EPA/530-SW-87-006
BATCH-TYPE ADSORPTION PROCEDURES FOR ESTIMATING
SOIL ATTENUATION OF CHEMICALS
Draft Technical Resource Document
for Public Comment
OFFICE OF SOLID WASTE AND EMERGENCY RESPONSE
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
Hazardous Waste Engineering Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
U.S. EnvironmentaJ Prtfeeftaa
Region 5, Library
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DISCLAIMER
This report was prepared by W. R. Roy, I. G. Krapac, S. F. J. Chou, and
R. A. Griffin of the Illinois State Geological Survey, Champaign, Illinois,
under Cooperative Agreement CR810245. The EPA Project Officer was M. H.
Roulier of the Hazardous Waste Engineering Research Laboratory, Cincinnati,
Ohio.
This is a draft report that is being released by EPA for public comment
on the accuracy and usefulness of the information in it. The report has
received extensive technical review, but the Agency's peer and administrative
review process has not yet been completed. Therefore, it does not necessarily
reflect the views or policies of the Agency. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use;
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FOREWARD
The Environmental Protection Agency was created because of increasing
public and governmental concern about the dangers of pollution to the health
and welfare of the American people. Noxious air, foul water, and spoiled
land are tragic testimony to the deterioration of our natural environment.
The complexity of the environment and the interplay of its components require
a concentrated and integrated attack on the problem.
The Office of Solid Waste is responsible for issuing regulations and
guidelines on the proper treatment, storage, and disposal of hazardous wastes,
in order to protect human health and the environment from the potential 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. Publications
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 assessing the capacity of soils and soil components of
liners for waste management facilities to attenuate 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.
Marcia Williams
Director, Office of Solid Waste
U.S. Environmental Protection Agency
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PREFACE
Subtitle C of the Resource Conservation and Recovery Act (RCRA) requires
the U.S. Environmental Protection Agency (EPA) to establish a Federal
hazardous waste management program. This program must ensure that hazard-
ous wastes are handled safely from generation until final disposition. EPA
issued a series 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 idnetified or listed as hazardous under 40 CFR Part
261. Part 264 standards are implemented through permits issued by author-
ized States or EPA according to 40 CFR Part 122 and Part 124 regulations.
Land treatment, storage, and diposal (LTSD) regulations in 40 CFR Part 264
issued on July 26, 1982, establish performance standards for hazardous
waste landfills, surface impoundments, land treatment units, and waste
piles.
EPA is developing three types of documents for preparers and reviewers
of permit applications for hazardous waste LTSD facilities. These types
include RCRA Technical Guidance Documents, Permit Guidance Manuals, and
Technical Resource Documents (TRD's).
The RCRA Tehnical Guidance Documents present design and operating
specifications or design evaluation techniques that generally comply with
or demonstrate compliance with the Design and Operating Requirements 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 appli-
cants and permit writers in addressing information requirements. 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.
The Technical Resource Documents present state-of-the-art summaries of
technologies and evaluation techniques determined by the Agency to consitiute
good engineering designs, practices, and procedures. They support the RCRA
Technical Guidance Documents and Permit Guidance Manuals in certain areas
(i.e., liners, leachate management, closure covers, and water balance) by
describing current technologies and methods for designing hazardous waste
facilities or for evaluating the performance of a facility design, although
emphasis is given to hazardous waste facilities, the information presented
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in these TRD's may be used for designing and operating nonhazardous waste
LTSD facilities as well. Whereas the RCRA Technical Guidance Documents and
Permit Guidance Manuals are directly related to the regulations, the informa-
tion in these TRD's covers a broader perspective and should not be used to
interpret the requirements of the regulations.
This document is a first edition draft being made available for public
review and comment. It has undergone review by recognized experts in the
technical areas covered, but Agency peer review processing has not yet been
completed. Public comment is desired on the accuracy and usefulness of the
information presented in this document. Comments received will be evaluated,
and suggestions for improvement will be incorporated, wherever feasible,
before publication of the second edition.
One original and two copies of all comments on this document should be
addressed to: RCRA Docket Clerk (Room S-212A), Office of Solid Waste (WH-562),
U.S. Environmental Protection Agency, 401 "M" Street, S.W., Washington, D.C.
20460. Comments should list the Docket Number (F-87-SACA-FFFFF) and identify
the document by title and number; e.g. "Batch-Type Adsorption Procedures for
Estimating Soil Attenuation of Chemicals" (EPA/530-SW-87-006).
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ABSTRACT
This document contains laboratory procedures and guidelines for
conducting adsorption experiments using batch equilibrium techniques to study
soil attenuation of chemicals dissolved in solution (solutes). The procedures
were designed for routine use, and may be used to generate data for the
construction of equilibrium adsorption isotherms or curves. Procedures for
inorganic and organic solutes, and volatile organic solutes are given.
The scientific basis and rationale for each procedural step is discussed
in detail, and was based on both the scientific literature and by procedural
development and testing by the authors and other cooperating laboratories,
using several different types of soil materials and solutes. The application
of major procedural steps and concepts is illustrated by examples, including
the application of batch adsorption data in calculations of solute movement
through compacted landfill liners, particularly for estimating the thickness
of liner required for pollutant retention.
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CONTENTS
page
Disclaimer ii
Foreward i i i
Preface iv
Abstract • vi
Figures viii
Tables xli
Acknowl edgment s x"iv
INTRODUCTION , 1
SECTION
1. Adsorption forces and mechanisms 5
2. Effects of adsorbent preparation 8
3. Effects of temperature . 17
4. Stability of nonionic solutes in solution 23
5. Effects of solution pH 28
6. Effects of ionic strength . 36
7. Effects of phase separation 42
8. Effects of the method of mixing 46
9; Selection of a soilrsolution ratio for ionic solutes. b2
10. Selection of a soil:solution ratio for nonionic
sol utes 60
11. Effects of the soil:solution ratio 68
12. Constant and variable soil:solution ratios 89
13. Determination of the equilibrium time 97
14. Construction of adsorption isotherms (curves) 108
15. Selection of adsorption equations 116
16. Application of adsorption data 119
17. Laboratory procedures for collecting adsorption
data 133
References 162
APPENDIX A. Summary and chemical composition of the adsorbent soils
and clays used in this study 173
B. Chemical composition of the metallic waste extract used
in this study and associated adsorption isotherms.... 179
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LIST OF FIGURES
page
Figure 1. Effect of air-drying on the concentration of
exchangeable manganese '. 10
Figure 2. Relationship between the concentration of
exchangeable potassium in the Harpster clay
loam and moisture content 10
Figure 3. Effect of oven-drying at 105°C on the concentration
of water soluble organic carbon in an Israeli
calcareous clay loam 12
Figure 4. Adsorption isotherm of acetophenone by fresh field
moist and air-dried samples of Crane Island alluvium 15
Figure 5. Arsenate adsorption isotherms by Catlin
at 15°C, 25°C and 35°C 19
Figure 6. Zinc, copper, and cadmium adsorption from a
DuPage County landfill leachate by kaolinite
at 25°C at various pH levels 29 .
Figure 7. Chromium (VI) adsorption by kaolinite at 25°C at
various pH levels 31
Figure 8. Langmuir-type maximum (mM/g) for the adsorption
of arsenic as As(V) and As(III) by amorphous
iron hydroxide 32
Figure 9. Effect of pH on the adsorption of triazines by a
Ca-montmorillonite sample 33
Figure 10. Effect of pH on the adsorption of different ionizable
organic solutes by an illite sample 33
Figure 11. The adsorption behavior 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 : 35
Figure 12. Ratio of concentration to activity versus ionic
strength for some common ions..... 37
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 43
Figure 14. Distribution of arsenic concentrations in solutions
that were either centrifuged or filtered 44
Figure 15. The National Bureau of Standards Rotary Extractor 48
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page
Figure 16. Distribution of arsenate concentrations after
24 hours of contact with different soil materials
as a function of soil solution ratio 55
Figure 17. Distribution of cadmium concentrations after
24 hours of contact with different soil materials
as a function of soil:solution ratio 56
Figure 18. Adsorption isotherm of o-xylene by Catlin at 23°C 63
Figure 19. Adsorption isotherms of dichloroethane and
tetrachloroethylene by Catlin at 23°C 64
Figure 20. Relationship between the linear Freundlich
constant (Kd) and soilrsolution ratio, as
a function of percent adsorption (lower range) 66
Figure 21. Relationship between the linear Freundlich
constant (Kd) and soilcsolution ratio, as
a function of percent adsorption (upper range) 67
Figure 22. Effect of soilrsolution ratio on cadmium
adsorption by a Sangamon paleosol sample 71
Figure 23. Cadmium adsorption by a Sangamon paleosol sample 73
Figure 24. Distribution of pH values of arsenate solutions
after 24 hours of contact with different soil
materials as a function of soil:solution ratio 75
Figure 25. Distribution of pH values of cadmium solutions
after 24 hours of contact with different soil
materials, as a function of soil:solution ratio 76
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 soil:solution ratio 78
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 79
Figure 28. Freundlich constant (Kf) for two PCB isomers vs.
sediment concentration with and without prewashing
to remove nonsettling particles 81
Figure 29. The Freundlich constant (K*) for the adsorption
of Aroclor 1242 by four different soils as a
function of soil :solution ratio 83
Figure 30. Aroclor 1242 adsorption isotherms by five soils
at various soil:solution ratios 84
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page
Figure 31. Adsorption of dieldrin, tetrachloroethylene,
and 1,2-dichloroethane by Catlin at various
soil:solution ratios 85
Figure 32. Adsorption of Aroclor 1242 by altered Vandalia
till and unaltered Vandalia till at various
soil:solution ratios 87
Figure 33. Distribution of exponents (1/n) and Freundlich
constants (K*) associated with arsenic, cadmium,
lead, and PCS (Aroclor 1242) adsorption isotherms 91
Figure 34. Cadmium adsorption isotherm with a Vandalia till
sample (unaltered) with the amount adsorbed
associated with each isotherm data shown 93
Figure 35. Distribution of cadmium adsorption data by a
Tifton sandy loam 94
Figure 36. Distribution of arsenate adsorption data by
different soil samples using different
soil:solution ratios 96
Figure 37. The adsorption behavior of cadmium by five soil
materials as a function of contact time 99
Figure 38. The adsorption behavior of arsenic by 11 different
soil materials as a function of contact time 10U
Figure 39. Determination of equilibration time of Ba, Pb,
and Zn from a laboratory extract of the Sandoval
Zinc slurry with the Sangamon Paleosol and
the Cecil clay sample 104
Figure 40. The adsorption behavior of o-xylene, dichloroethane,
and tetrachloroethylene by Catlin as a function
of contact time 106
Figure 41. The adsorption of arsenic by a kaolinite clay
sample as described by the traditional linear
Langmuir, double-reciprocal Langmuir, and
the Freundlich Equation 118
Figure 42. Lead adsorption by Cecil clay loam described by
a linear Freundlich equation through the origin 125
Figure 43. Predicted distance of lead migration in Cecil clay
loam based on three approaches 129
Figure 44. Flow diagram for the procedures for the
generation of batch adsorption data 134
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Figure B-l. Barium adsorption isotherm with the Sangamon
Paleosol from the metallic waste extract 181
Figure B-2. Lead adsorption isotherms of two soils using
the metallic waste extract 182
Figure B-3. Zinc adsorption isotherms of two soils using
the metallic waste extract 183
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LIST OF TABLES
Page
Table 1. Effect of drying on exchangeable Mn in four
Hawaiian soils 9
Table 2. pH of soil-water slurries (1:2 v/v) made with field-
moist samples compared to those that were
oven-dried at 110°C 13
Table 3. Effect of sample pretreatment on the Freundlich
partition coefficients (Kf) 14
Table 4. Effect of temperature on Freundlich adsorption
constants (Kf) for phenanthrene and a-naphthol 20
Table 5. Results of first ASTM sensitivity analysis 47
Table 6. Cadmium adsorption data from the 2nd ASTM
inter!aboratory sensitivity analysis using a NBS
rotary extractor as the mixing method 49
Table 7. Arsenic adsorption data from the 2nd ASTM
inter!aboratory sensitivity analysis using a NBS
rotary extractor as the mixing method 50
Table 8. Soil:solution ratio determination for the
Sangamon soil and Vandalia ablation till using
cadmium as the adsorbate 53
Table 9. Determination of soil:solution ratios for the
Sangamon Paleosol and the Cecil clay loam sample
using an extract of the Sandoval zinc slurry 58
Table 10. Determination of equilibration times for the
adsorption of arsenate by soil materials 102
Table 11. Determination of equilibration times for the adsorption
of Ba, Pb, and Zn from a Sandoval zinc slurry extract
by the Sangamon paleosol and Cecil clay 103
Table 12. Determination of equilibration time for the
adsorption of the PCB Aroclor 1242 by Catlin 107
Table 13. Data reduction for arsenic adsorption by a
kaolinite clay sample 110
Table 14. Lead adsorption data using a Pb(NO ) salt and
the Cecil clay !.t 123
Table 15. Summary of approaches to estimate minimum
liner thicknesses 132
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Table A-l. Summary of adsorbents 175
Table A-2. Summary of selected physicochemlcal characteristics
of clays and soils used in the development of TRD 176
Table A-3. Summary of major element composition (in oxide
form) of clay and soils used in the development
of TRD 177
Table A-4. Summary of trace element concentrations in the
clays and soils used in the development of the TRD 178
Table B-l. Chemical constituent concentrations obtained
by the ASTM-A (water shake extraction) performed
on the Sandoval zinc slurry 180
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ACKNOWLEDGMENTS
The authors wish to acknowledge the partial support of the U.S.
Environmental Protection Agency, Cincinnati, Ohio and Dr. Mike H. Roulier,
project officer of Cooperative Agreement CR810245-01. We also thank
Dr. Calvin C. Ainsworth, formerly with the Illinois State Geological Survey,
for his efforts during the first year of this project, Dr. Randy E. Hughes for
the mineralogical characterizations, Terence Beissel and Robert Arns for their
technical support, and members of the Analytical and Isotopic Chemistry
Section of the Illinois State Geological Survey (ISGS) for the adsorbent
characterizations. A number of laboratories contributed directly and
indirectly to this TRD through their participation in American Society for
Testing and Materials (ASTM) 034.02 on Waste Disposal round-robin testing of
batch adsorption procedures: Dr. Greg Boardman (VPI and State University,
Virginia), Dr. Chet Francis (Oak Ridge National Laboratory, Tennessee),
Dr. Marc Anderson (University of Wisconsin), Dr. William A. Sack (West
Virginia University), and Mr. Otis E. Michels (Daily and Associates Engineers,
Peoria, Illinois). Dr. John J. Hassett of the University of Illinois is
gratefully 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 Florida, and Dr.
Thomas C. Voice of Michigan State University are also acknowledged and
appreciated.
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INTRODUCTION
The capacity of geological materials to attenuate potential pollutants
has been studied by many researchers, especially during the last 30 years.
.One of the potential applications of information from such studies 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. This Technical Resource Document (TRD)
describes a number of laboratory batch procedures for assessing the capacity
of soils to adsorb (attenuate) chemicals 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 adsorption data in designing
soil liners for pollutant attenuation.
The batch adsorption or batch equilibration technique has often been used
in laboratory studies to assess the capacity of soils and soil components to
attenuate chemical constituents in solution. However, the batch procedures
that have been used vary considerably in terms of experimental conditions and
research objectives and, in some cases, may yield different results even when
the same soils, solutes and concentrations are studied.
In principle, the batch adsorption technique is relatively simple,
accounting, in part, for its popularity. This technique consists of mixing an
aqueous solution containing solutes of known composition and concentrations
with a given mass of adsorbent for a 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.
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While the approach is relatively simple, there are a number of experimental
parameters that may affect the adsorption of a given constituent. For
inorganic solutes, these parameters include contact time, temperature, method
of mixing, soilisolution 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;
Ainsworth et al., 1984; and Roy et al., 1984, 1985). For organic solutes,
similar parameters may also affect adsorption (Bailey and White, 1970; Grover
and Hance, 1970; Dao and Lavy, 1978; Koskinen and Cheng, 1983; and Horzempa
and DiToro, 1983). In addition, dissolved organic carbon, adsorbate
volatility, photodegradation, biodegradation, and compound stability can also
affect adsorption data associated with organic solutes (Harris and Warren,
1964; Scott et al., 1981; and Chou and Griffin, 1983).
Equilibration time, a basic experimental parameter in the batch
techniques cited above, has varied from 30 minutes to 2 weeks. Soilrsolution
ratios used in batch procedures have varied from very dilute systems
(1:100,000) to 1:1 pastes. These particular experimental conditions were
probably appropriate for the specific system under study, and appropriate for
the intended use of the data. However, these diverse differences in experi-
mental conditions may make comparisons of data between studies difficult.
