United States
Environmental Protection
Agency
Otfice of Solid Waste -
and Emergency Response
Washington, DC 20460
EPA/530/SW-87/006-F
April 1992
Technical Resource
Document
Batch-Type Procedures
For Estimating Soil
Adsorption of Chemicals
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EPA/530-SW-87-006-F
Technical Resource Document
BATCH-TYPE PROCEDURES FOR ESTIMATING
SOIL ADSORPTION OF CHEMICALS
by
W.R. Roy, I.G. Krapac, S.F.J. Chou, and R.A. Griffin
Illinois State Geological Survey
Champaign, Illinois 61820
Cooperative Agreement No. CR810245
Project Officer
M.H. Roulier
Waste Minimization, Destruction
and Disposal Research Division
Risk Reduction Engineering Laboratory
Cincinnati, Ohio 45268
Office of Solid Waste and Emergency Response
U.S. Environmental Protection Agency
Washington, D.C. 20460
Risk Reduction Engineering Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
1991 jCQi
f£& Printed on Recycled Paper
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DISCLAIMER
The studies reported in this document were funded in part by the United States Environmental Protection
Agency under Cooperative Agreement CR810245 with the Illinois State Geological Survey, Champaign, Il-
linois. This report has been subject to the Agency's peer and administrative review, and it has been ap-
proved for publication as an EPA document. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
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PREFACE
Subtitle C of the Resource Conservation and Recovery Act (RCRA) requires the U.S. Environmental Pro-
tection Agency (EPA) to establish a federal hazardous waste management program. This program must
ensure that hazardous wastes are handled safely from generation until final disposition. EPA issued a se-
ries of hazardous waste regulations under Subtitle C of RCRA that are published in 40 Code of Federal
Regulations (CFR) 260 through 265 and 122 through 124.
Parts 264 and 265 of 40 CFR contain standards applicable to owners and operators of all facilities that
treat, store, or dispose of hazardous wastes. Wastes are identified or listed as hazardous under 40 CFR
Part 261. The Part 264 standards are implemented through permits issued by authorized states or the
EPA in accordance to 40 CFR Part 122 and Part 124 regulations. Land treatment, storage, and disposal
(LTSD) regulations in 40 CFR Part 264 issued on July 26,1982, establish performance standards for haz-
ardous waste landfills, surface impoundments, land treatment units, and waste piles.
The Environmental Protection Agency is developing three types of documents for preparers and review-
ers of permit applications for hazardous waste LTSD facilities. These types include RCRA Technical Guid-
ance Documents, Permit Guidance Manuals, and Technical Resource Documents (TRDs). The RCRA
Technical Guidance Documents present design and operating specifications or design evaluation tech-
niques that generally comply with or demonstrate compliance with the Design and Operating Require-
ments and the Closure and Post-Closure Requirements of Part 264.
The Permit Guidance Manuals are being developed to describe the permit application information the
Agency seeks and to provide guidance to applicants and permit writers in addressing information require-
ments. These manuals will include a discussion of each step in the permitting process, and a description
of each set of specifications that must be considered for inclusion in the permit.
This document is a Technical Resource Document. It was prepared for the Risk Reduction Engineering
Laboratory, formerly the Hazardous Waste Engineering Research Laboratory of the Office of Research
and Development, at the request of and in cooperation with the Office of Solid Waste and Emergency Re-
sponse. This TRD was first issued as a draft for public comment under the title Batch-Type Adsorption
Procedures for Estimating Soil Attenuation of Chemicals (EPA/530-SW-87-006) in June 1987. The draft
TRD was also made available through the National Technical Information Service (Order No. PB87-
146155). All comments received on the draft TRD have been carefully considered and, if appropriate,
changes were made in this final document to address the public's concerns. With issuance of this docu-
ment, all previous drafts of the TRD are obsolete and should be discarded.
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FOREWORD
Today's rapidly developing and changing technologies and industrial products and practices frequently
carry with them the increased generation of materials that, if improperly dealt with, can threaten both pub-
lic health and the environment. The U.S. Environmental Protection Agency is charged by Congress with
protecting the Nation's land, air, and water resources. Under a mandate of national environmental laws,
the agency strives to formulate and implement actions leading to a compatible balance between human
activities and the ability of natural systems to support and nurture life. These laws direct the EPA to per-
form research to define our environmental problems, measure the impacts, and search for solutions.
The Risk Reduction Engineering Laboratory is responsible for planning, implementing, and managing
research, development, and demonstration programs to provide an authoritative, defensible engineering
basis in support of the policies, programs, and regulations of the EPA with respect to drinking water,
wastewater, pesticides, toxic substances, solid and hazardous wastes, and Superfund-related activities.
This publication is one of the products of that research and provides a vital communication link between
the researchers and the user community.
The Office of Solid Waste is responsible for issuing regulations and guidelines on the proper treatment,
storage, and disposal of hazardous wastes to protect human health and the environment from the poten-
tial harm associated with improper management of these wastes. These regulations are supplemented by
guidance manuals, technical guidelines, and technical resource documents, made available to assist the
regulated community and facility designers in understanding the scope of the regulatory program. Publica-
tions like this one provide facility designers with state-of-the-art information on design and performance
evaluation techniques.
This Technical Resource Document (TRD) describes a number of laboratory batch procedures for assess-
ing the capacity of soils and soil components of liners for waste management facilities to adsorb chemical
constituents from solution. Procedures for both organic and inorganic constituents are described, and
their scientific basis and rationale are documented. Examples are included to demonstrate the application
of the procedures and the use of the data in designing soil liners for pollutant retention.
E. Timothy Oppelt, Director
Risk Reduction Engineering Laboratory
IV
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ABSTRACT
This Technical Resource Document (TRD) contains laboratory procedures and guidelines for conducting
experiments using batch-equilibrium techniques to study soil adsorption of chemicals dissolved in solution
(solutes). The procedures were designed for routine use and can be used to generate data for construct-
ing equilibrium adsorption isotherms or curves. Procedures are given for inorganic and organic solutes
and volatile organic solutes.
The scientific basis for each procedural step is discussed in detail. Procedures were based on the scien-
tific literature and were developed and tested by the authors and cooperating laboratories. Examples are
given that show how to apply major procedural steps and concepts. Several types of soil materials and
solutes are featured, as well as the application of batch-adsorption data in calculations of solute move-
ment through compacted landfill liners, which is needed for estimating the thickness of liner required for
pollutant retention.
This TRD was submitted in May 1989 by the Illinois State Geological Survey in fulfillment of Cooperative
Agreement CR810245 with the U.S. Environmental Protection Agency. The TRD has been revised to ad-
dress issues that were raised during the public comment period on the draft TRD entitled Batch-Type Ad-
sorption Procedures for Estimating Soil Attenuation of Chemicals (EPA/530-SW-87-006); the revised TRD
also includes technical information that became available after the draft TRD was completed in May 1987.
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CONTENTS
Foreword
Preface
Abstract
Acknowledgments
Introduction
Chapter
1 Adsorption Forces and Mechanisms
2 Effects of Adsorbent Preparation
3 Effects of Temperature
4 Stability of Nonionic Organic Solutes in Solution
5 Effects of Solution pH
6 Effects of Ionic Strength
7 Effects of Phase Separation
8 Effects of Method of Mixing
9 Selection of a Soil:Solution Ratio for Ionic Solutes
10 Selection of a SoihSolution Ratio for Nonionic Solutes
11 Effects of the SoihSolution Ratio
12 Constant and Variable Soil:Solution Ratios
13 Determination of the Equilibration Time
14 Construction of Adsorption Isotherms
15 Selection of Adsorption Equations
16 Application of Batch-Adsorption Data
17 Laboratory Procedures for Generating Adsorption Data
References
Appendix
A Summary and Chemical Composition of the Adsorbent Soils
and Clays Used in This Study
B Composition of the Metallic Waste Extract Used
in This Study and Associated Adsorption Isotherms
1
3
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49
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75
87
95
99
vii
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FIGURES
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24
25
26
27
28
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30
31
Effect of air-drying on concentration of exchageable manganese
Relationship between concentration of exchangeable potassium in Harpster clay loam
and moisture content
Effect of oven-drying at 105°C on concentration of water-soluble organic carbon in
an Israeli calcareous clay loam
Adsorption isotherm of acetophenone by fresh field-moist and air-dried samples of Crane Island
alluvium
Arsenate adsorption isotherms by Catlin soil at 15°, 25°, and 35°C, and at pH 6.6
Zinc, copper, and cadmium adsorption from a Du Page County landfill leachate
by kaolinite at 25°C at various pH levels
Chromium (VI) adsorption by kaolinite at 25°C at various pH levels
Langmuir-type maximum for adsorption of arsenic as AS(V) and As(lil) by amorphous
iron hydroxide
Effect of pH on adsorption of triazines by a Ca-montmorillonite sample
Effect of pH on adsorption of different ionizabie organic solutes by an illite sample
Adsorption of PCB aroclor 1242 by a synthetic goethite, a Cecil clay, and EPA-14 soil samples
as a function of pH at 24°C
Ratio of concentration to activity versus ionic strength for some common ions
Effect of pore size and number of continuative filtrations of 100-mL aliquots of
HCB-saturated water on concentration of HCB in filtrates
Distribution of arsenic concentrations in solutions that were either centrifuged or filtered
National Bureau of Standards rotary extractor
Distribution of arsenic concentrations after 24 hours of contact with different soil
materials as a function of soihsolution ratio
Distribution of cadmium concentrations after 24 hours of contact with different soil
materials as a function of soil:solution ratio
Adsorption isotherm of o-xylene by Catlin soil at 23°C and at pH 6.1
Adsorption isotherms of dichloroethane and tetrachloroethylene by Catlin soil at 23°C
and at pH 6.1
Relationship between the linear Freundlich constant (Kd) and soihsolution ration
as a function of percent adsorption (lower range)
Relationship between the linear Freundlich constant (Kd) and soihsolution ration
as a function of percent adsorption (upper range)
Effect of soil:solution ratio on cadmium adsorption by a Sangamon Paleosol sample at
pH 6.1 and at 22°C
Cadmium adsorption by a Sangamon Paleosol sample
Distribution of pH values of arsenate solutions after 24 hours of contact with different
soil materials as a function of soihsolution ratio
Distribution of pH values of cadmium solutions after 24 hours of contact with different
soil materials as a function of soihsolution ratio
Distribution of pH values of solutions of zinc slurry extract after 24 hours of contact with different
soil materials as a function of soihsolution ratio
Distribution of the ionic strength of solution containing either arsenate or cadmium
after 24 hours of contact as a function of soihsolution ration
Freundlich constant (K) for two PCB isomers versus sediment concentration with
and without prewashing to remove nonsettling particles
Freundlich constant (K) for the adsorption of Aroclor 1242 by four different soils at 23°C
as a function of soihsolution ratio
Aroclor 1242 adsorption isotherms by 5 soils at 23°C using various soihsolution ratios
Adsorption of dieldrin, tetrachloroethylene, and 1,2-dichloroethane by Catlin soils at
various soihsolution ratios
5
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VIII
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FIGURES (continued)
32 Adsorption of Aroclor 1242 by altered Vandalia till and unaltered Vandalia till at 23°C using
various soihsolution ratios 48
33 Distribution of Freundlich constants (K) and exponents (1/n) associated with arsenic,
cadmium, lead, and PCB (Aroclor 1242) adsorption isotherms 50
34 Cadmium adsorption isotherm with a Vandalia till sample with the amount
adsorbed associated with each isotherm data point shown 51
35 Distribution of cadmium adsorption data for a Tifton sandy loam 47
36 Distribution of arsenate adsorption data at 23°C for different soil samples using different
soihsolution ratios 52
37 Adsorption of cadmium by 5 soil materials at 22°C as a function of contact time 54
38 Adsorption of arsenic by 11 soil materials as a function of contact time 54
39 Equilibration times of Ba, Pb, and Zn from a laboratory extract of the Sandoval zinc slurry
with the Sangamon Paleosol and the Cecil clay sample 56
40 Adsorption of o-xylene, dichloroethane, and tetrachloroethylene by Catlin soil as
a function of contact time 56
41 Adsorption of arsenic by a kaolinite clay sample at 25°C, described by traditional linear
Langmuir, double-reciprocal Langmuir, and Freundlich equation 66
42 Lead adsorption by Cecil clay loam at pH 45 and at 25°C, described by a linear Freundlich
equation forced through the origin 71
43 Predicted distance of lead migration in Cecil clay loam after 35 years, based on three
approaches 73
44 Flow diagram of the procedures for generating batch-adsorption data 76
B1 Barium adsorption isotherm at 21 °C with the Sangamon Paleosol from the metallic
waste extract "
B2 Lead adsorption isotherms at 24°C of two soils using the metallic waste extract 100
B3 Zinc adsorption isotherms at 24°C of two soils using the metallic waste extract 100
IX
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5
6
10
11
12
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15
A1
A2
A3
A4
B1
TABLES
Effect of drying on exchangeable manganese in for Hawaiian soils
pH of soil-water slurries (1:2 v/v) made with field-moist and oven-dried (110°C) samples
Effect of sample pretreatment on the Freundlich partition coefficients (/O of acetophenone
adsorption '
Effect of temperature on Freundlich adsorption constants (Kf) for phenanthrene
cc-naphthol
First ASTM sensitivity analysis for cadmium and arsenic
Cadmium adsorption data from second ASTM interlaboratory sensitivity analysis with
an NBS rotary extractor as the mixing device
Arsenic adsorption data from second ASTM interlaboratory sensitivity analysis with
an NBS rotary extractor as the mixing device
Soihsolution ratio determination for the Sangamon soil and Vandalia ablation till using cadmium
as the adsorbate
Soil: solution ratios for the Sangamon Paleosol and the Cecil clay loam sample using an extract
of the Sandoval zinc slurry
Equilibration times for adsorption of arsenate by soil materials
Equilibration times for adsorption of Ba, Pb, and Zn from a Sandoval zinc slurry extract bv the
Sangamon Paleosol and Cecil clay
Equilibration times for adsorption of the PCB Aroclor 1242 by Catlin soil
Data reduction for arsenic adsorption by a kaolinite clay sample
Lead adsorption data for a Pb(NO3)2 salt and the Cecil clay
Approaches for estimating minimum liner thicknesses on the basis of adsorption
Summary of adsorbents
Selected physicochemical characteristics of clays and soils used in developing this TRD
Major element composition (in oxide form) of clay and soils used in developing this TRD
Trace element concentrations in clays and soils used in developing this TRD
Chemical constituent concentrations obtained by the ASTM-A water shake extraction
performed on the Sandoval zinc slurry
7
7
12
27
29
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31
34
55
57
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95
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ACKNOWLEDGMENTS
The authors acknowledge the partial support of the U.S. Environmental Protection Agency, Cincinnati,
Ohio, and Dr. Michael H. Roulier, project officer of Cooperative Agreement CR810245. We thank Dr. Cal-
vin C. Ainsworth, formerly with the Illinois State Geological Survey (ISGS), for his efforts during the first
year of this project, Dr. Randall E. Hughes for the mineralogical characterizations, Terence Beissel and
Robert Arns for their technical support, and members of the Analytical Chemistry Section of the ISGS for
the adsorbent characterizations. Several laboratories contributed directly and indirectly to this TRD
through their participation in American Society for Testing and Materials (ASTM) D34.02 on Waste Dis-
posal round-robin testing of batch-adsorption procedures: Dr. Gregory Boardman (Virginia Polytechnic In-
stitute and State University), Dr. Chester Francis (Oak Ridge National Laboratory, Tennessee), Dr. Marc
Anderson (University of Wisconsin), Dr. William A. Sack (West Virginia University), and Otis E. Michels
(Daily and Associates Engineers, Peoria, Illinois). Dr. John J. Hassett of the University of Illinois is grate-
fully acknowledged for several informal discussions that helped to refine this document. The suggestions
made by Dr. Kenneth J. Williamson of Oregon State University, Dr. P.S.C. Rao of the University of Flor-
ida, and Dr. Thomas C. Voice of Michigan State University are also acknowledged and appreciated.
XI
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INTRODUCTION
The capacity of geological materials to attenuate potential pollutants has been studied by many re-
searchers, especially during the past 30 years. A potential application of such research is the design and
evaluation of compacted soil or clay liners for attenuation of chemical constituents of leachates from
waste-management facilities, such as landfills and surface impoundments. The batch-adsorption or static-
equilibration technique has often been used in laboratory studies to assess the capacity of soils and soil
components to remove chemical constituents from solution. Batch procedures vary considerably from one
another in terms of experimental conditions and research objectives and sometimes yield different results
even when the same soils, solutes, and concentrations are studied.
The simplicity of the batch-adsorption technique accounts in part for its popularity. With this technique, an
aqueous solution containing solutes of known composition and concentrations is mixed with a given mass
of adsorbent for a given period of time. The solution is then separated from the adsorbent and chemically
analyzed to determine changes in solute concentration. The amount of solute adsorbed by the adsorbent
is assumed to be the difference between the initial concentration (before contact with the adsorbent) and
the solute concentration after the mixing period. Although the approach is relatively simple, several experi-
mental parameters can affect the adsorption of a given constituent.
For inorganic solutes, these parameters include contact time, temperature, method of mixing, soil:solution
ratio, adsorbent moisture content, solution pH, hydrolysis, and the composition and concentration of other
dissolved constituents in the solution (White, 1966; Barrow and Shaw, 1975,1979; Helyar et al., 1976;
Hope and Syers, 1976; Griffin and Au, 1977; Barrow, 1978; Roy et al., 1984,1985). For organic solutes,
similar parameters may also affect adsorption (Bailey and White, 1970; Graver and Hance, 1970; Dao
and Lavy, 1978; Koskinen and Cheng, 1983; Horzempa and DiToro, 1983). Dissolved organic carbon, ad-
sorbate volatility, photodegradation, biodegradation, and compound stability can also affect adsorption
data associated with organic solutes (Harris and Warren, 1964; Scott et al., 1981; Chou and Griffin, 1983).
In the batch-technique studies cited above, equilibration time, a basic experimental parameter, has varied
from 30 minutes to 2 weeks. Soil:solution ratios used in batch procedures have varied from very dilute sys-
tems (1:100,000) to 1:1 pastes. These particular experimental conditions were probably appropriate for
the specific system under study and for the intended use of the data. However, this diversity in experimen-
tal conditions can make comparisons of data between studies difficult. Furthermore, few well-docu-
mented, comprehensive sources are available on conducting batch-adsorption experiments. The only
standardized batch-adsorption procedures designed for routine use are the procedural guidelines outlined
by EPA (1982) and the standard methods developed by the American Society for Testing and Materials
(ASTM) D-18, D-34, and E-47. Results from D-34.02 round-robin testing of batch-adsorption procedures
indicated coefficients of variation greater than 140% during initial testing; these were reduced to less than
10% by application of standard procedures and equipment between laboratories (Griffin et al., 1985).
This Technical Resource Document (TRD) incorporates the experience gained during those inter- labora-
tory testing programs and the interactions with the scientists and laboratories affiliated with ASTM. The
ASTM procedure D4646 was reviewed and voted on by the committee members. Comments were re-
ceived from 96 people active in research, government, industry, and waste management. This TRD de-
scribes a number of batch-adsorption procedures for inorganic and organic solutes, documents their
scientific basis, and recommends procedural steps. Examples are given that demonstrate how to apply
each procedural step. Chapter 16 shows how adsorption data can be used in designing or evaluating soil
liners for pollutant retention. Chapter 17 contains the actual procedures; before attempting them, the
reader should study the preceding chapters. Most of the procedural steps recommended here were tested
in the authors' laboratory with a variety of soils, solutions containing several solutes, and aqueous ex-
tracts of actual wastes. Characteristics of the soils, clays, and waste are described in the appendices.
The collection of accurate and meaningful adsorption data is not a simple task. Even though the proce-
dures described here are fairly easy to use and precise, some "scatter" in the data will inevitably occur,
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and the origins of the dev.at.ons will elude clear-cut explanations. The investigator is encouraged to perse-
vere and repeat the procedures as the situation demands. Perseverance is well warranted Groundwater
and surface water must coexist with the by-products of our civilization, and the acquisition of hiqh-qualitv
adsorption data is essential to the protection of water quality H"a»iy
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CHAPTER 1
ADSORPTION FORCES AND MECHANISMS
Before adsorption studies are discussed, a brief review is presented of the physicochemical forces and
mechanisms thought to be responsible for the adsorption of ions and molecules. Adsorption from solution
at the solid-liquid interface is a complex and imperfectly understood phenomenon.
These physicochemical forces can be broken down into eight categories (Reinbold et al., 1979; Griffin
and Roy, 1985; see references in Voice and Weber, 1983):
• London-van der Waals Attractive forces arise from momentary dipoles about atoms or molecules
caused by small perturbations of electronic motions. These dipoles induce small dipoles in neighboring at-
oms of opposite sign. Although the momentary dipoles and induced dipoles are constantly changing posi-
tion and sign, the net result is a weak attraction (4 to 8 kJ/mol for small molecules and atoms). These
forces are important in adsorption of organics and are generally attributed to the nonideal behavior in
gases. They also have been partially treated by quantum mechanical perturbation theory, which uses po-
larizabilities, ionization potentials, and the magnetic susceptibilities of the interacting atoms to explain vari-
ous phenomena such as adsorption.
• Coulomblc-electrostatic-chemical An electrostatic force results from a charged surface due to iso-
morphous substitution in the mineral lattice (permanent charge) or protonation of surface oxygen and OH
groups (pH-dependent charge) and an oppositely charged species, which maintains the electroneutrality
of the surface. In layer silicates, substitution of octahedrally or tetrahedrally coordinated cations by cations
of lower valence results in a net negative charge. This excess charge can bring about the formation of a
diffuse layer of positively charged atoms or molecules about the colloid; the density of this layer is greater
at the surface, and then decreases exponentially to the level of the bulk solution. This type of reaction is
important in adsorption of inorganic ions and ionized organic molecules.
• Hydrogen bonding A hydrogen atom is bonded to two or more other atoms; the "bond" is generally
conceived as an induced dipole phenomeon. No universal agreement has been reached on the best de-
scription of the hydrogen bond (Huheey, 1978), but it may be considered as the asymmetric electronic dis-
tribution of the 1s electron of the hydrogen atom by very electronegative atoms (e.g., F, O, S, Cl).
Hydrogen bonding is probably more than simply an exaggerated dipole-dipole or ion-dipole interaction,
since these concepts do not account for molecular geometry in some cases (see Huheey, 1978; Cotton
and Wilkinson, 1980). In reality, hydrogen bonds may be delocalized covalent bonds, i.e., resonance
bonds or multiple-center bonds (Huheey, 1978). The energy of this attraction ranges from 8 to 42 kJ/mol.
• Ligand exchange-anion penetration-coordination Many atoms or molecules form coordinated com-
plexes with ligands that range in complexity from simple linear molecules to extensive chelate complexes.
The coordinated complexes may carry a net charge that may be localized on some part of the complex.
These complexes may in turn be bonded to surfaces by hydrogen bonding or by polyvalent cation bridges
linking the complex to a charged surface. The possible geometrical arrangements of coordinated com-
plexes bonded to mineral faces are diverse. The bonded coordinated complexes may be displaced by
other coordinated complexes that better satisfy electroneutrality requirements (i.e., are stronger complex-
ing agents) while being constrained by steric limitations. The energy of ligand exchange reactions with in-
organic ions ranges from 8 to 60 kJ/mol.
• Chemisorption In this adsorption process an actual chemical bond, usually covalent, is formed be-
tween the molecule and surface atoms. A molecule undergoing chemisorption may lose its identity as the
atoms are rearranged, forming new compounds at the demand of the unsatisfied valences of the surface
atoms. The enthalpy of chemisorption (~A H>29 kJ/mol) is much greater than physical adsorption. The
basis of much catalytic activity at surfaces is chemisorption, which may organize molecules into forms
that readily undergo reactions. Chemisorption and physical adsorption are often difficult to distinguish
from one another because a chemisorbed layer may have a physically adsorbed layer upon it. Moreover,
some ligand exchange reactions are chemisorption processes.
