United States
Environmental Protection
Agency
Office of
Research and
Development
Off ice of Sol id Waste
and Emergency
Response
EPA/540/S-92/018
October 1992
&EPA Ground Water Issue
Behavior of Metals in Soils
Joan E. McLean* and Bert E. Bledsoe*
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 con-
tamination. 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
Superfund Technology Support Center for
Ground Water
Robert S. Kerr Environmental
Research Laboratory
Ada, Oklahoma
Technology Innovation Office
Office of Solid Waste and Emergency
Response, US EPA, Washington, DC
Walter W. Kovalick, Jr., Ph.D.
Director
-------
Table 1. Content of Various Elements in Soils (Lindsay, 1979)
Metal
Al
Fe
Mn
Cu
Cr
Cd
Zn
As
Se
Ni
Ag
Pb
Hg
Selected Average
for Soils
mg/kg
71,000
38,000
600
30
100
0.06
50
5
0.3
40
0.05
10
0.03
Common Range
for Soils
mg/kg
10,000-300,000
7,000-550,000
20-3,000
2-100
1-1000
0.01-0.70
1 0-300
1 .0-50
0.1-2
5-500
0.01-5
2-200
0.01-0.3
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 overtime, 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
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, oxidation-
reduction 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°,
ZnCI+, CdClj"), 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°, CdHCO3+, Cr(OH)4'. The associated atoms or
molecules are termed ligands. In the above examples, SO42',
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 SO42', Cl~, OH", PO43',
NO3" and CO32~. 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
-------
Mass Transfer
Ion
Exchange
and
Adsorption
Precipitation
and
Dissolution
of Solids
Free Metal
Concentration
in Soil Soltuion
Redpx
Reaction
Acid-Base
Reaction
Complex
Formation
Figure 1. Principal controls on free trace metal concentrations in
soils solution (Mattigod, et al., 1981).
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., CdCI3+, CdCI", CdCI2°).
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(lll), selenium,
Se(VI) and Se(IV), chromium, Cr(VI) and Cr(lll), and mercury,
Hg(ll) and Hg(l). 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.
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 0.45um 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
10um (Sposito, 1989). Gschwend and Reynolds (1987)
reported that colloidal particles of intermediate diameter,
0.1um to 1um, 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 co-
precipitation, 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
-------
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)
2-
PH
Figure 2. The solubility diagram for Pb in Nibley clay loam soil
(Santillan-Medrano and Jurinak, 1975).
T3
o
10
6.0 6.5 7.0 7.5 8.0 8.5 9.0
PH
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
-------
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 (Santillan-
Medrano and Jurinak, 1975; Cavallaro and McBride, 1978;
McBride, 1980; Jurinak and Bauer, 1956; McBride and
Bouldin, 1984; Dudley etal., 1988; Dudley etal., 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 AI3+
for Si4+ in the tetrahedral layers and/or substitution of Mg2+,
Fe2+, etc. for AI3+ 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-OH°]. 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;
Latterelletal., 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 etal., 1986;
O'Connor etal., 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.
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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
(AsO43' and AsO2'), selenium (SeO32' and SeO42'), and
chromium in one of its oxidation states (CrO42~). 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 (SeO32~) and
arsenate (AsO43~) are adsorbed to oxides and soils through
specific binding mechanisms (Rajan, 1979; Neal, et al.,
1987b). Selenite (SeO42~) 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(ll)
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(lll) > As(V) > Cd > Zn >
Pb>Cu>Cr(lll).
Table 2. Relative affinity of metals for soils and soil constituents
Soil or Soil
Constituent
Relative Order of Sorption
Reference
goethite
Fe oxide
montmorillonite
kaolinite
soils
soils
mineral soils
organic soils
soil
Cu>Pt»Zn>Co>Cd
Pb>Cu>Zn>Cd
Cd=Zn>Ni
Cd>Zn>Ni
Pb>Cu>Zn>Cd>Ni
Zn>Ni>Cd
Pb>Cu>Zn>Cd
Pb>Cu>Cd>Zn
Pb>Cu>Zn>Ni
Forbes et al., 1976
Benjamin and Leckie, 1981
Pulsand Bonn, 1988
Pulsand Bonn, 1988
Biddappa et al., 1981
Tiller etal., 1984
Elliott et al., 1986
Elliott et al., 1986
Harter, 1983
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Table 3. Characteristics of the soils (Korte et al., 1976)
Soil
Wagram
Ava
Kalkaska
Davidson
Molokai
Chalmers
Nicholson
Fanno
Mohave
Mohave ca
Anthony
Order
Ultisol
Alfisol
Spodosol
Ultisol
Oxisol
Mollisol
Alfisol
Alfisol
Aridisol
Aridisol
Entisol
pH
meq/100g
4.2
4.5
4.7
6.2
6.2
6.6
6.7
7
7.3
7.8
7.8
CEC
m2/g
2
19
10
9
14
26
37
33
10
12
6
Surface Area
%
8.0
61.5
8.9
61.3
67.3
125.6
120.5
122.1
38.3
127.5
19.8
Free
Fe oxides
%
0.6
4
1.8
17
23
3.1
5.6
3.7
1.7
2.5
1.8
Clay
4
31
5
61
52
35
49
46
11
40
15
Texture
loamy sand
silty clay loam
sand
clay
clay silty clay
loam
silty day
clay
sandy loam
clay loam
sandy loam
Increasing Mobility
t
Increasing
Attenuation
Capacity
Cu Pb Be Zn Cd Ni Hg
Figures. Relative mobility of cations through soil. (From Korte,
Skopp, Fuller, Niebla, and Alesii, 1976).