Moreover, there are currently no standardized batch adsorption procedures
designed for routine use with the exception of the procedural guidelines
outlined in EPA (1982) and the standard methods currently under development of
the American Society for Testing and Materials (ASTM) D-18, D-34, and E-47.
Results from recent D-34.02 round-robin testing of batch sorption procedures
indicated coefficients of variation of greater than 140% during initial
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testing, which were reduced to less than 10% by application of standard
procedures and equipment between laboratories (Griffin et al., 1985). The
experience gained during those interlaboratory testing programs and the
interactions with the scientists and laboratories affiliated with ASTM have
been incorporated into this document. The proposed ASTM 24-hour batch
adsorption procedure has been reviewed and voted upon by the committee
members. Comments have been received from 96 individuals who are active in
research, government, industry, and waste management. •
Furthermore, there are very few well documented and comprehensive sources
that can be consulted for conducting batch adsorption experiments. The
purpose of this Technical Resource Document (TRO) is to describe a number of
batch adsorption procedures for both inorganic and organic solutes, to
document their scientific basis and rationale, and to recommend procedural
steps that are best supported by current information. This TRD also contains
numerous examples to demonstrate the application of each procedural step.
Section 16 demonstrates how adsorption data can be used in designing or
evaluating soil liners for pollutant retention and discusses some cases where
this has been done. The last section of this report contains the actual
procedures written without narrative discussion. The reader should study the
preceding sections before attempting these procedures.
Most of the procedural steps recommended here have been tested in the
>
authors laboratory using a variety of soils, solutions containing several
solutes, and aqueous extracts of actual wastes. Characteristics of the soils,
clays, and waste are described in appendices to this document.
The information in Section 1 on adsorption forces and mechanisms was
taken from an open file report entitled "Interaction of Organic Solvents with
Saturated Soil Water Systems" that was written by Dr. R. A. Griffin and
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Or. W. R. Roy (April 1985) for the Environmental Institute for Waste
Management Studies, University of Alabama. Information and references for
other sections have also been drawn from that report.
The collection of accurate and meaningful adsorption data is not a simple
task. Even though the procedures described here were intended to be fairly
easy to use and precise, it is inevitable that some "scatter" in data will
occur, and the origins of the deviations will elude any clear-cut
explanations. The investigator is encouraged to persevere and repeat the •
procedures as the situation demands. The perseverance is well warranted as
the acquisition of high quality adsorption data is essential in predicting,
and thus protecting, the quality of ground and surface waters that must co-
exist with the by-products of our civilization.
W. R. Roy, I. G. Krapac
S. F. J. Chou, and R. A. Griffin
Illinois State Geological Survey
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SECTION 1: ADSORPTION FORCES AND MECHANISMS
Before undertaking adsorption studies, it may be informative to briefly
review the physical chemical forces and mechanisms that are 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 physical chemical forces may be broken
down into eight categories (after Reinbold et al., 1979; Griffin and Roy, 1985;
and a paper containing other useful references by Voice and Weber, 1983):
1. London-van der Waals. There are attractive forces that arise from momen-
tary dipoles about atoms or molecules caused by small perturbations of
electronic motions. These dipoles induce small dipoles in neighboring atoms
of opposite sign. Although the momentary dipoles and induced dipoles are
constantly changing position and sign, the net result is a weak attraction (4
to 8 KJoule/mole for small molecules and atoms). These forces are important
in adsorption of organics and are generally attributed t.o explaining the non-
ideal behavior in gases. They also have been partially treated by quantum
mechanical perturbation theory, using polarizabilities, ionization potentials,
and the magnetic susceptibilities of the interacting atoms to explain various
phenomena such as adsorption.
2. Coulombic-electrostatic-chemical. An electrostatic force resulting from a
charged surface due to isomorphous substitution in the mineral lattice (permanent
charge) or protonation of surface oxygen and OH groups (pH-dependent charge)
and an oppositely charged species to maintain the electroneutrality of the
surface is important in cation exchange reactions in soils. In layer sili-
cates, 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
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or molecules about the colloid whose density is greater at the surface, then
exponentially decreases to the level of the bulk solution. This type of reaction
is important in adsorption of both inorganic ions and ionized organic molecules.
3. Hydrogen bonding. This type of interaction is where a hydrogen atom is
bonded to two or more other atoms in that the "bond" is generally conceived as
an induced dipole phenomena. There is no universal agreement on the best
description of the hydrogen bond (Huheey, 1978), but it may be considered as
the asymmetric electronic distribution of the Is electron of the hydrogen atom
by very electronegative atoms (such as F, 0, S, Cl, etc.). There are reasons
to believe that more is involved in hydrogen bonding than simply an exaggerat-
ed dipole-dipole or an ion-dipole interaction due to the inability of these
concepts to account for molecular geometry in some cases (see Huheey, 1978;
Cotton and Wilkinson, 1980). H-bonds may be in reality delocalized covalent
bonds, i.e., resonance bonds or multiple-center bonds (Huheey, 1978). The
energy of this attraction ranges from 8 to 42 KJoule/mole.
4. Ligand exchange-anion penetration-coordination. Many atoms or molecules
form coordinated complexes with ligands that range in complexity from simple
linear molecules to extensive chelate complexes. The coordinated complexes
may carry a net charge which may be localized on some part of the complex.
These complexes may be in turn bonded to surfaces by H-bonding or by poly-
valent cation bridges linking the complex to a charged surface. -The possible
geometrical arrangements of coordinated complexes bonded to mineral faces is
diverse. The bonded coordinated complexes may be displaced by other coordin-
ated complexes that better satisfy electroneutrality requirements (i.e., are
stronger complexing agents) while being constrained by steric limitations.
The energy of ligand exchange reactions with inorganic ions ranges from 8 to
60 KJoule/mole.
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5. Chemisorption. In this adsorption process an actual chemical bond,
usually covalent, is formed between the molecule and the surface atoms. A
molecule undergoing Chemisorption may lose its identity as the atoms are re-
arranged, forming new compounds at the demand of the unsatisifed valences of
the surface atoms. The enthalpy of Chemisorption (~AH >29 KJoule/mole) is
much greater than physical adsorption. The basis of much catalytic activity
at surfaces is that Chemisorption may organize molecules into forms that
readily undergo reactions. It is often difficult to distinguish between
Chemisorption and physical adsorption because a chemisorbed layer may have a
physically adsorbed layer upon it. Moreover, some ligand exchange reactions
are Chemisorption processes.
6. Dipole-dipole or .orientation energy. This results from the attraction of
a permanent dipole for another permanent dipole. The resulting energy of
attraction is less than 8 KJoule/mole.
7. Induction or dipole-induced dipole. This results from the attraction of
an induced dipole brought about by either a permanent dipole or a charged site
or species. The energy of attraction is less than 8 KJoule/mole, but this
force often adds to coulombic interactions.
8. The hydrophobic effect. The exact nature of this adsorption mechanism is
uncertain. It is the view of some investigators 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 structured water molecules about 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
partitioning. Nonpolar organic solutes tend to migrate from the aqueous phase
to hydrophobic surfaces- on the adsorbent (Dzombak and Luthy, 1984, Chiou et
al., 1979, 1983; see also Griffin and Roy, 1985).
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SECTION 2: EFFECTS OF ADSORBENT PREPARATION
The process of preparing samples taken in the field for laboratory inves-
tigations can have a direct influence on analytical results. Adsorbent
samples (i.e., soils, clays, etc.) are usually dried so that they can be
homogenized 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 concen-
tration of exchangeable manganese in twenty-three Hawaiian soils tended to
increase as the samples were dried. A portion of their results is given in
Table 1. The changes that occurred between field moist and air drying were,
however, minimal compared to 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 moisture 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 micro-
organism respiration when they are re-wetted, and Stevenson (1956) concluded
that the degree of metabolic activity varies directly with the concentrations
of free ami no 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 re-moistened, a portion of the
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Table 1. Effect of drying on exchangeable Mn in four Hawaiian soils (from
Fujimoto and Sherman, 1945).
Location
of Soil
Kemoo
Koko Head
Kahuku
Waimanalo
pH(l:l)l
4.2
7.1
7.6
8.6
Field
Moisture
3.4
0.0
0.0
0.5
Air-dried
Mn concentr
4.5
4.3
0.4
0.4
Oven-dried2
621.2
29.4
11.7
1.2
Autoclaved3
374.8
NO*
367.9
NO
!pH of a 1:1 soi1:water suspension
20ven-dried for 24 hours at 105°C
SAutoclaved for 3 hours at 15 pounds pressure
''Not determined
-------
-10-
c
X
uu
70-
high-Mn soil
6 8 10 12 14
Air-drying time (weeks)
16
18
Figure 1. Effect of air-drying on the concentration of exchangeable manganese
(adapted from Fujimoto and Sherman, 1945).
280-
S 200 H
120-
x
uu
40-
10 20
moisture (%)
Figure 2. Relationship between the concentration of exchangeable potassium in
the Harpster clay loam and moisture content (adapted from Luebs et
al., 1956).
-------
L
-11-
organic matter dissolved and the magnitude of this decomposition depended
directly on the amount of organic matter present in the soil. Birch (1959)
later concluded that this decomposition following re-wetting was primarily due
to microbial decomposition of water soluble organic matter.
An alternative hypothesis was proposed by Raveh and Avinimelech (1978).
They envisioned that when organic macromolecules are in natural pedological
settings, they are aggregated by hydrogen bonds. When soils are dried, the
evaporation of water disrupts the H-bonds and the stability of the organic
matter decreases. 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 increase in the yellow or amber color of
extracts, attributable to the amount of organic matter made soluble by
drying. They also found that the amounts of Al, Fe, and Mn in NH OAc extracts
(pH 4.8) of a soil subjected to 40°C for 12 hours were greater than those
extracted from moist samples at field moisture 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 pH
units (Table 2). Raven and Avinimelech (1978) suggested that this increase in
acidity was due to the exposure of fresh organic surfaces that contained acidic
groups that were sterically hindered before drying. The increase in surface
acidity was also considered by Mortland and Raman (1968) who hypothesized a
different mechanism; as the samples are dried, adsorbed cations more strongly
polarize the residual water molecules, making them more acidic than free water.
-------
-12-
I
3
U
°c
(Q
6
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 Raven and Avnimelech, 1978).
-------
-13-
Table 2. pH of soil-water slurries (1:2 v/v) made with field-moist (FM)
samples compared to those that were oven-dried at 110°C (00) (from
van Lierop and MacKenzie, 1977).
Soil
Demers
J.I.v.
J.I.-l
J.I. -2
HDE
SB
Bigras v.
Leh. v.
Lamb.
Mac.
FM
4.0
4.2
5.5
6.2
4.5
4.1
4.2
4.2
6.7
6.3
OD
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
Other studies have demonstrated that drying samples lowers the ability of
a soil 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 studies.
Direct effects of adsorbent preparation have also been documented; Ashton
and Sheets (1959) found that the herbicide ethyl N,N-di-n-propylthiolcarbamate
(EPIC) was adsorbed as a vapor to a greater extent by air-dried soils than
soils that were moist. The adsorption of EPIC may have been suppressed at
higher moisture contents due to the competition of the EPIC 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 also suggested that competition between the atrazine and water
could account for this relationship.
Oven-drying may increase the hydrophobicity of soils which, in turn,
would enhance the adsorbent's affinity for hydrophobic solutes. It has been
-------
-14-
established that forest fires can increase the hydrophobicity of soil
materials near the surface. Heat-induced hydrophobicity studies by Debano et
al. (1976) suggested that temperatures as low as 98° to 118°C for an exposure
time of as little as 5 minutes can increase the hydrophobicity of a sample as
measured by water drop penetration time. It is not certain whether this heat-
induced hydrophobicity will influence adsorption results obtained by batch
techniques. Hassett et al. (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 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
samples tended to be similar to those from fresh field moist samples.
In contrast, Bartlett and James (1980) found that a soil sample which had
been oven-dried at 40°C adsorbed more phosphate during a six-hour equilibration
than samples which were kept moist. Harada and Wada (1974) reported that air-
drying their soil samples resulted in slight but significant increases in
Table 3. Effect of sample pretreatment on the Freundlich partition
coefficients (Kf) (.Hassett et al., 1980).
-------
C
n>
o> o. ^>
—•-JO.
• -'•in
•• n> o
Q- -J
I—' "D
VD (/> ft-
00 0> ->•
O 3 O
O
o n-
-«> 3-
0)
O)
3 O
n> -h
I—I ftl
o
—' fD
Oi rf
3 O
Q.T3
3-
Cu (D
—' 3
—• O
C 3
< 0>
o. n>
QJ m
X) 3-
rt-
(D -t.
O. ->•
(D
-h — •
-j cx
o
3 3
O
(u in
m
£1
c
o
X3
"
3 pH
2 f
r*t
5"
3
o
I
Amount adsorbed
o
I
oo
8
• D
> -n
i 3
t
Ul
I
(D Cu
r» 3
<-•• O.
Ct -J.
-j
i
-------
-16-
both the cation exchange capacity (CEC) and anion exchange capacity (AEC).
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
tetrahedra may have changed its stearic configuration to a form more conducive
to bonding.
In summary, drying adsorbent samples in order 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 either air-
drying or oven-drying may be viable methods of sample preparation if the
potential changes in adsorbent properties are understood and confronted.
However, understanding and confronting these changes may be research projects
in themselves.
• As a guideline for conducting batch adsorption studies, it would appear
that the 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 in contact with the atmosphere minimizes any changes that
may occur due to drying and is the most practical approach at this time. The
American Society for Testing and Materials defined air-drying as a process of
partial drying (of the sample) 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). It appears advisable to keep air-drying
to the minimum necessary to allow preparation of the sample, and to provide a
stable condition for measurements of the sample such as weighing. Air-drying
anaerobic soils and sediments will require special handling in order to prevent
the relatively reduced materials from oxidizing if exposed to the atmosphere.
Anaerobic materials can be "air-dried" in a glove box or in a glove bag that
is supported by a continuous supply of dry oxygen-free nitrogen or argon gas.
-------
-17-
SECTION 3: EFFECTS OF TEMPERATURE
Adsorption at the solid-liquid interface tends to occur when the attrac-
•
tive 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 thermodynami-
cally favorable change in the enthalpy (AH) (Hassett et al., 1981) or
sometimes by a favorable change in the entropy (AS) of the system where the
-TAS 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 solutes is thought to be primarily the result of a
thermodynamically favorable change in entropy (AS) involving little energy
transformation as heat. Thus it is valid to anticipate that the adsorption
behavior of ionic or polar solutes will show some temperature dependency,
whereas the adsorption of nonpolar solutes may not be greatly influenced by
the temperature of the system. The direction and magnitude of temperature
dependency will depend on the specific solute-soil system.
An early paper by Jurinak and Bauer (1956) reported that the 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°C to 35°C
and found that the zinc adsorbed endothermically; increased adsorption was
associated with higher temperatures.
Kinniburgh and Jackson (1981) reviewed the literature on cation adsorp-
tion by soils and concluded that the effects of temperature were usually
small, but in some cases they significantly influenced adsorption data.
The adsorption of phosphate by soils and soil materials is often
endothermic (Low and Black, 1950; Gardner and Jones, 1973; Griffin and
-------
-18-
Jurinak, 1973a; Singh and Jones, 1977; Taylor and Ellis, 1978). The adsorp-
tion 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 as total As 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.
In contrast to ionic species, Hassett et al. (1983) found that the
adsorption of the nonpolar solutes phenanthrene and a-naphthol by soils was
largely unaffected by temperature variations from 15°C to 35°C (Table 4). The
adsorption of 1, 2-dichlorobenzene by a soil sample studied by Chiou et al.
(1979) was insensitive to temperature differences between 3.5°C and 20°C, but
the adsorption of 1,1,1-trichloromethane was reduced at the lower temperature.
Weber et al. (1983) found that the adsorption of Aroclor 1254 by a
Saginaw River sediment was temperature-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 approxi-
mated by a Clausius-Clapeyron-type equation, integrated over a narrow
temperature range, viz.,
R In^/C^/U/T^l/y - AH' [1]
where C and C are the equilibrium concentration
of a solute at two different temperatures, T and T ,
and AH' is the apparent heat of adsorption.
Apparent heats of adsorption values may be used as estimations of the
amount of heat energy isothermally released or absorbed during the course of
-------
-19-
400
I
20 40 60
Equilibrium arsenic concentration (mg/L).
80
Figure 5. Arsenate adsorption isotherms by Catlin at 15°C, 25°C and 35°C, and
at pH 6.6.
-------
-20-
Table 4. Effect of temperature on Freundlich adsorption constants (Kf) for
phenanthrene and a-naphthol (Hassett et al., 1983).