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• Dipole-dfpole or orientation energy This interaction results from the attraction of a permanent dipole
tor another permanent dipole. The resulting energy of attraction is less than 8 kJ/mol.
• Induction or dipole-induced dipole This type of interaction results from the attraction of an induced
dipole brought about by either (1) a permanent dipole or (2) a charged site or species. The energy of at-
traction is less than 8 kJ/mol, but this force often adds to coulombic interactions.
• Hydrophobic effect The exact nature of this adsorption mechanism is uncertain. Some investigators
believe that hydrophobic adsorption is primarily an entropically driven mechanism brought about by the
destruction of the physical cavity occupied by the solute in the solvent, and from the partial loss of struc-
1! Au??f "^l60"'68Labout the solute' ordered by van der Waals forces (Horvath et al., 1976; Sinanoglu
and Abdulnur, 1965). Other researchers feel that the hydrophobic effect is the result of simple partitionino
Nonpolar organic solutes tend to migrate from the aqueous phase to hydrophobic surfaces on the adsorb-
ent (Dzombak and Luthy, 1984; Chiou et al., 1979,1983; Griffin and Roy, 1985).
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CHAPTER 2
EFFECTS OF ADSORBENT PREPARATION
The process of preparing samples taken in the field for laboratory investigations can directly influence ana-
lytical results. Adsorbent samples (e.g., soils and clays) are usually dried so that they can be homog-
enized and stored until needed. However, studies have shown that the method of drying the sample may
alter its chemical properties, which in turn can influence the results of batch-adsorption procedures.
An early paper by Fujimoto and Sherman (1945) concluded that the concentration of exchangeable man-
ganese in 23 Hawaiian soils tended to increase as the samples were dried. A portion of their results is
given in table 1. The differences in the amount of exchangeable manganese in field-moist and air-dried
samples were minimal compared with the changes that occurred upon oven-drying or autoclaving. They
also found that the amount of exchangeable manganese tended to increase as the duration of air-drying
increased until about 8 to 10 weeks (fig. 1). Luebs et al. (1956) found that the amount of exchangeable K+
in 13 Iowa soils increased when the soils were air-dried for 2 months. However, a reduction in the mois-
ture content of the soils from 25 to 10% was required before appreciable changes in exchangeable K+
could be detected (fig. 2).
Drying soil samples may also have an effect on the stability of the organic matter in soils. Air-drying soils
generally stimulates soil microorganism respiration when they are rewetted, and Stevenson (1956) con-
cluded that the degree of metabolic activity varies directly with the concentrations of free amino acids and
other nitrogenous materials released during air-drying.
Birch (1958), continuing the work of earlier investigators, found that when either oven-dried or air-dried
soils were remoistened, a portion of the organic matter dissolved, and that the magnitude of this decompo-
sition depended directly on the amount of organic matter present in the soil. He later concluded (Birch,
1959) that this decomposition after rewetting was caused primarily by microbial decomposition of water-
soluble organic matter.
An alternative hypothesis was proposed by Raveh and Avnimelech (1978). They postulated that organic
macromolecules in their natural pedological settings are aggregated by hydrogen bonds. When soils are
dried, the evaporation of water disrupts the hydrogen bonds, and the stability of the organic matter de-
high-Mn soil
16
18
280-
120-
40-
I
10
20
Air-drying time (weeks)
Figure 1 Effect of air-drying on the concentration of exchangeable
manganese (adapted from Fujimoto and Sherman, 1945).
moisture (%)
Figure 2 Relationship between the
concentration of exchangeable potassium
in the Harpster clay loam and moisture
content (adapted from Luebs et al., 1956).
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6-
4-
S
o
2-
1
10
Oven-drying time (weeks)
15
Figure 3 Effect of oven-drying at 105°C on the concentration of water-soluble
organic carbon in an Israeli calcareous clay loam (adapted from Raveh and
Avnimelech, 1978).
creases. They also observed that the amount of water-soluble carbon in aqueous extracts increased as
the length of oven-drying periods at 105°C increased (fig. 3).
According to Bartlett and James (1980), one of the most noticeable effects of air-drying soils is an in-
crease in the yellow or amber color of extracts, attributable to the amount of organic matter made soluble
by drying. They found that the amounts of Al, Fe, and Mn in NH4OAc extracts (pH 4.8) of a soil subjected
to 40 C for 12 hours were greater than those extracted from field-moist samples of the same soil.
Drying soil samples has been reported to change the pH of the soil (or soil reaction). Van Lierop and
MacKenzie (1977) found that oven-drying soil samples at 110°C tended to result in lower pHs relative to
the pHs of field-moist samples of the same soils. The change in pH varied from 0.3 to 1.1 (table 2) Raveh
and Aymmelech (1978) suggested that this increase in acidity was caused by the exposure of fresh
organic surfaces containing acidic groups that were sterically hindered before drying The increase in
nSfL3?^ ^ 3lSO c°n.sidered by Mortland and Raman (1968) who hypothesized a different mecha-
nism. As the samples are dried, adsorbed cations more strongly polarize the residual water molecules
making them more acidic than free water.
Other studies have demonstrated that drying samples lowers the ability of soils to oxidize chromium
(Bartlett and James, 1980), and can influence denitrification studies (Patten et al., 1980; Soulides and
Allison, 1961) and other soil chemical processes that may have an indirect effect on batch-adsorption
SlUdlGS.
- MM nave also been documented; Ashton and Sheets (1959) found that
the herbicide ethyl-A/,A/-d.-/7-ProPylth,olcarbamate (EPTC) was adsorbed as a vapor to a greater extent by
air-dried soils than soils that were moist. The adsorption of EPTC was suppressed at higher moisture con-
tents because of the competition of the EPTC vapor and water molecules for adsorption sites Dao and
Lavy (1978) observed that the adsorption of atrazine by Nebraskan soils decreased with an increase in
soil moisture. They suggested that competition between the atrazine and water could account for this rela-
'
Oven-drying may increase the hydrophobicity of a-soil, which in turn would enhance the adsorbent's af-
finity for hydrophobia solutes. Research has established that forest fires can increase the hydrophobicitv
of soil materials near ground surface. Heat-induced hydrophobicity studies by Debano et al (1976) sua-
gested that temperatures as low as 98° to 1 18°C for an exposure time of as little as 5 minutes can in-
crease the hydrophobicity of a sample as measured by water-drop penetration time. It is not certain
-------
Table 1 Effect of drying on exchangeable Mn in four Hawaiian soils
(from Fujimoto and Sherman, 1945)
Mn concentration (mg/L)
Soil
location
Kemoo
Koko Head
Kahuku
Waimanalo
pH (1:1 )•
4.2
7.1
7.6
8.6
Field-
moist
3.4
0.0
0.0
0.5
Air-
dried
4.5
4.3
0.4
0.4
Oven-
driedf
621.2
29.4
11.7
1.2
Auto-
clavedt
374.8
ND§
367.9
ND
*pH of a 1:1 soihwater suspension.
fOven-dried for 24 hours at 105°C.
^Autoclaved for 3 hours at 15 pounds pressure.
§ND, not determined.
Table 2 pH of soil-water slurries (1:2 v/v) made with field-moist and
oven-dried (110°C) samples (from van Lierop and MacKenzie, 1977)
PH
Soil
1
2
3
4
5
6
7
8
9
10
Field moist
4.0
4.2
5.5
6.2
4.5
4.1
4.2
4.2
6.7
6.3
Oven-dried
3.0
3.7
5.2
5.8
4.0
3.7
4.1
3.9
5.6
5.8
ApH
1.0
0.5
0.3
0.4
0.5
0.4
0.1
0.3
1.1
0.5
Table 3 Effect of sample pretreatment on the Freundlich partition
coefficients (Kf) of acetophenone adsorption (Hassett et al., 1980)
Freundlich partition coefficients
Sample
Sangamon
Crane Island
Fresh
1.07
0.90
Frozen
1.07
0.99
Air-
dried
0.95
1.04
Freeze-
dried
0.95
1.09
Oven-
dried
0.84
1.09
Mean*
0.98
1.03
1.00
1.01
0.97
*Means were not significantly different at the 5% level.
-------
whether this heat-induced hydrophobicity will influence adsorption results obtained by batch techniques.
Hassett et at. (1980), for example, found that the adsorption behavior of acetophenone on two alluvial silt
samples was not significantly affected by various drying techniques (table 3). Oven-drying a Sangamon
River sample appeared to generate a slightly lower Freundlich constant (Kf) relative to the values for fresh
(field-moist), frozen, air-dried, or freeze-dried samples, but the difference was not significant at the 5%
level of probability. As shown in figure 4, the distribution of isotherm points generated from air-dried sam-
ples tended to be similar to the distribution generated from fresh field-moist samples.
In contrast, Bartlett and James (1980) found that a soil sample that had been oven-dried at 40°C ad-
sorbed more phosphate during a 6-hoUr equilibration than samples that were kept moist. Harada and
Wada (1974) reported that air-drying their soil samples resulted in slight but significant increases in the
cation exchange capacity and anion exchange capacity. Bar-Yosef et al. (1969) found that oven-drying
kaolinite at 110°C reduced the amount of phosphate that could be desorbed relative to clay samples that
were not heat-treated. They thought that possibly during drying the phosphate tetrahedron may have
changed its stearic configuration to a form more conducive to bonding.
In summary, drying adsorbent samples to homogenize and store the samples until they are needed may
influence the results obtained by batch-adsorption studies. Bartlett and James (1980) concluded that air-
drying or oven-drying both may be viable methods of sample preparation if the potential changes in ad-
sorbent properties are understood and confronted. However, understanding and confronting these
changes may be research projects in themselves.
Recommendations Oven-drying of adsorbents is not an advisable technique to accelerate drying even
though air-drying may take several days with large bulk samples. Air-drying samples by the atmosphere
minimizes changes that may occur from drying and is the most practical approach at this time. The Ameri-
can Society for Testing and Materials (ASTM) defined air-drying as a process of partial drying (of the sam-
ple) to bring its moisture content near equilibrium with the atmosphere in the room in which further
reduction and division of the sample is to take place (ASTM, 1979). Air-drying should be kept to the mini-
mum necessary to allow preparation of the sample, and a stable condition should be provided for meas-
urements of the sample, such as weighing. Air-drying anaerobic soils and sediments requires special
800-
1 T
400 600 800
Equilibrium acetophenone concentration (mg/L)
Figure 4 Adsorption isotherm of acetophenone by fresh field-moist and
air-dried samples of Crane Island alluvium (adapted from Hassett et al., 1980).
8
-------
handling to prevent the relatively reduced materials from oxidizing if exposed to the atmosphere.
Anaerobic materials can be "air-dried" in a glove box or glove bag supported by a continuous supply of
dry oxygen-free nitrogen or argon gas.
-------
-------
CHAPTER 3
EFFECTS OF TEMPERATURE
Adsorption at the solid-liquid interface tends to occur when the attractive forces between the surface and
ionic solutes are greater than those between the solutes and the solvent (Zettlemoyer and Micale, 1971).
The adsorption of an ionic or polar solute is often the result of a thermodynamically favorable change in
the enthalpy (AH) (Hassett et al., 1981) or sometimes of a favorable change in the entropy (S) of the
system in which the -T AS term from the Gibbs-Helmholtz equation compensates for the positive value of
AH (Thomas, 1961), where T is the temperature of the system. The adsorption of nonpolar organic sol-
utes is thought to result primarily from a thermodynamically favorable change in entropy (AS) involving lit-
tle energy transformation as heat. Thus, the adsorption behavior of ionic or polar solutes will probably
show some temperature dependency, whereas the adsorption of nonpolar solutes may not be greatly influ-
enced by the temperature of the system. The direction and magnitude of temperature dependency will de-
pend on the specific solute-soil system.
An early paper by Jurinak and Bauer (1956) reported that adsorption of zinc by calcite was exothermic;
the amount of zinc adsorbed decreased with increasing temperature. In contrast, Kuo and Mikkelsen
(1979) studied the adsorption behavior of zinc by soils at temperatures ranging from 10° to 35°C and
found that zinc adsorbed endothermically; increased adsorption was associated with higher temperatures.
Kinniburgh and Jackson (1981) reviewed the literature on cation adsorption by soils and concluded that
the effects of temperature were usually small, but in some cases temperature significantly influenced ad-
sorption data.
The adsorption of phosphate by soils and soil materials is often endothermic (Low and Black, 1950; Gard-
ner and Jones, 1973; Griffin and Jurinak, 1973a; Singh and Jones, 1977; Taylor and Ellis, 1978; Roy et
400 -I
35 C
15"C
20 40 60
Equilibrium arsenic concentration (mg/L)
Figure 5 Arsenate adsorption isotherms by Catlin soil at 15°, 25°, and 35°C, and at pH 6.6.
11
-------
al., 1989). The adsorption of arsenate was also found to be an endothermic reaction (fig. 5). The amount
of arsenate adsorbed in equilibrium with a solution concentration of 50 mg/L total arsenic at 15°C
was about 31% (mass basis) less than that observed at 25°C and approximately 51% less than the
amount adsorbed at 35°C. Hassett et al. (1983) found that in contrast to the adsorption of ionic species,
the adsorption of the nonpolar solutes phenanthrene and oc-naphthol by soils was largely unaffected by
temperature variations from 15° to 35°C (table 4). The adsorption of 1 ,2-dichlorobenzene by a soil sample
studied by Chiou et al. (1979) was not affected much by temperature differences between 3.5° and 20°C,
but the adsorption of 1 ,1 ,1 -trichloromethane was reduced at the lower temperature.
Weber et al. (1 983) found that the adsorption of Aroclor 1 254 by a Saginaw River sediment was tempera-
ture-dependent; adsorption was reduced over a 10-degree temperature range. Moreover, Voice (1986,
written communication) demonstrated that the adsorption of 2,4,5,2',4',5'-hexadiclorobiphenyl by a Lake
Michigan sediment decreased with decreasing temperature over a 20-degree temperature range.
The effect of temperature on adsorption data is ultimately linked to the thermodynamics of the adsorption
process. This relationship may be approximated by a Clausius-Clapeyron-type equation, integrated over a
narrow temperature range,
AH' [1]
where Ct and Cz are the equilibrium concentrations of a solute at two different temperatures, 7, and T2,
and AH' is the apparent heat of adsorption.
Apparent heat-of-adsorption values can be used as estimates of the amount of heat energy isothermally
released or absorbed during the course of adsorption, although more correctly, these values probably re-
flect the heat of the overall reaction. Equation 1 can be rearranged as
Cfe/C, « exp((1/72-1/7|)AH'/H)
[2]
Equation 2 can be used to estimate the effects of temperature if the AH' of the specific adsorbent-solute
system is known. If AH' is small, such as with the adsorption of some hydrophobic organic solutes, then
the ratio of Cz to Ci will be close to 1. In other words, the solute concentration at 7, will be nearly the
same concentration as at 72l given that all other conditions are the same; such results for some organic
compounds are given by Hassett et al. (1983). In contrast, the adsorption of phosphate is often associ-
ated with relatively large AH' values. Consequently, phosphate adsorption may be sensitive to ambient
temperature fluctuations. Moreover, the larger the temperature fluctuations (the difference between 7,
and 72 in eq. 2), the greater the potential for the equilibrium solute concentrations to be affected. Thus, ad-
sorption experiments are usually conducted with temperature-controlled water baths or constant-tempera-
Table 4 Effect of temperature on Freundlich adsorption constants (Kf) for
phenanthrene and a-naphthol (from Hassett et al., 1983)
Solute
Phenanthrene
a-Naphthol
Soil
(EPA number)
5
15
5
15
Freundlich constant (Kf )
15°C
328
117
5.4
19
25°C-
304
151
5.5
25
35°C
340
126
7.7
31
12
-------
ture rooms. If such facilities are not available or are impractical, the laboratory work should be conducted
in rooms where the ambient temperature fluctuates by no more than 6°C (e.g., 22 ± 3°C). This 6-degree
range is based on the assumption that a "typical" heat-of-adsorption value for most solutes of environ-
mental significance is approximately 20 kJ/mol. This suggested range should be acceptable for most situ-
ations, but if adsorption of the solute results in a comparatively large heat of adsorption, more rigorous
temperature control may have to be implemented.
Recommendations Batch-adsorption procedures should be conducted under constant-temperature con-
ditions, if available, or in rooms where the ambient temperature is fairly constant (e.g., ± 3°C). When
batch experiments are performed, the temperature of the room should be recorded and treated as a po-
tential variable that can influence the data, and therefore may be useful in the interpretation of the results.
13
-------
-------
CHAPTER 4
STABILITY OF NONIONIC ORGANIC SOLUTES IN SOLUTION
In conducting batch-adsorption procedures, investigators must consider the physicochemical stability of
the solute in solution. Processes such as photodegradation, hydrolysis, and/or microbial degradation can
potentially contribute to a decrease in solute concentration concomitantly with adsorption, and these
changes may even occur before the solution contacts the adsorbent. A standard method is used to deter-
mine hydrolysis rate constants of organic compounds in water (ASTM, 1986a) and to conduct aqueous
photolysis tests (ASTM, 1986b). The following procedures can be used as simple screening tests to iden-
tify potential problems.
• Photolysis Photoreactive solutes that absorb light at wavelengths greater than 290 nm may be subject
to rapid photolysis in glass containers. For example, the half-life of hexachlorocyclopentadiene (C-56)
was found to be less than 5 minutes when exposed to sunlight (Chou and Griffin, 1983). Therefore, pre-
cautions should be taken to ensure that substances such as these are protected from light, not only sun-
light but laboratory lights as well. Appropriate measures include using amber glass, wrapping glassware
in aluminum foil, or adopting any other suitable technique that will eliminate photolysis transformations. A
simple aqueous screening test can help determine the stability of the solute(s) in the presence of light.
This procedure was designed to eliminate volatilization losses and ensure that only reductions in concen-
tration due to photolysis are measured during the test.
Photolysis test Place the initial stock solution into either a 30- or 50-mL borosilicate glass
hypovial, and fill the vial to eliminate any head space. Seal the vial with a Teflon-faced
septum and aluminum crimp cap to prevent volatilization, and place replicate samples in
sunlight for 2,4, and 6 hours. Analyze duplicate samples of the unexposed solute to de-
termine the concentration at time 0 and in two freshly opened hyppvials after 2,4, and 6
hours of exposure. Also determine the concentration of the solute in each of two control
vials (wrapped with aluminum foil or in amber glass vials) that have also been exposed.
Select an analytical method that is most applicable to the analysis of the specific solute
under study. Chromatographic methods are generally recommended because of their
chemical specificity in analyzing the parent compounds without interference from impuri-
ties. If the results indicate the solute is photoreactive, then all subsequent tests and ad-
sorption studies must be conducted under conditions that prevent exposure to light during
the reactions and analytical steps.
• Hydrolysis Hydrolysis is an important degradation path for certain classes of nonionic solutes, and an
investigator should know whether the solute under study will be subject to hydrolysis during the period of
the adsorption study. Otherwise, the amount of solute adsorbed by soils or sediments could be overesti-
mated if changes in solution concentration due to hydrolysis are not taken into account. Details of the hy-
drolysis reactions of various types of compounds can be found in many kinetics texts (e.g., Laidler, 1965;
Frost and Pearson, 1961). Discussions of hydrolysis from an environmental point of view have also been
published (Mabey and Mill, 1978; Tinsley, 1979). The temperature of a hydrolysis screening test proce-
dure should be kept constant. The temperature used in the hydrolysis test procedure should be the same
as that used in the adsorption experiments. The pH is also important and the hydrolysis screening test
should be carried out at the same pH range that will be used in the adsorption studies. Measures to pre-
vent photolysis should be implemented as previously discussed. In some cases, the hydrolysis of solutes
may be enhanced by the presence of other substances such as iron, which catalyze the rate of hexachlo-
rocyclopentadiene hydrolysis under conditions of low pH (Chou and Griffin, 1983). Therefore, the composi-
tion of the test solution must be considered.
Hydrolysis screening test Fill a 30- or 50-mL hypovial completely with the test solution to
eliminate any head space, then seal the vial with a Teflon-faced septum and aluminum
crimp cap. Place replicate samples in a constant-temperature room or water bath for 6,
12,24, and 48 hours. Select an analytical method that is most applicable to the analysis
of the specific compound under study and analyze duplicate samples for the concentra-
tion of the chemical substance at time 0 (control), and in two freshly opened hypovials af-
15
-------
ter 6,12,24, and 48 hours. If significant hydrolysis is indicated by the results of this test,
this must be considered in the interpretation of results from adsorption studies, and spe-
cial care should be given to the handling of flasks and to the analytical steps used.
• Microbial Degradation Microbial degradation can decrease the solution concentration of the solute,
leading to an overestimation of the amount adsorbed by the adsorbent. Therefore, for easily degraded
(labile) compounds, a batch technique will measure "apparent adsorption," which is a combination of ad-
sorption and degradation (and hydrolysis as indicated by the results of the hydrolysis screening test). The
influence of microbial degradation on "apparent adsorption" of phenol by soil was studied by Scott et al.
(1982). They found that Freundlich (Ki) values for the adsorption of phenol by nonsterile soils increased
linearly with time with a Palouse silt loam and increased exponentially with time with a Captina silt loam.
The Freundlich (Kf) values associated with adsorption by sterile soils remained essentially constant after
8 hours. A similar study for p-cresol was also reported by Boyd and King (1984). Their data indicated that
under aerobic conditions, p-cresol degradation began within 10 hours, and complete degradation oc-
curred within 48 hours or less for initial p-cresol concentrations of 5,10,20, and 50 u.g/L The adsorption
of organic compounds, such as phenol or other labile organics that are degraded within the time required
to attain adsorption equilibrium, cannot be evaluated accurately without accounting for or eliminating mi-
crobial degradation losses.
Biodegradation screening test The most common approach used to screen whether an
adsorbate undergoes biodegradation is to conduct kinetic studies by using sterile and
nonsterile soil. Before the kinetic studies, the weighed soil is placed into a reaction bottle
and then autoclaved three times at 2-day intervals, each time for 2 hours at 120°C and at
1.4-bar pressure (Scott et al., 1982). (See chapter 2 to help evaluate the possible
changes in adsorbent characteristics caused by autoclaving.)
Bulk solutions of the solute are prepared in sterilized distilled water and passed
through a sterilized 0.22-u.m membrane. Then, known amounts of the solutions are trans-
ferred to the sterilized and nonsterilized reaction bottles and sealed with sterilized Teflon-
faced septa and aluminum crimp caps. All samples are equilibrated at constant
temperature for 4,8,16, 24, and 48 hours. At the end of each equilibration period, the
solid-phase soil particles are separated from the solution phase by centrifuging duplicate
reaction bottles at 2,000 rpm for 1 hour. Aliquots of the supernatant solution are taken
with a syringe through a hole and septum in the caps of the bottles. Select an analytical
method that is most applicable to the analysis of the specific solute under study. The only
purpose of the suggested test procedure is to screen for biodegradability.
If the test indicates the solute would be biodegradable to a significant extent during
the period of the adsorption test, then the reaction times or temperatures may have to be
modified to reflect this result. The results of the adsorption study must then be interpreted
in the context of the solute equilibration time and the environmental significance of the
biodegradation of the solute relative to its adsorption affinity for soil materials.
16
-------
CHAPTER 5
EFFECTS OF SOLUTION pH
The adsorption behavior of ionic and ionizable inorganic and organic solutes by soils and soil materials is
often influenced by the pH of the soil-water system. In general, the adsorption of inorganic cations in-
creases with increasing pH (Kinniburgh and Jackson, 1981). For example, Griffin and Shimp (1976) re-
ported that the amount of lead adsorbed by kaolinite from a landfill leachate was pH dependent; the
amount of lead removed from solution increased with increasing pH. In their batch-adsorption experi-
ments, as with similar studies, the pH of the soil solutions was periodically adjusted to the indicated pH by
the addition of either dilute acids or bases. A sharp change in slope of the isotherms between pH 4 and 6
was attributed to the precipitation of PbCOs. The reduced adsorption at the lower pH values was attrib-
uted to the increase in competition for adsorption sites by H+ and by Al resulting from the dissolution of
the clay. Similar examples for Cd, Cu, and Zn (fig. 6) show that higher pH values have been associated
with greater removal from solution.
