobile fraction of metals in

soils under specific site conditions is one that simulates that

soil or soil-waste solution chemistry.

TCLP

The Toxicity Characterization Leaching Procedure (TCLP)

(USEPA, 1986) is a single extraction procedure, using 0.1M

acetic acid, developed to simulate the leaching a waste might

undergo if disposed of in a municipal landfill. This method is

frequently used to determine the leaching potential of cationic

metals in landfill situations where, due to microbial

degradation of the waste under anaerobic conditions, acetic

acid is produced. While this procedure is appropriate for

demonstrating whether an excavated metal contaminated soil

is defined as hazardous for disposal at a landfill, its application

Cd Cu

Ni Zn

Exchange

Carbonate

Residual

Organic

Oxide

Exchange

Exchange

Exchange

Residual

Residual

Residual

Oxide

Oxide

Organic

Organic

Organic

Carbonate

Carbonate

Oxide Carbonate
16

Table 4. Some bibliographic data on the extraction of heavy metals present in soils and sediments

 (Colvet, et al., 1990)

Authors Exchangeable Fraction associated Fraction associated Fraction associated Total amount and

with carbonates with oxides with organic matter residual fraction

McLaren Crawford11 0.05 N CaCI2 2.5% CH3COOH 0.1 M (COOH)2 1 M K4P2O7 HF

 (1973) +0.175 M (COONH4 ) 2

pH = 3.5

Stover et al.12 1 M KNO3 0.1 M EDTA 0.1 M Na4P2O7 1 M HNO3

 (1976) +NaF pH = 6.5

Gatehouse et al.13 1 M CH3COONH4 0.1 M NH2OH 30% H2O2 HF-HClO4

 (1977) +CH3COOH pH = 4.5 +1 M CH3COONH4 pH = 4.5

Filipek and Owen14 1 M CH3COOH 0.25 M NH2OH, HCl Acidified HNO3-HF-HClO4

 (1979) in 25% (v/v) CH3 COOH 30% H2O2

Tessier et al.

3 1 M MgCl2 or 1 M CH3COONa 0.04 M NH2OH, HCl 0.02 M HNO3+ HF-HClO4

 (1979) 1 M CH3COONa + 1 M CH3COOH in 25% (v/v) CH3COOH 30% H2O2, pH = 2

at pH = 8.2 at pH = 5.0 at 96 ± 3°C at 85 ± 2°C, 2 h

 or 0.3 M Na2S2O4 +30% H2O2+HNO3, pH = 2

 +0.175 M Na—citrate at 85 ± 2°C, 3 h

 +0.025 M citric acid 3.2 M CH3COONH4

in 20% HNO3

Förstner et al. 2 0.2 M BaCI2 0.1 M NH2OH, HNO3 + 30% H2O2 + NH4OH HF-HClO4

 (1979) 25% (v/v) CH3COOH + HCI

Schalscha et al.4 1 M KNO3 0.5 M NaF 0.1 M EDTA pH = 6.5 0.1 M Na4P2O5 1 M HNO3

 (1980) pH = 6.5 double extraction

Garcia-Miragaya15 1 N CaCI2 2.5% CH3COOH 0.05 M EDTA 0.1 N Na4P2O5 HF

 (1981) pH = 7

Badri and Aston16 1 M CH3COONH4 0.25 M NH2OH, HCI 30% H2O2

 (1981 ) +0.5 M (CH3COO)2 Mg pH = 2 + 1 M CH3COONH4

Förstner et al.17 1 M CH3COONH4 (1) 0.1 M NH2OH, ClH 30% H2O2, HNO3 HNO3 at 180°C