Freundlich constant (Kf)
Solute
Soil
15°C
25°C
35°C
phenanthrene
a-Naphthol
•
5
15
5
15
328
117
5.4
19
304
151
5.5
25
340
126
7.7
31
-------
-21-
adsorption, although it is probably more correct to view such values as heats
of the overall reaction. Eq. [1] can be rearranged as
^ « exp ((1/T - l/T ) AH-/R) [2]
Eq. [2] can be used to estimate the effects of temperature if AH- of the
specific adsorbent-solute system is known. If the magnitude of AH- is small
such as with the adsorption of some hydrophobic organic solutes, then the
ratio of C to C will be close to 1. In other words, the solute
2 i
concentration at temperature 1 will be nearly the same concentration as at
temperature 2, 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 associated with relatively large AH-
values. Consequently phosphate adsorption may be sensitive to ambient
temperature fluctuations. Moreover, if the temperature fluctuations are large
(the difference between T and T in eq. [2]), there is a greater potential
for the equilibrium solute concentrations to be affected by temperature
changes or fluctuations. To avoid this experimental artifact, adsorption
experiments are usually conducted with temperature-controlled water baths or
constant temperature rooms. If such facilities are not available or are
impractical, it is suggested that the laboratory work be conducted in rooms
where the ambient temperature fluctuates by no more than 6°C (i.e., 22 ±
3°C). This 6-degree range was based on the assumption that a "typical" heat
of adsorption value for most solutes of environmental significance is
approximately < 20 kJoule/mole, based on the discussions in Section 1. This
suggested range should be acceptable for most situations, but in cases where
the adsorption of the solute results in a comparatively large heat of -
adsorption, more rigorous temperature control may have to be implemented.
-------
-22-
• In summary, it is recommended that these batch adsorption procedures
should be conducted under constant temperature conditions, if available, or in
rooms where the ambient temperature is fairly constant (e.g., 22 + 3°C). It
is also recommended that when the batch experiments are performed, the temper-
ature of the room should be recorded and treated as a potential variable that
may influence the data or as one that may be useful in the interpretation of
the results.
-------
-23-
SECTION 4: STABILITY OF NONIONIC ORGANIC SOLUTES IN SOLUTION
In conducting a batch adsorption procedure, it is important to 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 is contacted
with the adsorbent.
1. Photolysis - Photoreactive solutes which absorb light at wavelengths
greatef 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). There-
fore, it is recommended that precautions be taken to ensure that substances'
such as these are protected from light, not only sunlight but laboratory
lights as well. Appropriate measures include use of amber glass, wrapping
glassware in aluminum foil, or any other suitable technique that will
eliminate the possiblity of photolysis transformations via exposure to
light. A simple aqueous screening test is presented here to 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
concentration due to photolysis are measured during the test.
Photolysis Test:
In this screening test, place the initial stock solution into either a 30-
ml_ or 50-mL borosilicate glass hypo-vial and fill the vial to eliminate
any head space. Then 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 determine the concentration at time = 0 and in two
-------
-24-
freshly opened hypo-vials after 2, 4, and 6 hours of exposure,
respectively. 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 similarly exposed as the samples. Select an
analytical method which 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 impurities. If the results indicate the solute
is photoreactive, then all subsequent tests and adsorption studies must be
conducted under conditions which prevent exposure to light during the
reactions and analytical steps.
2. Hydrolysis - Hydrolysis is an important degradation path for certain
classes of nonionic solutes, and it is necessary to know whether the solute
under study is subject to hydrolysis during the period of the adsorption
study. Otherwise, the amount of solute adsorbed by soils or sediments could
be over-estimated if changes in solution concentration due to hydrolysis are
not taken into account. Details of the hydrolysis 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).
It is important that the temperature of a hydrolysis screening test
procedure be kept constant. The temperature used in the hydrolysis test
procedure should be the same temperature to be used in the adsorption
experiments. The pH is also important and it is recommended that the
hydrolysis screening test be carried out at the same pH range that will be
used in the adsorption studies. The prevention of photolysis is to be
implemented as previously discussed. In some cases, the hydrolysis of solutes
-------
-25-
may be enhanced by the presence of other substances such as iron which
catalyzed the rate of hexachlorocyclopentadiene hydrolysis'under conditions of
low pH (Chou and Griffin, 1983). Therefore, the composition of the test
solution must be considered.
Hydrolysis Screening Test:
Fill either a 30-mL or 50-mL Hypo-vial 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 which is most applicable to the analysis of the specific compound
under study and analyze duplicate samples of the concentration of the .
chemical substance at time = 0 (control), and in two, freshly opened Hypo-
vials after 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
special care should be given to the handling of flasks and to the analytical
steps employed.
3. Microbial Degradation
Microbial degradation can also decrease the solution concentration of the
solute thus leading to an overestimation of the amount adsorbed by the ad-
sorbent. Therefore, for easily .degraded (labile) compounds, a batch technique
will measure "apparent adsorption," which is in reality a combination of
adsorption and degradation (and hydrolysis as indicated by the results of test
2). The influence of microbial degradation on "apparent adsorption" of phenol
by soil was studied by Scott et al. (1982). They found that Freundlich Kf
values for the adsorption of phenol by nonsterile soil increased linearly with
time with a Palouse silt loam and increased exponentially with time with
-------
-26-
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 was initiated within 10
hours, and complete degradation occurred within 48 hours or less for initial
p-cresol concentrations of 5, 10, 20, and 50 u9/L. The adsorption of organic
compounds, such as phenol or other labile organics which are degraded within
the time required to attain adsorption equilibrium, cannot be evaluated
accurately without accounting for or eliminating microbial 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 non-
sterile soil. Prior to 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 the Section on Effects of Adsorbent Preparation to help
evaluate the possible changes in adsorbent characteristics caused by
autoclaving.)
Bulk solutions of the solute are prepared in distilled water and passed
through a sterilized 0.22-ym membrane to sterilize the solutions, then known
amounts of the solutions are transferred to the sterilized and non-sterilized
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 super-
natant solution are taken with a syringe through a hole and septum in the caps
-------
-27-
of the bottles. Select an analytical method which is most applicable to the
analysis of the specific solute under study. The major purpose of the
suggested test procedure is only for screening for biodegradability.
If the test indicates the solute to 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.
-------
-28-
SECTION 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 increases with
increasing pH (Kinniburgh and Jackson, 1981). For example, Griffin and Shimp
(1976) reported 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 experiments, 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 precipita-
tion of PbCO . The reduced adsorption at the lower pH values was attributed
to the increase in competition for adsorption sites by H+ and by A13+
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. Relatively small differences in pH (~ one-half of a pH
unit) can result in major differences in the amount of solute adsorbed.
The pH of the soil solution has also been shown to have a direct effect
on the adsorption of anionic solutes. In contrast to cationic solutes, anion
adsorption is generally enhanced in acidic environments, however, 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 adsorption 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 passes through a
maximum value between pH 4 and 5. Griffin et al. (1977a) found that the
-------
-29-
8 -
6 -
6.5
r
50 100 150 200 250
Equilibrium concentration (mg/L)
300
350
Figure 6. Zinc, copper, and cadmium adsorption from a DuPage County landfill
leachate by kaolinite at 25°C at various pH levels (Frost and
Griffin, 1977)
-------
-30-
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.
The adsorption of arsenic as arsenate (As(V)) is also pH-dependent with
lower pHs resulting in greater adsorption (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
be the result of the increased positive charge due to the increased protona-
tion 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(III)) 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 increased as the pH
decreased (Fig. 9). At low pHs, the triazine solutes may have been increas-
ingly protonated, 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 (the
herbicides MCPA, 2,4-D, DNBP, and 2,4,5-T) by clays. The adsorption of ONBP
(Fig. 10), for example, sharply decreased as the pH of the system increased
from approximately pH 4.7 to pH 6. In alkaline solutions (pH > 7), ONBP
adsorption was reduced due to negative adsorption which occurred, i.e., the
DNBP was repelled by the clay. In this pH range, DNBP occurred largely as
neutral molecules since the pK of the organic solute was 4.35. The adsorption
-------
-31-
-. 100
O)
•g
JD
1
03
o 50
c
3
O
PH
Figure 7. Chromium (VI) adsorption by kaolinite at 25°C at various pH
levels. The chromium concentrations shown are the initial
concentrations added (modified from Griffin et al., 1977a).
-------
01
£
o
1.6-
1.4-
1.2-
1.0-
E
3
x 0.8-
.!= 0.6-
I
0.4-
0.2-
0.0-
As(lll)
1 T
6 7
pH
~T
9
ISGS 1985
T
10
i
co
IM
I
Figure 8. Langmuir-type maximum (mM/g) for the adsorption of arsenic as As(V) and As(III) by
amorphous iron hydroxide (Pierce and Moore, 1982).
-------
CO
c
n>
t—>
o
-p. O -t
x_x _. -«,
c n>
o rt o
Z fl> rt
CD in
-o o
—«< ~*
Ql TO
Q. 0» DC
a* 3
•o o
Ct -*• 3
fD —'
0. —' rt
_i. 3-
-h ct n>
T ft)
O Q>
3 t/> Q.
O> (/>
~n 3 o
T -O -J
_i. —i-o
uj n> <-»•
(/) ->.
o> '-^ o
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=> ro -t>
o.»
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oo i -••
o o -ti
— •«• -»i
r+ n>
<• ' — » -J
ro n>
l_i>_x 3
10 rt
o» ro
ro» -••
- — 4i O
• • 3
CTI -••
I IM
•H D)
• cr
Amount adsorbed (jUmole/gt
co
O O>
-D 3
n
3
O.
IO
c
fD
VO
fD
O
•*"""*• "^>
D>
Q.T3
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Q.
r+
-h 3"
-J fD
3 Q>
Q-
3C trt
O> O
3 -J
o -u
n> ri-
ft)
N
3
fD
O)
o
Cu
I
o
3
rt
O
10
O
% adsorbed
O)
o
00
o
o
o
I
co
CO
rt
CD
-------
-34-
of benzidine also followed a similar pattern. The ionization constants of
benzidine are 4.3 and 3.3 (pKb and pKb , respectively). Consequently,
Zierath et al. (1980) found that the amount of benzidine adsorbed by two soils
decreased when the solution pH was increased from a pH of 5 to pH 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 adsorption, although the cationic form should be adsorbed to a
much greater extent due to Coulombic interactions.
The adsorption behavior of neutral, nonpolar hydrophobic organic solutes
appears to be largely unaffected by the pH of the soil-water system. Hassett
et al. (1980) found no'correlation between the adsorption behavior of poly-
cyclic aromatic hydrocarbons (PAH) and the pHs of 14 spils 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.
• In summary, the potential influence of pH on the results generated by
batch adsorption procedures will depend on the system under study. It is
recommended that the equilibrium pH of the soil-solute mixtures be determined
prior to separating the solution from the soil or soil component suspension
and routinely given along with the adsorption data. In the case of anaerobic
adsorbent-solute systems, pH 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 pH data may render the adsorption
data impossible to interpret in a meaningful way.
-------
-35-
250-
Goethite
200-
01
_l
e
EPA-14
100-
Cecil clay
50-
ISGS 1985
10
I
12
pH
Figure 11. The adsorption behavior 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.
-------
-36-
SECTION 6: EFFECTS OF IONIC STRENGTH
The Ionic strength of the solution in batch adsorption procedures may
have several direct and indirect effects on the results. The extent of these
r
effects will depend on both the magnitude of the ionic strength and on the
concentration, composition, and charge of the ionic constituents constituting
the ionic strength of the adsorbent-liquid system.
Without regard to the specific composition of the solution, the ionic
strength may directly affect batch adsorption data in two ways: 1) changes in
solute activity, and 2) changes in the thickness (and therefore properties) of
the diffuse electrical double layers associated with colloidal particles. The
activity of most solutes tends to decrease as the ionic strength of the
solution increases due to the shielding effect arising from neighboring
ions. However, beyond a threshold ionic strength (often in very concentrated
solutions such as brines), the activity of some ionic constituents reverses
itself and steadily increases, finally yielding activities exceeding their
original concentration (Fig. 12). This phenomenon is relevant to batch
adsorption data since the use of actual solute concentrations rather than
activities of ions may not yield calculated results that agree with observed
results due to the departure of concentration from ideality in non-dilute
systems. Discussion of this topic may be found elsewhere (Atkins, 1982; Bonn
et al., 1979; Bolt and Bruggenwert, 1978; Garrels and Christ, 1965; and Stumm
and Morgan, 1981).
It is a basic tenet of Diffuse Double Layer Theory that the physical
thickness of the electrical double layer composed of adsorbed cations about a
colloidal particle is inversely proportional to the ionic strength of the bulk
solution. This phenomenon may not only affect exchange and adsorption reac-
tions at the solid-liquid interface, but may control the physicochemical
-------
-37-
0.0 -
Na1
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.
-------
-38-
properties of the material at the macroscopic level,, such as hydraulic
conductivity.
In attempts to minimize changes in ionic strength in the construction of
adsorption isotherms, some investigators 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 rationale for the choice has rarely been explained or
justified (Ryden and Syers, 1975).
The addition of a background electrolyte has been observed to have no
measureable effect in some soil-solute systems while 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 adsorption by gibbsite was independent of
ionic strength in the range of 0.002 M to 0.02 M when the ionic strength was
controlled by NaCl, KC1, and MgCl . However, 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 M to 1 M NaCl. 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 M to 1.0 M.
Common to many studies is the observation that polyvalent cation salts
promote phosphate adsorption relative to that from distilled water (Barrow,
1972; Fox and Searle, 1978; Heylar et al., 1976; El Mahi and Mustafa, 1980;
and White, 1980). Helyar et al. (1976) speculated that Ca2+ may act as a
potential determining ion while others (El Mahi and Mustafa, 1980) suspected
that the solubility of solid phosphate compounds was exceeded (see also
Anderson et al., 1981)'.
-------
• :. L .,
-39-
The relationship between ionic strength and the adsorption of organic
solutes has also been examined. Increasing 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 2,4-0 (2,4-dichlorophenoxyacetic acid) (Moreale
and Van Bladel, 1980). The increase in adsorption of the weakly acidic
herbicides cited here was attributed to a decrease in pH. A decrease in pH
would increase the proportion of the molecular species, which could then be
adsorbed. On the other hand, Choi and Aomine (1974) found that increasing the
ionic strength at constant pH decreased the adsorption of pentachlorophenol (a
weak acid: pKa = 4.5), and the amount of decrease was dependent on the anion
used in adjusting the ionic strength of the solution containing the penta-
chlorophenol. In batch adsorption studies, Abernathy and Davidson (1971)
found that the adsorption of fluometuron (l,l-dimethyl-3-(a,a,a-trifluoro-m-
tolyl)urea) was decreased and prometryn (2,4-bis(isopropylamino)-6-
(methylthio)-s-triazine was increased by increasing the CaCl concentration
from 0.01 to 0.5N.
In experiments designed to evaluate the effect of solution ionic strength
on 2,4,5,2',4',5'-hexachlorobiphenyl (HCBP) adsorption, Horzempa and OiToro
(1983) found that the Freundlich constant (Kf) appeared to be only slightly
influenced by increasing NaCl concentration from 10"1* M to 10"2 M. However,
in similar experiments CaCl2 significantly affected the Kf values over the
same concentration range.
The use of background electrolytes may also promote competitive inter-
actions between the ions derived from the background electrolyte and the
solute(s) under study. (Competitive interactions are discussed in Section
-------
-40-
11.) For example, Griffin and Au (1977) found that the adsorption of Pb by
montmorillonite was reduced when 0.1 M Ca(C10 ) 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 relative to that in a
distilled water system. Other "side reactions" may take place that can
complicate batch adsorption data; Na-Ca and Na-Mg exchange reactions on
bentonite were unaffected by CIO " in a study by Sposito et al. (1983) while
Cl" appeared to become a reactant in the exchange reactions, rather than
serving as an "inert" background electrolyte. The formation of CaCl"1" and
MgCl + 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 governed by the philosophy that they should be
simple and designed primarily for routine use. Thus the use 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 soluble constituents in the
particular clay or soil under investigation.
• It is recommended that the electrical conductivity (EC) of the
equilibrated soil-solution be measured so that the ionic strength of the
solution can be calculated by the relationship given by Griffin and Jurinak
(1973b), viz.,
-------
-41-
I = 0.0127 x EC(dS/m) [3]
where I is the ionic strength in units of moles/L. In the case of 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 may
render the adsorption data difficult to interpret.