The pH of the soil solution has also been shown to have a direct effect on the adsorption of anionic sol-
utes. In contrast to cationic solutes, anion adsorption is generally enhanced in acidic environments, al-
though some anionic solutes are adsorbed to a greater extent in alkaline systems. Parfitt (1978)
generalized that sulfate adsorption by soils becomes essentially insignificant above pH 8, while the ad-
sorption maxima of boric acid and silicic acid appears to correspond to a pH of approximately 9. White
(1980) generalized that phosphate adsorption by goethite decreased uniformally between pH 3 and 12,
while the magnitude of phosphate adsorption by alumina passed through a maximum value between pH 4
and 5. Griffin et al. (1977a) found that the adsorption of chromium (VI) at low concentrations by kaolinite
passed through a maximum value between pH 4 and 5 (fig. 7). No adsorption occurred above pH 8.5.
r
The adsorption of arsenic as arsenate (As(V)) is also pH dependent with lower pHs resulting in greater ad-
sorption (fig. 8). The adsorption of molybdate by soils also appears to exhibit a maximum value at pH 4
(Parfitt, 1978). This trend, characteristic of most inorganic oxyanions, is thought to result from the in-
creased positive charge due to the increased protonation of surface hydroxyls associated with the edges
of colloidal particles and hydrous metal oxides in acidic environments. The adsorption behavior of arsenic
as arsenite (As(lll)) may (Griffin et al., 1977b) or may not (Pierce and Moore, 1982; fig. 8 of this report) be
strongly dependent on pH.
The adsorption of ionizable organic solutes is also influenced by the pH of the soil solution. For example,
Frissel and Bolt (1962), Weber (1966), and Hance (1969) showed that the adsorption of the triazine in-
creased as the pH decreased (fig. 9). At low pHs, the triazine solutes may have been increasingly proton-
ated, which increased the magnitude of coulombic interaction with negatively charged sites on clay
surfaces. McGlamery and Slife (1966) found that the adsorption of atrazine by the Drummer clay loam
was influenced more by pH than by temperature. Frissel and Bolt (1962) also presented data illustrating
the pH dependency of the adsorption of other ionizable organic compounds (herbicides MCPA, 2,4-D,
DNBP, and 2,4,5-T) by clays. The adsorption of DNBP (fig. 10), for example, sharply decreased as the
pH of the system increased from approximately pH 4.7 to 6. In alkaline solutions (pH > 7), DNBP adsorp-
tion was reduced because of negative adsorption, i.e., clay repelled DNBP. In this pH range, DNBP oc-
curred largely as neutral molecules, since the pK of the organic solute was 4.35. The adsorption of
benzidine also followed a similar pattern. The ionization constants of benzidine are 4.3 and 3.3 (pKb1 and
pKb2). Consequently, Zierath et al. (1980) found that the amount of benzidine adsorbed by two soils de-
creased when the solution pH was increased from a pH of 5 to 11. Benzidine can exist in solution as both
ionized (cationic) species and a neutral species. As the pH of the suspensions was increased, a larger
portion of the total amount of benzidine existed as the neutral form. Both species are subject to adsorp-
tion, although the cationic form should be adsorbed to a much greater extent due to coulombic interac-
tions.
The adsorption behavior of neutral, nonpolar hydrophobic organic solutes appears to be largely unaf-
fected by the pH of the soil-water system. Hassett et al. (1980) found no correlation between the adsorp-
17
-------
8 -
50
100 150 200 250
Equilibrium concentration (mg/L)
300
350
Figure 6 Zinc, copper, and cadmium adsorption from a Du Page County landfill leachate
by kaolinite at 25°C at various pH levels (adapted from Frost and Griffin, 1977).
tfon behavior of polycyclic aromatic hydrocarbons (PAH) and the pHs of 14 soils ranging from pH 4.5 to
8.3. Correlations between the adsorption constants and the actual pH of the solutions were not attempted.
In the present study, the adsorption of the PCB Aroclor 1242 was not significantly influenced by the pH of
three different soil suspensions (fig. 11). The linear Freundlich constants (Kd) were essentially constant
over the range of pH 3 to approximately 10. However, humic and fulvic acids are considered to be the ma-
jor constituents of organic matter in soil (Stevenson, 1982). Ghosh and Schnitzer (1980) postulated that
the macromolecular structure of humic and fulvic acids may be influenced by solution pH and ionic
strength. They may change from rigid spherocolloids with a limited adsorption capacity to flexible, linear
18
-------
Figure 7 Chromium (VI) adsorption by kaolinite at 25°C at various pH levels.
Chromium concentrations shown (4,8, and 17 mg/L) are the initial concentrations
added (modified from Griffin et al., 1977a).
1.6-
1.4-
1.2-
I °-8-
E
i 0.6
I
to
_l
0.4-
0.2-
0.0
10
pH
Figure 8 Langmuir-type maximum (mM/g) for the adsorption of arsenic as As(V) and As(lll) by amorphous
iron hydroxide (from Pierce and Moore, 1982).
polyelectrolytes with a greater capacity to retain nonpplar organic compounds. The influence of pH and
ionic strength on the adsorption of hydrophobic organics is a current area of research.
Recommendations The potential influence of pH on the results generated by batch-adsorption proce-
dures depends on the system under study. The equilibrium pH of the soil-solute mixtures should be deter-
mined before separating the solution from the soil or soil component suspension. The measurement
should be given along with the adsorption data. For anaerobic adsorbent-solute systems, pH measure-
ments should be conducted in a glove box or bag so that the suspensions do not oxidize when the con-
tainers are opened. The failure to measure and report pH data may render the adsorption data impossible
to interpret in a meaningful way.
19
-------
100
80
n
SS 40-
20-
6
i
a
5 6
pH
Figure 9 Effect of pH on the adsorption of triazines by Figure 10 Effect of pH on the adsorption of different
a Ca-montmorillonite sample (adapted from Hance, 1969). ionizable organic solutes by an illite sample (1) 2,4-D,
(2) 2,4,5-T, (3) MCPA, and (4) DNBP (adapted from Frissel
and Bolt, 1962).
200-
"S
.J
e
£-150-
constant,
•5
=5 100-
3
it
50-
0-
••^"^
f EPA-14
Cecil clay
10
pH
12
Figure 11 Adsorption of the PCB Aroclor 1242 by a synthetic goethite, a Cecil clay,
and EPA-14 soil samples as a function of pH at 24°C.
20
-------
CHAPTER 6
EFFECTS OF IONIC STRENGTH
The ionic strength of the solution may have several direct and indirect effects on adsorption data. The ex-
tent of these effects depends on the magnitude of the ionic strength and on the concentration, composi-
tion, and charge of the ionic constituents.
Ionic strength may affect adsorption data in two ways: (1) by changing solute activity, and (2) by changing
the thickness (and therefore properties) of the diffuse electrical double layers associated with colloidal par-
ticles. Because of the shielding effect of neighboring ions, the activity of most solutes tends to decrease
as the ionic strength of the solution increases. However, beyond a threshold ionic strength (often in very
concentrated solutions such as brines), the activities of some ionic constituents reverse themselves and
steadily increase, finally yielding activities exceeding those of their original concentrations (fig. 12). Thus,
adsorption data based on solute concentrations rather than ionic activity may produce calculated results
that disagree with observed results because of the departure of concentrations from ideality in nondilute
systems. Discussion of this topic may be found elsewhere (Atkins, 1982; Bohn et al., 1979; Bolt and
Bruggenwert, 1978; Garrels and Christ, 1965; Stumm and Morgan, 1981).
A basic tenet of Diffuse Double Layer Theory states that the physical thickness of the electrical double
layer composed of adsorbed cations around a colloidal particle is inversely proportional to the ionic
strength of the bulk solution. This phenomenon can affect not only exchange and adsorption reactions at
the solid-liquid interface, but may control the physicochemical properties of the material, such as hydraulic
conductivity, at the macroscopic level.
1.2
1.0
I 0.8
it
•S 0.6
%
c
o
E 0.4
0.2
0.0
Na"*
Ca2 +
SO4
I
j_
j_
0.005 0.01
0.05 0.1 0.2
Ionic strength (M/L)
0.5
1.0 2.0
5.0
Figure 12 Ratio of concentration to activity (i.e., single ion activity coefficient) versus ionic
strength for some common ions.
21
-------
Some Investigators, attempting to minimize changes in ionic strength in the construction of adsorption iso-
therms, have added a water-soluble compound to serve as a background electrolyte (sometimes referred
to as a support medium or background ionic medium) to the solutions containing the solute(s) under
study. The selection of background electrolytes and concentration has varied considerably, and the ration-
ale for the choice has rarely been explained (Ryden and Syers, 1975). The addition of a background
electrolyte has been observed to have no measurable effect in some soil-solute systems, whereas both
synergistic and antagonistic effects have been observed in other systems. The effect of ionic strength on
phosphate adsorption has received much attention. Helyar et al. (1976) concluded that phosphate adsorp-
tion by gibbsite was independent of ionic strength in the range of 0.002 to 0.02 M when the ionic strength
was controlled by NaCI, KCI, and MgCI2. Ryden and Syers (1975) and Ryden et al. (1977) reported that
phosphate adsorption by two soils in a 40-hour interval increased as the ionic strength of the solutions
was increased by the addition of 10~3 to 1 M NaCI. The adsorption of selenite by goethite was reported by
Hingston et al. (1968) as being insensitive to ionic strength in the range of 0.01 to 1.0 M.
In many studies, polyvalent cation salts promoted phosphate adsorption (Barrow, 1972; El Mahi and
Mustafa, 1980; Fox and Searle, 1978; Heylar et al., 1976; White, 1980). Helyar et al. (1976) speculated
that Ca2+ may act as a potential determining ion; El Mahi and Mustafa (1980) suspected that the solubility
of solid phosphate compounds was exceeded (see Anderson et al., 1981).
The relationship between ionic strength and the adsorption of organic solutes has also been examined. In-
creasing the ionic strength from less than 0.01 to 0.1 N resulted in a significant increase in adsorption of
2,4,5-T (Koskinen and Cheng, 1983). This trend has been observed with other weakly acidic herbicides,
such as picloram (4-amino-3,5,6-trichloropicolinic acid) (Farmer and Aochi, 1974) and 2,4-D (2,4-dichloro-
phenoxyacetic acid) (Moreale and Van Bladel, 1980). The increase in adsorption of these weakly acidic
herbicides was attributed to a decrease in pH. A decrease in pH would increase the proportion of the mo-
lecular species, which could then be adsorbed. Choi and Aomine (1974) found that increasing the ionic
strength at constant pH decreased the adsorption of pentachlorophenol (weak acid: pKa = 4.5). The
amount of decrease was dependent on the anion used to adjust the ionic strength of the solution contain-
ing the pentachlorophenol. In batch-adsorption studies, Abernathy and Davidson (1971) found that the ad-
sorption of fluometuron (1,1-dimethyl-3-(a,a,a-trifluoro-/77-tolyl)urea) was decreased and prometryn
(2,4-i>/s(isopropylamino)-6-(methylthio)-s-triazine) was increased by increasing the CaCI2 concentration
from 0.01 to 0.5 N.
In experiments designed to evaluate the effect of solution ionic strength on 2,4,5,2',4',5' -hexachloro-
biphenyl (HCBP) adsorption, Horzempa and DiToro (1983) found that the Freundlich constant (Kf) ap-
peared to be only slightly influenced by increasing NaCI concentration from 10"4 to 10"2 M. But in similar
experiments, CaCI2 significantly affected the Kf values over the same concentration range.
The use of background electrolytes may also promote competitive interactions between the ions derived
from the background electrolyte and the solute(s) under study. (Competitive interactions are discussed in
chapter 11.) For example, Griffin and Au (1977) found that the adsorption of Pb by montmorillonite was re-
duced when 0.1 M Ca(CIO4)2 was used as a background electrolyte. The excess Ca2+ in solution was
also adsorbed by the clay, reducing the number of adsorption sites available to Pb. Other "side reactions"
may take place that can complicate batch-adsorption data; Na-Ca and Na-Mg exchange reactions on ben-
ton'rte were unaffected by CIO4 in a study by Sposito et al. (1983), whereas Cl~ appeared to serve as a re-
actant in the exchange reactions, rather than serving as an "inert" background electrolyte. The formation
of CaCI* and MgCI* complexes may have caused the observed exchange behavior. The appropriateness
of the use of a background electrolyte depends on three factors: (1) the specific conceptual model of the
adsorbent-solute system envisioned by the investigator, (2) the chemical nature of the system itself, and
(3) the overall objectives of the investigation and the intended use of the data.
The position taken in developing the batch-adsorption procedures presented in this document was gov-
erned by the philosophy that they should be simple and designed primarily for routine use. Thus, the use
22
-------
of a background electrolyte was rejected in anticipation that the inherent ionic strength of the solutions will
be influenced by the chemical constituents occurring in the leachate or extract, and those derived from sol-
uble constituents in the particular clay or soil under investigation.
Recommendation Measure the electrical conductivity of the equilibrated soil-solution so that the ionic
strength of the solution can be calculated by the relationship given by Griffin and Jurinak (1973b),
/=0.0127xEC [3]
where / is ionic strength in moles per liter, and EC is electrical conductivity in decisiemens per meter. For
anaerobic adsorbent-solute systems, EC measurements should be conducted in a glove box or bag so
that the suspensions do not oxidize when the containers are opened. The failure to measure and report
EC data and/or ionic strength can make adsorption data difficult to interpret.
23
-------
-------
CHAPTER 7
EFFECTS OF PHASE SEPARATION
A search of the literature indicated that very few researchers have used a filtration technique to separate
the liquid and solid phases before analyzing the liquid phase in batch-adsorption studies. This is probably
due to the potential of the filter membranes to retain significant quantities of the solute, particularly or-
ganic compounds. Luh and Baker (1970) found that a correction factor was necessary to account for re-
tention of C-labeled materials on the filters used in their study. The factor was reasonably constant, but
the filtration technique was abandoned for a centrifugation technique, which uses gravity to separate the
solids from the liquid phase. In a preliminary test, Yaron and Saltzman (1972) also abandoned the filtra-
tion technique because the filter paper retained parathion. In similar studies, Griffin and Chou (1980)
found that cellulose acetate membranes (0.45- and 0.22-u.m pore size) adsorbed significant amounts of
polybrominated biphenyls (PBBs) or hexachlorobenzene (HCB). The problem could be overcome but re-
quired a tedious presaturation technique. They showed that continuously passing nine 100-mL portions of
HCB-saturated water through the membranes saturated the adsorption sites and yielded constant and re-
producible values for the concentration of the compound passing through the membranes (fig. 13). Pre-
saturating the membranes by soaking in HCB-saturated water yielded results that were not significantly
different from results obtained by passing solution through the membrane.
The effects of centrifugation and filtration on arsenic concentrations were investigated (fig. 14). No signifi-
cant differences were found between filtration and centrifugation with respect to solute concentrations.
We concluded that laboratories performing adsorption studies could be given the option of either filtration
or centrifugation without impairing the general usefulness of the results, as long as the affinity of the filtra-
tion membrane for the solute was evaluated adequately; failure to do so could lead to erroneous results.
Recommendation Solid and liquid phases should be separated by centrifugation unless the investigator
can clearly demonstrate that the filtration technique does not significantly affect the results.
A Wlillipore membrane (0.45 jUm)
• Millipore membrane (0.22 jUm)
® Membrane presaturated by soaking
r
5678
Number of continuative filtrations
T
10
11
T
12
Figure 13 Effect of pore size and number of continuative filtrations of 100-mL aliquots of HCB-saturated
water on the concentration of HCB in filtrates (Griffin and Chou, 1980).
25
-------
130-
110
120
Arsenic concentration (mg/L)
CENTRIFUGATION
130
Figure 14 Distribution of arsenic concentrations in solutions that were
either centrifuged or filtered. Values obtained by the two methods were
statistically not significantly different (adapted from Griffin et al., 1985).
26
-------
CHAPTER 8
EFFECTS OF THE METHOD OF MIXING
In theory, the mechanical device used to mix the solid-liquid mixture during the equilibration interval
should have no effect on the equilibrium distribution of solutes and adsorbates. But some studies show
that the mixing method can influence the resulting adsorption data. In a study on phosphate adsorption,
Barrow and Shaw (1979) compared three mixing devices: a reciprocating shaker, a rotating tumbler, and
a roller. They found that the amount of phosphate adsorbed was greatest when a reciprocating shaker
was used and tended to be less with a roller. They concluded that the more vigorous the agitation, the
greater the breakdown of soil particles and the more "new" sites exposed for phosphate adsorption.
Barrow and Shaw also concluded that the efficacies of the three agitation devices to thoroughly mix the
suspensions may have contributed to the differences.
In the development of ASTM procedure D4646 (ASTM, 1987) described by Griffin et al. (1985), a first-gen-
eration procedure was formulated involving the ASTM-A water-shake extraction method (ASTM, 1979). A
round-robin sensitivity analysis of this early procedure was performed by several laboratories. The mixing
method influenced the amount of cadmium and arsenic adsorbed by a Catlin silt loam sample. When
shaking was more vigorous, greater amounts of solute were adsorbed. Results from the first sensitivity
Table 5 First ASTM sensitivity analysis for cadmium and arsenic at high and low initial
concentrations with shakers and a paddle stirrer as the mixing devices
Initial concentration*
Replicate
Lab no.
A
B
C1
C2
D
E
Mean
±SD
%CV
1
2
3
1
2
3
1
2
3
1
2
3
1
2
3
1
2
3
200 |j.g/mL
Cd
16.87
13.24
11.36
83.8
88.2
86.7
1.88
1.77
1.70
26.5
21.5
10.0
3.2
3.2
2.9
2.7
2.9
2.7
21.2
±30.8
145.4
As
186
186
185
128
131
127
162
168
175
130
134
130
153.5
±25.6
16.7
10 jig/mL Shaker rate
Cd
0.080
0.034
0.022
0.166
0.159
0.176
<0.01
<0.01
<0.01
0.064
0.057
0.096
<0.01
<0.01
<0.01
0.007
0.008
0.008
0.073b
±0.064
87.7
As Strokes/min Throw (cm)
59 8
7.92
7.77 70 4
7.91
0.43
0.38 100 3
0.46
5.00
4.78 70 3
5.81
Paddle stirrer used
0.53
0.55 Not known
0.55
3.51
±3.32
94.5
* Represents postprocedure solute concentrations.
t Does not include values lower than the detection limit.
27
-------
analysis are reported in table 5. Large differences in concentrations between the laboratories yielded inter-
laboratory coefficients of variation in excess of 145%. This first round of interlaboratory study clearly
shows why a standard adsorption procedure was needed.
To improve the consistency of interlaboratory results, a National Bureau of Standards (NBS) rotary extrac-
tor was tested as the mixing system (fig. 15). A second sensitivity analysis was carried out (tables 6 and
7) in which each of the participating laboratories used an NBS rotating extractor. The coefficient of vari-
ation between the mean values for each laboratory reflects in part the precision of the mixing method. The
coefficient of variation of Rvalues based on initial cadmium and arsenic.concentrations of 10 and 200
mg/L were less than 8% and 12% for cadmium and arsenic, respectively. These results can be compared
with those from the first round, in which shakers were predominantly used: the coefficients of variation
were as great as 145% for similar concentrations (table 5). Because all other parts of the procedure were
the same in both cases, the mixing method was concluded to be a primary contributor to the variation be-
tween the interlaboratory means. The NBS rotary extractor was adopted as the method of choice because
of the much lower coefficient of variation between laboratory means.
Recommendations For all adsorption experiments, an NBS rotary extractor or its equivalent should be
used during each phase of the construction of an adsorption curve, i.e., determining a soil:solution ratio
(chapter 8), equilibration time (chapter 3), and of course the adsorption curves themselves. Adsorption
data generated with other mixing devices may be valid; however, these data should not be routinely ac-
cepted unless the investigator can document that the device used yields data comparable to those from
an NBS rotary extractor or its equivalent. This documentation will help standardize results among
laboratories.
2-Liter plastic or glass bottles
1 /15-Horsepower electric motor
29RPM
Figure 15 National Bureau of Standards rotary extractor (Diamondstone et al., 1982).
28
-------
Table 6 Cadmium adsorption data from second. ASTM interlaboratory sensitivity analysis
Replicate
Lab no.
A 1
2
3
B 1
2
3
C 1
2
3
D 1
2
3
Mean
±SD
%CV
Table 7 Arsenic
Initial
cone
(ng/mL)
200
200
200
200
200
200
200
200
200
190
190
190
adsorption
with an NBS rotary extractor
Replicate
Lab no. (
A 1
/» i
2
3
B 1
U >
2
3
C 1
^^ i
3
D 1
2
3
Mean
±SD
%CV
Initial
cone
;^g/mL)
205
205
205
200
200
200
200
200
200
200
200
200
24-h
cone
(ng/mL)
35.7
36.2
34.6
31.8
35.8
36.8
35.6
35.6
35.0
31.0
30.0
31.0
34.1
±2.3
7.1
data from the
as the mixing
24-h
cone
(|4,g/mL)
180.3
180.3
182.0
175.5
178.0
170.7
186.3
177.3
175.0
160.0
180.0
180.0
177.1
±6.65
3.76
(mL%)
92.0
90.5
95.6
105.8
91.7
88.7
92.4
92.4
94.3
102.5
106.6
102.5
96.3
±5.2
6.6
second ASTM
device
"d
(mL/g)
2.74
2.74
2.53
2.79
2.47
3.43
1.47
2.56
2.85
5.0
2.22
2.22
2.75
±0.85
30.9
Initial
cone
(ixg/mL)
10.1
10.1
10.1
10.0
10.0
10.0
10.0
10.0
10.0
9.8
9.8
9.8
24-h
cone
(ng/mL)
0.114
0.126
0.125
0.110
0.135
0.165
0.127
0.127
0.132
0.130
0.110
0.120
0.127
±0.01
7.94
interteboratory sensitivity
Initial
cone
(|ig/mL)
10.0
10.0
10.0
10.0
10.0
10.0
10.0
10.0
10.0
12.0
12.0
12.0
24-h
cone
(|ig/mL)
5.76
5.85
5.89
5.52
5.40
5.48
5.57
5.64
5.59
6.80
6.80
6.90
5.93
±0.56
9.47
(mL/g)
1734
1567
1580
1798
1461
1214
1554
1554
1495
1487
1761
1613
1568
±156
9.97
analysis
Rd
(mL/g)
14.72
14.18
13.95
16.23
17.03
16.49
15.90
15.29
15.77
15.29
15.29
14.78
15.42
±0.92
5.99
29
-------
-------
CHAPTER 9
SELECTION OF A SOIL:SOLUTION RATIO FOR IONIC SOLUTES
The term "soil to solution ratio" refers to the ratio of the mass of the adsorbent sample to the volume of liq-
uid. For the purposes of these procedures, 1 mL of solution, regardless of its composition, weighs 1 g. To
construct an adsorption isotherm (curve), soihsolution ratios must be determined that will permit enough
solute to be adsorbed to result in measurable, statistically significant differences in solution concentration.
In these procedures, increasing the soihsolution ratio from a "lower" to a "higher" ratio, such as from 1:1
to 1:100, means that the volume of solution increases relative to the weight of the soil material. If the
soihsolution ratio is too low, i.e., too much adsorbent or too little solution, most of the solute may be ad-
sorbed, forcing the investigator to attempt to measure small differences in concentration between small
quantities of solute. If the ratio is too high, i.e., not enough adsorbent for a given volume, the changes in
the initial solute concentration may be very small, forcing the investigator to measure small differences in
concentration between large amounts of solute. Unfortunately, with inorganic and polar organic com-
pounds, a suitable soihsolution ratio cannot be determined a priori. The soihsolution ratio of the ASTM pro-
cedure D4646 is 1:20 (ASTM, 1987). However, a single ratio cannot be used satisfactorily in all cases.