 (1981) pH = 7 +0.0I M HNO3, pH = 2 pH = 2 at 85°C

(2) 0.2 M (COONH4)2 extraction with

 +0.2 M (COOH)2, pH = 3 1 M CH3COONH4

Greffard et al.6 resin-H+ (1) (COONa)2 30% H2O2 at 40°C

 (1981) (2) (COONa)2 + UV

Sposito et al.10 0.5 M KNO3 0.5 M Na2 EDTA 0.5 M NaOH 4 M HNO3 at 80°C

Dekeyser et al.18 1M CH3COONH4 (1) 0.1 M NH2OH, HCl HNO3-HF-HCl

 (1983) pH = 4.5 (2) 0.2 M (COONH4)2

 (HCOOH)2, pH = 3.3

 obscurité

(3) Same as (2)+UV

Kuo et al.7 1 M MgCI2 (1) (COONa)2 6% NaClO4 at 85°C HNO3-HCIO4

 (1983) (2) Citrate dithionite

 bicarbonate

Meguelatti et al.5 1 M BaCI2 1 M CH3COOH 0.1 M NH2OH 30% H2O2 HF-HCI

 (1983) +0.6 M CH3COONa +25% (v/v) CH3COOH +0.02 M HNO3

+3.2 M CH3COONH4

Shuman19 1 M Mg(NO3)2 (1) 0.1 M NH2OH, HCI 0.7 M NaOCI HF-HNO3-HCI

 (1985) pH = 7 pH = 2 pH = 8.5

(2) 0.2 M (COONH4)2

 +0.2 M (COOH)2, pH = 3

(3) Same as (2) + ascorbic

 acid

Gibson and Farmer20 1 M CH3COONH4 1 M CH3COONa (1) 0.1 M NH2OH, CIH 30% H2O2 Aqua regia

 (1986) pH = 7 pH = 5 +0.01 M HNO3 +0.02 M HNO3 + HF

(2) 1 M NH2OH, CIH in at 85 °C

 25% (v/v) CH3COOH
17

for evaluating the mobility of metals under field conditions has

been questioned (Dragun et al., 1990). Production of acetic

acid does not commonly occur in soils. In certain soil-waste

systems, leaching tests using acetic acid may be appropriate,

but it is not universally representative of the leaching solution

for soil-waste systems. The acetic acid leaching procedure

was developed for cationic metals. The procedure is not

appropriate for extraction of anionic metals. Bartlett (1991)

reported that this procedure actually causes the reduction of

Cr(VI) to Cr(III) leading to a false measurement of the

leachability of Cr(VI) in soil. A more appropriate leaching

solution would mimic the specific waste or waste-soil matrix.

Hickey and Kittrick (1984) used an acetate buffer solution in

their sequential extraction scheme to remove metals

associated with carbonates. This is a similar solution to the

TCLP solution except that it is buffered to pH 5. This buffered

solution fully dissolves the carbonate minerals in the soil. The

unbuffered acetic acid solution used in the TCLP solution

cannot maintain a low enough pH in calcareous soils to

dissolve carbonates. The metals extracted by the TCLP

solution are not related to any definable geochemical fraction

and the fraction of metals extracted using this procedure have

not been correlated with the mobile fraction of metals in soil.

Evaluating the Behavior of Metals in Soils

Sorption studies

Soil sorption studies are commonly performed to evaluate the

extent of metal retention by a soil or soil constituent. Sorption

studies are often used in an attempt to generate the

equilibrium distribution coefficient (Kd), the ratio of metal

sorbed to metal in solution at equilibrium, which may be

utilized in transport models. Sorption studies are also used for

comparison of the relative retention of several metals by a soil

or the relative retention of a metal by several soils, and are

used extensively in correlation studies to determine the

relative importance of a soil's chemical and physical properties

for metal retention. Sorption studies also can be used to

evaluate the effect that changing a soil solution parameter,

e.g., adjustment of pH, ionic strength, addition of competing

cations, or addition of inorganic or organic ligands, has on

metal retention by a soil.

In a sorption study, the soil is reacted with solutions containing

varying quantities of the metal(s) of interest for a specified

time period using either batch or column techniques. The

concentration range used in the study should overlap the

concentration of environmental concern. A background

electrolyte solution also should be used to simulate normal

soil’s solution chemistry or the waste matrix and to equalize

the ionic strength across all soils. The reaction time should

approach thermodynamic equilibrium, usually determined by a

preliminary kinetics experiment. After the specified time

period the soil and solution are separated by centrifugation

and/or filtration. The soil and/or solution phases are then

analyzed by atomic absorption spectrophotometry or

inductively coupled plasma emission spectrometry. With

these techniques it is not possible to distinguish between true

adsorption and precipitation reactions. For that reason the

term sorption will be used.