-------
-42-
SECTION 7: EFFECTS OF PHASE SEPARATION
In a search of the literature, very few researchers were found to have
used a filtration technique to separate the liquid and solid phases prior to
the analysis of 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 organic compounds. Luh and Baker
(1970) found that a correction factor was necessary to account for retention
of i^C-tagged materials on the filters used in their study. The factor was
reasonably constant, but the filtration technique was abandoned in favor of a
centrifugation technique which avoided the problem by using gravitational
forces to separate the solids from the liquid phase. In a preliminary test,
Yaron and Saltzman (1972) also abandoned the filtration technique due to the
filter paper retaining parathion. In similar studies, Griffin and Chou (1980)
found that cellulose acetate membranes (0.45- and 0.22-ym pore size) adsorbed
significant "amounts of polybrominated biphenyls (PBBs) or hexachlorobenzene
(HCB). The problem could -be overcome but required 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 reproducible values for the concentration of the compound
passing through the membranes (Fig. 13). Figure 13 also indicated that
presaturation of 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). In this case, there were no significant
differences between filtration and centrifugation with respect to solute
concentrations. It was concluded that laboratories performing adsorption
-------
-43-
6 -I
Millipore membrane (0.45 JLim)
Millipore membrane (0.22 (1m)
Membrane presaturated by soaking
T
5678
Number of continuative filtrations
T
10
T
11
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).
-------
-44-
130-1
I
110 120
Arsenic concentration (mg/L)
CENTRIFUGATION
r
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).
-------
-45-
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 filtration membrane for the solute was evaluated adequately;
failure to do so may lead to erroneous results.
• As a guideline for conducting batch adsorption studies, it is recommended
that the solid and liquid phases be separated by centrifugation unless the
investigator can clearly demonstrate that the use of filtration techniques
does not significantly affect the results.
-------
-46-
SECTION 8: EFFECTS OF THE METHOD OF MIXING
In theory, the equilibrium distribution of solutes and adsorbates should
be independent of the mechanical device used to mix the solid-liquid mixture
during the equilibration interval. However, there have been some indications
in past studies that the method of mixing can influence the resulting adsorp-
tion data. For example, Barrow and Shaw (1979) compared three mixing methods
in a study concerned with phosphate adsorption: 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 phospate
adsorption tended to be less when a roller was used to mix the suspensions.
Barrow and Shaw (1979) felt that this trend was an experimental artifact,
related to the vigor of mixing. They envisioned that the differences were due
to particle breakdown; the more vigorous the agitation, the greater the soil
particles were broken down exposing "new" adsorption sites available to
phosphate for adsorption. They also acknowledged that the efficacy of the
three agitation devices, with respect to their ability to thoroughly mix the
suspensions, may have contributed to the differences.
In the development of the ASTM 24-hour Batch-Type Distribution Ratio (Rd)
procedure described by Griffin et al. (1985), a first generation procedure was
formulated around the ASTM-A, Water Shake Extraction Method (ASTM, 1979). A
round-robin sensitivity analysis of this early procedure performed by a number
of participating laboratories (see Acknowledgments) found that the method of
mixing 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. The results from the first sensitivity analysis are reported in
Table 5. Large differences in concentrations between the laboratories yielded
-------
-47-
Table 5. Results of first ASTM sensitivity analysis for cadmium (Cd) and
arsenic (As) at high (200yg/mL) and low (10 ug/mL) initial
concentrations where shakers and a paddle stirrer were used as the
mixing method.
1 ah
LdU
A
B
C
C
D
E
Overall
S
C.V.(%)
C amr* 1 o
oalilfJ I c
1
2
3
1
2
3
1
2
3
1
2
3
1
2
3
1
2
- 3
mean
H-
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
igh
As
186
186
185
128
131
127
162
168
175
130
134
130
153
25
16
Lov
Cd
,.n/ml 2_ ______
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
.5 0.073b
.6 0.064
.7 87.7
tf
As
7.92
7.77
7.91
0.43
0.38
0.46
5.00
4.78
5.81
0.53
0.55
0.55
3.51
3.32
94.5
Shaker Rate
Strokes/min Throw (inches)
59 3"
70 1.5"
100 1.25".
'
70 1.25"
Paddle stirrer used
Not known
-
a Represents post-procedure solute concentrations.
b Does not include values less than the detection limit.
interlaboratory coefficients of variation (% C.V.) in excess of 145 percent.
This first round of interlaboratory study was a clear example of why a
standard adsorption procedure was needed.
To improve the consistency of interlaboratory results, a National Bureau
of Standards (NBS) rotary extractor was tested as the mixing system (Fig.
15). A second sensitivity analysis was carried out (Tables 6 and 7) where
-------
2-Liter plastic or glass bottles
1 /15-Horsepower electric motor
CD
Figure 15. The National Bureau of Standards Rotary Extractor (Diamondstone et al., 1982).
-------
-49-
Table 6. Cadmium adsorption data from the 2nd ASTM interlaboratory
sensitivity analysis using a NBS rotary extractor as the mixing
method.
Lab
A. Rep
1
2
3
B.
1
2
3
C.
1
2
3
0.
1
2
3
Initial 24 hr
cone. cone.
- ug/mL -
200
200
200
200
200
200
200
200
200
190
190
190
X
S
C.V. (%)
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
Rd
mL/g
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
Initial 24 hr
cone. cone.
- yg/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
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.
Rd
mL/g
1734
1567
1580
1798
1461
1214
1554
1554
1495
1487
1761
1613
1568
±156
9.97
-------
-50-
Table 7. Arsenic adsorption data from the 2nd ASTM inter!aboratory
sensitivity analysis using a NBS rotary extractor as the mixing
method.
Lab
A.
B.
C.
0.
Rep
1
2
3
1
2
3
1
2
3
1
2
3
Initial
cone.
205
205
205
200
200
200
200
200
200
200
200
200
X
S
C.V.(%)
24 hr
cone.
ug/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
Rd
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.
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 hr
cone.
yg/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
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
-------
L
-51-
each of the participating laboratories used an NBS rotating extractor. The
coefficient of variation (% C.V.) between the mean values for each laboratory
reflects in part the precision of the mixing method. The coefficient of
variation of Rd values based on initial cadmium and arsenic concentrations of
10 mg/L and 200 mg/L were less than 8 percent and 12 percent for cadmium and
arsenic, respectively. These results can be compared with those from the
first round using predominantly shakers, which were as great as 145 percent
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 between 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.
• It is strongly recommended that all adsorption experiments, including both
inorganic and organic systems, use an NBS rotary extractor or equivalent
during each phase of the construction of an adsorption curve (i.e.,
determining a soil:solution ratio (Section 9), equilibration time (Section
13), and of course the adsorption curves themselves). Adsorption data
generated with other mixing devices may be valid, but to insure standardized
results between laboratories, these data should not be routinely accepted
unless the investigator can document that these other devices yielded data
comparable to those from an NBS rotary extractor or equivalent.
-------
L
-52-
SECTION 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 liquid. For the purposes of these proce-
dures, it shall be assumed that one milliliter of solution, regardless of its
composition, weighs one gram. In order to construct an adsorption isotherm
(curve), it is necessary to determine soilrsolution ratios that will permit
enough solute to be adsorbed to result in measurable, statistically signifi-
cant differences in solution concentration. In these procedures, increasing
the soil:solution ratio from a "low ratio" to "higher ratios", such as 1:1 to
1:100, means that the volume of solution increases relative to the weight of
the soil material. If the soil:solution ratio was too low, i.e., too much
adsorbent or too little solution, the majority of the solute initially in
solution may be adsorbed, forcing the investigator to attempt to analytically
measure small differences in concentration between concentrations that are
low. On the other hand, if the ratio was 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 concen-
tration between large concentrations. Unfortunately, with inorganic and polar
organic compounds, a suitable soil :solution ratio cannot be determined^
priori. The soil:solution ratio of the ASTM 24-hour Rd procedure is 1:20
(Griffin et al., 1985). 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 Section 17. A value of
10% to about 30% adsorption for the highest solute concentration used is a
useful criterion for selecting a soil-.solution ratio. This will give a dis-
cernible decrease in solute concentration that is statistically acceptable
-------
-53-
with respect to the initial concentration. Justification for this guideline
is given in Section 12. An example of this type of approach is given in Table 8.
Using a 1:4 soil-.solution 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 soil:solution ratio was used to generate data at lower
concentrations 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
Table 8. Soil:solution ratio determination for the Sangamon soil and Vandalia
ablation till using cadmium as the adsorbate.
Initial concentration
= 200 mg/L
SANGAMON
Soil : solution
Ratio .
1:4
1:10
1:20
1:40
1:60
1:100
1:200
1:500
Cd
yg/g
722
1631
2792
4246
5165
6250
7500
9250
adsorbed
%
95.2
86.1
73.7
56.0
45.4
33.0
19.8
9.8
wg/g
635
1359
2143
3012
3441
3880
4560
4900
VANDALIA
(ABLATION)
Cd adsorbed
%
94.1
76.2
44.3
25.4
19.1
13.0
8.0
4.6
Initial concentration = 10 mg/L
1:100
1:200
1:500
1:1000
957
1736
3178
4325
91.1
82.7
60.5
41.2
840
1474
2215
__
80.0
70.2
42.2
__
-------
', ft.
-54-
the 200 mg/L Cd solution was used, only 9.8% was adsorbed at the same
soilrsolution ratio. Essentially, the object is to select 'a soilrsolution
ratio that will serve as 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%, and at the same time the amount of
cadmium remaining from a low concentration (10 mg/L) solution was also within
analytical detection limits.
This rationale for selecting soilrsolution ratios is illustrated graphi-
cally in Figs. 16 and 17. In each figure, the amount of solute remaining in
solution after 24 hours is plotted against the soilrsolution ratio. The .
speckled area approximates the desired'solute concentration after 24 hours of
mixing given that about 10% to 30% of the solute is adsorbed. When the data
points or lines connecting the data points fall within this speckled area or
"adsorption target zone," the corresponding soilrsolution ratio will usually
yield satisfactory results. The adsorption behavior of seven soil materials
with respect to arsenic is shown in Figure 16. In this case, a IrlO ratio was
chosen to construct adsorption isotherms with six of the seven adsorbents.
Figure 17 illustrates the same concept with six samples using cadmium as the
solute. A IrlO ratio was chosen for the Tifton loamy sand, although any ratio
between IrlO to Ir4 would have probably yielded satisfactory results. A Ir20
ratio was chosen to study cadmium adsorption by the Cecil clay loam sample
while a IrlOO ratio appeared to be feasible for the remaining four soil
materials.
Comparison of the two figures indicates that the adsorption of arsenic
was essentially a linear function of the soilrsolution ratio whereas the
adsorption behavior of cadmium appeared to be influenced by the soilrsolution
-------
U3
C
Solution concentration of As (mg/L.)
rt
-«.
O
-J
o. -«•
-i. cr
-h C
-J O
(t 3
3
r+ O
-t>
in
O o>
5
n>
O>
o>
o
o
3
O
n>
<-»•
-»
Of
rt
-•> o
C 3
3 t/>
O
o <-••
3 n>
ro
o 3-
-i. o
— • C
•• -J
l/l (/)
o
— • o
C -i>
rt
-•• O
O O
3 3
rt-
OJ
O
(1:100 Soil:solution ratio
i
01
en
i
-------
-56-
180
160-
140-1
120-
100-
80-
60-
40-
20-
Tifton loamy sand
i
1 :100 1:40 1:20 1-10 1:5
Soil:solution ratio (mass/volume)
I
1 :4
ISGS 1985
Figure 17. Distribution of cadmium concentrations after 24 hours of contact
with different soil materials as a function of soil:solution
ratio.
-------
-57-
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 section.
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 aqueous extract of
the waste contained several aqueous constituents of interest, and a suitable
soil:solution ratio had to be determined for each solute. It would be ideal
if one single ratio could be used for all of the solutes with each given soil
but, for example, the concentration of zinc in the extract (550 mg/L) was much
larger than that of barium (2.26 mg/L).
A 1:20'soil:solution ratio for the Sangamon sample (Table 9) resulted in
32.5% of the zinc being adsorbed, but using the same ratio also resulted in
96.2% of the lead in solution being adsorbed which resulted in the solution
concentration of lead being very close to detection limits. When the "stock"
extract was diluted to construct an adsorption isotherm, the adsorption
behavior of lead could not be described using this soil:solution ratio (1:20)
since most of the equilibrium concentrations of lead would be below analytical
detection limits. Thus a 1:20 ratio was selected to construct a zinc adsorp-
tion isotherm, while 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 with this system. A 1:20 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.
-------
-58-
Table 9. Determination of soil:solution ratios for the Sangamon Paleosol and the
Cecil clay loam sample using an extract of Sandoval zinc slurry.
Solution
Cone. (mg/L)
0.84
1.15
1.80
2.00
2.19
2.25
2.26
0.15
0.55
2.64
4.70
8.18
11.4
14.6
269
365
485
494
532
542
541
SANGAMON
% Adsorbed
Ba
62.9
49.1
20.3
11.5
3.1
0.4
-
Pb
99.0
96.2
81.2
67.8
44.0
21.9
-
Zn
50.3
32.5
10.4
8.9
1.7
0
-
SOIL
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
CECIL CLAY
% Adsorbed
Ba
8.7
4.4
5.0
0.4
0.4
-
Pb
69.3
52.5
26.5
21.1
13.6
11.6
-
Zn
53.5
35.3
21.3
10.0
4.6
-
LOAM
Soil : Solution
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
* Soil:solution ratio selected for the kinetic experiments and the adsorption
isotherms.
-------
-59-
A 1:4 soil:solution was chosen to study zinc adsorption by Cecil clay
loam (Table 9), although any ratio between 1:3 to about 1:8 could also be
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 Section 11, there are guidelines for selecting ratios
within the acceptable range. Thus three different soil:solution 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).
-------
-60-
SECTION 10: SELECTION OF A SOIL:SOLUTION RATIO FOR NONIONIC SOLUTES
While finding a suitable soil :solution ratio for ionic and polar solutes
requires laboratory work, there is a simple calculation that can be used to
estimate a suitable ratio for nonionic solutes, particularly hydrophobic
organic species. This estimation technique requires a value for the organic
carbon content of the adsorbent and for the organic carbon partition coeffic-
ient (Koc) of the solute (McCall , 1981).
A derivation of this estimation technique begins with:
let K - y9s solute/g soil . ^s/9 rd1
d ug solute/g solution ug /g L^J
w w
where Kd is equivalent to the Freundlich constant Kf (refer to Section 14) in
the special case where the isotherm in linear (i.e., i/n is unity), and
v9s/9 is the mass of solute adsorbed per gram of the adsorbent, and
u9w/9 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 approximately equal to its volume (i.e., 1 ml « 1 g),
then R is the soil :solution ratio. Eq. [4] becomes
Since ug$ + ygw should equal the total mass of solute initially added (yg°)
assuming that losses due to volatilization or microbial degradation are
negligible, then
u9
- u9s) R
or
C6]
y9s
R = (u9° - u9s) Kd C7]
-------
-61-
Thus, it is possible to select an appropriate soil isolation ratio (R)
based on an estimate of the Kd value of the specific solute'-adsorbent
system. An estimation of Kd can be calculated if the organic carbon content
(OC) of the adsorbent and the KQC of the solute are known by
Kd = KQC (%OC)/100 [8]
The organic carbon partition coefficient (Koc) 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; and Roy and Griffin, 1985). Many of the KQC values
that have been reported were based on empirical equations that relate the
solubility (S) of the solute in water to its organic carbon partition
coefficient (Koc), such as the expression given by Hassett et al. (-1983),
viz.,
log KQC = 3.95 - 0.62 log S (mg/L) [9]
A similar linear relationship has been observed relating the octanol-water
partition coefficient to its organic carbon partition coefficient, such as the
version given in Hassett et al. (1983), viz.,
log Koc = 0.088 + log KQW [10]
A compilation of octanol-water partition coefficients was published by Leo et
al. (1971). The historical evolution 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 mixture containing dichloroethane,
tetrachloroethylene, and o-xylene by a Catlin silt loam sample was studied.
In order to construct adsorption isotherms, suitable soilrsolution ratios for
each solute had to be determined. The organic carbon content of this soil
-------
-62-
sample was 4.04%. An estimate of a Kd value for each solute was based on its
water solubility using eq. [9],
The solubility of dichloroethane and tetrachloroethylene is 8450 mg/L 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). Using eqs. [8]
and [9], the calculated Kd values of dichloroethane, tetrachloroethylene, and
o-xylene were approximately 1.3, 13.4, and 14.7, respectively. Recall from
Section 9, a soilrsolution ratio corresponding to about 10 to 30% adsorption
is a useful criterion for selection a suitable ratio. Thus, assuming that 20%
adsorption will fall into the "target zone" for each of the organic solutes,
u9s/u9° is set equal to 20. Then the soilrsolution ratio for each solute may
be calculated by arbitrarily setting yg° equal to 100. For example,
tetrachloroethylene:
R - 20 1
UUU-2U) 13.4 "53T6"
o-xylene:
R - 20 1
(100-20) 14.7 58.8
There is no reason to work with such awkward numbers for the actual
measurements. These soil: solution ratios could be simplified to 1:50 and
1:60. As discussed in Section 9, it would be fortuitous when a single
soil:solution ratio could be used to generate an adsorption isotherm for every
solute of interest in a multicomponent mixture. In this example, a 1:50 ratio
was selected for the mixture to generate adsorption isotherms for each solute
as shown in Figs. 18 and 19; a single ratio was suitable in this case. This
also illustrates that there may be a range of suitable ratios for some organic
solutes, but Section 11 should be consulted for guidelines for selecting
suitable ratios.