An empirical, systematic procedure to determine a suitable ratio for a given soil-water and concentration
range is given in chapter 17. A value of 10% to about 30% adsorption for the highest solute concentration
used is a useful criterion for selecting a soihsolution ratio. This will give a discernible decrease in solute
concentration that is statistically acceptable with respect to the initial concentration. Justification for this
guideline is given in chapter 12. An example of this type of approach is given in table 8. With a 1:4 soihso-
lution ratio, more than 90% of the cadmium initially added (200 mg/L) was adsorbed by both a Sangamon
Paleosol and Vandalia till sample. If the 1:4 soihsolution ratio was used to generate data at lower concen-
trations than the 200 mg/L used in this example, the equilibrium cadmium concentrations would be below
analytical detection limits. In contrast, when a 1:500 ratio was used at the lower concentration (10 mg/L),
about 60% of the cadmium initially added was adsorbed by the Sangamon sample. However, when the
200 mg/L Cd solution was used, only 9.8% was adsorbed at the same soihsolution ratio. Essentially, the
object is to select a soihsolution ratio that is a compromise. In this case, a ratio of 1:100 was chosen to
generate an adsorption isotherm because the amount of cadmium adsorbed from the high concentration
range of the isotherm (200 mg/L solution) was approximately between 10% and 30%; at the same time
Table 8 Soiksolution ratio determination for the Sangamon soil and Vandalia
ablation till using cadmium as the adsorbate __^
Sangamon
Cd adsorbed
Vandalia (ablation)
Cd adsorbed
Soiksolution ratio
Initial concentration, 200 mg/L
1:4
1:10
1:20
1:40
1:60
1:100
1:200
1:500
Initial concentration, 10 mg/L
1:100
1:200
1:500
1:1000
722
1631
2792
4246
5165
6250
7500
9250
957
1736
3178
4325
95.2
86.1
73.7
56.0
45.4
33.0
19.8
9.8
91.1
82.7
60.5
41.2
635
1359
2143
3012
3441
3880
4560
4900
840
1474
2215
—
94.1
76.2
44.3
25.4
19.1
13.0
8.0
4.6
80.0
70.2
42.2
—
31
-------
remaining fr°m a low °°ncentrati°n 00 mg/L) solution was also within analytical
This rationale for selecting soilsolution ratios is illustrated in figures 16 and 17. In each figure, the amount
of solute remaining in solution after 24 hours is plotted against the soihsolution ratio. The speckled area
approximates the desired solute concentration after 24 hours of mixing based on 10% to 30% solute ad-
sorption. When the data points or lines connecting the data points fall within this speckled area or "adsorp-
tion target zone," the corresponding soihsolution ratio will usually yield satisfactory results. The adsorption
behavior of seven soil materials with respect to arsenic is shown in figure 16; here, a 1 -10 ratio was cho-
sen to construct adsorption isotherms with six of the seven adsorbents. Figure 17 illustrates the same con-
cept with six samples using cadmium as the solute. A1:10 ratio was chosen for the Tifton loamy sand
but any ratio between 1:10 and 1:4 probably would have yielded satisfactory results. A 1 -20 ratio was '
Vandalia Till
(ablation phase)
Sangamon Paleosol
1:100 1:40 1:20
1:10
Soihsolution ratio (mass/volume)
1:5
32
-------
180-f
160-
Tifton loamy sand
1 :100 1 :40 1:20 1:10 1 :5
Soil:solution ratio (mass/volume)
Figure 17 Distribution of cadmium concentrations after 24 hours of contact
with different soil materials as a function of soil:solution ratio.
Comparison of the two figures indicates that the adsorption of arsenic was essentially a linear function of
the soihsolution ratio, whereas the adsorption behavior of cadmium appeared to be influenced by the
soihsolution ratio: as the ratio of soil to solution decreases, progressively less cadmium was adsorbed per
gram of adsorbent. The significance of this trend is discussed in the next chapter.
The same type of rationale may be applied to solutions containing more than one solute of interest. A
laboratory extract of a metallic waste sample (see appendix B) will help to illustrate this point. The aque-
ous extract of the waste contained several aqueous constituents of interest, and a suitable soihsolution ra-
tio had to be determined for each solute. If one single ratio could be used for all of the solutes with each
given soil, that would be ideal, but this is not always possible. For example, the concentration of zinc in
the extract (550 mg/L) was much larger than that of barium (2.26 mg/L).
A1:20 soil-.solution ratio for the Sangamon sample (table 9) resulted in 32.5% of the zinc being adsorbed,
but the same ratio also resulted in 96.2% of the lead in solution being adsorbed; the solution concentra-
tion of lead approached the limits of detection. When the "stock" extract was diluted to construct an ad-
33
-------
sorptfon isotherm, the adsorption behavior of lead could not be described using this soiksolution ratio
(1:20) because most of the equilibrium concentrations of lead would be below analytical detection limits
Thus a 1:20 ratio was selected to construct a zinc adsorption isotherm, and a 1:100 ratio appeared to be
useful for deriving Pb and Ba adsorption data.
Barium was adsorbed by Cecil clay loam but not to a significant extent (table 9). Since a 1:1 ratio did not
result in at least 10% adsorption, no additional experiments were done. A1:20 ratio was selected for lead
adsorption by Cecil clay loam (table 9), although any ratio between 1:20 and 1:60 would probably have
been acceptable.
A1:4 soil:solution was chosen to study zinc adsorption by Cecil clay loam (table 9), although any ratio be-
tween 1:3 and about 1:8 could have been used. In some cases, there is a range of suitable soil-solution
ratios for a given soil, but even this range of values must be found experimentally. However, as discussed
in chapter 11, there are guidelines for selecting ratios within the acceptable range. Thus, three different
soil:so!ution ratios (1:4,1:20,1:100) were used to construct barium, lead, and zinc adsorption isotherms
with the two soil samples (results shown in appendix B).
Table 9 Soiksolution ratios for the Sangamon Paleosol and the Cecil clay loam
sample using an extract of Sandoval zinc slurry
Sangamon soil
Cecil clay loam
Solution
cone (mg/L)
Barium
0.84
1.15
1.80
2.00
2.19
2.25
2.26
Lead
0.15
0.55
2.64
4.70
8.18
11.4
14.6
Zinc
269
365
485
494
532
542
541
%Adsorbed
62.9
49.1
20.3
11.5
3.1
0.4
99.0
96.2
81.2
67.8
44.0
21.9
50.3
32.5
10.4
8.9
1.7
0
Soil:solution
ratio
1:10
1:20
1:60
1:100*
1:200
1:500
blank
1:10
1:20
1:60
1:100*
1:200
1:500
blank
1:10
1:20*
1:60
1:100
1:200
1:500
blank
Solution
cone (mg/L)
2.09
2.19
2.24
2.24
2.24
2.27
4.51
6.98
10.8
11.6
12.7
13.0
14.7
262
365
444
486
515
552
%Adsorbed
8.7
4.4
5.0
0.4
0.4
69.3
52.5
26.5
21.1
13.6
11.6
53.5
35.3
21.3
10.0
4.6
Soiksolution
ratio
1:1
1:2
1:4
1:10
1:20
blank
1:10
1:20*
1:60
1:100
1:200
1:500
blank
1:1
1:2
1:4*
1:10
1:20
blank
* Soiksolution ratio selected for the kinetic experiments and the adsorption isotherms.
34
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CHAPTER 10
SELECTION OF A SOIL:SOLUTION RATIO FOR NONIONIC SOLUTES
Finding a suitable soilisolution ratio for ionic and polar solutes requires laboratory work, but a simple cal-
culation can be used to estimate a suitable ratio for nonionic solutes, particularly hydrophobic organic spe-
cies. This estimation technique requires a value for the organic carbon content of the adsorbent and for
the organic carbon partition coefficient (K^) of the solute (see ASTM, 1988).
A derivation of this estimation technique begins with
|igssolute/g soil _ |j.gs/g
u.gwsolute/g solution ~ u.gw/g
[4]
where Kd is equivalent to the Freundlich constant (Kf) (see chapter 14) in the special case where the iso-
therm is linear (i.e., 1/n is unity), ugs/g is the mass of solute adsorbed per gram of the adsorbent, and
u.gw/g is the mass of solute per gram of solution.
Also, let R = g adsorbent/g aqueous solution. If we assume that the weight of the solution is approxi-
mately equal to its volume (i.e., 1 mL «1 g), then R is the soihsolution ratio. Equation 4 becomes
[5]
Since u.gs + u.gw should equal the total mass of solute initially added (u,g°), assuming that losses from vola-
tilization or microbial degradation are negligible, then
K _
"
or
R =
MOs
[6]
[7]
Thus, it is possible to select an appropriate soil:solution ratio (R) based on an estimate of the Kd value of
the specific solute-adsorbent system. An estimate of Kd can be calculated if the organic carbon content
(OC) of the adsorbent and the K^0 of the solute are known by
[8]
The organic carbon partition coefficients (/^c) of many hydrophobic and other organic solutes have been
compiled and are given elsewhere (Kenaga, 1980; Kenaga and Goring, 1980; Banerjee et al., 1980;
Hassett et al., 1983; Griffin and Roy, 1985; Roy and Griffin, 1985; see Nirmalakhandan and Speece,
1988, for solubility data). Many of the Koc values reported were based on empirical equations that relate
the solubility (S) of the solute in water to its organic carbon partition coefficient (K^), such as the expres-
sion given by Hassett et al. (1983),
log /(oc = 3.95 - 0.62 log S (mg/L) [9]
A similar linear relationship has been observed between the octanol-water partition coefficient (/Cow) and
its organic carbon partition coefficient (see Hassett et al., 1983).
35
-------
log Kbc = 0.088 + log KOVI
[10]
A compilation of octanol-water partition coefficients was published by Leo et al. (1971). The historical evo-
lution of these concepts was discussed by Griffin and Roy (1985).
To illustrate the application of this estimation technique, the adsorption behavior of a ternary-solute mix-
ture containing dichloroethane, tetrachloroethylene, and o-xylene by a Catlin silt loam sample was stud-
ied. To construct adsorption isotherms, suitable soihsolution ratios for each solute had to be determined.
The organic carbon content of this soil sample was 4.04%. An estimate of a Rvalue for each solute was
based on its water solubility using equation 9.
The solubility of dichloroethane and tetrachloroethylene is 8,450 and 200 mg/L, respectively, at 25°C
(Chiou et al., 1979), and the solubility of o-xylene is approximately 175 mg/L at 25°C (McAuliffe, 1966).
From equations 8 and 9, the calculated /^values of dichloroethane, tetrachloroethylene, and o-xylene
were approximately 1.3,13.4, and 14.7, respectively. Recall from chapter 9, a soihsolution ratio corre-
sponding to about 10% to 30% adsorption is a useful criterion for selecting a suitable ratio. Thus, assum-
ing that a 20% adsorption will fall into the "target zone" for each of the organic solutes, set u.gs/u,g° equal
to 20. Then the soihsolution ratio for each solute may be calculated by arbitrarily setting ug° equal to 100.
For example,
tetrachloroethylene: R =
20
1
(100-20)13.4 53.6
o-oxylene: R =
20
1
(100-20)14.7 58.8
These awkward soihsolution ratios can be simplified to 1:50 and 1:60. If possible (see chapter 9), use a
single soihsolution ratio to generate an adsorption isotherm for every solute of interest in a multicompo-
nent mixture. In this example, a single ratio (1:50) was suitable for generating adsorption isotherms for
each solute (figs. 18 and 19). There may be a range of suitable ratios for some organic solutes. Consult
chapter 11 for guidelines on selecting suitable ratios.
•5 18
0.4 0.8 1.2 1.6
Solution concentration (mg/L)
2.0
Figure 18 Adsorption isotherm of o-xylene by the
Catlin soil at 23'C and at pH 6.1.
0 0.02 0.04
Solution concentration (mg/L)
Figure 19 Adsorption isotherms of dichloroethane
and tetrachloroethylene by the Catlin soil at 23°C and
at pH 6.1.
36
-------
This estimation technique can be generalized and shown as a relationship between the linear Freundlich
constant (Kd) and the soil:solution ratio (/?) as a function of different amounts of adsorption on a percent-
age basis (figs. 20 and 21). McCall et al. (1981) demonstrated that equation 5 could be rearranged as
[11]
This equation was used to generate figures 20 and 21, which should serve as convenient guides for se-
lecting soihsolution ratios. For example, the solubility of carbon tetrachloride in water is 800 mg/L at
25°C. The K"oc value, estimated using equation 9, was 140. Using a Catlin silt loam sample with an or-
ganic carbon content of 4.04%, a Rvalue was calculated as
140(4.04)
100
[12]
1:180-
60 80 100 120 140
Figure 20 Relationship between the linear Freundlich constant (Kd) and soil solution
ratio as a function of percent adsorption (lower range).
37
-------
From figure 20, a soilrsolution ratio of about 1:10 should yield approximately 30% adsorption.
The organic carbon content of soils and sediments may control the extent of adsorption of nonpolar
organic solutes provided that the organic carbon content is greater than approximately 0.1% (Schwarzen-
bach and Westall, 1981). At lower organic carbon contents, the extent of adsorption may be underesti-
mated by the empirical regressions, equations 9 and 10, because they do not account for the adsorption
properties of mineral surfaces (see MacKay et al., 1986).
1:900-
0 200 400 600 800 1000 1200
Linear Freundlich constant (Kd)
Figure 21 Relationship between the linear Freundlich constant (Kd) and soihsolution
-.itio as a function of percent adsorption (upper range).
38
-------
CHAPTER 11
EFFECTS OF THE SOIL:SOLUTION RATIO
The soihsolution ratio may be one of the most important experimental variables to consider when con-
structing an adsorption isotherm and evaluating the adsorption data, particularly when comparing results
from different investigators who used different ratios. As shown in figure 18, increasing the amount of ad-
sorbent while holding the volume of solution constant had the effect of increasing the mass as well as sur-
face area on which the arsenate ions could be adsorbed. Hence, logic suggests that as the amount of
adsorbent is increased, the amount of arsenic left in solution after exposure should decrease in an essen-
tially uniform manner as shown in figure 16.
Figure 17 demonstrated a nonlinear response; the amount of cadmium left in solution after 24 hours ap-
peared to be approaching a constant value as the amount of adsorbent was increased (i.e., the soihsolu-
tion ratio was decreased). There is no single explanation for all systems for this nonlinear response, or
what White (1966) called the "soihsolution ratio effect." The soihsolution ratio effect does not negate the
selection of a soihsolution ratio, but the consequences of that selection must be considered. The influence
of this phenomenon on phosphate adsorption has received a great deal of attention, although the reports
on the effects conflict (Barrow and Shaw, 1979). Phosphate adsorption was increased by the use of high
soihsolution ratios in the studies of Fordham (1963), Barrow et al. (1965), and White (1966). Hope and Sy-
ers (1976) found that high ratios resulted in lower phosphate adsorption. An early paper by Kurtz et al.
(1946) found no soihsolution ratio effect (i.e., a linear response) when studying phosphate adsorption by
Illinois soils.
White (1966) attempted to reconcile his results by arguing that the system was not at equilibrium. How-
ever, this line of reasoning contradicted his rationale for selecting an equilibration interval. Larsen and
Widdowson (1964) had concluded 2 years earlier that the soihsolution ratio effect was due to an increase
in microbial activity as the mass of the soil was increased.
Hope and Syers (1976) argued that different soihsolution ratios affected only the rate at which phosphate
was removed from solution. They found that the change in solution phosphate concentration was propor-
tional to the reciprocal of time. Thus, when the reciprocal-time scale was extrapolated to zero, i.e., infinite
time, the effects of different soihsolution ratios disappeared; the isotherms merged into a single point.
They concluded from this analysis that about 2 to 3 months of equilibration would be necessary for soihso-
lution effects to essentially disappear. Adsorption data would then be essentially independent of the
soihsolution ratio, approaching the expected linear response.
This hypothesis was challenged by Barrow and Shaw (1979), who found that the reciprocal-time analysis
did not explain the soihsolution ratio effects observed in their study. They concluded that such effects
were related to particle breakdown during shaking. As more soil was used (i.e., as the ratio decreased),
more particles broke down, exposing new adsorption sites available to phosphate. However, this concept
does not explain the results shown in figure 17.
The soihsolution ratio effect often has been attributed to the competitive interactions between a given sol-
ute and species that are concomitantly desorbed or exchanged during the partitioning of solutes and ad-
sorbates. As the amount of adsorbent is increased, a larger source of these potentially competing
constituents becomes available. The net effect is that the magnitude of adsorption decreases (given equal
initial concentrations). For example, Griffin and Au (1977) found that the adsorption of lead progressively
decreased as the sample size of a calcium-saturated montmorillonite was increased. As the amount of ad-
sorbent was increased, the amount of calcium that was desorbed or exchanged from the clay also in-
creased and competed with lead for adsorption sites.
A similar phenomenon was observed in this study. Using CdCIa, we investigated the adsorption charac-
teristics of a Sangamon Paleosol. A pronounced soihsolution ratio effect on cadmium adsorption occurred
39
-------
1:200
20
40 60 80 100 120
Equilibrium cadmium concentration (mg/L)
140
160
Figure 22 Effect of soihsolution ratio on cadmium adsorption by a Sangamon
Paleosol sample at pH 6.1 and at 22°C. The solid dots were derived by usina a
1:100 ratio (Roy et al., 1984).
(fig. 22). The curvilinear distribution of data points was derived by using a 1:100 soihsolution ratio How-
ever, when different soihsolution ratios were used, the resulting data did not follow the same pattern but
felkin a nearly straight line that intersected the adsorption curve obtained where 1:100 ratios were used
Ca andMg probably were exchanging with cadmium and thus reducing cadmium adsorption Hence
the greater the amount of sample, the larger the amount of Ca2+ and Mg2+ capable of competing with cad-
mium. At any given equilibrium concentration of cadmium, higher soihsolution ratios (i.e., less absorbent
per volume of liquid) were associated with increased cadmium adsorption.
The Sangamon sample contained about 50% expandable clays and 40% illite (appendix A) Work by
Bittel and Miller (1974) indicated that selectivity coefficients for Ca2+ and Cd2+ exchange reactions with
rnontmorillonite, iilite, and kaolin'rte were between 0.8 and 1.3 (on a concentration basis), suggesting that
these clay minerals have no strong affinity for one cation versus the other over a pH range of approxi-
mately 5 to 7 (cf. Bolt and Bruggenwert, 1978). Calcium will readily exchange with cadmium and vice
versa. If the adsorption data are plotted as cadmium absorbed relative to Cd2+/(Ca2+ + Mg2+) on a molar
basis (fig. 23), the different soihsolution ratios coalesce into one adsorption curve.
The soihsolution ratio can also influence the chemical composition of the system, which in turn can di-
rectly or indirectly affect adsorption data. It is a well-established practice to generate aqueous extracts of
soil samples to make qualitative assessments for soil management. Reitemeir (1945) reviewed the litera-
ture on effects of dilution on ionic concentration in soil solutions and attempted to generalize the results:
Nonsaline soils:
• Solution potassium increases with dilution,
• Calcium and magnesium in solution frequently increase with dilution, while the ratio of
Ca:Mg changes, and
• Phosphorus usually increases proportionally to dilution.
40
-------
6-
5-
4-
2-
1 -
1 MOO
•
1:60
1:40
• •
*^1:20
•1:10
• 1:4
2.0 4.0 6.0 8.0 10.0 12.0 14.0 16.0
Ratio of equilibrium molar concentrations of [Cd] / [Ca + Wig]
18.0
20.0
Figure 23 Cadmium adsorption by a Sangamon Paleosol sample. The adsorption curve
shown is a transformation of figure 22, taking competitive interactions of Ca2+ and Mg2+
into account (Roy et al., 1984).
Alkali, calcareous, and gypsiferous soils:
• In virtually all cases, dilution results in increased amounts of Ca, Mg, Na, K, SO4, P, and Si.
Thus the ionic concentrations in soil solutions and soil extracts are not inversely proportional to the
amount of water present.
The pH of the soil-liquid suspension will also be affected by the soihsolution ratio. The relationship be-
tween pH and adsorption is discussed in chapter 5. The pH of a soil suspension in a batch-adsorption
procedure is controlled by three factors:
• the "natural" pH of the adsorbent and its buffering capacity to maintain that pH,
• the pH and composition of the liquid phase, and
• adsorption reactions that directly or indirectly change the HsO* and/or OHT
concentration in solution.
The first two factors are illustrated by figures 24 and 25. The equilibrium pH of solutions mixed with eight
soil materials are plotted against the soil:soiution ratio. In figure 24, the soil materials were exposed to a
sodium arsenate solution containing 200 mg/L As with an initial pH of 4.65. Consequently, at progres-
sively higher ratios (i.e., more dilute systems), the pH of the solutions became progressively closer to that
of the arsenate solution. Thus, at ratios of approximately 1:20 or higher, the pH of the arsenate solution
dominated the pH of the suspensions. At lower soihsolution ratios, the equilibrium pH of each solution be-
came more like that of the soil, the relative strength of this tendency depending on the pH-buffering capac-
ity of the soil.
In the second example (fig. 25), the soil materials were exposed to a cadmium chloride solution contain-
ing 200 mg/L Cd with an initial pH of 5.45. A1:20 ratio for a kaolinite clay sample was associated with a
solution pH of 7.05; a 1:4 ratio resulted in a solution pH of 7.45. Thus an isotherm generated with a 1:4 ra-
tio may yield lower amounts of cadmium adsorption than one using a 1:20 ratio simply because the pH of
the former was more basic for reasons discussed in chapter 5.
41
-------
!u nattered)
VandaliaTill (ablation)
— Solute solution tends to dominate pH of mixture
Tifton foamy sand
4-
Soil tends to dominate pH of mixture
1:100 1:40 1:20
1:10
Soil solution ratio (mass/volume)
—I—
1:5
1:4
^t* °f ar1enate SOlutions (earning the same initial
24 hours of contact with different soil materials as a
1m r£ nftho " P, y J °bSfrVed Wlth complex ^''component extracts or leachates. The equilib-
r2^ ?n 1th 7 extracVappendix B>was P'otted ^ainst the soil:solution ratio using two so s
££? u TS> u°Wer soil:solution ratios tended to be associated with pHs lower than that of the
extract. However, the pH tended to be constant when a 1:10 or smaller ratio was used
The soilrsolution ratio will often influence the ionic strength of the solution. This is to be expected ThP
ionic strennth nf am/ end if inn io rontrnllpH hu tho ™n^,Tt.",*-~« -i u 1 ^ . expeciea. l ne
is controlled by the concentration and charge of both the solute(s) under
and other aqueous ions derived from the dissolution of soluble miner-
er|t- The ionic strength of the solutions in contact with the Tifton
Cecil clay loam tended to decrease as the soihsolution ratio decreased (fia 27) This
to two factors: (1) as the ratio decreased, more arsenic or cadmium was removed
: , u7 Jl W !? '°wered tne ionic strength, and (2) these two soils contained a low content of watPr
ffta 27?w°pTr± that,contributked to the ioni« strength upon dissolution. The other t^solatSs
fig. 27) were shghtly calcareous by comparison, and consequently lower ratios resulted in an increase in
onic strength because of the dissolution of slightly soluble minerals. Discernible decreases^S5?to
removal of cadmium were masked by the dissolution of carbonates. Whether these changes or d^er
wSX^
42
-------
7-
6-
Kaolinite
— Solute solution tends to dominate pH of mixture
pH -j- pH of Cd solution
5-
4-
Tifton loamy sand
Cecil clay loam
Soil tends to dominate pH of mixture
i
o.