Increasing Mobility
t
Increasing
Attenuation
Capacity
Figure 6. Relative mobility of anions through soil. (From Korte,
Skopp, Fuller, Niebla, and Alesii, 1976).
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 0.01 M CaCI2. 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 0.01 M 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
-------
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 ug/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 forZn
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+, CdCI+, CdCI2°, CdCI3-.
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 (MeOhT) 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 SO42', the
adsorption of Cd on soil and soil constituents was inhibited
(O'Connor, etal., 1984; Hirsch etal., 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 toO.SM NaCI), 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~, NO32~, or SO42~ ions.
Under these conditions (anion concentration of 0.1 M), 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 Cl~
at the concentration of CaCI2 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 etal., 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
-------
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 mech-
anisms 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, SeO32', 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.
100
80
•S 60
o
tn
c 40
*? 20
Initial [Selenite] =pmol kg"1
Ionic Strength = 50 mol m~3 NaCI
• = Altamont
n = Ciervo
O = Los Banos
A = Panhill
T = Panoche
•. 9
4.0
5.0
6.0
7.0
pH
8.0
9.0
10.0
Figure 8. Selenite adsorption envelope for five alluvial soils. The
intitial total selenite concentration was approximately 2
l kg-1 (Neal, et al., 1987a).
10
s
o
E
E
D
E
'x
ro
o
fc
o
in
DekalbA
Hagerstown A
Dekalb B
Hagerstown B
pH
Figure 7. Effect of soil pH level on maximum Pb, Cu, Zn, and Ni
retention by Dekalb and Hagerstown A and B horizons.
Nil and Ni2 refer to two apparent sorption maxima.
(Harter, 1983).
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 AI3+ 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
-------
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.
S
o
EXPERIMENTAL DATA
• Cu and ligands
O Cu-ligands absent
Cjg 4.3 mM
GEOCHEM MODELS
— Cu and model ligands
--• Cu-ligands absent
Cu(OH)2 (s)
PH
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(lll) precipitates as
a highly adsorptive solid phase (ferric hydroxide), while Fe(ll)
is very soluble and does not retain other metals. The
reduction of Fe(lll) to Fe(ll) 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(lll) 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 CaCI2, 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
10
-------
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 etal. (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. Mickey 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.
11
-------
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 (Mickey 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 CI-.
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 CaCI2. Puls etal. (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, AgCI,
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
(Hg°), mercurous ions (Hg22+) 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, HgCI2 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, Hg°. 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) (AsO43-), or as arsenite, As(lll) (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(lll), is also strongly pH-
dependent. Griffin and Shimp (1978) observed an increase in
12
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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(lll) by iron oxide occurred at pH 7. Elkhatib
et al. (1984b) found adsorption of As(lll) 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 methyl-
arsines 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 volatil-
ization (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(lll), 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 (Se°), selenite
(SeO32~), 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(lll) and the hexavalent chromium,
Cr(VI). Forms of Cr(VI) in soils are as chromate ion, HCrO4'
predominant at pH<6.5, or CrO42~, predominant at pH 6.5, and
as dichromate, Cr2O72~ 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(lll) 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).
13
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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 CrO42~ at acidic and
neutral pH (Davis and Leckie, 1980; Zachara etal., 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 SO42~ 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(lll) forms hydroxy complexes
in natural water, including Cr(OH)2+, Cr(OH)2+, Cr(OH)3°, 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(lll)
was found to be the least mobile (Griffin and Shimp, 1978).