-------
-63-
20
» 18H
O)
— 16
14-1
•8 12
i, 10
SL
•a
c
o
8-
6-
4-
2-
o-xylene
Kd = 10.41
0.4 0.8 1.2 1.6
Solution concentration (mg/L)
2.0
Figure 18. Adsorption isotherm of o-xylene by Catlin at 23°C, and at pH 6.1,
-------
-64-
0.8-
O)
a.
0.6-
o
to
"0.4-1
o
a
•§
0.2-
tetrachloroethylene
Kd = 17.28
0.02
0.04
0.06
Solution concentration (mg/L)
Figure 19. Adsorption isotherms of dichloroethane and tetrachloroethylene by
Catlin at 23°C, and at pH 6.1.
-------
-65-
This estimation technique can be generalized and shown as a relationship
between the linear Freundlich constant (K^) and the soil:solution ratio (R),
as a function of different amounts of adsorption on a percentage basis (Figs.
20 and 21). McCall et al. (1981) demonstrated that eq. [5] could be
rearranged as:
£- ((u9°/pgs) - 1) Kd [11]
which was used to generate Figs. 20 and 21. These figures should serve as
convenient guides for selecting soilrsolution ratios. For example, the
solubility of carbon tetrachloride in water is 800 mg/L at 25°C. The Koc
value, estimated using eq. [9], was 140. Using a Catlin silt loam sample with
an organic carbon content of 4.04%, a Kd value was calculated as
K = HO (4.04) = 6 ri2-,
*d run ° L1ZJ
Reading from Figure 20, a soilrsolution ratio of about 1:10 should yield
approximately 30% adsorption.
-------
-66-
1:180-
1 .20 -
20 40
120 140
linear Freundlich constant (Kd)
Figure 20. Relationship between the linear Freundlich constant (Kd) and soil:
solution ratio, as a function of percent adsorption (lower range).
-------
-67-
1:900-
1:800-
200
400 600 800
Linear Freundlich constant (Kd)
1000 1200
Figure 21. Relationship between the linear Freundlich constant (Kd) and soil
solution ratio, as a function of percent adsorption (upper range)
-------
-68-
SECTION 11: EFFECTS OF THE SOIL-.SOLUTION RATIO
The soil :solution ratio may be one of the most important experimental
variables to consider when constructing an adsorption isotherm and evaluating
the adsorption data, particularly when comparing results from different
investigators using different ratios. In Figure 18, increasing the amount of
adsorbent while holding the volume of solution constant had the effect of
increasing the mass as well as surface area on which the arsenate ions could
be adsorbed. Hence, intuition suggests that as the amount of adsorbent is
increased, the amount of arsenic left in solution after exposure should
decrease in an essentially uniform manner as shown in Fig. 16.
Figure 17 demonstrated a non-linear response; the amount of cadmium left
in solution after 24 h'ours appeared to be approaching a constant value as the
amount of adsorbent was increased (i.e., the soil: solution ratio was
decreased). There is no single explanation for all systems for this non-
linear response or what White (1966) called the "soi1:solution ratio
effect." This phenomenon does not negate the selection of a soi1:solution
ratio, but the consequences of that selection must be considered.
The soil:solution ratio effect and the adsorption of phosphate has
probably received the most attention, although there are conflicting reports
concerning its effects (Barrow and Shaw, 1979). Phosphate adsorption was
increased by the use of high soil:solution ratios in the studies of Fordham
(1963), Barrow et al. (1965), and White (1966). Hope and Syers (1976) found
that high ratios resulted in lower phosphate adsorption. An early paper by
Kurtz et al. (1946) found no soil:solution ratio effect (i.e., a linear
response) when studying phosphate adsorption by Illinois soils.
-------
-69-
White (1966) attempted to reconcile his results by arguing that the
system was not at equilibrium. However, this line of reasoning contradicted
his rationale for selecting an equilibration time. Larsen and Widdowson
(1964) had concluded two years earlier that the soil:solution 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 soil:solution ratios _affected
only the rate at which phosphate was removed from solution. They found that
the change in solution phosphate concentration when mixed with their soils was
proportional to the reciprocal of time. Thus, when the reciprocal-time scale
was extrapolated to zero, i.e., infinite time, the effects of different
soil:solution 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 in order for soilrsolution effects to essentially disappear
and thus the adsorption data would be essentially independent of the
soil:solution ratio, approaching the expected linear response.
This hypothesis was challenged by Barrow and Shaw (1979) who found that
the reciprocal-time analysis used by Hope and Syers (1976) did not explain the
soil:solution ratio effects observed in their study. Barrow and Shaw (1979)
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.
While the mechanisms proposed above may be operative in some systems, the
soil:solution ratio effect has often been attributed to the competitive
interactions between a given solute and species that are concommitantly
desorbed or exchanged during the partitioning of solutes and adsorbates. As
-------
-70-
the amount of adsorbent is increased, there is a larger source of these
potentially competing constitutents. The net effect is that the magnitude of
adsorption (given equal initial concentrations) decreases. 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 adsorbent was increased, the amount of calcium
that was desorbed or exchanged from the clay also increased and competed with
lead for adsorption sites.
A similar phenomenon was observed in this study in which the adsorption
characteristics of a Sangamon paleosol were investigated using CdCl . There
was a strong soilrsolution ratio effect on cadmium adsorption (Fig. 22). The
curvilinear distribution of data points was derived by using a 1:100 soil:
solution ratio. However, when different soil:solution ratios were used, the
resulting data did not follow the same pattern but fell on a nearly straight
line that intersected the adsorption curve obtained where 1:100 ratios were
used.
It was suspected that Ca2+ and Mg2+ 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 cadmium. At any
given equilibrium concentration of cadmium, higher soi1:solution ratios (i.e.,
less adsorbent 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 montmorillonite, illite
and kaolinite were between 0.8 and 1.3 (on a concentration basis), suggesting
-------
-71-
1:200
20
40 60 30 100 120
Equilibrium cadmium concentration (mg/U
140 160
Figure 22. Effect of soilisolution ratio on cadmium adsorption by a Sangamon
paleosol sample at pH 6.1, and at 22°C. The solid dots were
derived by using a 1:100 ratio (Roy et al.t 1984).
-------
-72-
that these clay minerals have no strong affinity for one cation versus the
other over a pH-range of approximately pH 5 to pH 7 (c f. Bolt and
Bruggenwert, 1978). Calcium will readily exchange with cadmium and vice
versa. If the adsorption data are plotted as cadmium adsorbed relative to
Cd2+/(Ca2+ + Mg2+) on a molar basis (Fig. 23), the different soilrsolution
ratios coalesced into one adsorption curve.
The soil:solution ratio can also influence the chemical composition of
the system which in turn can directly 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 literature on the effects of dilution on ionic concentration in
soil solutions and attempted to generalize the results:
Nonsaline soils:
1. Solution potassium increased with dilution
-2. Calcium and magnesium in solution frequently increases with
dilution while the ratio of Ca:Mg changes
3. Phosphorus usually increases proportionally to dilution
Alkali, calcareous, and gypsiferous soils:
In virtually all cases, dilution results in increased amounts of
Ca, Mg, Na, K, SO , P, and Si
Thus the ionic concentrations in soil solutions and soil extracts are not
inversely proportioned to the amount of water present.
The pH of the soil-liquid suspension will also be affected by the
soi1:solution ratio. The relationship between pH and adsorption is discussed
in Section 5. The pH of a soil suspension in a batch adsorption procedure
will be controlled by three factors:
-------
-73-
. T
6-
• . 1.100
5-
1.60
2
•a
- 3H
o
<
2-
1 -
1 40
•1 .20
1 10
2.0 4.0 6.0 8.0 10.0 12.0 14.0 16.0 18.0 20.0
Ratio of equilibrium molar concentrations of [Cd] / [Ca + Mg] -cs '985
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).
-------
-74-
1. the "natural" pH of the adsorbent and its buffering capacity to
maintain that pH
2. the pH and composition of the liquid phase
3. adsorption reactions that directly or indirectly change the H 0+
and/or OH" 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:
solution ratio. In Figure 24, the soil materials were exposed to an sodium
arsenate solution containing 200 my/L As with an initial pH of 4.65.
Consequently, at progressively 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 soil:solution
ratios, the equilibrium pH of each solution became more like that of the soil,
the relative strength of this tendency depending on the pH buffering capacity
of the soil.
In the second example (Fig. 25) the soil materials were exposed to a
cadmium chloride soltuion containing 200 mg/L Cd with an initial pH of 5.45.
A 1:20 ratio for a kaolinite clay sample (Fig. 25) was associated with a
solution pH of 7.05, while a 1:4 ratio resulted in a solution pH of 7.45, an
increase of 0.4 pH units. Thus an isotherm generated with a 1:4 ratio may
yield lower amounts of cadmium adsorption than one using a 1:20 ratio simply
because the pH of the former tended to be more basic for reasons discussed in
Section 5.
A similar type of relationship may be observed with complex, multi-
component extracts or leachates. The equilibrium pH of the zinc slurry
extract (Appendix B) was plotted against soil-.solution ratio using two soils
-------
-75-
8-
(unaltered)
pH
4-
-- Solute solution tends to dominate pH of mixture
Tifton loamy sand ^
Cecil clay loam
Soil tends to dominate pH of mixture _- .
1•100 1 40 1 20
1:10
Soil: solution ratio (mass/volume)
1 :5
—I"
1 '4
Figure 24. Distribution of pH values of arsenate solutions (containing the
same initial arsenate concentration) after 24 hours of contact
with different soil materials as a function of soiltsolution
ratio.
-------
-76-
8-
7-
6-
(unaltered)
Sangamon Paleosol
Vandalia Till (ablation phase)
Solute solution tends to dominate pH of mixture
-j- pH of Cd solution
5-
4-
Tifton loamy sand
Cecil clay loam
Soil tends to dominate pH of mixture
i—i 1—
1:100 V40 1:20
—I
1-10
Soil solution 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 soil:solution ratio.
-------
-77-
(Fig. 26). In both cases, lower soil isolation 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 soil :solution ratio will often influence the ionic strength of the
solution. This is to be expected since the ionic strength of any solution
would be controlled by the concentration and charge of both the solute(s)
under study, desorbed or exchanged ions, and/or other aqueous ions derived
from the dissolution of soluble minerals that naturally occur in the
adsorbent. The ionic strength of the solutions in contact with the Tifton
loamy sand and the Cecil clay loam tended to decrease as the soil:solution
ratio decreased (Fig. 27). This trend was attributed to two factors: 1) as
the ratio decreased, more arsenic or cadmium was removed from solution which
lowered the ionic strength, and 2) these two soils contained a low content of
water soluble compounds that contributed to the ionic strength upon dis-
solution. The other three soil materials (Fig. 27) were slightly calcareous
by comparison and consequently lower ratios resulted in an increase in ionic
strength due to dissolution of slightly soluble minerals. Discernible
decreases caused by the removal of cadmium were masked by the dissolution of
carbonates. Whether these changes or differences in ionic strength will have
a major impact on the adsorption data is difficult to generalize (see Section
6). No routine adsorption procedure designed to be relatively simple can
address this problem completely. Defining the relationship between ionic
strength, soil:solution ratio, and adsorption for any soil-solute(s) system
may be a large project in its own right.
The adsorption of organic solutes may also be influenced by the
soil:solution ratio used in batch procedures. Grover and Hance (1970) found
that the Freundlich constant (Kf) decreased significantly by a factor of 2.6
-------
-78-
7-
a
T T
1-100 1:40
1:20
1:5
1 :4
1:2
Soil:solution ratio (mass/volume)
Figure 26. Distribution of pH values
after 24 hours of contact
soil:solution ratio.
of solutions
with two soi1
of the zinc
samples as
slurry extract
a function of
-------
-79-
7-
6-
u
1 4H
3-
Oilute solutions
-0.6
Very dilute solutions
1:100 1:40 1:20
1:10
Soil: solution ratio (mass/volume)
1:5
-0.5
-0.4
3
13
§
^0.3
-0.2
1:4
SOS 1985
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.
-------
-80-
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 a likely
cause for the differences in the extent of adsorption was related to the
aggregate size of the soil. In a comparison of the relative soil particle
sizes at three soil:solution ratios they placed 10 g of soil which had been
passed through a No. 10 mesh sieve into flasks. The flasks were shaken with
2.5, 10, and 100 ml of a 0.1 M CaCl solution, mixed by shaking gently end-
over-end for 30 seconds, and then allowed to stand. They found that the
dispersion of soil aggregates was greater at the 1:10 soil:solution ratio than
at the 1:0.25 ratio; the 1:1 ratio was intermediate. A similar sedimentation
behavior was also observed in the absence of 0.1 M CaCl . Thus they concluded
2
the extent of adsorption of linuron and atrazine is related to the aggregate
size of the soil.
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 effect in this case
appeared to result from the presence of soluble microparticles derived from
the soil which also tended to retain the solutes (see also Voice and Weber,
1985). They concluded that soil:solution effects reported in the literature
may have been due to incomplete phase separation during centrifugation or to
accumulative relative errors in measuring concentrations.
Similar conclusions were also reached by Gschwend and Wu (1985). If
precautions are taken to eliminate or account for nonsettling (or nonfilter-
able) microparticles or organic macromolecules, which remain in the aqueous
phase during batch adsorption procedures, the observed partition coefficients
(Kf or KQC) were found to remain constant over a wide range of soil:solution
ratios. Figure 28 showed that a succession of prewashing treatments of
-------
T)
IQ
03
ro
00
o -n
(/) T O T
O fD 3 o>
3~ i O C
*D in 3 Q.
3 ;j- cl- —i
O. _i. T _•.
3 0» o
O* lf"* <~t" 23"
3 -«•
Q. rt O o
030
C T £ ^
3 ^"^ fy
»-• O 3- 3
vo < r*
OO (T) -—
CD O ^-~
in
0) .in -ti
—• cr
--•Or*
3 —• s:
UD O
CD
T3
O) O>
rt Q.
O Z in
—• -•• O
(T> r* 3
in 3" n>
. O -,
c in
c-t-
O» -—- in
Q. O «
PJ —'
T3 O in
el- in fD
(D fD CL
Q. Q. -••
3
-t» in (D
-J "< 3
O 3 r*
3 cr
O
in
Freundlich constant, Kj (mL/g)
en
o
o -
8
o* _,
O
O
O
o
o
1:
it
fi
01
•
O
3;
€'
01
K
lorobipl
I
00
I:
I:
Ij
-------
-82-
sediments greatly reduced the effects of the nonsettling particles (NSP).
When prewashed sediments were used for batch equilibration experiments, the
observed Kf remained virtually constant over the range of soil-.solution 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 differ-
ence in Kf with and without prewashing clearly reflected the great sensitivity
of very strongly adsorbed compounds to small NSP concentrations in the aqueous
phase.
Voice and Weber (1985) concluded that while soluble microparticles could
play the major role in the soilisolution ratio effect with regard to the
adsorption of organic solutes, they felt that it could not account for all of
the data given in the literature. They proposed a hypothesis where the -
soiltsolution 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.
In other organic solute-adsorbent systems, the adsorption behavior of the
solute was not influenced by the soilrsolution ratio. Bowman and Sans (1985)
reported that the adsorbent concentration (soil:solution ratio) did not appear
to significantly affect 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 soil':solution
ratio (Fig. 29). The Freundich constant was essentially constant over a wide
range of soiltsolution ratios. When different soi1:solution ratios were used
in the construction of adsorption isotherms, the resulting data tended to plot
on the same line (Figures 30 and 31) and the slopes of the adsorption
isotherms were nearly unity. In some cases, a curvilinear distribution of
-------
-83-
800-
700-
_§
,r 300
200-
100-
Catlin
EPA-14
Cecil clay
Sangamon Paleosol
1 120 1 60 1 40 1:30 1:24
Soil. solution ratio (mass/volume)
1 20
Figure 29. The Freundlich constant (Kf) for the adsorption of Aroclor 1242 by
four different soils at 23 C as a function of soilisolution ratio.