1—1 1 1—
1:100 1:40 1:20
1:10
Soil isolation ratio (mass/volume)
—I—
1:5
1:4
Figure 25 Distribution of pH values of cadmium solutions (containing the same initial
cadmium concentration) after 24 hours of contact with different soil materials as a
function of soihsolution ratio.
. pH of zinc slurry extract
> Sangamon Paleosol
4-
1 r
1:100 1:40 1:20
T
1:10 1:5
Soihsolution ratio (mass/volume)
1:4 1:2
Figure 26 Distribution of pH values of solutions of the zinc slurry extract after 24
hours of contact with two soil samples as a function of soiksolution ratio.
43
-------
s
-na w
1:10
Soil :solution ratio (mass/volume)
Figure 27 Distribution of the ionic strength of solution containing either arsenate or cadmium after
24 hours of contact as a function of soil:solution ratio.
the relationship between ionic strength, soihsolution ratio, and adsorption for any soil-solute(s) system
may be a major project in its own right.
The adsorption of organic solutes can also be influenced by the soil:solution ratio used in batch proce-
dures. Grover and Hance (1970) found that the Freundlich constant (K» decreased significantly by a fac-
tor of 2.6 as the soil:solution ratio was decreased from 1:10 to 1:0.25 in a study concerned with linuron
and atrazine adsorption. They suggested that the differences in adsorption were related to the aggregate
size of the soil. In a comparison of the relative soil particle sizes at three soihsolution ratios, they placed
into flasks 10 g of soil that had been passed through a no. 10 mesh sieve. Added to the flasks were 2.5,
10, and 100 ml of a 0.1 M CaCI2 solution. The solutions were mixed by shaking the flasks gently end
over end for 30 seconds, and the solutions were then allowed to stand. Grover and Hance found that the
dispersion of soil aggregates was greater at the 1:10 soihsolution ratio than at the 1:0.25 ratio; the 1:1 ra-
tio was intermediate. A similar sedimentation behavior was observed in the absence of 0.1 M CaCI2.
Thus, they concluded that the extent of adsorption of linuron and atrazine is related to the aggregate size
of the soil.
44
-------
Voice et al. (1983) reported that the solids concentration seemed to significantly affect the adsorption of
several hydrophobic pollutants by Lake Michigan sediments. They concluded that the soil solution ratio ef-
fect in this case appeared to result from the presence of soluble microparticles derived from the soil,
which also tended to retain the solutes (see Voice and Weber, 1985). They concluded that soihsolution
effects reported in the literature may have resulted from incomplete phase separation during centrifuga-
tion or from accumulative relative errors in measuring concentrations.
Similar conclusions were also reached by Gschwend and Wu (1985). If precautions were taken to elimi-
nate or account for nonsettling (or nonfilterable) microparticles or organic macromolecules, which re-
mained in the aqueous phase during batch-adsorption procedures, the observed partition coefficients (Kf
or K"oc) were found to remain constant over a wide range of soihsolution ratios. A succession of prewash-
ing treatments of sediments greatly reduced the effects of the nonsettling particles (fig. 28). When pre-
washed sediments were used for batch equilibration experiments, the observed Kf remained
virtually constant over the range of soihsolution ratios tested. This relationship was most dramatically
shown for the partitioning of the hydrophobic compound, 2,3,4,5,6,2',5'-heptachlorobiphenyl, and the dif-
ference in Kf with and without prewashing clearly reflected the great sensitivity of very strongly adsorbed
compounds to small nonsettling particle concentrations in the aqueous phase.
Voice and Weber (1985) concluded that although soluble microparticles could play the major role in the
soihsolution ratio effect with organic solutes, microparticles could not account for all of the data given in
the literature. Voice and Weber hypothesized that the soihsolution effect was the result of a "complexation
phenomenon," whereby organic matter in the solution phase forms complexes with the solute. The solute
can exist as a complexed and uncomplexed state in solution, and possibly in other solution states.
10°-
I
I
10'
,3-
800-1
700-
2, 4, 5, 2', 5'-Pentachlorobiphenyl
200-
100-
2, 3, 4, 5, 6, 2', 5'-Heptachlorobiphenyl
T
T
Catlin
EPA-14
Cecil clay
Sangamon Paleosol
T
T
1:10,000 1:1,000 1:100
Soihsolution ratio (mass/volume)
—I
1:30
1:120 1:60 1:40 1:30 1:24
Soihsolution ratio (mass/volume)
1:20
Figure 28 Freundlich constant (Kf) for two PCB isomers Figure 29 Freundlich constant (Kf) for the adsorption of
versus sediment concentration with (open symbols) and Aroclor 1242 by four different soils at 23°C as a function
without (closed symbols) prewashing to remove nonset- of soil solution ratio.
tling particles (adapted from Gschwend and Wu, 1985).
45
-------
In other organic solute-adsorbent systems, the adsorption behavior of the solute was not influenced by
the soil:solution ratio. Bowman and Sans (1985) reported that the adsorbent concentration (soil solution
ratio) did not appear to affect significantly the partitioning of several pesticides in sediment-water systems
over a fairly wide range of values.
The adsorption of Aroclor 1242 was not influenced by the soihsolution ratio (fig. 29). The Freundlich con-
stant was essentially constant over a wide range of soil:solution ratios. When different soihsolution ratios
were used in the construction of adsorption isotherms, the resulting data tended to plot on the same line
(figs. 30 and 31), and the slopes of the adsorption isotherms were near unity. In some cases, a curvilinear
distribution of data points was derived by using different soil:solution ratios with some adsorbents (fig. 32).
45-
40
1:500
0.02
0.04 0.06 0.08 0.10
Equilibrium Aroclor 1242 concentration (mg/L)
0.12
0.14
Figure 30 Aroclor 1242 adsorption isotherms by 5 soils at 23°C using various soil:solution ratios.
46
^
-------
However, the application of different soihsolution ratios still yielded a single, consistent relationship be-
tween the amount of Aroclor 1242 in solution and the amount retained by the tills at equilibrium (fig. 32).
In summary, the selection of a soihsolution ratio may or may not have a profound effect on adsorption
data. The soihsolution ratio may influence the pH, ionic strength, and chemical composition of the
suspension, which in turn may influence adsorption data. In some cases, such as competitive interac-
tions, the soihsolution ratio effect can be rationalized, but in other systems, the ratio effect presents prob-
lems, particularly for procedures intended for the routine collection of batch-adsorption data. Voice et al.
(1983) commented that some combination of techniques or new methodologies may evolve to handle the
ratio effect, but no simple solutions are readily apparent.
1:250
0.02
~T
0.04 0.06 0.08 0.10
Equilibrium concentration (mg/U
0.12
Figure 31 Adsorption of dieldrin, tetrachloroethylene, and 1,2-dichloroethane by the Catlin soil
at 23°C using various soil solution ratios.
47
-------
Recommendations The specific soilrsolution ratios provided in chapter 17, section 8.3, should be
adopted as standard ratios for the construction of adsorption isotherms. For example, if a 1:8 ratio is satis-
factory for the generation of adsorption data, the investigator should use a 1:10 ratio, which is one of the
standard ratios. For many systems, there will be a range of suitable ratios. The user should not arbitrarily
select any ratio within this range but should select the closest standard ratio. These standard ratios range
from 1:4 to 1:10,000 and should accommodate most situations. Adherence to this recommendation will
enable direct comparisons of adsorption data generated by different investigators. Adsorption data based
on ad hoc ratios may provide a basis for limited comparison; however, unless a particular solute-adsorb-
ent system can be shown not to be subject to soil solution ratio effects, there will always be some doubt
that the results are comparable.
25-
20-
215-
8
Ti
Z
o 10-
5-
1:250
Vandalia till (altered)
1:100
1:6
0.02
0.04 0.06 0.08 0.10
Equilibrium Aroclor 1242 concentration (mg/L)
0.12
Figure 32 Adsorption of Aroclor 1242 by altered Vandalia till and unaltered Vandalia till at 23°C
using various soihsolution ratios.
48
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CHAPTER 12
CONSTANT AND VARIABLE SOIL:SOLUTION RATIOS
Two experimental techniques are used to generate batch-adsorption data:
• Constant soil:solution ratio method: mixing a batch of aqueous solutions—each solution containing pro-
gressively decreasing solute concentrations—with absorbent, keeping the amount of adsorbent (by
weight) constant in all solutions.
• Variable soihsolution ratio method: mixing a batch of solutions, all containing the same solute concentra-
tion, with progressively increasing amounts of absorbent.
The first technique presumably is based on a standard ratio selected from the procedures given in chap-
ters 9,10, and 17. The second is similar to the technique used to select a soihsolution ratio for ionic sol-
utes (chapter 9). Although one might expect that both techniques would yield the same results and thus
could be used interchangeably, this does not hold true for all systems.
In the first technique, the initial (stock) solute solution—a solution prepared in the laboratory or a leachate
taken from the field—is progressively diluted, forming a batch of diluted solutions that are added to con-
tainers holding the same amounts of soil material. As discussed in chapter 11, the soihsolution ratio may
affect the adsorption data. Figure 22 showed that using soihsolution ratios ranging from 1:200 to 1:4
yielded adsorption data that were in poor agreement with the isotherm generated from a 1:100 ratio. In
this case, the phenomenon was attributed to competitive interactions between Cd2+ and desorbed Ca2+
and Mg2+. Where the data were replotted (fig. 23) with these competitive interactions taken into account,
the adsorption data coalesced into a single consistent relationship. This replotting technique does not
work in all cases. Techniques for modeling competitive adsorption are emerging (see Murali and Aylmore,
1983a,b,c; Roy et al., 1989) and are currently too complicated to use in routine batch procedures. Also,
not all "soihsolution ratio effects" can be attributed to competition (see chapter 11). This effect was ob-
served during the development of these procedures; applying the variable soihsolution ratio technique
yielded results (amounts of cadmium and lead adsorbed) similar to or lower than the amounts obtained by
the constant soihsolution procedure.
For this reason, an isotherm produced by variable soihsolution ratios was considered the more environ-
mentally conservative isotherm. In this document an isotherm produced in this way is called an environ-
mentally conservative isotherm (ECI). The ECI has two major advantages over an isotherm produced by
the use of a fixed soihsolution ratio: (1) if the solute-adsorbent system reaches equilibrium in 24 hours—
or, more correctly, satisfies the conditions of the operational definition of equilibrium (chapter 13)—then
the data generated in selecting a soihsolution ratio can be used to construct an isotherm; and (2) the
effects of competition and other processes are implicitly accounted for, although their exact nature is
unknown.
Figure 33 provides further evidence that using a variable soihsolution ratio yields environmentally conser-
vative estimates of adsorption. The adsorption data were modeled with the Freundlich equation (chapter
14), yielding the isotherm constants shown. The isotherms associated with the constant soihsolution ratio
technique are called constant soihsolution ratio isotherms (CSI). The results shown in figure 33 may be
generalized as
and
1 //?ECI ^
This type of analysis indicates that an isotherm generated by using different adsorbent masses generally
yields lower predictions of solute adsorption and is thus viewed as being environmentally conservative.
The ECI is recommended as the method of choice for routine use.
49
-------
1 1-
o
UI
0 0.02
Kf (CSI), L/mg
B
1.0-1
0.8-
O 0.6-
ui
0.4-
0.2-
0.0-
• Cadmium
O Arsenic
A Lead
A PCS
0.0
0.2
0.4
1/n(CSI)
0.6
0.8
1.0
Figure 33 Distribution of (A) Freundlich constants (Kf) and (B) exponents
(1/n) associated with arsenic, cadmium, lead, and PCS (Aroclor 1242)
adsorption isotherms.
Experimental data produced during the process of selecting a soihsolution ratio may be used to construct
an isotherm if the system equilibrated within 24 hours (equilibration time is discussed in the next chapter).
However, some of the data points inevitably will be associated with situations in which less than 10% of
the solute was adsorbed. In chapter 9, we recommended that a soihsolution ratio be chosen from which at
least 10% or greater adsorption occurred. To understand the importance of this recommendation, con-
sider a situation in which an investigator conducted experiments for selecting a soihsolution ratio, then
used data points from the entire concentration range and attempted to construct adsorption isotherms
(figs. 34 and 35). As shown in figures 34 and 35, the data points associated with less than 10% adsorp-
tion did not conform to the general pattern established by data associated with greater than 10% adsorp-
tion. Eliminating these data points yielded better results, i.e., more reasonable revalues (chapter 14).
Rgure 35 shows an extreme case; nearly all data were associated with less than 10% adsorption. There
was Itttle point in fitting this data set with an isotherm equation.
Although the ECI is useful for many situations, it cannot be universally applied to every situation. The ECI
may be limited to cases in which (1) the adsorbent has a relatively high affinity for the solute, and (2) the
initial solute concentration is relatively low. The ECI technique is often used with sparingly-soluble organic
solutes in which the initial solute concentration is low.
The ECI technique was used to derive arsenic adsorption isotherms with the soil adsorbents used to de-
velop these procedures. Soihsolution ratios of 1:4 and higher were used, and the initial concentration of ar-
senic was 200 mg/L. Varying the amount of adsorbent over this range of soihsolution ratios did not
50
-------
0.0
0.4
1 T
0.8 1.2 1.6
Equilibrium cadmium concentration (mg/L)
r
2.0
1
2.4
Figure 34 Cadmium adsorption isotherm at 22°C for a Vandalia till sample (unaltered) with the amount
adsorbed associated with each isotherm data point shown. The mean pH of the soil-solute suspensions
was 6.8.
240-
200-
-a 160-
1
T3
J3
I 120-
(O
c
=1
<
80-
40-
0
0
5.9%
2.7
\_ 10.2%
x^^f *
24.6% adsorbed ^"^
X
X
X
r2 = 0.04
X
/
'
,
,
' c
/ £
/ I
/ s
/ 8
' 1
1 i II
40 80 120 160
0
.7%
••^
0.32%
I
200
Equilibrium cadmium concentration (mg/L)
Figure 35 Distribution of cadmium adsorption data at 22°C for a
Tifton sandy loam with the amount adsorbed associated with each
isotherm data point shown. The mean pH of the soil-solute suspensions
was 6.8.
51
-------
change the equilibrium arsenic concentration substantially (fig. 36). The relatively small changes in arse-
nic equilibrium concentrations caused the data points to be somewhat clustered, leaving an area between
the on'gin of the isotherm and the lowermost arsenic equilibrium concentration without data points. There-
fore, regression of these data sets using isotherm equations could potentially lead to large errors. More-
over, lower soil:solution ratios cannot be used to fill in the gaps; the use of ratios much lower than 1:4
would eventually produce a very thick suspension or paste that could not be efficiently mixed, separated,
or analyzed. This "ratio gap" problem is accentuated as the initial solute concentration increases; the
"cluster" simply migrates to the right side of the isotherm. For these reasons, the constant soihsolution ra-
tio isotherm (CSI) is also recommended for application as an alternate procedure, given that the ECI tech-
nique does not produce useful or applicable results, as figure 36 illustrates.
0.6-
0.5-
0.4-
8
n
0.2-
0.1-
o.o-
1:20 • (13% ads rbed)
1:10
Catlin silt loam
1:4
1:18 (11%)
/ Till (Vandalia ablation)
1:10(18%)
1:5
Inaccessible region
Kaolinite clay
1:5
Till (altered Vandalia)
1:10 (11%)
40
—T~
80
—1—
120
—T~
160
200
Equilibrium arsenic concentration (mg/L)
Figure 36 Distribution of arsenate adsorption data at 23°C for different soil samples using
different soihsolution ratios. The pH values of each soil-solute system were similar to those
given for each soil in appendix A.
52
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CHAPTER 13
DETERMINATION OF THE EQUILIBRATION TIME
Adsorption at the solid-liquid interface is a thermodynamic process, and adsorption measurements are
taken when the system has equilibrated. The equilibration time in batch-adsorption experiments is the
time interval in which the system reaches chemical equilibrium and the concentrations of the products
and reactants cease to change with respect to time,
ft = ° [13]
Past studies have used many different equilibration times. Lawrence and Tosine (1976) used 30 minutes
to equilibrate PCBs with soil, whereas Jones et al. (1979) allowed a soil-phosphate mixture to equilibrate
for 6 days before separating the liquid from the soil. The equilibration times used in most studies were
probably based on preliminary kinetic studies. However, there is a clear danger in assuming that the
equilibration time reported by one investigator is valid for another system even when the adsorbent-solute
systems are similar. Equilibration time is an experimental variable that must be determined for any sys-
tem before an adsorption isotherm (curve) is constructed.
The ASTM procedure D4646 (ASTM, 1987) can be used to characterize the affinity of a soil or clay for sol-
utes after 24 hours, but 24 hours may or may not be long enough for chemical equilibrium to develop (as
indicated earlier, some investigators have used equilibration times of days or even weeks) Adsorption is
generally regarded as a fast reaction, and subsequent removal of solute from solution may be attributed
to other processes. Adsorption processes at solid-liquid interfaces are often initially rapid; further reduc-
tion in solute concentration continues at a decreasing rate, asymptotically approaching a constant concen-
tration. In some cases, equilibrium is never clearly attained. The ambiguity in the definition and
measurement of equilibration times has been acknowledged as a major problem in adsorption studies
(Anderson et al., 1981). For most systems involving complicated adsorbents such as soils, it is very diffi-
cult to determine when adsorption processes dominate and when they become less important as other
processes, such as ion penetration or precipitation, become significant. The EPA (EPA, 1982) suggested
that the equilibration time should be the minimum amount of time needed to establish a rate of change of
the solute concentration in solution equal to or less than 5% per 24-hour interval. This definition is an op-
erational definition of equilibrium and is equivalent to a steady state. Cast in a form similar to that of equa-
tion 13 it may be written as
AC
— < 0.05 per 24-hour interval
[14]
The efficacy of this operational definition for equilibrium was evaluated using seven soil materials Each of
the adsorbents was exposed to arsenic and cadmium solutions, initially containing 200 mg/L for periods
of up to 72 hours. The solutions were analyzed, and the rate of removal of the solute was determined All
of these particular soil-solute systems were in equilibrium after 24 hours as defined by this operational
definition. Figure 37 presents representative data for cadmium adsorption, and figure 38 shows the ad-
sorption behavior of arsenate for 11 different soil materials. In this example, it is not obvious in some
cases when the rate of change of the solute concentration is equal to or less than 5% per 24-hour inter-
val. It may be more convenient to analyze kinetic data as shown in table 10: an equilibration time of 24
hours was selected for three of the samples, and a period of 48 hours was used to equilibrate arsenate
with a Vandalia till sample (ablation phase). In each example, the calculated %AC represents the change
in concentration during the preceding 24 hours. During the next 24-hour interval, the arsenic concentra-
tion continued to decrease, but only by 1.14%. It is this slow and relatively small decrease in solute con-
centration that follows the more rapid and pronounced decrease that is frequently a problem. The applica-
tion of this operational definition of equilibrium assumes that this additional 1.14% decrease is negligible
53
-------
idmium adsorbed per sram of soil (mg/g
> ro *> at m
» - i
Kaolinite
L ^-*— -~^_ Sangamon paleosol
1
T _-___*_^j
f " ]
f
Vandalia till (ablation) f
( Vandalia till (altered)
_ Vandalia till (unaltered) _
6 "6 10 20 30 40 50
Time (hours)
60 70
Figure 37 Adsorption of cadmium by 5 soil materials at 22°C as a function of contact time.
200
190
30 40
Time (hours)
50
60
70
Figure 38 Adsorption of arsenic by 11 soil materials as a function of contact time.
54
-------
Table 10 Equilibration times for adsorption of arsenate by soil materials
Time
(h)
0
1
4
8
16
24f
36
48
72
0
1
4
8
16
24
36
48t
72
%AC *
Cecil
—
—
—
—
—
8.17
—
0.56
1.08
Vandalia
—
—
—
—
—
18.97
—
5.51
3.15
Solution
cone (mg/L)
clay loam
193.4
182.9
181.5
180.0
177.6
177.6
177.1
176.6
174.7
till (ablation)
199.3
179.5
173.2
169.4
164.0
161.5
157.3
152.6
147.8
%AC
Solution
cone (mg/L)
Kaolinite
—
—
—
—
—
16.36
—
1.14
0.55
Sangamon
—
—
—
—
—
33.13
—
2.20
2.25
199.3
171.4
168.6
168.1
166.2
166.7
164.8
164.8
163.9
Paleosol
199.3
165.3
158.2
158.1
152.6
149.7
148.3
146.4
143.1
* %AC = (C, - CgJ/C,, where C, is solution concentration at time *
and C2 is the concentration at time t + 24 hours.
t Equilibration time selected for the adsorption isotherms.
and may be attributable to processes other than adsorption. Therefore, this solute-adsorbent system is de-
fined as being at steady state after 24 hours of contact.
The application of this definition with multicomponent solutions is exemplified by the metallic waste slurry
(appendix B). To determine the equilibration time, we conducted preliminary kinetic experiments using the
soil:solution ratios previously determined (chapter 9). Barium was adsorbed by the Sangamon Paleosol
sample and this system appeared to reach equilibrium within 24 hours. The rate of change in solute con-
centration for the first 24-hour period was 12.2% (table 11). After 24 hours, the rate of change was less
than 5% for each subsequent 24-hour interval. Similarly, the solution concentrations of lead and zinc were
also constant after 24 hours (fig. 39)—constant in the sense that the rate of change in solution concentra-
tion of these two solutes per 24-hour interval was less than 5% (table 11).
The zinc-Cecil clay system appeared to reach the operational equilibrium within 24 hours (fig 39) The
rate of change in zinc concentration during the 24- to 48-hour interval was 0.2% (table 11). Lead did not
equilibrate with the Cecil clay until about 48 hours; the rate of change in lead concentration during the 48-
to 72-hour interval was 1.5% (table 11). Thus, an equilibrium interval of 24 hours was used to construct
adsorption isotherms except that for the adsorption of lead by the Cecil clay (a 48-hour interval was used).
Solution concentrations of o-xylene, dichloroethane, and tetrachloroethylene tended to change by
amounts less than about 5% after 24 hours when in contact with Catlin (fig. 40). The rate of adsorption of
the PCB Aroclor 1242 by a Catlin sample was nil after the initial 24 hours of contact (table 12)
55
-------
560-
1
520-
480-
3
J 440-
c
| 400-
e
g 360-
u
c
•2 A
3
s
10-
O-J
in
3
O
JZ
I I
w
CO
1 *
' 1 1
|
S
Zn-Cecll clay
Zn-Sangamon
Pb-Sangamon
\ Pb-Cecil clay
"— \
Ba-Sangamon
i — ,-
I I 1
i
I
10
20
30
40
Time (hours)
50
60
70
Figure 39 Equilibration times of Ba, Pb, and Zn from a laboratory extract of the
Sandoval zinc slurry with the Sangamon Paleosol and the Cecil clay sample.
500-
400-
300-
.0
s.
12-
8-
4-
Dichloroethane
J3-xy)ene
10
20 30
Time (hours)
40
50
Figure 40 Adsorption of o-xylene, dichloroethane, and tetrachloroethylene at
23°C by Catlin soil as a function of contact time.
Recommendation The equilibration time should be the minimum time needed to establish a rate of
change of the solute concentration in solution that is equal to or less than 5% for a 24-hour interval. The
equilibration time should be determined for each solute-adsorbent system before adsorption isotherms
are constructed. The typical equilibration time is 24 hours.