Two techniques, batch and column studies, may be used to

generate sorption isotherms. The batch technique involves

placing the soil and the solutions containing the various

concentrations of the metals into a vessel and mixing the

samples for a prescribed time period. This is the most

commonly used technique because of its ease of laboratory

operation and ease of data handling. The disadvantages of

the technique are 1) results are sensitive to the soil:solution

ratio used, 2) soil:solution ratios in actual soil systems cannot

be done in batch studies, so scaling of data from batch studies

to soils systems is uncertain, 3) results are sensitive to the

mixing rate used, 4) separation techniques may affect results,

and 5) many investigators have found that batch generated

sorption coefficients are not adequate to describe the behavior

of metals in flow through systems.

The column method consists of packing a glass or plastic

column with soil. The solutions containing various

concentrations of the metals of interest are pumped through

the columns and the effluents are collected and analyzed by

AA or ICP. Breakthrough occurs when the effluent

concentration equals the influent concentration. The

advantages of this technique are 1) low soil:solution ratios can

be used, 2) separation of the soil and solution phase is not

required, 3) mechanical mixing is not required and 4) column

studies more closely simulate field conditions than batch

methods. The disadvantages are 1) results depend on flow

rates used, 2) columns are difficult to set-up and maintain, 3)

uniform packing of the column is difficult often leading to

channel flow, and 4) fewer columns can be operated at one

time compared with the number of batch reactors.

Equilibrium sorption is described by a sorption isotherm. A

sorption isotherm is the relationship between the amount of

metal sorbed and the equilibrium concentration of the metal

or, more correctly, the activity of the free metal in the soil

solution. A typical sorption isotherm is shown in Figure 11. If

the relationship is linear over the concentration range studied

then the sorption process can be described by a single

coefficient, the distribution coefficient, Kd. For metals,

however, the relationship is seldom linear and other equations

with two or more coefficients must be used to describe the

data.

Curvilinear isotherm

Linear isotherm

S = CKd

C mg/L

S mg/kg

5

4

3

2

1

0 100 200

Figure 11. Sorption isotherms.
18

Equations most frequently used, because of their relative

simplicity, to describe the curvilinear sorption behavior of

metals in soil are the Langmuir and the Freundlich equations.

The Langmuir equation was developed to model gas

adsorption on solid surfaces. The derivation of the equation

was based on the assumption that adsorption is independent

of surface coverage, that there is no interaction between

adsorbed ions, and that only a monolayer of adsorption occurs

on the surface. These conditions are not typically met with

metals sorption on soils.

The linearized form of the Langmuir equation is:

 C = C + 1 (1)

 S M Mb

where C is the concentration or activity of the free metal in

solution, S is the quantity of the metal ion sorbed by the soil

(i.e., mg metal sorbed/Kg soil), M is the maximum sorption

capacity of the soil, and b is the coefficient related to bonding

energy. When C/S is plotted as a function of C, the slope is

the reciprocal of the sorption capacity, M, and the intercept is

1/Mb.

The Freundlich expression is an empirically derived equation

to describe the logarithmic decrease in adsorption energy with

increasing surface coverage. The linearized form of the

Freundlich equation is:

log S =(N)log C + log K (2)

where S and C have the same definition as above and N and

K are constants fitted from the experimental data. When the

slope, N, equals 1, the equation simplifies to:

S=CKd or Kd = S/C (3)

where Kd is the distribution coefficient. In most studies

reported in the literature for metal sorption, the slope of the

Freundlich isotherm is seldom equal to 1 and the simplified

expression and its single term, Kd, are not appropriate to

describe the data.

Figure 12 illustrates the use of the Langmuir expression to

describe Cu sorption by a soil (Cavallaro and McBride, 1978).