-------
-84-
45-
40 -
1 .500
1.50
Cecil clay
Vandalia Till (ablation)
1 100
Sangamon paleosoi
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 five soils at 23°C using
various soil:solution ratios.
-------
-85-
• 1-250
0.02
0.04
0.06 0.08
Equilibrium concentration (mg/L)
0.10
0.12
Figure 31. Adsorption of dieldrin, tetrachloroethylene, and.1,2-
dichloroethane by Catlin at 23°C using various soil:so1ution
ratios.
-------
-86-
data points was derived by using different soil:so1ution ratios with some
adsorbents (Figure 32). However, the application of different soil solution
ratios still yielded a single, consistent relationship between the amount of
Aroclor 1242 in solution and the amount retained by the tills at equilibrium
(Figure 32).
In summary, the selection of a soil:solution ratio may or may not have a
profound effect on adsorption data. The soil:solution 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 inter-
actions, the soilrsolution ratio effect can be rationalized, but in other
systems, the ratio effect presents problems, particularly for procedures
intended for the routine collection of batch adsorption data. Voice et al.
(1983) commented that it is possible that some combination of techniques or
new methodologies will evolve (to handle the ratio effect), but no simple
solutions are readily apparent.
• In Section 17, specific soil:solution ratios are suggested for the
construction of adsorption isotherms. It is strongly recommended that these
ratios and only these ratios be used to ensure that different users will use
the same ratios regardless of the solute-adsorbent system under study. Thus
these ratios (Section 17.8.3) could be regarded as "standard soil:solution
ratios." For example, if it appears that a 1:8 ratio is satisfactory for the
generation of adsorption data, the investigator should attempt to use a 1:10
ratio, i.e., one of the "standard ratios". As shown in Section 9, 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
-------
-87-
25-
1.250
Vandalia till (altered)
0.02
0.04 0.06 0.08 0.10 0.12
Equilibrium Aroclor 1242 concentration (mg/Li scs -aas
Figure 32. Adsorption of Aroclor 1242 by altered Vandalia till and unaltered
Vandalia till at 23°C using various soil:solution ratios.
-------
-88-
enable the direct comparisions of adsorption data generated by different
investigators. Adsorption data based on ad hoc ratios may provide a basis for
limited comparisons, but there will always be some doubt that the results are
comparable unless it can be clearly shown that a particular solute-adsorbent
system is not subject to soilrsolution ratio effects.
-------
-89-
SECTION 12: CONSTANT AND VARIABLE SOIL:SOLUTION RATIOS
Basically, there are two experimental techniques in generating batch
adsorption data:
1. Mixing a batch of solutions, arranged in progressively decreasing
solution concentrations, where each solution is mixed with the same
(constant) weight of adsorbent.
2. Mixing a batch of solutions, all containing the same initial solute
concentration, with progressively increasing amounts of adsorbent.
The first technique obviously makes use of a single or constant soil:
solution ratio, presumedly a standard ratio selected using the procedures
given in Sections 9 or 10 and 17.8.3. The latter technique makes use of
different soil:solutian ratios in a manner very similar to the technique for
selecting a soil: solution ratio for ionic solutes (Section 9). Intuitively,
one would expect that either technique would yield the same result, and that
either could be used. While these generalizations are true in some cases,
they are not valid for all systems.
When using the constant soil:solution ratio technique, the initial or
stock solute solution, albeit 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 the containers, each with the same amount
of soil material. However, as discussed in Section 11, the soil:solution
ratio used may influence the adsorption data. Figure 22 showed that using
different soi 1-.solution ratios, ranging from 1:200 to 1:4 yielded adsorption
data that were in poor agreement with the isotherm generated when a fixed
(1:100) ratio was used. In this case, this phenomenon was attributed to
competitive interactions between Cd2+, and desorbed Ca2+ and Mg2+ and, as
shown in Figure 23, where the data was replotted taking into account these
-------
-90-
competitive interactions, the adsorption data coalesced into a single
consistent relationship. However, this replotting technique will not work in
all cases. The techniques for modeling competitive adsorption are currently
emerging (see for example, Murali and Aylmore, 1983 a, b, c; and Roy et al.,
1986), and are currently too complicated for use in routine batch
procedures. Moreover, not all "soilrsolution ratio effects" can be attributed
to competition (see Section 11). This dichotomy was characteristic of several
of the soils and soil components used in developing these procedures; the
application of the variable soiltsolution ratio technique yielded results
(amounts of cadmium and lead adsorbed) that were either similar to those using
a fixed ratio or tended to be lower. On the basis of this trend, an isotherm
produced using variable soil:solution ratios was viewed as the more environ-
mentally conservative. Hence, an isotherm produced in this 'manner is called
(in this document) an Environmentally Conservative Isotherm (ECI).
The Environmentally Conservative Isotherm (ECI)
The ECI has two major advantages over an isotherm where a fixed
soil:solution ratio is used: 1) if the solute-adsorbent system reached
equilibrium in 24 hours, or more correctly satisfied the conditions of the
operational definition of equilibrium (Section 13), then the data generated in
selecting a soilrsolution ratio can be used to construct an isotherm, and 2)
the effects of competition and other processes are implicitly accounted for
without knowing their exact nature.
Further documentation that using a variable soil:solution ratio yields
environmentally conservative estimates may be shown by Figure 33. The
adsorption data were modeled with the Freundlich equation (Section 14)
yielding the isotherm constants shown. The isotherms associated with the
-------
-91-
I.O-i
0.8-
5 0.6H
LU
0.2-
00
00 0.2 04 0.6 0.8 1 0
1 n iCSI)
• Cadmium
O Arsenic
A Lead
2-
1-
1
2
Kf ICSI), L/mg
Figure 33. Distribution of (A) exponents (1/n) and (B) Freundlich constants
(Kf) associated with arsenic, cadmium, lead, and PC8 (Aroclor
1242) adsorption isotherms.
-------
-92-
constant soil :solution ratio technique have been called Constant Soil :solution
ratio Isotherms (CSI). The results shown in Figure 33 may be generalized as
Kf EC I < Kf CSI
and ^" >
and hence, based on this type of analysis, an isotherm generated using
different adsorbent masses yields generally lower predictions of solute
adsorption and was thus viewed as being environmentally conservative. The ECI
is recommended as the method of choice in this document for routine use.
Experimental data produced during the process of selecting a
soil : solution ratio may be used to construct an isotherm if the system
equilibrated within 24 hours (equilibration time is discussed in the next
section). However, it is inevitable that some of the data points will be
associated with situations where less than 10% of the solute was adsorbed. In
Section 9, it was recommended that one should choose a soil :solution ratio
where at least 10% or greater adsorption occurred. To illustrate why this
recommendation was made, consider a situation where an investigator conducted
the experiments for selecting a soil :solution ratio then used data points from
the entire concentration range regardless of the accuracy of the determination
and attempted to construct adsorption isotherms (Figures 34 and 35). In each
case shown, the data points that were associated with less than 10% adsorption
did not conform to the general pattern established by the data associated with
greater than 10% adsorption. Elimination of these data points yielded more
satisfactory results, i.e., more reasonable r2 values (see Section 14).
Figure 35 represents an extreme case; nearly all of the data were associated
with less than 10% adsorption. As shown, fitting this data set with an
isotherm equation had little meaning.
-------
-93-
30-
0.4
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 with a Vandalia till sample
(unaltered) with the amount adsorbed associated with each isotherm
data shown. The mean pH of the soil-solute suspensions was 6.8.
-------
-94-
240
200
-S 160-
120 -
80-
40-
24 6ao adsorbed
X
X
= 004
—i r- 1 r-
4° 80 120 160
Equilibrium cadmium concentration (mg/ U)
200
Figure 35. Distribution of cadmium adsorption data at 22°C by a Tifton sandy
loam. The solid line is the Freundlich equation through the data;
the dashed line is the presumed shape of the adsorption
isotherm. The average pH of the soil-solute suspensions was 4.8.
-------
-95-
While the ECI is useful for many situations, it can not be universally
applied to every situation. The ECI may be limited to cases where (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 where the initial solute concentration is
low.
The ECI technique was used to derive arsenic adsorption isotherms with
the soil adsorbents used in the development of these procedures.
Soilrsolution ratios of 1:4 and higher were used, and the initial concentra-
tion of arsenic was 200 mg/L. Figure 36 illustrates the results; varying the
amount of adsorbent over this range of soil:solution ratios did not change the
equilibrium arsenic concentration substantially. The relatively small
changes in arsenic equilibrium concentrations caused the data points to be
somewhat clustered together, leaving an area between the origin of the
isotherm and the lower-most arsenic equilibrium concentration without data
points. It should be intuitively apparent that regression of these data sets
using isotherm equations could lead to potentially large errors. Moreover, it
was not possible to use lower soil:solution ratios to fill in the gaps; the
use of ratios much less 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. It is for these reasons that the constant soil:solution ratio
isotherm (CSI) is also recommended for application as an alternate procedure,
given that the (ECI) technique does not produce useful or applicable
results. Such as the situation shown in Figure 36.
-------
-96-
0.6-
0.5-
0.4-
•O
33
£1
0.3 -
Inaccessible region
0.2 -
0.1 -
o.o-
1 : 20 • (13% adsorbed)
•
1:10 *.*
S
Catlin silt loam
1 .4
Till (Vandalia ablation)
1-4
1-4
!£*^V
Kaolinite i
1.18
S
13
10 (18%)
(11%)
1 4
1 5 1-9
I 1
1 -7
Till (altered Vandalia)
10 (11%)
40 80 120
Equilibrium arsenic concentration Img/L)
160
200
ISGS '985
Figure 36. Distribution of arsenate adsorption data at 23°C by different soil
samples using different soi1:solution ratios. The pH values of
each soil-solute system were similar to those given in Appendix A
for each soil.
-------
-97-
SECTION 13: DETERMINATION OF THE EQUILIBRATION TIME
The equilibration time in batch adsorption experiments is the time
interval in which the system reaches chemical equilibrium and the concen-
trations of the products and reactants cease to change with respect to time,
viz.,
||= 0 [13]
Adsorption at the solid-liquid interface is a thermodynamic process and
adsorption measurements are taken when the system has equilibrated or, in
other words, when the reaction(s) between the adsorbent and solute has gone to
completion.
In past studies, many different equilibration times have been used. For
example, Lawrence and Tosine (1976) used 30 minutes to equilibrate PCBs with
soil, while Jones et al. (1979) allowed a soi1-phosphate mixture to equili-
brate for six days before separating the liquid from the soil. The
equilibration times given in most studies were probably valid and were 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 though it is a similar adsorbent-solute system. Equilibration
time is an experimental variable that must be determined for any system prior
to undertaking the construction of an adsorption isotherm (curve).
The proposed ASTM 24-hour R^ procedure provides a measurement of the
affinity of a soil or clay for solutes after 24 hours (Griffin et al.,
1985). However, 24 hours may or may not be long enough for the development of
chemical equilibrium. As indicated earlier, some investigators have used
equilibration times of days or even weeks. However, adsorption per se is
generally regarded as a fast reaction, and subsequent removal of a solute from
solution may be attributed to other processes. Adsorption processes at
-------
-98-
solid-liquid interfaces are often initially rapid, while further reduction in
solute concentration continues at a decreasing rate, asymptotically
approaching a constant concentration. In some cases, equilibrium was 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 difficult to determine when adsorption
processes dominate then become less important as other processes, such as ion
penetration or precipitation, become significant. The EPA (U.S. EPA, 1982)
suggested that the equilibrium 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.
Thus this definition is an operational definition of equilibrium and is
equivalent to a steady state. Cast in a form similar to that of eq. [13] it
may be written as
-|£- < 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 found to be in
equilibrium after 24 hours as defined by this operational definition. Figure
37 presents representative data for cadmium adsorption and Figure 38 shows the
adsorption behavior of arsenate by 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 or less than 5% per 24-hour interval. It may be
-------
-99-
0>
O)
E
^
I
O
Vandalia till (ablation)
Vandalia till (unaltered)
70
Time (hours)
Figure 37. The adsorption behavior of cadmium by five soil materials at 22°C
as a function of contact time.
-------
-100-
200
190-
Ceal clay loam ,
1
110
90
70
50
30
0
EPA-14
•r
^
1 I
1 1
' C,o,c,,,SC, i
1 :
i i i i i *
10
20
30 40
Time (hours)
50
60
70
Figure 38. The adsorption behavior of arsenic at 23°C by 11 different soil
materials as a function of contact time.
-------
-101-
more convenient to analyze kinetic data in a manner shown in Table 10. As
shown, an equilibration time of 24 hours was selected for three of the
samples, while 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. After
the first 24-hour interval, the amount of arsenic mixed with the kaolinite
sample continued to decrease, but only by 1.14/6 during the next 24-hour
interval. It is this slow and relatively small decrease in solute
concentration which follows the more rapid and pronounced decrease that is
frequently a problem. The application of this operational definition of
equilibrium assumes that this additional 1.14% decrease is negligible and may
be attributable to processes other than adsorption. Therefore, this solute-
adsorbent system is defined as being at steady state after 24 hours of
contact.
The application of this procedure with multicomponent solutions is
exemplified by the metallic waste slurry (Appendix 8). In order to determine
the time interval necessary for the development of chemical equilibrium of
each system, preliminary kinetic experiments were carried out using the soil:
solution ratios previously determined (Section 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 concentration for the first 24-hour
period was 12.2% (Table 11). After 24 hours, the rate of change was less than
5 percent 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 concentration of
these two solutes per 24-hour interval was less than 5 percent (Table 11).
-------
-102-
Table 10. Determination of equilibration times for the adsorption of arsenate
by soil materials.
Time
(hr)
0
1
4
8
16
*24
36
48
72
Vandal
Time
(hr)
0
1
4
8
16
24
36
*48
72
1%AC = (Cx
Cecil clay loam
XAC1 Solution
Cone. (mg/L)
193.4
182.9
181.5
180.0
177.6
8.17 177.6
177.1
0.56 176.6
1.08 174.7
ia Till (alation)
&AC Solution
. Cone. (mg/L)
199.3
179.5
173.2
169.4
164.0
18.97 161.5
157.3
5.51 152.6
3.15 147.8
- C )/C where C = solution
P — -hho /*r\n/*£
Kaolinate
Time
(hr)
0
1
4
8
16
*24
36
48
72
Sangarnon
Time
(hr)
0
1
4
8
16
%AC
—
-.
-
_
_
16.36
_
1.14
0.55
Solution
Cone. (mg/L)
199.3
171.4
168.6
168.1
166.2
166.7
164.8
164.8
163.9
Paleosol
%AC
_
_
_
_
_
*24 33.13
36
48
72
concentration at
in^^a-^Trtrt a-fr" +• 4 nt/-\
.
2.20
2.25
time
•». j. '
Solution
Cone. (mg/L)
199.3
165.3
158.2
158.1
152.6
149.7
148.3
146.4
143.1
t, and
") A t>*S^llM*»
equilibration time selected for the adsorption isotherms
-------
-103-
Table 11. Determination of equilibration times for the adsorption of Ba, Pb,
and Zn from a Sandoval zinc slurry extract by the Sangamon paleosol
and Cecil clay.
Time
(hr)
-
0
1
8
*24
31
48
72
—
0
1
8
*24
31
48
72
_
0
1
8
*24
31
48
72
%Ad
Ba - -
-
-
-
12.2
-
-1.0
-1.5
Pb - -
-
-
-
66.8
-
8.9
5.1
Zn -• -
-
-
-
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
(hr)
0
1
8
24
31
*48
72
0
1
8
*24
31
48
72
%Ad Solution
Cone. (mg/L)
Pb
15.4
7.72
7.27
54.5 7.01
6.72
6.3 6.57
1.5 6.47
Zn
549
421 '
430
20.9 434
430
0.2 433
0.2 432
= (C - C )/C where C = solution concentration at time t, and
C = concentration at time t + 24 hours
* equilibration time selected for the adsorption isotherms
-------
-104-
580-
560-
520-
_ 480-
3
J 440-
c
0
§ 400-
e
I 360-
u
c
o
I '"
10-
(
I
5
I ^
1
L
L^
^^v.^
^
?
,
i
hr
i— ,
1
i i
3 10 20
3
O
£
03
V
>— • .
1 1
30 40
3
0
CM
r»
Zn-Cecil clay
Zn-Sangamon
Pb-Sangamon
1 8a-Sangamon
1
l 1 1
50 60 70
Time (hours)
SGS 1985
Figure 39. Determination of equilibration time of Ba, Pb, and In from a
laboratory extract of the Sandoval Zinc slurry with the Sangamon
Paleosol and the Cecil clay sample.