56
-------
Table 11 Equilibration times for adsorption of Ba, Pb, and Zn from
a Sandoval zinc slurry extract by the Sangamon Paleosol and Cecil clay
Time
(h)
Barium
0
1
8
24*
31
48
72
Lead
0
1
8
24*
31
48
72
Zinc
0
1
8
24*
31
48
72
%AC,
—
—
—
12.2
—
-1.0
-1.5
—
—
—
66.8
—
8.9
5.1
—
—
—
33.4
—
1.1
0.3
Solution
cone (mg/L)
2.30
2.00
2.02
2.02
2.02
2.04
2.07
15.5
5.68
5.01
5.14
4.81
4.68
4.92
563
387
375
375
371
371
370
Time
(h)
Lead
0
1
8
24
31
48*
72
Zinc
0
1
8
24*
31
48
72
Solution
%AC^ cone (mg/L)
— 15.4
— 7.72
— 7.27
54.5 7.01
— 6.72
6.3 6.57
1.5 6.47
— 549
— 421
— 430
20.9 434
— 430
0.2 433
0.2 432
* %AC = (C, - C2)/C,, where G, is solution concentration at time f,
and C2 is concentration at time f + 24 hours.
t Equilibration time selected for the adsorption isotherms.
Table 12 Equilibration times for adsorption of the PCB
Aroclor 1242 by Catlin soil
Time (h)
%AC
Solution cone (mg/L)
0
2
4
6
8
24*
48
_
—
—
—
—
94.31
0.00
0.220
0.020
0.018
0.017
0.017
0.013
0.013
*Equilibration time selected for an adsorption isotherm.
57
-------
-------
CHAPTER 14
CONSTRUCTION OF ADSORPTION ISOTHERMS (CURVES)
An adsorption isotherm or curve is a graphic representation showing the amount of solute adsorbed by an
adsorbent as a function of the equilibrium concentration of the solute. This relationship is quantitatively de-
fined by some type of partition function or adsorption isotherm equation that is statistically applied to the
adsorption data to generalize the adsorption data.
In studies concerned with the adsorption of gases by solids, more than 40 equations have been used to
describe the data. Historically, only a few of the equations have been found to be applicable to solid-liquid
systems. Only the two most commonly used and simplest of these adsorption equations will be discussed
here—the Freundlich and Langmuir isotherms. Neither may be appropriate for a given system. The
reader may wish to consult a paper by Kinniburgh (1986) on the applicability of other adsorption equa-
tions.
The Freundlich Equation
Probably the oldest, most widely used adsorption equation for solid-liquid systems is the Freundlich ad-
sorption equation, named after H. Freundlich (Freundlich, 1909),
x _ isrA/n
f15]
where x is the amount or concentration of the solute adsorbed, m is the mass of the adsorbent, C is the
equilibrium concentration of the solute, and K>and 1/n are constants.
The Freundlich equation was originally proposed as an empirical expression without a theoretical founda-
tion. However, some investigators have referred to the Freundlich constant (Kf) as being related to the ca-
pacity or affinity of the adsorbent; the exponential term may be an indicator of the intensity of adsorption
or how the capacity of the adsorbent varies with the equilibrium solute concentration (see Suffet and
McGuire, 1980).
Other investigators attempted to show that the Freundlich equation has a theoretical basis. A number of
derivations of the Freundlich equation were based on the Gibbs adsorption equation (Chakravarti and
Dhar, 1927; Rideal, 1930; Freundlich, 1930; Halsey and Taylor, 1947; see Hayward and Trapnell, 1964;
Kipling, 1965). Zeldowitsch (1935) demonstrated that the Freundlich equation could be explained in terms
of a nonhomogeneous surface. Sips (1948) established in a rigorous fashion a general relationship be-
tween surface heterogeneity and the Freundlich equation, a derivation Sposito (1980) partially adapted to
his system to derive a Freundlich-type expression for trace-level exchange reactions.
The Freundlich equation is frequently used, probably because it is simple. It contains two constants; both
are positive-value numbers that can be solved statistically when expressed in logarithmic form:
log(x/m) = \ogKf+-[/n\og C
[16]
By taking the logarithms of both sides of equation 15, the constants K>and 1/n may be solved, via equa-
tion 16, as a simple linear regression,
y, = a + bx,
[17]
59
-------
where log(x//77)/ » //
logK> » a
11 n - b
log C/ » x/
The technique for solving a linear regression can be found in any introductory statistics textbook and is
also a common feature of most moderately priced electronic calculators. (Note: linear regressions are
sometimes referred to as the line of best fit or method of least squares.) For the sake of completeness,
the constants may be solved (with n* as the number of pairs of data points) using
X//77/)
1 n*(£log C/x log x/mf) - (£log
[18]
The following example is given to illustrate the application of the Freundlich equation. Previous work
showed that the adsorption of arsenate by kaolinite could be characterized by using a 1:10 soil: solution
ratio (chapter 9) and that the system reached a steady state after 24 hours. Under these experimental
conditions, 17 dilutions of a stock KH2AsO4 solution were mixed with an NBS rotary extractor with kao-
linite for 24 hours. Table 13 contains all the data needed to construct an isotherm and also includes the
Table 13 Data reduction for arsenic adsorption at 25°C by a kaolinite clay
sample (volume of solution, 200 ml_)
Initial
cone
(mg/L)
4.89
10.0
15.2
19.9
19.9
19.9
29.9
40.3
49.4
80.5
80.5
80.5
98.8
121.0
137.7
160.3
160.3
Equilibrium
cone
(mg/L)
1.20
3.56
6.78
10.1
10.1
10.3
17.6
25.0
33.4
58.4
59.5
58.9
76.3
92.6
109.4
128.3
129.7
Adsorbent
wife)
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
20.42
Amount
adsorbed
O^g/g)
36*
64
84
98
98
96
123
153
160
221
210
216
225
284
283
320
306
PH
8.30
8.26
8.26
8.19
8.23
8.25
8.16
8.03
8.02
7.77
7.80
7.83
7.69
7.56
7.50
7.27
7.26
EC (dS/m)
160
168
170
185
185
185
205
221
240
305
313
305
350
385
413
434
430
* Sample calculation:
_x_ _ (Initial cone. - equil. cone.) x volume of solution
m ~ weight of adsorbent
(4.89 mg/L-
20 mg/L)x 0.200 L =
mg/g =
60
-------
linite for 24 hours. Table 13 contains all the data needed to construct an isotherm and also includes the
pH and electrical conductivity (EC) of each solution, determined as recommended at the ends of chapters
5 and 6.
In this example (table 13),
and thus
log Kf = 1.536 1//7 = 0.452
•%• = 34.328 (As)0'452
[19]
where As is the equilibrium concentration of arsenic in solution (mg/L). The units mg/L are equivalent to
u.g/mL, and therefore the units of Kf are mL1/"u.g(1"1/n)/g~1 from
x/m(u.g/g) =
As (u.g1/n/mL1/n)
The 1/nterm has no units. The selection of the units for x/m and the equilibrium solute concentration will
determine the units of Kf in a given situation. When Mn < 1 , the units used must be considered when ad-
sorption constants are compared from different sources (see Bowman, 1981 ; Hassett et al., 1983).
Thus, equation 19 becomes a predictive equation capable of describing the adsorption data. The reader
may wish to use the data given in table 13 to verify equation 19. For example, equation 19 should not be
used to predict x/m at equilibrium concentrations greater than 130 mg/L; to do so requires the collection of
data in this higher concentration range. The validity of this cautionary note becomes apparent when one
considers that the Freundlich equation predicts infinite adsorption at infinite concentrations; hence, any
soil or clay would have an unlimited capacity to retain chemicals dissolved in water. Not only would an infi-
nite capacity be thermodynamically inconsistent, but experience has shown that the extent of adsorption
is ultimately limited by the surface area (or some portion of the surface) of the adsorbent. Thus, there are
two drawbacks in using the Freundlich equation: (1) it cannot be extrapolated with confidence beyond the
experimental range used in its construction, and (2) it will not yield a maximum capacity term, which in
many cases is a convenient single-value number that estimates the maximum amount of adsorption be-
yond which the soil or clay is saturated and no further net adsorption can be expected.
The Langmuir Equation
The Langmuir equation has given rise to a number of Langmuir-type expressions that have been widely
used to describe adsorption data for solid-liquid systems. The most commonly used expression may be
generalized as
KLMC
m
[20]
where x is the amount or concentration of the solute adsorbed, m is the mass of the adsorbent, C is the
equilibrium concentration of the solute, and KL and Mare constants.
Langmuir (1918) derived an expression similar to equation 20 to describe the adsorption of gases on sol-
ids (flat surfaces of glass, mica, and platinum). He generalized that the Freundlich equation was unable to
describe the adsorption of gases when the range of pressures was large. Langmuir's original derivation
was based on the premise that during the adsorption of gases, a dynamic equilibrium is established in
which the rate of condensation (adsorption) is equal to the rate of evaporation (desorption). Derivations of
the Langmuir and Langmuir-type equations for gas-solid interactions are given elsewhere (Langmuir,
1918; Hayward and Trapnell, 1964; Ponec et al., 1974). Langmuir-type expressions for ion exchange reac-
tions in soils have also been derived (Sposito, 1979; Elprince and Sposito, 1981).
61
-------
The applicability of Langmuir-type equations to solid-liquid systems has been a controversial topic in re-
cent years (see Harter and Baker, 1977; Veith and Sposito, 1977; Barrow, 1978; Sposito, 1982). How-
ever, this controversy is concerned not with the ability of the equation to simply describe the adsorption
data, but with interpretations of adsorption mechanisms and energetics that are based on the results of
applying Langmuir-type expressions.
Some investigators have concluded that the Langmuir constant (KL) is somehow related to the bonding
energy between the adsorbed ion and the adsorbent, but that specific functional relationship is uncertain.
The constant Min equation 20 is generally accepted as the adsorption maximum of the adsorbent with re-
spect to the specific solute and is interpreted as the maximum amount or concentration that an adsorbent
can retain.
Langmuir-type equations are frequently used because of their ease of application. Like the Freundlich
equation, such equations contain only two constants, both of which are positive-value numbers that can
be statistically solved when equation 20 is cast in a linear form. Two linearized expressions are possible:
x/m
x/m
1 . C
KLM M
1 , J_
KLMC M
[21]
[22]
The linearized form of equation 21 is sometimes referred to as the "traditional linear Langmuir equation,"
and equation 22 is called the "double-reciprocal Langmuir equation." The latter is more suitable for situ-
ations in which the distribution of equilibrium concentrations tends to be skewed towards the lower end of
the range of the equilibrium concentrations. As indicated above, linearized Langmuir-type expressions
such as equations 21 and 22 are equivalent to a simple linear regression,
y,= a + bx,
where the traditional linear Langmuir equation is
y\
a
b
and the double-reciprocal form is
y\
a
b
X,
= (C/x/m),
= 1/Ki.M
= 1/M
= C,
(1/X//77);
MM
VKLM
1/C,
The techniques for solving either equations 21 or 22 are the same as those used to solve the linear form
of the Freundlich equation (eq. 16). From the data set given in table 13, application of the linear Langmuir-
type equations yields:
Traditional Linear Langmuir
a =
KLM
= 0.0792
[23]
62
-------
and thus
Double-Reciprocal Langmuir
= - = 0.0028
_x _ 3.568 x10"2(353.856)C
m ~ 1 + 3.568x10~2(C)
a = -J; = 0.0050
[24]
[25]
[26]
b =
;•= 0.0297
[27]
and thus
_x = 0.1702(198.098) C
/n ~ 1+0.1702(0)
In this example, the units for the adsorption maximum are the same as for x/m (ug/g), and the units for K,
are liters per milligram:
M(ng/g) C(mg/L)
i+/C(L/mg)C(mg/L)
[28]
The selection of units for x/m and the equilibrium solute concentration determines the units for M and KL.
Equations 25 and 28 are predictive expressions that can describe the adsorption of arsenic by kaolinite.
The reader should work through these examples to verify the results. In the previous examples, the iso-
therm constants were derived by linear regression. Kinniburgh (1986) recommended that isotherm con-
stants be solved by nonlinear regression (nonlinear least squares) to obtain more accurate values than
those derived by linear regression. A short BASIC program using a nonlinear least-squares method for de-
termining Langmuir constants was written by Persoff and Thomas (1988).
63
-------
-------
CHAPTER 15
SELECTION OF ADSORPTION EQUATIONS
Three isotherm regressions were used to describe the example data set given in table 12. Given the se-
lection of different models, one equation usually will describe the results with the greatest accuracy No
clear consensus has been reached on which equation (Freundlich or Langmuir-type) is the most reliable
for simply fitting data. Barrow (1978) objected to the application of Langmuir-type expressions but his ob-
jection was based on theoretical considerations. Singh (1984) compared five adsorption equations and
found that the Freundlich equation was the most accurate in describing the adsorption of SO42' by soils
Polyzopoulos et al. (1984) compared four adsorption equations in a study concerned with phosphate ad-
sorption by soil. They found that Langmuir-type or Freundlich expressions described the data with compa-
rable success. K
Generally the choice of an equation is based on the coefficient of determination (r2) obtained in a given
case and the equation's simplicity (Polyzopoulos et al., 1984). The Freundlich and Langmuir equations
each contain only two constants and are easily solved.
The coefficient of determination (sometimes called the goodness of fit) is a measure of how closely the re-
gression line fits the data, and may be calculated using equation 29:
, £(y/-y)2
I(y/-y)2 [29]
where y, is the value of the dependent variable predicted by the regression, y, is the value actually meas-
ured, and y is the arithmetic mean of all y,. The value of r2 will always be between 0 and 1 inclusive If all
of the points are close to the regression line or, in this example, if all of the adsorption data plot closely to
the statistically constructed adsorption isotherm, the corresponding r2 will be close to 1 The application
of equations 16, 21, and 22 to the data set in table 12 yielded dissimilar r2 values:
Freundlich 0.996
traditional linear Langmuir 0.954
double-reciprocal Langmuir 0.916
When the coefficient of determination is used as a criterion, the Freundlich equation best described the
adsorption data, although the traditional linear Langmuir expression would also yield satisfactory results
Figure 41 clearly shows that the double-reciprocal linear Langmuir equation did not fit the adsorption data
well and that the traditional linear form tended to overpredict adsorption in the upper part of the isotherm
Obviously the high r value associated with the Freundlich equation is reflected by the closeness of fit of
the isotherm with the data.
Obtaining a reliable fit of adsorption data with the chosen equation (so that r2 values are close to 1) is a
major concern in the construction of adsorption isotherms. However, in some cases, a low r2 value will be
obtained regardless of the equation used, raising concerns that the adsorption constants actually have lit-
tle meaning. Probably the simplest statistical test for such situations is to use f-statistics to examine
whether the sample correlation coefficient (r) is significantly different from a population correlation coeffi-
cient (p) where p = 0. This test appears in most introductory statistics textbooks and will not be discussed
here.
65
-------
320-
280-
240-
_ 200-
"H
I
160-
120-
Traditional linear
Langmuir equation
(r2 - 0.954)
Freundlich equation
(r2 = 0.996)
Double-reciprocal
near Langmuir equation
r2 =0.916)
60
SO
100
120
Equilibrium arsenic concentration (mg/L)
Figure 41 Adsorption of arsenic by a kaolinite clay sample at 25°C, described by
the traditional linear Langmuir, double-reciprocal Langmuir, and Freundlich equation.
The mean pH of the soil-solute suspensions was 7.8.
66
-------
CHAPTER 16
APPLICATION OF BATCH-ADSORPTION DATA
Adsorption data are used in describing the partitioning of chemicals between soils and water, and have
been used successfully as input parameters in many models describing the movement of chemicals in
soil (Dragun, 1988). Batch-adsorption data have also been applied successfully to groundwater systems.
For example, Curtis et al. (1986) found that the rates of movement of halogenated organic solutes in a
sandy aquifer in Canada were in good agreement with those predicted from adsorption data. In a study
described by CH2M Hill, Inc. (1986), data on the distribution and concentration of organic solutes at a
field site in Indiana were in good agreement with data predicted from laboratory adsorption studies.
Miller et al. (1989) found that isotherms generated with a batch technique were very similar to those de-
rived from flow-through column experiments for the adsorption of anions by soils. Adsorption tended to be
greater in the flow systems, possibly because of precipitation or reduced competition between the solutes
and desorbed antecedent species.
This chapter is a brief introduction to the application of batch-adsorption data in calculations of solute
movement through compacted landfill liners. These calculations are used particularly for estimating the
minimum thickness of liner required to prevent pollutant movement beyond a certain depth of the liner for
a specified period of time. As leachate moves through a liner, the movement of chemical solutes in the
leachate may be retarded if they are adsorbed by the liner. We may define R as the ratio of the velocity of
the leachate to that of the solute,
R « leachate/ Solute [30]
The flterm is called the retardation function or factor. When the solute is not retained by the liner, /?
equals 1 : the solute moves at the same velocity as the leachate. Increasing degrees of adsorption yield
larger values for R. The retardation factor may also be defined by an empirical relationship (Freeze and
Cherry, 1979, and references cited therein) as
8
where pb is the dry bulk density of the liner, Kd is a distribution coefficient, and 6 is the volumetric water
content of the liner. The distribution coefficient is a parameter that describes the partitioning of solutes be-
tween the leachate and liner soil materials at equilibrium. The distribution coefficient may be defined as
f - **
where S is equal to x/m (the amount adsorbed per mass of adsorbent), and C is the equilibrium concentra-
tion of the solute. In other words, equation 32 is the slope of an adsorption isotherm.
Before equation 31 can be used, a functional relationship for dS/dC must be determined. The possible so-
lutions range from simple assumptions to complex numerical solutions. The simplest case is one in which
the adsorption of the solute conforms to a Freundlich equation (chapter 14) isotherm where the 1/n term
is unity,
— = S = KfC"" = K,C [33]
67
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Such an isotherm is termed linear; a plot of S versus C is a straight line. The slope of this type of plot
yields Kd,
-= = KforKd
[34]
hence,
[35]
The retardation factor is unitless; if Kd is in milliliters per gram, then the units of the term
p6 (g/cm3) K"d(mL/g)/0 (cmVcm3)
cancel because 1 cm3 = 1 ml.
When a linear isotherm is used, the Freundlich constant (Kf} reduces to the simple partition constant
(Kd), a single-value number used to calculate solute-adsorbate partitioning at any equilibrium concentra-
tion of the solute. Because of its mathematical simplicity, this approach (the linear isotherm assumption)
has been widely used and may be valid for many dilute systems. When the adsorption isotherm of a sol-
ute is a nonlinear function (\ln *= 1), the retardation factor is concentration-dependent:
hence,
_d_
dC
R(C) = 1 +
n
-1
en
[36]
[37]
Equation 37 is complicated by the fact that the numerical value of R depends on the concentration of the
solute. Solute movement may be seriously underestimated if, when dealing with nonlinear isotherms, in-
vestigators assume that a constant retardation factor is valid for a given system.
Rao (1974) developed an empirical technique to estimate a weighted-mean adsorption partition coeffi-
cient (Kd) for the Freundlich equation. In this technique, the rate of adsorption with respect to concentra-
tion (3S/0C) is normalized by the total amount of solute in a given concentration range,
Kri =
fdS
~*r{
1 "v/
i»Cn
\dC
n
dC
A/n
f'
dC
Co
[38]
The solute concentration C0 isjhe highest concentration (before contact with the adsorbent). If Kf is in
units of mL1/nu,g(1"1/n)/g, then K^may be expressed in milliliters per gram, since
mi 1
mL
9
A weighted-mean retardation factor (R) may be calculated as
68
-------
R = 1 +
8
[39]
In a study concerned with pesticide adsorption by a soil sample, Davidson et al. (1980) found that the er-
ror introduced by assuming linear adsorption isotherms was not serious at low concentrations
(<10 mg/L) but became significant at higher concentrations. Van Genuchten et al. (1977) proposed
an alternative method for isotherm linearization that the reader may wish to examine.
To demonstrate possible applications of these concepts, the following examples are presented to illus-
trate how batch-adsorption data are used to estimate clay liner thickness.
In this hypothetical example, the metallic waste described in appendix B is to be placed into a disposal ba-
sin lined with Cecil clay loam (see appendix A). The soil, which was graded, blended, and compacted,
has a saturated hydraulic conductivity of 10"7 cm/sec. The major concern of the company operating the
disposal facility is the possible uncontrolled movement of a leachate plume containing high concentra-
tions of lead in solution. In a preliminary analysis, this company conducted batch-adsorption experiments
using a Pb(NO3)2 salt and samples of the Cecil soil (table 14). The question posed is, what must the mini-
mum thickness of the liner be to attenuate the lead from solution over a 5-year operating life and a 30-
year post-closure period?
Several approaches can be used to answer this question. For each approach, the mean pore velocity of
the leachate through the liner must be calculated by using Darcy's law as
V = Ksati/n
[40]
where /Csat is the saturated hydraulic conductivity of the liner, /is the hydraulic gradient (dH/dZ), and neis
the effective (water-conducting) porosity of the liner.
If we assume saturated conditions, subject to steady-state flow through an isotropic liner over time t, and
neglect the effects of dispersion and diffusion, equation 40 can be combined with equation 31 to yield
Z= tKsali/Rng [41]
where Zis the estimated vertical distance of migration of the solute in centimeters, and t is time in sec-
onds.
Equation 41 treats solute movement as a piston-flow problem: a chemically uniform slug of leachate mov-
ing downward. This equation is simple and may readily be used to estimate the minimum thickness of a
liner. The application of the equation is simplified by assuming that the isotherm is linear. In this example
(table 14 and figure 42), a linear regression of the data through the origin (Steel and Torrie, 1960) yielded
•£ = S = 342 (Pb)
Moreover, the liner is assumed to have the following properties:
ne =
r =
Pb =
"sat =
/ =
1.1038 x109 sec =
0.09 cm3/cm3
0.36 cm3/cm3
1.7g/cm3
1 x10"7 cm/sec
dH/dZ= 1 cm/cm, and that
35 years
69
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With these assumptions, the retardation factor becomes
R = 1+-
and solving equation 40 becomes
Z = (1.1038 x 109) (1 x 10~7) (1)71619(0.09) = 0.8 cm
On the basis of this approach, the compacted liner would have to be at least 1 cm thick to adsorb lead
over a 35-year period. But although the application of a linear isotherm yields a reasonable coefficient of
determination (r = 0.95), inspection of figure 42 indicates that this approach overestimates lead adsorp-
tion at high lead concentrations and underestimates adsorption at lower concentrations. The adsorption of
lead (table 14) is more accurately described by a Freundlich equation,
- = S = 291(Pb)
°-492
As a second level of refinement, the nonlinearity of the isotherm is considered using equation 38 to esti-
mate a weighted-mean retardation factor (Davidson et al., 1980). An appropriate value for CQ was deter-
mined from a laboratory extract of the metallic waste sample (appendix B), which suggests that the
maximum amount of lead that initially will come in contact with the liner is approximately 15 mg/L Pb A re-
vised retardation factor is derived from equation 38:
and the minimum thickness, based on the weighted-mean retardation factor, is
Z = (1.1038 x 109) (1 x 10~7) (1)7348(0.09) = 3.5 cm ,
Thus, when the nohlinearity of the isotherm is considered, the minimum thickness of the liner is estimated
to be about 4 cm. As a third level of refinement, the chemical composition of the leachate was consid-
ered. The first two estimates were based on lead adsorption from a pure Pb(NO3)2 solution. Laboratory ex-
Table 14 Lead adsorption data for a Pb(NO3)2 salt and the Cecil clay
(volume of solution, 200 mL; adsorbent weight, 10.18 g)
Initial
cone
(mg/L)
2.07
5.11
5.11
6.22
7.28
10.2
10.2
12.4
14.6
14.6
Equilibrium
cone
(mg/L)
0.05
0.11
0.11
0.16
0.22
0.41
0.43
0.65
0.94
0.94
Amount
adsorbed
(x/m) as \ig/g
61
100
100
121
141
196
195
235
273
273
PH
4.79
4.74
4.75
4.74
4.73
4.68
4.67
4.66
4.62
4.62
EC
(dS/m)
27
33
35
34
33
39
40
45
45
43
70
-------
280
240
200
160
o
+->
c
3
o
120
80
40
0.0 0.2 0.4 0.6 0.8 1.0
Equilibrium lead concentration (mg/l)
Figure 42 Lead adsorption by Cecil clay loam at pH 4.5 and at 25°C,
described by a linear Freundlich equation forced through the origin.
tracts, of the waste also contained large concentrations of zinc (appendix B). The adsorption of lead from
the extracts was significantly less than that from the pure Pb(NO3)2 solution, presumably because of com-
petitive interactions between Zn2+ and Pb2+ for adsorption sites. The net effect is that lead could be more
mobile in the presence of zinc. The adsorption of lead by Cecil from the laboratory extract of the waste
can be described by
= s = 70 (Pb)
°-481
71
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If the minimum liner thickness is recalculated using these isotherm constants and equations 38 and 40,
the thickness is estimated to be about 15 cm, again assuming that the initial lead concentration in the
leachate is 15 mg/L Clearly, migration distance estimates based on adsorption data from pure, single-sol-
ute tests may underestimate the minimum thickness of liners because these estimates fail to account for
competitive interactions that may significantly reduce adsorption. At the next level in refining the esti-
mated liner thickness, the effects of dispersion and diffusion are considered. In saturated homogeneous
materials that are subjected to steady-state flow conditions along a flow path z, the change in solute con-
centration as a function of time may be generalized (Ogata, 1970; Bear, 1972; Boast, 1973; Freeze and
Cherry, 1979) as
ac2
'az2'
ac pjb as
'• dz ~ e ar
[42]
where
Pb
6
S
t
concentration of the solute,
effective diffusion-dispersion coefficient (distance/time) along the flow path z,
mean convective flow velocity (distance/time) along the flow path z,
bulk density (wt/vol) of the material,
volumetric water content (vol/vol),
amount of solute adsorbed per mass of adsorbent (x/m), and
time
Equation 41 can be rearranged as
[43]
«* l/^. w«-
where R is the retardation factor
The analytical solution to this second-order differential equation (Ogata, 1970) is given by
C 1 I" , ( z- Vt* } (Vz} . ( z+Vt*
-pr = — erfc ———^ +exp \-=r erfc ^
Co 21 l2(D2/*)0-5l \Dz\ 2(Dzif*)a5
[44]
where CICQ = ratio of the solute concentration at time t and distance z to the intial solute
concentration C0,
erfc = complementary error function,
V - average linear pore water velocity (cm/sec),
DZ s vertical dispersion coefficient (cm2/sec),
t * = retarded time (actual time divided by the retardation factor of R or R), and
z = vertical distance of migration (cm).