Hydroxy species of Cr(lll) precipitate at pH 4.5 and complete
precipitation of the hydroxy species occurs at pH 5.5.
Hexavalent chromium can be reduced to Cr(lll) 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 (Gary 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(ll) 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(lll) 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(lll). Chromium (III) complexed with soluble organic ligands
will remain in the soil solution (James and Bartlett, 1983a). In
addition to decreased Cr(lll) adsorption, added organic matter
also may facilitate oxidation of Cr(lll) 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 metal-
ligand 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 overtime. 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
14
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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; Mickey 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,
CaCI2, 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.
Mickey 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 MgCI2), metals
associated with carbonates (acetate buffer, pH 5), metals
associated with Mn and Fe oxides (0.04M NH4OH-HCI),
metals associated with organic matter (0.02M HNO3+H2O2),
and residual metals (HF+HCIO4). 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
Cd
Cu
Exchange
Carbonate
Organic
Ni
Exchange
Carbonate
Exchange
Organic
Figure 10. Proportions of Cd, Cu, Ni, and Zn in each of the
operationally defined geochemical fractions of the
experimental samples (Mickey 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
15
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Table 4. Some bibliographic data on the extraction of heavy metals present in soils and sediments
(Colvet, et al., 1990)
Authors
McLaren Crawford11
(1973)
Stover et al.12
(1976)
Gatehouse et al.13
(1977)
Filipek and Owen14
(1979)
Tessier et al.3
(1979)
Fb'rstner et al. 2
(1979)
Schalscha et al.4
(1980)
Garcia-Miragaya15
(1981)
Badri and Aston16
(1981 )
Fb'rstner et al.17
(1981)
Greffard et al.6
(1981)
Sposito et al.10
Dekeyser et al.18
(1983)
Kuo et al.7
(1983)
Meguelatti et al.5
(1983)
Shuman19
(1985)
Gibson and Farmer20
(1986)
Exchangeable Fraction associated
with carbonates
0.05 N CaCI2 2.5% CH3COOH
1 M KNO3
+ NaF
1 M CH3COONH4
+CH3COOH pH = 4.5
1 M CH3COOH
1 M MgCI2 or 1 M CH3COONa
1 M CH3COONa + 1 M CH3COOH
at pH = 8.2 at pH = 5.0
0.2 M BaCI2
1 M KNO3 0.5 M NaF
pH = 6.5
1 N CaCI2 2.5% CH3COOH
1 M CH3COONH4
+0.5 M (CH3COO)2 Mg
1 M CH3COONH4
pH = 7
resin-H+
0.5 M KNO3
1M CH3COONH4
pH = 4.5
1 M MgCI2
1 M BaCI2 1 M CH3COOH
+0.6 M CH3COONa
1 M Mg(NO3)2
pH = 7
1 M CH3COONH4 1 M CH3COONa
pH = 7 pH = 5
Fraction associated
with oxides
0.1 M (COOH)2
+0.175M (COONH4)2
pH = 3.5
0.1 M EDTA
pH = 6.5
0.1 M NH2OH
+ 1 M CH3COONH4 pH = 4.5
0.25 M NH2OH, HCI
in 25% (v/v) CH3 COOH
0.04 M NH2OH, HCI
in 25% (v/v) CH3COOH
at 96 ± 3°C
or 0.3 M Na2S2O4
+0.175 M Na— citrate
+0.025 M citric acid
0.1 M NH2OH, HNO3 +
25% (v/v) CH3COOH + HCI
0.1 M EDTApH = 6.5
double extraction
0.05 M EDTA
pH = 7
0.25 M NH2OH, HCI
pH = 2
(1) 0.1 M NH2OH, CIH
+O.OI M HNO3, pH = 2
(2) 0.2 M (COONH4)2
+0.2 M (COOH)2, pH = 3
(1) (COONa)2
(2) (COONa)2 + UV
0.5 M Na2EDTA
(1) 0.1 M NH2OH, HCI
(2) 0.2 M (COONH4)2
(HCOOH)2, pH = 3.3
obscurite
(3) Same as (2)+UV
(1) (COONa)2
(2) Citrate dithionite
bicarbonate
0.1 M NH2OH
+25% (v/v) CH3COOH
(1)0.1 M NH2OH, HCI
pH = 2
(2) 0.2 M (COONH4)2
+0.2 M (COOH)2, pH = 3
(3) Same as (2) + ascorbic
acid
(1)0.1 M NH2OH, CIH
+0.01 M HNO3
(2) 1 M NH2OH, CIH in
25% (v/v) CH3COOH
Fraction associated Total amount and
with organic matter residual fraction
1 M K4P2O7 HF
0.1 M Na4P2O7 1 M HNO3
30% H2O2 HF-HCIO4
Acidified HNO3-HF-HCIO4
30% H2O2
0.02 M HNO3+ HF-HCIO4
30% H2O2, pH = 2
at 85 ± 2°C, 2 h
+30% H2O2+HNO3, pH = 2
at 85 ± 2°C, 3 h
3.2 M CH3COONH4
in 20% HNO3
30% H2O2 + NH4OH HF-HCIO4
0.1 M Na4P2O5 1 M HNO3
0.1 N Na4P2O5 HF
30% H2O2
+ 1 M CH3COONH4
30% H2O2, HNO3 HNO3at180°C
pH = 2 at 85°C
extraction with
1 M CH3COONH4
30% H2O2 at 40°C
0.5 M NaOH 4 M HNO3 at 80°C
HNO3-HF-HCI
6% NaCIO4 at 85°C HNO3-HCIO4
30% H2O2 HF-HCI
+0.02 M HNO3
+3.2 M CH3COONH4
0.7 M NaOCI HF-HNO3-HCI