-------
-105-
Zinc was adsorbed by the Cecil clay, and this 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 with the exception of lead adsorption by the Cecil clay
in which a 48-hour interval was used.
Solution concentrations of o-xylene, dichloroethane and tetrachloro-
ethylene tended to change by amounts less than about 5% when in contact with
Catlin (Fig. 40). As shown in Table 12, the rate of adsorption of the PCB
Aroclor 1242 by a Catlin sample was nil after the initial 24 hours of contact.
• In summary, it is recommended that the equilibration time should be the
minimum amount of time needed to establish a rate of change of the solute
concentration in solution that is equal to or less than 5% per a 24-hour
interval. This minimum time, typically 24 hours, should be determined for
each solute-adsorbent system prior to the construction of adsorption
isotherms.
-------
(O
c
n>
rt <-»• —|
-*• n> 3-
3 rt CD
n> -»
» ft> ft!
O Q.
ft) rr en
rt-'O
O T
XJ T T3
1C O r+
o> -••
o>
3 3-
n> eu
ro
co o
o -h
o
o
cr i
fu CO
r» 3
— n>
(/> O
3-
c
3
O
rt
O 3
3 n>
o
-tl ft!
3
O O-
o
3
O
CTi
-------
-107-
Table 12. Determination of equilibration time for the
adsorption of the PCB Aroclor 1242 by Catlin.
Time
(hr)
0
2
4
6
8
*24
48
Solution
%AC cone. (mg/L)
0.220
0.020
0.018
0.017
0.017
94.31 0.013
0.00 0.013
* Equilibration time selected for an adsorption
isotherm.
-------
-108-
SECTION 14: CONSTRUCTION OF ADSORPTION ISOTHERMS (CURVES)
An adsorption isotherm or curve is a graphical representation that shows
the amount of solute adsorbed by an adsorbent as a function of the equilibrium
concentration of the solute. This relationship is quantitatively defined by
some type of partition function or adsorption isotherm equation that is
statistically applied to the adsorption data in order to generalize the
adsorption data.
In studies concerned with the adsorption of gases by solids, over 40
different 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 most commonly used and simplest of these adsorption
equations will be discussed here.
THE FREUNOLICH EQUATION
Probably the oldest known adsorption equation that has been widely used
for solid-liquid systems is the Freundlich adsorption equation, viz.,
x - K r1/11 rit;-i
m " KfC M
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 Kf and
!/n are constants.
Freundlich (1909) used this expression extensively, but it was first
proposed by van Bemmelen in 1888. The Freundlich equation was originally
proposed as an empirical expression without a theoretical foundation.
However, some investigators have referred to the Freundlich constant Kf as
being related to the capacity or affinity of the adsorbent and the exponential
term as an indicator of the intensity, or how the capacity of the adsorbent
varies with the equilibrium solute concentration (various references cited in
Suffet and McGuire, 1980).
-------
-109-
Other investigations have 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 and Halsey and Taylor, 1947; see also Hayward
and Trapnell, 1964, and Kipling, 1965). Zeldowitsch (1935) demonstrated that
the Freundlich equation could be explained in terms of a non-homogeneous
surface. Sips (1948) established in a rigorous fashion a general relationship
between 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 an often used expression, probably because of
its ease of application: it contains two constants that are both positive-
valued numbers that may be statistically solved when the expression is cast in
the logarithmic form, viz.,
log(x/m) = log Kf + 1/nlogC [16]
By taking the logarithms of both sides of eq. [15], the constants Kf and i/n
may be solved, via eq. [16], as a simple linear regression. The following
example is given to illustrate the application of the Freundlich equation.
From previous work, it was determined that the adsorption of arsenate by
kaolinite could be characterized by using a 1:10 soil:solution ratio (Section
9), and that the system reached a steady state after 24 hours. Using these
experimental conditions, 17 dilutions of a stock KH AsO solution were mixed
with an NBS rotary extractor with kaolinite for 24 hours. Table 13 contains
all the data needed to construct an isotherm, as well as the individual pH and
electrical conductivity (EC) of each solution as recommended at the end of
Sections 5 and 6.
-------
-110-
Table 13. Data reduction for arsenic adsorption at 25°C by a kaolinite clay
sample. The volume of solution was 200 mL
Initial" Equilibrium
concentration concentration
(mg/L) (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
* sample cal
X
m
. (
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
culation:
(Initial cone.
Adsorbent
weight
(g)
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
- equil . cone
Amount
adsorbed
(x/m) as yg/g
36*
64
84
98
98
96
123
153
160
221
'210
216
225
284
283
320
306
.) x volume of
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
solution
EC
(dS/m)
160
168
170
185
185
185
205
221
240
305
313
305
350
385
413
434
430
weight of adsorbent
4.89 mg/L - 1.20
mg/L) x 0.200L = n_n,fi , = ., ,n
20.42 g
As indicated earlier, the Freundlich equation may be solved when cast in
a logarithmic form that is equivalent to a simple linear regression, viz.,
= a +
[17]
-------
-111-
where log(x/m)^ = y^
log Kf = a
i/n = b
log C1 = x.j
The technique for solving a linear regression may be found in any intro-
ductory statistics text, and is also a common feature of most intermediate-
priced electronic calculators. (Note that linear regressions are sometimes
referred to as the line of best fit, or method of least squares.) However,
for the sake of completeness, the constants may be solved using
, n* (z log CH x log x/m - ) - (E log C, ) (E log x/m.)
b - I - ( - ] - ] - 1 - L.) [is]
n* (E(log C.)2) - (Z log C.)2
Kf '
Z log C.
where n* is the number of pairs of data points.
In this example (Table 13)
Log Kf = 1.536 i/n = 0.452
and thus,
£= 34.328 (As)0'452 [19]
where (As) is the equilibrium concentration of arsenic in solution
(mg/L).
Thus, eq. [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 eq. [19].
Eq. [19] will statistically predict solute-adsorbate partitioning over an
equilibrium concentration range of 0 to approximately 130 mg As/L. This
expression (eq. [19]), as well as any Freundlich expression, should never be
extrapolated beyond the experimental range used in its construction. In other
-------
-112-
words, eq. [19] should not be used to predict x/m at equilibrium concentra-
tions greater than 130 mg/L; to do so will require 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, and hence that any soil or clay would
have an unlimited capacity to retain chemicals dissolved in water. Not only
would an infinite capacity be thermodynamically inconsistent, 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 draw-
backs with respect to using the Freundlich equation: (1) it cannot be extra-
polated with confidence beyond the experimental range used in its construc-
tion, and (2) it will not yield a maximum capacity term which in many cases is
a convenient single-valued number that estimates the maximum amount of
adsorption beyond 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
x KLMC
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 Kj_ and
M are constants.
Langmuir (1918) derived an expression similar to eq. [20] to describe the
adsorption of gases on solids (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.
-------
-113-
Langmuir's original derivation was based on the premise that during the
adsorption of gases, a dynamic equilibrium is established where the rate of
condensation (adsorption) is equal to the rate of evaporation (desorption).
Derivations of the Langmuir and Langmuir-type equations for gas-solid inter-
actions are given elsewhere (Langmuir, 1918, Hayward and Trapnell, 1964; and
Ponec et al., 1974). Langmuir-type expressions for ion exchange reactions in
soils have also been derived (Sposito, 1979; and Elprince and Sposito, 1981).
The applicability of Langmuir-type equations in solid-liquid systems has
been a controversial topic in recent years (see Harter and Baker, 1977; Veith
and Sposito, 1977; Barrow, 1978; and Sposito, 1982). However, this contro-
versy is concerned with interpretations with respect to adsorption mechanisms
and energetics based on the results of applying Langmuir-type expressions
rather than the ability of the equation to simply describe the adsorption
data.
It appears to be the general consensus of several investigators that the
Langmuir constant (K^) is somehow related to the bonding energy between the
adsorbed ion and the adsorbent but the specific functional relationship is
uncertain. The constant M in eq. [20] is also generally accepted as the
adsorption maximum of the adsorbent with respect to the specific solute, and
it is interpreted as the maximum amount or concentration that an adsorbent can
retain.
Langmuir-type equations are often used because of their ease of
application. Like the Freundlich equation, it contains only two constants
that are both positive-valued numbers that may be statistically solved when
eq. [20] is cast in a linear form. Two linearized expressions are possible:
/^ i /%
x/m " M " ^ -"
-------
-114-
xjm - K[t + TT
The linearized form of eq. [21] is sometimes referred to as the "tradi-
tional linear Langmuir equation," while eq. [22] is called the "double-
reciprocal Langmuir equation." The latter is more suitable to situations
where 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 eqs.
[21] and [22] are equivalent to a simple linear regression, viz.,
YT = a + bxi [17]
whereas in the case of the traditional linear Langmuir equation
yn- = (C/x/nOi
a = 1/KLM
b = 1/M
xi ' ci
and in the case of the double-reciprocal form
a = 1/M
b = 1/KLM
The techniques for solving either eqs. [21] or [22] are the same as those
applied to solve the linear form of the Freundlich equation (eq. [16]). Using
the data set given in Table 13, applying the linear Langmuir-type equations
i
yields:
Traditional Linear Langmuir:
a = ^p = 0.0792 [23]
-------
-115-
and thus
and thus,
0.0028 [24]
x_ = 3.568 x 10"2 (353.856) C
m 1 + 3.568 x 10"Z (C)
Double-Reciprocal Plot:
a = . = 0.0050 [26]
b = ^ = 0.0297 [27]
x_ Q.1702 (198.098)C
m 1 + 0.1702(C)
Thus, eqs. [25] and [28] are also predictive expressions capable of describing
the adsorption of arsenic by kaolinite. The reader should also work through
these examples to verify the results.
-------
-116-
SECTION 15: SELECTION OF ADSORPTION EQUATIONS
In brief, there are three isotherm regressions to describe the example
data set given in Table 12. However, given the selection of different
equations, usually one of the equations will describe the results with the
greatest accuracy. In terms of simply fitting adsorption data, there appears
to be no clear consensus as to which equation (Freundlich or Langmuir-type)
generally is most reliable. Barrow (1978) objected to the application of
Langmuir-type expressions, but his objection 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 SO 2~ by soils. Polyzopoulos et al. (1984) compared four
adsorption equations in a study concerned with phosphate adsorption by soil.
They found that either a Langmuir-type or Freundlich expression could describe
the data with comparable success.
Generally the choice among equations is based on the coefficient of
determination (r2) obtained in a given case along with a given equation's
simplicity in form (Polyzopoulos et al., 1984). The Freundlich and Langmuir
equations both contain only two constants and are both easily solved.
The coefficient of determination (sometimes called the "goodness of fit")
is a measure of how closely the regression line fits that data, and may be
calculated using eq. [29]:
2 s (*i - y)2
p = ] [29]
E (y1 - y)2
where y. is the value of the dependent variable predicted by the regression,
y.,- is the value actually measured, and y is the arithmetic mean of all y^.
The value of r2 will always be between zero and one, inclusive. If all of the
points are close to the regression line or, in this case, if all of the
-------
-117-
adsorption data plot closely to the statistically-constructed adsorption
isotherm, the corresponding r2 will be close to one., The application of eqs.
[16], [21], and [22] to the data set given in Table 12 yielded dissimilar r2
values:
Freundlich 0.996
traditional linear Langmuir 0.954
double-reciprocal Langmuir 0.916
Using the coefficient of determination as a criterion, the Freundlich equation
best describes the adsorption data although the traditional linear Langmuir
expression would also yield satisfactory results. Inspection of Figure 41
clearly shows that the double-reciprocal linear Langmuir equation did not fit
the adsorption data well, while the traditional linear form tended to
overpredict adsorption in the upper part of the isotherm. Obviously the high
r2 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
such that r2 values are close to one is a major concern when constructing
adsorption isotherms. However, it is inevitable that in some cases a low r2
value may be obtained regardless of the equation used giving rise to concerns
that the adsorption constants have little meaning. Probably the simplest
statistical test that can be used in these situations is to use t-statistics
to examine whether the sample correlation coefficient (r) is significantly
different from a population correlation coefficient (p) where p = 0. This
test should be given in most introductory statistics text books and will not
be discussed here.
-------
-118-
320
Traditional linear
Langmuir equation
Double-reciprocal
linear Langmuir equation
ir2 =0916)
• SGS 985
30
Equilibrium arsenic concentration (mg/L)
—1—
100
—I—
120
Figure 41. The adsorption of arsenic by a kaolinite clay sample at 25°C as
described by the traditional linear Langmuir, double-reciprocal
Langmuir, and the Freundlich Equation. The mean pH of the soil-
solute suspensions was 7.8.
-------
-119-
SECTION 16: APPLICATION OF BATCH ADSORPTION DATA
This section was included to serve as a brief introduction to the applica-
tion of batch adsorption data in calculations of solute movement through
compacted landfill liners, particularly for estimating the minimum thickness of
liner required to prevent pollutant movement beyond a certain depth for a speci-
fied period of time. As leachate moves through a liner, the movement of chemical
solutes in the leachate may be retarded if adsorbed by the liner. We may define
R as the ratio of the velocity of the leachate to that of the solute, viz.,
R * vleachate/vsolute ™
The R term is called the retardation function or factor. When the solute is
not retained by the liner, R equals one; 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
R - 1 + C31]
9
where p^ is the dry bulk density of the liner
K^ is a distribution coefficient, and
9 is the volumetric water content of the liner.
The distribution coefficient is a parameter that describes the partitioning of
solutes between the leachate and the liner soil materials at equilibrium. The
distribution coefficent may be defined as
£ s Kd
where S is equal to x/m (the amount adsorbed per mass
of adsorbent), and
C is the equilibrium concentration of the solute.
In other words, eq [32] is the slope of an adsorption isotherm.
-------
-120-
In order to use eq. [31], a functional relationship for dS/dC must be
determined. The possible solutions range from simple assumptions to complex
numerical solutions. The simplest case is where the adsorption of the solute
conforms to a Freundlich equation (Section 14) isotherm where the Vn term is
unity, viz
£ " S ' KfC1/n = KfC [33]
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;
§ = Kf or Kd [34]
P^K .
hence, R » 1 + -2_£ [35]
9
In the case of a linear isotherm, the Freundlich constant (Kf) reduces to
the simple partition constant (Kd), a single-valued number that is used to
calculate solute-adsorbate partitioning at any equilibrium concentration 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 solute is a nonlinear function
(Vn * 1), the retardation factor is concentration-dependent:
dS _ d ,K ri/ru _ Kf ri/n - 1
" " ~3C (KfC ) ' ~ C ,
n K C1/n ' l
ph ^f^
hence, R(C) = 1 + ° T [37]
Eq. [37] is complicated by the fact that the numerical value of R will
depend on the concentration of the solute. Solute movement may be seriously
underestimated by assuming that a constant retardation factor is valid for a
-------
an
-121-
given system when dealing with nonlinear isotherms. Rao (1974) proposed
empirical estimation technique to solve eq. [37] which may be written as
K Co1/n " l
R - 1 + ^-r [38]
o
where R is a weighted-mean value and Co is the highest initial (before contact
with the adsorbent) concentration of the solute.
In a study concerned with pesticide adsorption by a soil sample, Davidson
et al. (1976) found that the error introduced by assuming linear adsorption
isotherms was not serious at low concentrations (< lOmg/L) but becomes
significant at higher concentrations. Van Genuchten et al. (1977) proposed an
alternative method for isotherm linearization that the reader may wish to
examine.
In order to demonstrate possible applications of these concepts, the
following examples are presented to illustrate how batch adsorption data are
used to estimate clay liner thickness.
In this hypothetical example, the metallic waste described in Appendix 8
is to be placed into a disposal basin that has been lined with Cecil clay loam
(refer to Appendix A). The soil has been graded, blended, and compacted and
has a saturated hydraulic conductivity of 10~7cm/sec. The major concern of
the company operating the disposal facility is the possible uncontrolled move-
ment of a leachate plume containing high concentrations of lead in solution.
In a preliminary analysis, this company conducted batch adsorption experiments
using a Pb(NO ) salt, and samples of the Cecil soil (Table 14). The question
that is posed is that for a 5-year operating life and a 30-year post-closure
period, what is the minimum thickness that the liner must be in order to
attenuate the lead from solution?
-------
-122-
There are several approaches that may be used to answer this question.
For each approach, the mean pore velocity of the leachate through the liner
must be calculated, and this may be done using Darcy's Law as
V = Ksati/ne [39]
where Ksat is the saturated hydraulic conductivity of the liner,
i is the hydraulic gradient (dH/dZ), and
ne is the effective (water conducting) porosity of the liner.
If we assume saturated conditions, subjected to steady state flow through
an isotropic liner over times t, and neglect the effects of dispersion and
diffusion, eq. [39] can be combined with eq. [31] to yield
Z = t Ksat i/R ne [40]
where Z is the estimated vertical distance of migration
of the solute (in cm),
t is time in seconds, and all other variables have
been defined previously.