Furthermore, Dz - aV+ D*, where a is the dispersivity (cm) and D* is the effective diffusion coefficient in
porous media (cm2/sec).
In the following examples, the three previous liner thickness estimations were recalculated using equation
43. The only additional information needed to conduct this analysis was a dispersivity value. The disper-
sivity has been found to be scale-dependent and is estimated to be about 10% of the distance measure-
ment of the analysis (Gelhar and Axness, 1981). A diffusion coefficient of Pb2+ in soil of 1 x I0"6cm2/sec
was used in this analysis (Daniel et al., 1988). In figure 43, the relative concentration (C/C0) is shown as a
function of distance of migration after 35 years. Case A represents the first situation, in which the adsorp-
tion of lead, a Pb(NO3)2 salt, was assumed to be depicted by a linear isotherm. Case B corresponds to
the second calculation, in which a weighted-mean retardation factor was used with the Pb(NO3)2 solute-
72
-------
c
o
01
EC
8 0.4. -
0.2 _
0.0
12 16 20
Distance of migration (cm)
24
28
32
Figure 43 Predicted distance of lead migration in Cecil clay loam after 35 years, based on three approaches:
case A (linear isotherm assumption, Pb(NO3)2 salt); case B (weighted-mean retardation factor, Pb(NO3)2 salt);
and case C (weighted-mean retardation factor, multicomponent waste extract).
soil system. Case C represents the adsorption of lead from the multicomponent-waste extract, coupled
with the corresponding weighted-mean retardation factor. Case C, which takes into account dispersion,
indicates that lead may move farther than predicted by an elementary piston-flow model (eq. 40). The ef-
fects of diffusion on the predicted migration distances were negligible (not shown).
An element of interpretation is involved in evaluating graphs (see fig. 43) for the purpose of estimating
liner thickness. A judgment must be made as to which C/C0 ratio, for practical considerations, translates
into the minimum significant concentration. In this hypothetical example, the regulatory agency decided
that a lead concentration of <0.05 mg/L (the U.S. drinking water standard for lead) would be an opera-
tional definition of the compliance concentration.
If the initial lead concentration is 15 mg/L, the lead concentration of <0.05 mg/L is predicted to occur at a
depth of 5 cm in case A and at 10 cm in case B. The results for case C represent the fourth level of refine-
ment in this analysis, yielding the most accurate liner thickness estimate. After 35 years, the concentra-
tion of lead in solution would be reduced to <0.05 mg/L at a depth of 35 cm on the basis of these calcu-
lations. Consequently, the minimum liner thickness would be 35 cm. The actual thickness necessary in a
field application must be somewhat greater to allow for nonequilibrium conditions and the normal engi-
neering safety factors. The application of batch-adsorption data provides an estimation of boundary condi-
tions, i.e., the minimum thickness.
In summary, the minimum liner thickness for a hypothetical liner varied from 1 to 35 cm, depending on the
approach (table 15). Liner thickness estimates can be refined further if the adsorption data can be inte-
grated with other information, such as the solubility of solid phases, oxidation-reduction equilibria, co-
solvent effects, and the design and performance of on-site earthen liners. This information would include
73
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seepage rate through the cover, fraction of seepage that will pass through the liner, and other water flux
information that would allow calculation of the distribution of a pollutant in soil as a function of time and
space.
Table 15 Approaches for estimating minimum liner thicknesses on the basis of adsorption
Flow model
Piston flow*
Piston flow
Advection dipersionf
Advection dispersion
Piston flow
Advection dispersion
Isotherm
treatment
linear
nonlinear
linear
nonlinear
nonlinear
nonlinear
Solute system
single solute
single solute
single solute
single solute
mixture:}:
mixture
Minimum liner
thickness (cm)
1
4
5
10
15
35
* Represented by equation 41.
t Represented by equation 44.
t Laboratory extract.
74
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CHAPTER 17
LABORATORY PROCEDURES FOR GENERATING ADSORPTION DATA
Procedures for the determination of the soil:solution ratio, equilibration time, and other parameters neces-
sary for the construction of adsorption isotherms are contained in this chapter:
page
1 Scope of Application 69
2 Summary of Methods 69
3 Interferences 69
4 Terminology and Definitions 70
5 Equipment and Procedural Requirements 71
6 Volatile Organic Solutes: Experimental Considerations 73
7 Preparation of Adsorbents 74
8 Determination of Soil:Solution Ratios for Ionic Solutes 75
9 Determination of Soil:Solution Ratios for Nonionic Solutes 77
10 Determination of Equilibration Time 77
11 Construction of the Environmentally Conservative Isotherm 78
12 Construction of the Constant Soil:SoIution Ratio Isotherm 79
The rationale for these procedures, presented in previous chapters, should be studied before attempting
to use them. This chapter refers to parts of the TRD that elucidate topics relevant to a specific procedural
step. The flow diagram (fig. 44) summarizes the procedures and their interrelationships.
1 Scope of Application
1.1 The extent of adsorption of a chemical (solute) from solution by an adsorbent (sediment, soil,
clay) at equilibrium can be estimated using these procedures.
1.2 These methods apply to the generation of adsorption isotherms or curves for inorganic and or-
ganic (volatile and nonvolatile) compounds; these isotherms indicate how the extent of adsorption var-
ies with the equilibrium concentration of the solute.
1.3 Contingencies within these methods allow for the construction of adsorption isotherms at various
solute concentration ranges.
1.4 These methods can be used for constructing adsorption isotherms to study the adsorption behav-
ior of solutes in synthetic waste solutions, laboratory extracts, or field leachates including aerobic and
anaerobic solid-liquid systems.
2 Summary of Methods
The experimental design of these methods is based on a batch technique as opposed to a column ap-
proach. Two general techniques for obtaining adsorption data are incorporated in these methods. The first
technique involves mixing a batch of solutions, each with the same volume but containing serial dilutions
of the initial solute concentrations, with a fixed mass of adsorbent in each reaction vessel. The second
technique involves mixing a batch of solutions, each with the same volume and initial concentration of the
solute, with different amounts of the adsorbent. In either case, the change in solute concentrations after
contact with the adsorbent provides the basis for the construction of adsorption isotherms (chapter 12).
The appropriate soihsolution ratios and equilibration times are determined to maximize the accuracy of
the adsorption isotherm and to complement analytical capabilities.
3 Interferences
Solutes of unknown stability must be handled with care to determine whether precipitation, hydrolysis,
photodegradation, microbial degradation, oxidation-reduction (e.g., Cr3* to Cr ), or other physicochemical
processes are operating at a significant rate within the time frame of the procedure. The instability and
hence loss of the solute from solution may affect the outcome of this procedure (see chapter 4). The com-
75
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Generation or collection
of solution containing
test solute
Determination of solute
solution stability
(hydrolysis, photodegradation,
microbial degradation,
and volatility)
Determination of Interactions
between solute solution and
laboratory equipment
Determination of equilibration time
Construction of constant
soil to solution isotherm (CSI)
Collection of adsorbent
Determination of
percent moisture
Determination of soil to
solution ratios
Preparation of adsorbent
Air drying
Reduction of aggregates
Splitting and subsampling
Construction of environmentally
conservative isotherm (ECI)
Figure 44 Flow diagram of the procedures for generating batch adsorption data.
patibility of the method and the solute of interest may be assessed by determining the differences be-
tween the initial solute concentration and the final blank concentration of the solute. If this difference is
greater than 3%, then the adsorption data generated must be carefully evaluated (see 8.5.11).
4 Terminology and Definitions
4.1 Solute—chemical species (e.g., ion, molecule) in solution
4.2 Solute solutions include:
76
-------
4.2.1 A solution of reagent water containing a known amount of a solute derived from laboratory
reagents.
4.2.2 A solution containing a variety of solutes extracted from a material in a laboratory setting by
the use of methods such as the ASTM-A or ASTM-B extraction procedures. (Note: neither the
EPA Extraction Procedure nor the proposed Toxicity Characteristic Leaching Procedure is recom-
mended. These procedures were designed for waste classification and were not intended to pro-
duce solutions that mimic in-situ leachates.)
4.2.3 A solution containing a variety of solutes collected in a field situation representing a
leachate or waste effluent.
4.3 Adsorption—a physicochemical process whereby solutes are retained by an adsorbent and
concentrated at solid-liquid interfaces (chapter 1).
4.4 Adsorbate—chemical species adsorbed by an adsorbent.
4.5 Adsorbent—substance that adsorbs the solute from solution.
Equipment and Procedural Requirements
5.1 Laboratory equipment
5.1.1 Agitation equipment: the National Bureau of Standards extractor (rotating tumbler) or
equivalent will be used exclusively as the agitation apparatus (chapter 8).
5.1.2 Rotation rate: with procedures involving inorganic, volatile, and nonvolatile organic com-
pounds, the rotary extractor will be operated at 29 ± 2 rpm.
5.1.3 Glove box or bags: when anaerobic adsorbent-solute systems are being handled, these
procedures may have to be conducted in air-tight enclosures filled with an oxygen-free inert gas
(e.g., N2, Ar) to prevent or retard oxidation.
5.2 Phase-separation equipment
5.2.1 Inorganic compounds: a filtration apparatus made of materials compatible with the solu-
tions being filtered and equipped with a 0.45-u.m pore-size membrane filter or a constant-tempera-
ture centrifuge capable of separating >0.1-u.m particles will be used to separate the solid phase
from the solid-liquid suspensions.
5.2.2 Filtration membrane: if filtration is used, the affinity of the filtration membrane for the solute
must be evaluated to prevent errors in the results.
5.2.3 Organic compounds: a constant-temperature centrifuge, compatible with the reaction con-
tainers and capable of separating >0.1-u.m particles, should be used for organic solutes. The
transfer of the organic solute solutions from the reaction containers to centrifuge containers is not
an acceptable procedure because of adsorption, volatilization, and other losses. The reaction con-
tainer should be used as the centrifugation container. Filtration of organic solutions is .rjoj a recom-
mended practice (chapter 7).
5.2.4 Calculation of centrifugation time may be calculated by using equation 1,
t =
9riln(/Vf?f)
2o>2r2(pp-p)
[1]
77
-------
where
t » time(min),
tl = viscosity of water (8.95 x 1 0"3 g/sec-cm at 25°C),
r - partical radius (cm),
pp = partical density (g/cm3)
p - density of solution (g/cm3),
rpm m revolutions per minute,
Rt = distance (cm) from the center of the centrifuge rotor to
the top of solution in centrifuge tube, and
Rb = distance (cm) from the center of the centrifuge rotor to bottom of
the centrifuge tube.
Removal of particles that are as small as 0.1 urn in radius and have a particle density of 2 65 g/cm3 from
a solution with a density of 1 g/cm3 may be estimated using equation 2,
3 71 x 108
5.3 Reaction containers
5.3.1 Inorganic solutes: containers compatible with the rotary extractor should be used with inor-
ganic solutes. The containers shall be composed of materials that adsorb negligible amounts of
the solute. They must have a watertight closure made of chemically inert materials (polypro-pvl-
ene, Teflon, or similar material). The size of the container should be such that the solid and liquid
phases will fill about 80% to 90% of the container.
5.3.2 Nonvolatile organic solutes: amber glass serum bottles and stainless steel centrifuge tubes
or bottles compatible with the rotary extractor and centrifuge are suggested to be used in conjunc-
tion with nonvolatile organic solutes. The container must have a watertight closure made of chemi-
cally inert materials (Teflon, plastic, or similar material). The size of the container must be com-
patible with the centrifuge, and be such that the volume of the solid and liquid phases should fill
80% to 90% of the container.
5.3.3 Volatile organic solutes: amber glass, 125-mL serum bottles (Wheaton no 223787 or
equivalent) fitted with Teflon septa (Pierce no. 12813 Tuf-Bond Discs or equivalent) will be used
with volatile organic solutes. The size of this serum bottle (125 ml) is compatible with several
types and brands of centrifuges. This size provides sufficient volume such that the volume of the
solid and liquid should occupy 1QQ% of the container (i.e., there should be no head space).
5.3.4 Note that the commonly available materials for containers can be ranked starting with the
material most inert with respect to the adsorption of hydrophobic solutes (T.C. Voice written com-
munication, 1986):
Corex,
Pyrex (not much different from Corex),
silanized serum bottles,
other glasses, and
stainless steel, Teflon, and plastic.
5.4 Reagents
5.4.1 Reagent-grade chemicals will be used in all experiments and must conform to the specifica-
tions of the American Chemical Society. Other grades may be used, provided that the reaqent is
pure enough to be used without lessening the accuracy of the determination
78
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5.4.2 Unless otherwise indicated, references to water mean type IV reagent water, as defined in
the Handbook for Analytical Quality Control in Water and Wastewater (EPA-600/4-79-019).
5.5 Solute Solution and Adsorbent Requirements
5.5.1 To construct adsorption isotherms for inorganic solutes, a minimum of 5 liters of solute solu-
tion would be required, based on the use of 200-mL samples of the solute solution with 250-mL
reaction containers. Investigators using reaction containers that are a different size should adjust
the estimated total volume of solution proportionately.
5.5.2 To construct adsorption isotherms for organic solutes, approximately 9 liters of solute solu-
tion would be required, based on the use of 100-mL samples of the solute solution with 125-mL
reaction containers. Investigators using different-sized reaction containers should adjust the esti-
mated total volume of solution proportionately.
5.5.3 The mass of adsorbent required to complete this procedure will vary depending on the vol-
ume of reaction containers, soihsolution ratios, and related factors. Based on 250-mL reaction
containers and the minimum soihsolution ratio of 1:4 (50 g adsorbent per 200 mL of solute solu-
tion), about 2 kg of adsorbent would be required.
5.5.4 This procedure should take about 5 to 9 days, excluding time for analysis.
Volatile Organic Solutes: Experimental Considerations
6.1 Stock solutions could either be purchased as certified solutions or prepared from pure standard
materials (liquid or gaseous phases). Solutions should be prepared in methanol. The use of pipettes
to transfer solutions cannot be recommended; glass syringes should be used to prevent losses due to
volatization. Because of the toxicity of some volatile organic compounds, solutions should be pre-
pared and transferred in a fume hood, and a NIOSH/MESA-approved toxic-gas respirator should be
used by the analyst.
6.2 Preparation of stock volatile solute solutions
6.2.1 Place approximately 9 mL of methanol into a 10-mL ground glass stoppered volumetric
flask. Allow the flask to stand open until all methanol-wetted surfaces have dried. Weigh the flask
with the remaining methanol to the nearest 0.01 mg. Using a syringe, immediately add the test
solute, until the change in weight of the flask corresponds to the desired concentration of the test
solute in the methanol. Be sure that the drops of solute fall directly into the methanol without con-
tacting the neck or sides of the flask. Dilute to volume with methanol, put the stopper on the flask,
and mix by inverting it several times.
6.2.2 Transfer the stock solution into a Teflon-sealed screw-cap vial. Store, with little or no head
space, at approximately 4°C. All stock solutions must be replaced after 1 month, or sooner if a
comparison with quality-control standards indicates a loss of accuracy.
6.2.3 Stabilize the temperature of the stock solution at 20°C before preparing secondary solu-
tions.
6.2.4 Store all solutions so that head space within the container is zero or minimal.
6.3 Preparation of volatile organic compound solutions
6.3.1 Place 990 mL of type IV water that has been boiled and cooled to 20°C into each of a se-
ries of clean 1-L amber-glass bottles. (Generally eight solute concentrations are required for com-
pletion of the adsorption procedures.) Seal the bottles with open-top screw caps fitted with
Teflon-lined septa.
6.3.2 Inject measured amounts of the stock solution (prepared as described in section 6.2) into
each of the bottles. Mix by inverting the bottles several times but avoid excessive shaking, which
may result in partial loss of the solute.
79
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6.3.3 Solutions stored in containers with head space are not stable and should be discarded 1
hour after preparation if not used in an experiment.
6.4 Filling of reaction containers
6.4.1 Immediately upon completing the steps described in section 6.3, pour each solute solution
carefully— to minimize agitation— into preweighed reaction containers or other containers that
have a predetermined volume (section 6.5) and contain specific amounts of adsorbent Fill the
containers completely; allow no head space. Shake gently to remove trapped air from the adsorb-
ent. Place the Teflon-faced septum and aluminum seal on the container and invert to be sure no
head space remains.
6.5 Determination of Reaction Container Volume
6.5.1 When transferring the solutions prepared as described in section 6.3 into the reaction con-
tainers, pour them quickly but gently into containers of predetermined weight or volume.
6.5.2 Because the volume of solute solution is not measured during transfer into the reaction con-
tainers, this volume is determined indirectly.
6.5.3 Reaction vessels containing the same amount of adsorbent as those described in section
6.4 to determine the container volume for each soihsolution ratio. Pipette type IV water is pipetted
into each container until there is no head space. With a calibrated syringe, measure the water
added to each of the containers. The amount of solute solution referred to in section 6 4 should
be the volume as determined in this section (6.5).
6.5.4 Alternatively, determine the volume of solution added (described in section 6 4) by weigh-
ing | the container with the adsorbent before and after adding the solution. The weight can be con-
verted to a volume if the density of the added solution is known.
6.6 Throughout all experiments, use blanks to determine effects of adsorption/desorption from con-
tainers as well as losses due to volatilization. Refer to section 8.5.1 1 for discussion of blanks.
6.7 For more information on preparing solutions for volatile constituents or the analyses of these con-
fnSs^^^
7 Preparation of Adsorbents
7.1 Spread samples of adsorbents such as soils, clays or sediments out on a flat surface in a laver
no more than 2 to 3 cm deep. Then allow them to air dry, out of direct sunlight, until they are in
equilibrium with the moisture content of the room atmosphere. The sample should be dried enough to
facilitate processing and subsampling. Do not oven-dry samples (chapter 2). Process anaerobic sam-
ples in a similar manner for these and subsequent steps, but these operations should be conducted in
a glove box or glove bag filled with an oxygen-free inert gas (e.g., N2 or Ar) to prevent oxidation
7.2 Weigh the entire sample after it has been air-dried. Pass the sample through a 2-mm-screen
sieve. Using a clean mortar and rubber-tipped pestle, crush large aggregates without grinding the
sample. Aggregates such as pebbles and stones that cannot be crushed should be removed col-
lected, and weighed.
7.3 Mix the sieved material until the sample is homogeneous. To obtain subsamples of size use a rif-
rSno Der7 • °m|- °SeJ unbiased sP|ittin9 Procedure (Annual Book ofASTM Standards: method
So£&™^ to Testing Size' in part 1 4: or method D2013-72- p
7.4 Determine the moisture content of the air-dried sample by using method D221 6, Laboratory De-
termination of Moisture Content of Soils, from Annual Book ofASTM Standards, part 1 9.
7.5 Determine the mass of the sample, corrected for moisture content, required for study.
80
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7.5.1 Determine the air-dry soil (adsorbent) mass equivalent to the desired mass of oven-dried
soil:
A = Ms[1 +(M/100)]
where A = air dry soil mass (g),
Ms = mass of oven-dried soil desired (g), and
M = percent moisture
8 Determination of SoihSolution Ratios for Ionic Solutes
8.1 A series of soihsolution ratios ranging from 1:4 to 1:500 should be tested and evaluated for the
construction of adsorption isotherms (chapters 9 and 11).
8.2 The soil:solution ratio is defined as the oven-dry equivalent mass of adsorbent in grams (section
7.5) per volume in milliliters of solution.
8.3 Recommended soihsolution ratios are 1:4,1:10,1:20,1:40,1:60,1:100,1:200,1:500. Ratios
greater than 1:500 are rarely needed for most ionic solutes. In circumstances requiring soihsolution ra-
tios greater than 1:500 that meet the criteria outlined in section 8.5.14, use the ratios 1:1000,1:2000,
1:5000 and 1:10,000. The determination of a soiksolution ratio may be an iterative process, whereby
the eight ratios between 1:4 and 1:500 are tested before attempting the extremely "dilute" systems
(1:1000, and higher). Using an iterative process will reduce the amount of solute solution used, and
will help ensure that enough solution will exist to complete the entire procedure. Ratios less than 1:4
should not be used because of limitations in mixing.
8.4 An example of how different soihsolution ratios are made is given below for an air-dried sample
with a moisture content of 3%:
Soihsolution
ratio (g/mL)
1:4
1:10
1:20
1:40
1:60
1:100
1:200
1:500
Air-dry
weight (g)
51.5
20.6
10.3
5.15
3.43
2.06
1.03
0.412
Oven-dry equivalent
of adsorbent (g)
50.0
20.0
10.0
5.00
3.33
2.00
1.00
0.400
Volume of solution
containing solute (ml_)
200
200
200
200
200
200
200
200
8.5 Soil solution procedure
8.5.1 Calculate the masses of adsorbent samples for the various soihsolution ratios based on an
oven-dried equivalent weight (section 7.5), such that for nonvolatile solutes, the volume of adsorb-
ent plus solution occupies 80% to 90% of the container and for volatile solutes 100%.
8.5.2 Weigh the samples of adsorbent to be used in the soihsolution series. If handling anaerobic
adsorbent-solute systems, conduct steps 8.5.2 to 8.5.7 in a glove box or bag before placing the
containers on the rotary extractor.
8.5.3 Place the weighed samples into clean, labeled containers.
8.5.4 Pipette the solution containing the solutes (stock solution) into each container holding the
adsorbent. The volume of solution should be identical in all containers.
81
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8.5.5 Pipette the stock solution into a container holding no adsorbent. This sample will be the
"blank" and designated as CB. For each set of tests a minimum of one blank, and preferably
three blanks, should be tested simultaneously and under identical conditions as the samples.