The equation describes the behavior of Cu over all

concentrations used in this study. Often, however, nonlinear

behavior over the concentration range studied is observed

with the use of either the Langmuir or Freundlich equations.

In Figure 13, the Langmuir expression was used to describe

the sorption behavior of Cd and Zn by hydrous manganese

oxide (Zasoski and Burau, 1988). This non-linear behavior

when using the Langmuir equation has been noted by

numerous researchers using various metals and soils and soil

constituents (Benjamin and Leckie, 1981; Shuman, 1975;

Loganathan and Burau, 1973). Non-linearity of metal sorption

using the Freundlich equation has also been noted by Zasoski

and Burau (1988) Benjamin and Leckie (1981), Catts and

Langmuir (1986), O'Connor et al. (1984), O'Connor et al.

(1983), and Elrashdi and O'Connor (1982). This non-linear

behavior has been interpreted to indicate multiple sites of

sorption that have different energies of retention. The

mechanisms at low concentrations have been attributed to

specific adsorption, whereas the mechanisms at higher

concentrations have been considered to be exchange

1 2 4 6 8 10 12 14

4

8

12

16

20

24 Lansing A

r

2 = .99

[Cu2+] x 106 )M)

c/x/m x 107 (M/ma/100g)

Figure 12. Langmuir adsorption isotherm for Cu2+ adsorption on

 the Lansing A soil (Cavallaro and McBride 1978).

0.8

0.6

0.4

0.2

0.0

1 35

Zn

X/M Cd

(mol kg-1)

(X/M) Meeq

(m3kg-1 • 103)

7 9 35

-1

Figure 13. Langmuir plots of Cd and Zn sorbed on δ- MnO2 for

 the noncompetitive pH 4 data (Zasoski and Burau,

 1988).

reactions or precipitation. These results illustrate the

importance of generating sorption data over the concentration

range of interest for a particular application. Large error in

predicting sorptive behavior may result from using data

generated in one system and applied to a system with higher

or lower metal concentration.

Several researchers have, however, suggested that other

equations, for example the two-surface Langmuir equation

(Sposito, 1982; Travis and Etnier, 1981) or the competitive

Langmuir equation (Griffin and Au, 1977; Travis and Etnier,
19

1981), be used to describe the non-linear behavior

encountered with the Langmuir equation (Sposito, 1982).

The Langmuir and Freundlich isotherm expressions have

proven valuable in interpreting metal behavior in soils. The

adsorption isotherm equations were, however, developed for

modeling gas adsorption on solids. The sorption of metals by

soils violates many of the assumptions associated with these

equations. Also, the mechanism described by these

equations is adsorption, but it is impossible in a soil system to

distinguish between adsorption and precipitation reactions.

Adsorption isotherm equations should not be used to indicate

adsorption mechanisms without collaborative evidence, but

they can be used for an empirical description of the data.

Harter (1984) warned against over interpreting the sorption

maximum and "bonding energy" determined using the

Langmuir equation. The applicability of adsorption isotherm

equations to the interpretation of soil chemical phenomena is

a subject of controversy. For further discussion of this

controversy see Elprince and Sposito (1981), Griffin and Au

(1977), Veith and Sposito (1977), Sposito (1979), Harter and

Baker (1977), and Harter (1984).

Desorption

Desorption studies are often performed to determine the

reversibility of the sorption reactions. This gives an idea of the

strength of the association of the metal with the soil surface.

An example of the reversibility of a sorption reaction (Figure

14) is taken from Dudley et al. (1988). In this study, two

calcareous soils were reacted with various concentrations of

Cd. The soils were then desorbed with CaCl2. For the low

carbonate soil (Kidman) virtually all of the sorbed Cd was

desorbed by the Ca. Only 10-15 percent of the sorbed Cd on

the highly calcareous Skumpah soil was desorbed by Ca.

These results suggest that Cd was held by the Kidman soil as

an exchangeable cation, whereas in the Skumpah, Cd was

specifically adsorbed by the CaCO3.