pH = 8.5
30% H2O2 Aqua regia
+0.02 M HNO3 + HF
at 85 °C
16
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for evaluating the mobility of metals under field conditions has
been questioned (Dragun etal., 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(lll) 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.
Mickey 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. Atypical 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.
4-
O) o .
E 31
Linear isotherm
S = CKd
100
C mg/L
200
Figure 11. Sorption isotherms.
17
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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:
M
1
Mb
(1)
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
6 8 10
[Cu2*] x 106 )M)
12 14
Figure 12. Langmuir adsorption isotherm for Cu2+ adsorption on
the Lansing A soil (Cavallaro and McBride 1978).
0.8
0.6
X/M
(mol kg'1)
0.4
0.2
0.0
. Cd
35
(X/M) Meeqn
(m3kg-1 • 103)
Figure 13. Langmuir plots of Cd and Zn sorbed on 5- 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,
18
-------
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 CaCI2. 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
n
O
I
O 50-
(0
O
TJ
TJ
O
0 200 400 600 800 1000
Cd in solid phase before desorption |iM
Figure 14. Desorption of Cd from Kidman and Skumpah soil by
0.01 MCaCI2 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
CaCI2 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
overtime. Several equations have been used to describe the
kinetics of sorption reactions of ions on soils and soil
constituents. These equations include: first-order, second-
order, Elovich equation, parabolic diffusion equation, and
power function equation. An excellent review of kinetic
processes in soil systems is given by Sparks (1989).
19
-------
o
o
o
0 24 48 72 96 120 144 168 192 216 240 264 288 312 336 360
Time (hours)
Figure 15. Change with time in reduced concentration of metals
in suspensions of the Skumpah soils. C0 was the
concentration of metal ions at time = 0 (Dudley, et al.,
1988).
Summary
Metals added to soil will normally be retained at the soil
surface. Movement of metals into other environmental
compartments, i.e., ground water, surface water, or the
atmosphere, should be minimal as long as the retention
capacity of the soil is not exceeded. The extent of movement
of a metal in the soil system is intimately related to the solution
and surface chemistry of the soil and to the specific properties
of the metal and associated waste matrix.
The retention mechanisms for metals added to soil include
adsorption of the metal by the soil solid surfaces and
precipitation. The retention of cationic metals by soil has been
correlated with such soil properties as pH, redox potential,
surface area, cation exchange capacity, organic mater
content, clay content, iron and manganese oxide content, and
carbonate content. Anion retention has been correlated with
pH, iron and manganese oxide content, and redox potential.
In addition to soil properties, consideration must be given to
the type of metal and its concentration and to the presence of
competing ions, complexing ligands, and the pH and redox
potential of the soil-waste matrix. Transport of metals
associate with various wastes may be enhanced due to (Puls
etal., 1991):
1. facilitated transport caused by metal association with
mobile colloidal size particles,
2. formation of metal organic and inorganic complexes that
do not sorb to soil solid surfaces,
3. competition with other constituents of waste, both organic
and inorganic, forsorption sites, and
4. deceased availability of surface sites caused by the
presence of a complex waste matrix.
Because of the wide range of soil characteristics and various
forms by which metals can be added to soil, evaluating the
extent of metal retention by a soil is site/soil/waste specific.
Changes in the soil environment over time, such as the
degradation of the organic waste matrix, changes in pH, redox
potential, or soil solution composition, due to various
remedation 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.
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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