8.5.6 Close the bottles, ensuring a watertight seal, and place on a rotary tumbler for mixing.
8.5.7 Collect, preserve, and analyze an aliquot of the stock solution to determine the initial con-
centration of the solute(s) before contact with reaction containers, adsorbent, phase separation
materials, and other surfaces. This sample will be designated as C0. The volume and preserva-
tion techniques of the aliquot will vary depending on the solute and analytical method.
8.5.8 Continuously agitate samples at 29 ± 2 rpm for 24 ± 0.5 hours at room temperature (22 ±
8.5.9 After 24 hours of agitation, open containers. If the suspensions are anaerobic, return the
containers to a glove box or bags prior to opening the containers and make all measurements in
the inert atmosphere of the glove box or bag. Observe and record the solution temperature oH
and any changes in the adsorbent or solution. '
8.5.10 Separate the solid and liquid phases of each sample, using either centrifugation or filtra-
tion (section 5.2). Determine the electrical conductivity of an aliquot of each supernate (see chap-
ter 6). Collect and preserve aliquots of each supernate of sufficient volume to determine the
solute concentration.
8.5.11 After analysis of all the solutions generated by the soil:solution procedure, compare the in-
itial solute concentration(s) and blank samples to determine whether there was adsorption or
desorption of the solute onto or from surfaces other than the adsorbent. If the difference between
the blank and initial solute concentrations is greater than 3%, the adsorption data must be cor-
rected. To make this assessment, determine the percent difference between the initial concentra-
tion and the blank solute concentration:
%P=(CO:CB)XIOO
where
%D
C0
CB
percent difference,
initial solute concentration (mg/L, u,g/L), and
solute concentration (mg/L, u.g/L) in blank solution.
If %D is a negative value, the solute concentration in the blank was greater than the initial solute
concentration. Subtract the difference in concentration from all adsorption data, excludina the
stock or initial concentration value.
8.5.12 Using the analyzed initial solute concentration and the final solute concentration for the
vanous soil:solution ratios tested, calculate the percent of solute adsorbed:
where
%A
Co
= percent adsorbed,
= initial solute concentration (mg/L, u/L), and
= solute concentration after contact with the adsorbent.
A negative value implies a contamination problem. Examine the laboratory technique and/or
cleaning procedures. The adsorbent may contain previously adsorbed constituents that are
desorbing into solution.
82
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8.5.13 Select a soil solution ratio indicating between 10% and 30% adsorption of the highest sol-
ute concentration. Use this ratio to determine the equilibration time (chapter 10) and to generate
data for construction of a constant soihsolution isotherm (CSI). Often, several soihsolution ratios
will generate between 10% and 30% solute adsorption. Selection of a specific soil-to-solution ra-
tio is the investigator's prerogative, with the limitation that it be one listed in section 8.3 (chapters
9 and 11).
9 Determination of Soil:Solution Ratios for Nonionic Solutes
9.1 For ionic solutes, a suitable soihsolution ratio must be determined empirically, but a useful soil:so-
lution ratio for nonionic solutes (hydrophobic organics) can be calculated if the organic carbon content
of the adsorbent and the water solubility of the solute are known. The equations and their derivations
i for determining the soil:solution ratios for nonionic solutes are given in chapter 10.
9.2 The soil solution ratios listed in section 8.3 most closely matching the calculated soil:solution ratio
shall be used throughout this procedure. If the calculated ratio is in the middle of two ratios listed in
section 8.3, the lower ratio (greatest mass of absorbent per milliliter of solute) is recommended to ob-
tain the highest precision and accuracy.
10 Determination of Equilibration Time
10.1 To determine equilibration time, use the soil:solution ratio as determined in section 8.5.13 for in-
organic solutes and in section 9 for hydrophobic organic solutes.
10.2 Use a minimum of four agitation intervals to determine the equilibration time. Recommended in-
tervals are 1,24,48, and 72 hours, and represent the amount of time the solution and adsorbent are
in contact.
10.3 Weigh the adsorbent on an oven-dry basis (section 7.5) and place into clean, labeled contain-
ers. If handling anaerobic systems, perform steps 10.3 and 10.8 in a glove box or bag before placing
the containers on the rotary extractor.
10.4 Pipette the solute solution into the various containers at the times designated in step 10.2. Im-
mediately cap the container and place on the rotary extractor and agitate at 29 ± 2 rpm at room tem-
perature (22 ± 3°C).
10.5 Pipette the solute solution into a container containing no adsorbent. Agitate this blank sample
for 72 hours.
10.6 Collect, preserve, and analyze an aliquot of the stock solute solution.
10.7 Remove the containers at the designated times from the rotary extractor and record the solution
temperature, pH, and any changes in the adsorbent or solution. If handling anaerobic suspensions, re-
turn the containers to a glove box or bag before opening the containers.
10.8 Separate solid and liquid phases using centrifugation or filtration (section 5.2). Determine the
electrical conductivity of an aliquot of each supernate. Collect and preserve aliquots of each super-
nate in sufficient quantity for determining the solute concentration.
10.9 Determine the rate of change in the solute concentrations at the various times by
%AC =
- C2)
-,x100
where %AC
C2
percent change
concentration of the solute at time t, and
concentration of the solute after 1,24,48, or 72 hours.
83
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10.10 The equilibrium time is defined as the minimum amount of time needed to establish a rate of
change of the solute concentration equal to or less than 5% per a 24-hour interval (see chapter 13).
11 Construction of the Environmentally Conservative Isotherm (ECI)
10) prSedures* be ^"''68 ** ^ Soil:solution (sections 8 and 9> and equilibrium (section
112 If the equilibrium time as determined in section 10.9 is equal to or less than 24 hours, use the
data obtained from the soil:solution procedure to construct an ECI. However, if the equilibrium time is
greater than 24 hours, the soil:solution ratio determination procedure must be repeated at the eauilib-
ECI 6d ^ SeCti°n 1°'9' (Refert° °haPter 12 f°r diSCUSSi°n °f the advanta9es °
ilf hBT » !! SeCti0? 9 yjeld? a Single soil:solution ^io for nonionic solutes, select additional ratios
that bracket the cacuated rat.o. Use a minimum of eight soilrsolution ratios selected from those listed
in section 83. Evaluate these ratios as outlined in section 8.5. When volatile solutes are under study
refer to section 6 for experimental considerations. y'
omtn IT™?1 ,°f ^ Sta PuintS !° Construct an ECL Solution ratios resulting in 'ess than 1 0%
adsorption of the solute should not be used to construct the ECI (refer to chapter 1 2 for justification)
tiona? datT ^^ "** °btained' vary the recommended soihsolution ratios to generate addi-
1 1 .5 Using the data generated by the soihsolution procedure, calculate the amount of solute ad-
sorbed per mass of adsorbent.
1 1 .5.1 Determine the amount of solute adsorbed per mass of adsorbent by
where x/tn
m
C0
C
V
m
amount of solute adsorbed per unit mass of adsorbent,
mass of adsorbent (oven-dried basis) in grams added to reaction container
initial solute concentration (determinedanalytically) before exposure to
adsorbent,
solute concentration after exposure to adsorbent at equilibrium, and
volume of solute solution added to reaction container.
alt iCi°SStrU?2; * a" ECI reqUJ-eS (1) an Vm value for each soil:solution ratio that meets the crite-
ria in 11.2, and (2) the corresponding equilibrium concentration value (C) of the solute.
11.7 Construction of an ECI
1 1 .7.1 Using linear graph paper, plot the equilibrium concentration (C), log C, or C/x/m on the co-
ordinate (x axis) and the corresponding x/m, log x/m, or C value as the
(y axis). Refer to chapter 1 2 for an example.
11*1' Th9
or a La^muir-type
11.7.3 The linear expression of the Freundlich equation is
log (x/m) = Kf+-\/n\ogC
84
-------
where x/m = amount of solute adsorbed per unit mass of adsorbent,
Kf = a constant,
1/n = a constant (sometimes written as A/), and
C = equilibrium concentration of solute after contact with adsorbent.
A linear regression can be used to fit a curve through the data plotted in 11.5.1, where the inter-
cept equals log K/and the slope equals 1/n. An example in which the Freundlich equation is used
is given in chapter 14.
11.7.4 A linear expression of the Langmuir-type equation is:
1
C
•MM
where xlm
KL
M
C
x/m KLM M
amount of solute adsorbed per unit mass of adsorbent,
a constant,
a constant, and
equilibrium concentration of the solute after exposure to adsorbent.
A linear regression can be used to fit a curve through the data plotted in 11.7.1, where the inter-
cept equals 1/Ki.M and the slope equals 1/M. Examples using Langmuir-type equations are given
in chapter 14.
11.7.5 Calculate the coefficient of determination (r2) of the regression. Examples are given in
chapter 15.
11.7.6 Use the equation that results in the coefficient of determination value closest to 1.0 to gen-
erate a curve through the data plotted in 11.7.1.
11.7.7 The data plotted in 11.7.1 and the curve of best fit from 11.7.6 represent an ECI.
11.7.8 Report the following information with the ECI:
• temperature at which the tests were conducted,
• pH and electrical conductivity (EC) of all solute solutions,
• concentrations of stock (Co) and blank (Ce) solute solutions and the factor, if any, used to cor-
rect data,
• soil:solution ratios, corresponding quantity of solute solution and mass of adsorbent, initial
(Co) and final (C) solute concentration, and the percent of solute adsorbed,
• %AC for each equilibration time,
• equation for the line of best fit and corresponding r value, and
• complete description of the adsorbent
12 Construction of the Constant SoikSolution Ratio Isotherm (CSI)
12.1 The CSI requires the initial solute concentration to vary and the mass of adsorbent to remain
constant—unlike the ECI, which holds the initial concentration of the solute constant and varies the
mass of adsorbent in each containers. (Chapter 12 presents advantages and limitations of both tech-
niques.)
12.2 Use the recommended soihsolution ratio, %A between 10% and 30% (section 8), and equilib-
rium time, %AC < 5% per 24-hour interval (section 10), in the construction of a CSI.
12.3 Weigh the adsorbent (mass prescribed by the soiksolution ratio) into clean, labeled containers.
If handling anaerobic adsorbent-solute systems, conduct steps 12.3 to 12.5 in a glove box or glove
bag.
85
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12.4 Make a series of approximately eight dilutions of the stock solute solution so that the series
shows a progress.ve decrease in solute concentration. In the most dilute solution, the solute concen-
Si10"-,? £U !?8 s"fficje"iso that tne amount remaining in solution is above detection limits after con-
tact with the adsorbent. The volume of each diluted solution necessary for construction of the CSI will
depend upon the size of the reaction container used.
12.4.1 The dilution of complex solutions may cause changes in pH and/or redox potential with
the subsequent precipitation of the solute(s) (see chapter 11). Try to limit such reactions, or if not
possible, use the procedures in section 11 to create adsorption isotherms.
12.5 Immediately afterthe dilutions of the stock solute solution, pipette the diluted solutions into the
containers holding the adsorbent. Each solution should have a corresponding container and the vol-
ume of solution in all containers should be equal.
12.6 Place the containers on the rotary extractor at 29 ± 2 rpm at room temperature (22 ± 3°C) Aai-
^ n^ t'me determined in section 1 °- Collect and preserve aliquots of the stock solute solution
and all dilutions using accepted techniques (e.g., Standard Methods for the Examination of Water'
and Wastewater, American Public Health Association, 1985,16th ed., Washington, DC,
p. 38-45).
12.7 After the agitation period, remove the containers from the rotary extractor and open If the sus-
pensions are anaerobic, return the containers to a glove box or bag, then open the containers Ob-
serve and record the solution temperature, pH, and any changes in the adsorbent or solution.
12.8 Separate the solid and liquid phases using either centrifugation or filtration (section 5 2) Deter-
mine the electrical conductivity of an aliquot of each supernate. Collect and preserve aliquots of each
supernate of sufficient volume for the solute concentration determinations.
i!2'9, ??!(m'1? tlle solute concentration'" the stock solution, the dilute solutions before (C0 in equa-
w I , ,- and,aftei: (C!n equation 11 -5-1) exposure to the adsorbent. If significant differences in the
blank solutions (section 8.5.11) are ascertained, the adsorption data must be corrected.
12.10 Using the data generated where the various solute concentrations were exposed to the same
mass of adsorbent, calculate the amount of solute adsorbed per mass of adsorbent (x/m) Refer to
equation 11.5.1 for calculation of x/m.
12.11 Construction of the CSI requires (1) an x/m value for each solute concentration and (2) the cor-
responding equilibrium concentration value (C) of the solute.
12.12 For constructing the CSI, follow the same procedure and reporting requirements as the ECI
Refer to section 11.5 for directions on construction of the ECI/CSI.
86
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93
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APPENDIX A
SUMMARY AND CHEMICAL COMPOSITION OF THE ADSORBENT SOILS
AND CLAYS USED IN THIS STUDY
Eleven different soil materials were used as adsorbents during the development of the batch-adsorption
procedures. Adsorbents and sample locations are summarized in table A1; relevant physicochemical char-
acteristics and mineralogy are summarized in table A2. Also characterized is the chemical composition of
the materials: major elements in table A3 and trace constituents in table A4.
The eleven clays and soils represented a wide range in physicochemical properties and characteristics.
• The Catlin soil is a dark prairie soil (Mollisol) with a relatively high organic matter content in the sur-
face horizon. The clay-size fraction is predominantly illite. The Catlin, an important soil for agricul-
ture, is one of the Mollisols that dominate in the Great Plains.
• The two Cecil soils, Tifton and soil EPA-14, are Ultisols—highly weathered and acidic soils that are
dominated by kaolinite, and iron and aluminum hydroxides. Most of the soils in the southeastern part
of the United States are Ultisols.
• The Vandalia till, an Illinoian-age deposit, is fairly representative of midwestern glacial tills. It is a
sandy till, gray, and calcareous where unweathered; the predominate clay is illite.
• At the sampling site (table A1), the Sangamon Paleosol was a buried soil that had formed in the Van-
dalia till and was overlain by glacial loess. The Sangamon Paleosol, Vandalia (ablation phase), al-
tered (oxidized) Vandalia, and unaltered (unoxidized) Vandalia tills are a common stratigraphic
sequence in Illinois. This sequence is also present at the Wilsonville hazardous-waste site at Wilson-
ville, Illinois.
The soil sample designated as EPA-14 was used by Hassett et al. (1980,1981) and Zierath et al. (1980)
in studies concerned with the adsorption of hydrophobic solutes. The Cecil clay sample from South Caro-
lina was used by Roy et al. (1986) in a study concerned with the adsorption of anionic mixtures. The kao-
linite and illite clay samples have also been used in previous studies (Griffin and Shimp, 1976; Frost and
Griffin, 1977).
Table A1 Summary of adsorbents
Adsorbent
Sample location
Soil horizon
Classification
Catlin silt loam
Cecil clay
Cecil clay loam
EPA-14
Illite
Kaolinite
Sangamon Paleosol
Tifton loamy sand
Vandalia till member
ablation
altered
unaltered
Champaign, Illinois
Spartanburg, South Carolina
Cecil, Georgia
Ceredo, West Virginia
Elizabeth, Illinois
Pike County, Illinois
Macoupin County, Illinois
(near Sawyerville)
Tifton, Georgia
Glasford Formation
Macoupin County, Illinois
(near Sawyerville)
(near Sawyerville)
(near Eagerville)
AI
B2,
Ap
A
—
—
Bt
Ap
B3
C2
C4
Typic Argiudoll
Typic Hapludult
Typic Hapludult
unknown
—
—
unknown
Plinthic Paleudult
—
—
—
95
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Table A2 Selected physicochemical characteristics of clays and soils used in developing this TRD
Adsorbent
Catlin silt
loam
Cecil clay
Cecil clay
Innm
loam
EPA-14
Illito clay
Kaolinite clay
Sangamon
Paleosol
Tifton loamy
sand
Vandalia till
altered
unaltered
ablation
phase
PH*
6.1
4.5
4.6
4.5
7.9
8.1
6.1
4.7
7.4
7.5
6.4
%
Sand
11
31
32
2
0
0
45
85
45
45
56
%
Silt
69
12
17
63
0
0
25
9
38
40
21
% %Organic
Clay carbon
21
58
51
34
100
100
30
5
17
15
23
4.04
0.34
ND§
0.48
1.81
0.51
0.10
ND
0.18
0.34
0.10
CECf
(meq/
100 g)
18.1
3.7
3.8
18.9
20.5
15.1
16.7
1.9
6.6
4.9
10.5
Surface
area$
(m2/g)
14.8
36.9
29.7
145fl
ND
34.2
22.9
1.7
7.3
5.6
10.6
Clay analysis (%)
Illite
55-67
<5
5-6
13
70
8
33-36
0
71-77
75-82
32-58
Kaol-
inite
5-15
68-92
79-92
37
0
87
7-14
73-96
3-10
4-19
2-6
Expand-
ables
24-30
3-32
2-16
14
0
5
50-60
4-27
18-19
6-9
32-29
Other
clay-
sized
minerals
chlorite
gibbsite,
goethite,
hematite
goethite,
hematite
gibbsite
30% mixed
layer
quartz
goethite
goethite
* pH of a 1:1 soikwater suspension.
t Catkin exchange capacity.
i Surface area by N2 adsorption using BET method.
§ No data available.
fl Surface area by ethylene glycol (from Hassett et al., 1981).
Table A3 Major element composition (in oxide form) of clay and soils used in developing this TRD
Adsorbent
Si02 Ti02 AI203 Fe203 CaO MgO Na.,,0 K2O P2O5
Catlin silt loam
Cecil clay
Cecil clay loam
EPA-14*
Illite clayf
Kaolinite clayf
Sangamon Paleosol
Tifton loamy sand
Vandalia
altered
unaltered
ablation phase
72.5
44.8
66.2
ND
48.5
46.6
82.7
96.4
61.3
59.1
83.5
0.73
1.15
0.94
ND
0.67
2.45
0.43
0.27
0.33
0.33
0.35
10.8
30.0
20.4
ND
24.6
41.9
10.2
1.3
6.7
6.5
7.9
4.0
10.4
6.8
6.99
4.11
0.94
2.9
0.5
2.1
2.4
2.5
0.9
<0.1
<0.1
0.71
3.27
0.57
0.50
<0.1
9.4
9.7
0.6
0.71
0.19
0.19
ND
1.73
0.30
0.65
0.03
4.66
4.95
0.54
0.84
0.07
0.04
0.21
0.14
0.13
0.45
0.01
0.52
0.49
0.56
2.14
0.54
0.65
2.94
10.23
1.49
1.49
0.05
2.03
2.08
1.98
0.1
0.1
ND
ND
ND
<0.1
<0.1
<0.1
<0.1
<0.1
* Data from Hassett et al., 1981.
t Data from Griffin and Shimp, 1976.
96
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Table A4 Trace element concentrations in the clays and soils used in developing this TRD
As
B
Ba
Be
Br
Cd
Ce
Cr
Co
Cu
Cs
Eu
Ga
Hf
La
Li
Lu
Mn
Ni
Pb
Rb
Sb
Sc
Se
Sm
Sr
Ta
Tb
Th
U
W
Yb
Zn
Catlin
S.I.
10
250
721
3
8
<1
62
73
14
20
4
1
10
12
36
29
0.6
834
<8
20
82
1
9
<2
6
90
1
1
8
5
2
3
88
Cecil
c.
40
30
117
1
19
<1
123
206
6
45
10
1
37
8
63
29
0.4
—
70
44
59
1
25
3
8
<5
2
1
24
7
5
3
40
Cecil
C.I.
4
25
166
2
3
<1
81
73
3
17
5
1
26
18
47
18
0.6
93
<8
14
74
0.4
13
<2
8
<5
2
1
16
7
2
3
37
EPA- Illite Kaol-
14* c. inite c.
10 — t —
— 44 46
450 — —
— — —
_ _ _
— 19 <3
87 — —
— — —
11 — —
— — —
8 — —
•j _ __
23 — —
8 — —
46 — —
— — —
— — —
216 <390 29
— — —
— 94 46
200 — —
6 — —
16 — —
2 — —
— — —
<80 — —
1 ^_
1 — —
•]g
— — —
— — —
3 — —
— 38 20
Sang-
amon
6
230
500
2
<1
<1.3
38
52
15
11
3
1
12
—
28
27
0.5
970
<9
19
79
0.6
8
<1
5
55
—
—
5
<3
—
2
71
Tifton .
S.I.
1
170
44
<0.5
2
<1.3
50
20
1
<4
2
0.4
2
26
19
4
0.4
90
<9
<10
18
0.3
4
<2
3
<5
1
1
6
1
<1
2
<2
altered
6
172
359
2
<7
<1.3
25
39
8
15
3
1
9
—
19
23
0.3
388
<9
F14
68
0.4
6
<2
3
75
—
—
4
<3
—
2
42
Vandalia till
unaltered
7
150
347
2
3
<1.3
24
41
9
19
3
1
7
—
19
25
0.3
<400
<9
13
68
0.4
6
<1
3
75
—
—
4
<2
—
2
73
ablation
5
250
460
2
<2
<1.3
29
43
8
12
3
1
8
—
24
22
0.4
352
<9
<9
88
0.3
7
<1
4
62
—
—
4
<2
—
2
44
* Data from Hassett et al., 1981.
t No data available.
97
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APPENDIX B
COMPOSITION OF THE METALLIC WASTE EXTRACT
USED IN THIS STUDY AND ASSOCIATED ADSORPTION ISOTHERMS
The basic batch-adsorption procedure for ionic solutes was tested and refined through the use of a metal-
lic waste sample collected November 1,1984, from the Sandoval Zinc Company near Sandoval, Illinois.
Grab samples were taken from a dry slurry lagoon that was used to store metallic scrubber sludges (Gibb
and Cartwright, 1982). Samples were taken from the surface and at a depth of about 1 m, then compos-
ited and air-dried. The relatively fine-grained material was then mixed and poured through a 2-mm sieve.
The laboratory work began by making 20 L of an extract of the metal-rich waste using the ASTM-A water-
shake extraction procedure (ASTM, 1979). The aqueous extract contained about 0.05% Zn (table B1) and
lesser quantities of Ba, Ca, K, and Pb. The extract was slightly acidic (pH 6.27) and was used as the
stock solution for all of the adsorption experiments. The Sangamon Paleosol sample and the Cecil clay
were selected for study because these two soils represented widely different physicochemical materials.
Adsorption isotherms (figs. B1 to B3) were generated using the procedures described in this volume.
Table B1 Chemical constituent concentrations (mg/L) obtained by ASTM-A
water shake extraction performed on the Sandoval zinc slurry
pH 6.27
EC(dS/m) 0.17
Al <0.05
As <0.08
B <0.08
Ba 2.25
Be <0.01
Ca 17.7
Cd 0.45
Co
Cr
Cu
Fe
K
Mg
Mn
Mo
Na
<0.45
<0.08
<0.01
<0.05
6.57
0.89
0.64
<0.02
<0.75
Ni
P
Pb
Sb
Se
Si
Sn
V
Zn
0.12
<0.05
15.0
<0.05
<0.04
<0.20
<0.03
<0.08
550
20-
10-
./
0.5
1
1.0
r
1.5
I
2.0
Equilibrium barium concentration (mg/L)
Figure B1 Barium adsorption isotherm at 21 °C with the Sangamon Paleosol from
the metallic waste extract. The average pH of the soil-solute suspensions was 5.6.
99
-------
1.2-
1.0-
„ 0.8-
8 0.6-
0.4-
0.2-
Sangatnon Paleosol
Cecil clay
~I 1 1 1 T
12345
Equilibrium lead concentration (mg/L)
Figure B2 Lead adsorption isotherms at 24°C of two soils using the metallic waste
extract. The average pH of the Sangamon Paleosol suspensions was 5.6, and pH
4.3 for the Cecil clay.
0 100 200 300 400 500
Equilibrium zinc concentration (mg/L)
Figure B3 Zinc adsorption isotherms at 24°C of two soils using the metallic waste
extract. The average pH of the Sangamon Paieosol suspensions was 5.9, and pH
4.3 for the Cecil Clay.
100
•U.S. Government Printing Office: 1992— 648-003/41820
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