Desorption studies are performed after completion of the

sorption study. They can be carried out using either batch or

Figure 14. Desorption of Cd from Kidman and Skumpah soil by

 0.01 M CaCl2 at a soil:solution ratio of 1:25 (Dudley, et

 al., 1988).

column techniques. For the batch technique the soils used in

the sorption are reacted with a salt solution, typically 0.01N

CaCl2 or a matrix representative of the soil-waste system

being studied. Samples are shaken for a specified time

period. The soil and liquid are then separated by

centrifugation and/or filtration and the solution is analyzed for

the metals by AA or ICP. This process is repeated several

times. For a column study, the metal equilibrated soil column

is flushed with an appropriate solution until the system

reaches steady state conditions.

Kinetics

Attention has been mainly given in the literature to equilibrium

processes in soils but soil processes are never at equilibrium.

Soil systems are dynamic and are thus constantly changing.

Most kinetic studies have been performed to establish the

proper time interval for use in equilibrium sorption/desorption

studies. Most studies assume that ion exchange processes

are rapid in soils and that 16 to 24 hours mixing periods,

common time periods used in sorption studies, are adequate.

This assumption may not be appropriate if other reactions in

addition to simple ion exchange, i.e., specific adsorption and

precipitation, are involved in metals retention (Harter and

Lehmann, 1983). McBride (1980) found that the initial

adsorption of Cd on calcite was very rapid, while CdCO3

precipitation of higher Cd2+ concentrations was slow.

Lehmann and Harter (1984) used kinetics of desorption to

study the strength of Cu bonding to a soil. A plot of

concentration of Cu in solution versus time indicates an initial

rapid release of the Cu followed by a slow reaction. They

interpreted these results to indicate that Cu was held at two

sites: the rapidly released Cu being loosely held on the soil

surfaces and the slowly released Cu being tightly bound.

Each metal-soil system should be tested to determine the time

necessary for the individual system to come to equilibrium.

Figure 15 illustrates the different time periods required for

equilibrium for three metals sorbed by a calcareous soil

(Dudley, et al., 1988). Copper reached equilibrium within a

few hours whereas Cd and Zn did not approach steady state

conditions for 144 hours. Use of the time interval appropriate

for Cu equilibrium for this soil would mean that only 50 percent

of the Cd and Zn adsorbed under steady state conditions

would have been determined.

Kinetic studies are being more widely performed because of

their importance in determining the transport of metals in soil

systems. Many mathematical transport models now allow a

kinetic term for sorption. Equilibrium studies predict whether a

reaction will occur but give no indication of the time necessary

for the reaction to take place. Kinetic studies also contribute

to an understanding of reaction mechanisms not discernible

from thermodynamic studies (Zasoski and Burau, 1988; Harter

and Smith, 1981; Sparks, 1989).

Kinetic studies are similar to sorption procedures, using either

batch or column techniques, except samples are collected

over time. Several equations have been used to describe the

kinetics of sorption reactions of ions on soils and soil

constituents. These equations include: first-order, secondorder, Elovich equation, parabolic diffusion equation, and

power function equation. An excellent review of kinetic

processes in soil systems is given by Sparks (1989).

0

0

50

100

150

200 400 600 800 1000

Kidman

Skumpah

Cd desorbed by Ca 

μ

M

Cd in solid phase before desorption μM
20

may enhance metal mobility. The extent of vertical

contamination is intimately related to the soil solution and

surface chemistry of the soil matrix with reference to the metal

and waste matrix.

Laboratory methods for evaluating the behavior of metals in

soils are available in the literature. Thermodynamic

equilibrium computer models are also available to assist with

this evaluation. The advantages and disadvantages of some

of the available procedures have been presented in this

document.

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1.0

0.9

0.8

0.7

0.6

0.5

0.4

0.3

0.2

0.1

0.0

0 24 48 72 96 120 144 168 192 216 240 264 288 312 336 360

= Zn

= Cd

= Cu

Fi 17 Ch ith ti i d d t ti

Time (hours)

C/Co

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The retention mechanisms for metals added to soil include

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