Eq. [40] treats solute movement as a piston-flow problem; a chemically uniform
slug of leachate moving downward. This expression is simple, and may readily
be used to estimate the minimum thickness of a liner. To simplify its
application, it is often assumed 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
-------
-123-
Table 14. Lead adsorption data using a Pb(NO ) salt and the Cecil clay. The
volume of solution was 200 ml and the adsorbent weight was 10.18
grams.
Initial
Concentration
(mg/L)
2.07
5.11
5.11
6.22
7.28
10.2
10.2
12.4
14.6
14.6
Equilibrium
Concentration
(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 ug/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
4b
43
-------
-124-
1* S = 342(Pb)
Moreover, it is assumed that the liner has the following properties
ne = 0.09 cm3/cm3
9 = 0.36 cm3/cm3
pb = 1.7 g/cm3
Ksat = 1 x 10"7cm/sec
i = dH/dz = 1 cm/cm
and
35 years = 1.1038 x!Q9 seconds.
With these assumptions, the retardation factor becomes
and solving eq. [40] becomes
Z = (1.1038 x 109)(1 x 10-7) (1)/1619(0.09)
Z = 0.8 cm
Thus, based on this approach, the compacted liner would have to have a minimum
thickness of only about 1 cm to attenuate lead over a 35-year period. However,
while the application of a linear isotherm yields a reasonable coefficient of
determination (r2 = 0.95), inspection of Figure 42 indicates that this
approach over-estimates lead adsorption 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
-------
-125-
O>
31
c
3
O
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 through the origin,
-------
-126-
As a second level of refinement, the nonlinearity of the isotherm is
considered using eq [38] to estimate a weighted-mean retardation factor
(Davidson et al., 1976). An appropriate value for Co was determined from a
laboratory extract of the metallic waste sample (Appendix B) which suggested
that the maximum amount of lead that will initially come in contact with the
liner is approximately 15 mg/L Pb. Making use of eq. [38], a revised
retardation factor becomes
R = 1 + -
0736
= 348
and the minimum thickness, based on the weighted-mean retardation factor is
Z = (1.1038 x 109 )(1 x 1Q-7)(1)/348(0.09)
=3.5 cm
Consequently by considering the nonlinearlity of the isotherm, 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 considered. The first two estimates were based on lead adsorption from a
pure Pb(NO ) solution. Laboratory extracts of the waste also contained large
concentrations of zinc (Appendix B). The adsorption of lead from the extract
was found to be significantly less than that from the pure Pb(NO ) solution,
presumably due to competitive interactions between Zn2+ and Pb2"*" for
adsorption sites. The net effect is that lead is more mobile in the presence
of zinc. The adsorption of lead by Cecil from the laboratory extract of the
waste was found experimentally to be described by:
£= S = 70 (Pb)0-481
-------
-127-
If the minimum liner thickness is recalculated using these isotherm constants
and eqs. [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 using pure,
single-solute data may underestimate the minimum thickness of liners because
they fail to account for competitive interactions which may significantly
reduce adsorption.
At the next level in refining the estimated 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 concentration as a function of time may be generalized
(Ogata, 1970; Bear, 1972; Boast, 1973; and Freeze and Cherry, 1979) as
3C _ Q 3C2 - 9C _ % 35
3t Z 77 Z a! 9 3t
o*-
where C is the concentration of the solute,
Dz is the effective diffusion-dispersion coefficient
(distance 2/time) along the flow path z,
Vz is the mean convective flow velocity (distance/time) along the
flow path z,
pb is the bulk density (weight/volume) of the material,
9 is the volumetric water content (vol./vol.),
S is the amount of solute adsorbed per mass of adsorbent (x/m), and
t is time.
Eq. [41] may be rearranged as
R IT • Dz *4 - ~'t H
o£
where R is the retardation factor.
-------
-128-
The analytical solution to this second-order differential equation (Ogata,
1970) is given by
S- = i [erfc( Z " Vt* ) + exp({£)erfc( z * Vt* .)] [43]
Co 2 0.5 D 0.5
where C/Co is the ratio of the solute concentration at time t
and distance z to the initial solute concentration, Co;
erfc is the complementary error function,
V is the average linear pore water velocity (cm/sec),
Dz is the vertical dispersion coefficient (cm2/sec),
t* is the retarded time (actual time divided by the retardation
factor R or R),
and z is vertical distance of migration (cm),
furthermore,
Dz = aV + D* where o is the dispersivity (cm) and D* is
the diffusion coefficient in water (cm2/sec).
In the following examples, the three previous liner thickness estimations
were recalculated using eq. [43]. The only additional information needed to
conduct this analysis was to assign a value to the dispersivity. The
dispersivity (a) has been found to be scale dependent and is estimated to be
about 10% of the distance measurement of the analysis (Gelhar and Axness,
1981). A diffusion coefficient of Pb2+ in water of 2 x 10"? cm2/sec was used
in this analysis (Russel , 1961). The results are shown in Fig. 43 where the
relative concentration (C/Co) is shown as a function of distance of migration
after 35 years. Case A represents the first situation where the adsorption of
lead using a Pb(NO ) salt was assumed to be described by a linear isotherm.
Case B corresponds to the second calculation where a weighted-mean retardation
-------
-129-
o
CJ
o
c
-------
-130-
factor was used with the Pb(N03)2 solute-soil system.. Case C is based on the
adsorption of lead from the multicomponent waste extract coupled with the
corresponding weighted-mean retardation factor. In this example, taking into
account dispersion indicates that the lead may move further than that
predicted by an elementary piston-flow model (eq. [40]). The effects of
diffusion on the predicted migration distances were negligible (not shown).
There is an element of interpretation when evaluating graphs such as
Fig. 43 with respect to making liner estimations. A judgment that must be
made is to decide which C/Co ratio, for all practical considerations,
translates into the minimum significant concentration. In this hypothetical
example, the regulatory agency decided that a lead concentration of less than
0.05 mg/L (the U.S. drinking water standard for lead) would be an operational
definition of the compliance concentration.
Assuming that 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 refinement in this analysis yielding the most accurate liner thickness
estimation. After 35 years, the concentration of lead in solution would be
reduced to < 0.05 mg/L at a depth of 35 cm, based on these calculations.
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 engineering safety factors. The
application of batch adsorption data provides an estimation of boundary
conditions, i.e. the minimum thickness.
-------
-131-
• In summary, the minimum liner thickness for a hypothetical liner varied
from 1 cm to 35 cm, depending on the approach (Table 15). Liner thickness
estimations can be refined further if the adsorption data can also be
integrated with other information about the design and performance of a site
to more accurately predict retention and release of pollutants from an earthen
liner. This information would include 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.
-------
-132-
Table 15. Summary of approaches to estimate minimum liner thicknesses.
Flow model
Piston-Flow1
Piston-flow
Advection-2
dispersion
Advection-
dispersion
Piston-flow
Advection-
dispersion
Isotherm
Treatment
linear
nonlinear
linear
nonlinear
nonlinear
nonlinear
Solute
System
single-solute
% single-solute
single-solute
single-solute
mixture3
mixture
Minimum Li
Thickness
1
4
5
10
15
35
ner
(cm)
1 Represented by eq. [40].
2 Represented by eq. [43].
3 Laboratory extract.
-------
-133-
SECTION 17: LABORATORY PROCEDURES FOR GENERATION OF ADSORPTION DATA
This section of the TRO contains the procedures for the determination of
the soilrsolution ratio, equilibration time, and other parameters necessary
for the construction of adsorption isotherms. The rationale behind these
procedures is discussed in the previous sections of this report and should be
studied before attempting these procedures. Throughout this section,
references are made to other portions of the TRD which elucidate, through
discussion and/or example, topics which are relevant to the specific
procedural step. It is recommended that those sections be reviewed for
further clarification. The following flow diagram (Fig. 44) summarizes the
procedures and their inter-relationships.
17.1. SCOPE OF APPLICATION
17.1.1 The extent of adsorption of a chemical (solute) from solution
by an adsorbent (i.e., sediment, soil, clay) at equilibrium is
measured using these procedures.
17.1.2 These methods are applicable for the generation of adsorption
isotherms or curves for inorganic and organic (volatile and
nonvolatile) compounds to indicate how the extent of adsorption
varies with the equilibrium concentration of the solute.
17.1.3 Contingencies within these methods allow for the construction
of adsorption isotherms at various solute concentration ranges.
17.1.4 These methods can be used for constructing adsorption isotherms
to study the adsorption behavior of solutes in synthetic waste
solutions, laboratory extracts, or field leachates including
both aerobic and anaerobic systems.
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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 soil to
solution ratios
Determination of equilibration time
Construction of constant
soil to solution isotherm (CSI)
Nonionic solutes loinic solutes
Collection of adsorbent
Preparation of adsorbent
T
Air drying
Reduction of aggregates
Splitting and subsamplmg
Determination of
percent moisture
Construction of environmentally
conservative isotherm (ECI)
Application of adsorption data
Figure 44. Flow diagram for the procedures for the generation of batch
adsorption data.
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17.2. SUMMARY OF METHODS
The experimental design of these methods is based on a batch technique
as opposed to a column approach. 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 progressively decreasing initial solute concentrations with
a fixed mass of adsorbent in each reaction vessel. The second
technique is to mix 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 (see Section 12). The appropriate soil:solution
ratios and equilibration times are determined to maximize the accuracy
of the adsorption isotherm and to compliment analytical capabilities.
17.3. INTERFERENCES
17.3.1 When dealing with solutes of unknown stability, care must be
taken to determine if hydrolysis, photodegradation, microbial
degradation, oxidation-reduction (i.e., Cr3+ to Cr6+) or other
physicochemical processes are operating at a. significant rate
within the time frame of the procedure. The stability and
hence loss from solution may affect the outcome of this
procedure if. the aforementioned reactions are significant (see
Section 4). The compatibility of the method and the solute of
interest may be assessed by determining the differences between
the initial solute concentration and the final blank concentra-
tion of the solute. If this difference is greater than 3%,
then the adsorption data generated must be carefully evaluated
(see 17.8.5.11).
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17.4. TERMINOLOGY-DEFINITIONS
17.4.1 Solute - chemical species (e.g., ion, molecule, etc.) in
solution
17.4.2 Solute solutions shall be considered:
17.4.2.1 A solution of reagent water containing a known amount
of a solute derived from laboratory reagents.
17.4.2.2 A solution containing a variety of solutes extracted
from a material in a laboratory setting using methods
such as the ASTM-A or ASTM-B extraction procedures.1
17.4.2.3 A solution containing a variety of solutes collected
in a field situation representing a leachate or waste
effluent.
17.4.3 Adsorption - a physicochemical process whereby solutes are
retained by an adsorbent and are concentrated at solid-liquid
interfaces (see Section 1).
17.4.4 Adsorbate - chemical species adsorbed by an adsorbent
17.4.5 Adsorbent - substance that adsorbs the solute from solution.
17.5. LABORATORY EQUIPMENT
17.5.1.1 Agitation Equipment - the National Bureau of Stan-
dards extractor (rotating tumbler) or equivalent will
be exclusively used as the agitation apparatus (see
Section 8).
1 Neither the EPA Extraction Procedure (EP) or the proposed Toxicity
Characteristic Leaching Procedure (TCLP) are recommended. These procedures
were designed for waste classification, and were not intended to produce
solutions that mimic in situ leachates.
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17.5.1.2 Rotation Rate - When performing the procedures involving
inorganic, volatile, and nonvolatile organic compounds
the rotatary extractor will be operated at 29 ± 2 rpm.
17.5.1.3 Glove box or glove bags - When handling anaerobic
adsorbent-solute systems, it may be necessary to
conduct these procedures in air-tight enclosures
filled with an oxygen free inert gas (e.g. N , Ar) to
prevent or retard oxidation.
17.5.2 PHASE SEPARATION EQUIPMENT
17.5.2.1 Inorganic Compounds - A filtration apparatus made of
materials compatible with the solutions being filter-
ed and equipped with a 0.4b-micron pore size membrane
filter or a constant temperature centrifuge capable
of separating >0.1 micron particles will be used for
separation of the solid phase from the solid-liquid
suspensions.
17.5.2.2 If filtration is used, the affinity of the filtration
membrane for the solute must be evaluted. Failure to
do so may lead to erroneous results.
17.5.2.3 Organic Compounds - A constant temperature centrifuge
compatible with the reaction containers and capable of
separating >0.1 micron particles should be used for
separation of the suspension of phases when the solute
of interest is organic. The transfer of the organic
solute solutions from the reaction containers to cen-
trifuge containers is not an acceptable procedure due
to adsorption, volatilization, and other losses; the
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reaction container should be used as the centrifuga-
tion container. Filtration of organic solutions is
not a recommended practice (see Section 7).
17.5.2.4 Calculation of centrifugation time may be facilitated
by using eq. [1], viz.,
9nln(R,/R )
^ t [1]
2o) r (p-p)
where tt2 =
t = time (in minutes)
n = viscosity of water (8.95xlO"3g/sec-cm at 25°C)
r = particle radius (in cm)
PP = particle density (g/cm3)
p = density of solution (g/cm3)
rpm = revolutions per minute
Rj. = distance (in cm) from the center of the centri-
fuge rotor to the top of solution in centrifuge tube.
Rb = distance (in cm) from the center of the centri-
fuge rotor to bottom of the centrifuge tube.
To remove particles down to O.lym in radius and having
a particle density of 2.65 g/cm3 from a solution having
a density of 1 g/cm3 may be estimated using eq. [2], viz.,
3
t (min) = 3'71 x 210 ln(R /R ) [2]
(rpm)
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17.5.3 REACTION CONTAINERS
17.5.3.1 Inorganic Solutes - Containers compatible with the
rotary extractor should be used in conjunction with
inorganic solutes. The containers shall be composed
of materials that adsorb negligible amounts of the
solute. The containers must have a water-tight
closure made of chemically inert materials (i.e.,
polypropylene, teflon, etc.). The size of the con-
tainer should provide that the volume of the solid
and liquid will occupy about 80% to 90% of the container.
17.5.3.2 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 conjunction
with nonvolatile organic solutes. The container must
have a water-tight closure made of chemically inert
materials (i.e., teflon, plastic, etc.). The size of
the container must be compatible with the centrifuge
and provide that the volume of the solid and liquid
should occupy about 80% to 90% of the container.
17.5.3.3 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 in conjunction with volatile
organic solutes. The size of the serum bottle (125
ml) was found to be compatible with several types and
brands of centrifuges. This size provides sufficient
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volume such that the volume of the solid and liquid
should occupy 100% of the container (i.e., there
should be no head space).
17.5.3.4 As an advisory guide for hydrophobic solutes,
commonly available materials for containers can be
ranked starting with the material that is most inert
with respect to adsorption (T. C. Voice, written
communication).
Corex
Pyrex (not appreciably different from Corex)
Silanized serum bottles
Other types of glass
Stainless steel (unacceptable, >95% adsoprtion of PC8s)
Teflon (unacceptable, >95% adsorption of PCBs)
Plastic (unacceptable, >95% adsorption of PC8s)
17.5.4 REAGENTS
17.5.4.1 Reagent grade chemicals will be used in all experi-
ments. Unless otherwide indicated, it is intended
that all reagents shall conform to the specifications
of the American Chemical Society, where such specifi-
cations are available. Other grades may be used,
provided it is first ascertained that the reagent is
of sufficient purity to permit its use without
lessening the accuracy of the determination.
17.5.4.2 Unless otherwise indicated, references to water shall
be understood to mean type IV reagent water, as de-
fined in the Handbook for analytical quality control
in water and wastewater laboratories, EPA-600/4-79-019.
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17.5.5 SOLUTE SOLUTION, ADSORBENT AND TIME (PROCEDURAL) REQUIREMENTS
17.5.5.1 To construct adsorption isotherms for inorganic
solutes using these procedures, a minimum of 5 liters
of solute solution would be required based on the use
of 200-mL samples of the solute solution with 250-mL
reaction containers. Investigators using different
sized reaction containers should adjust the estimated
total volume of solution proportionately.
17.5.5.2 To construct adsorption isotherms using these pro-
cedures for organic solutes, approximately 9 liters
of solute solution 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 estimated
total volume of solution proportionately.
17.5.5.3 The mass of adsorbent required for completion of this
procedure will vary depending on the volume of
reaction containers, soil: solution ratios, etc.
Based on 250-mL reaction containers and the minimum
soilrsolution ratio of 1:4 (50 g adsorbent per 200 mL
of solute solution), about 2 Kg of adsorbent would be
required.
17.5.5.4 Approximately 5 to 9 days, excluding analytical time,
will be required to complete this procedure.
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17.6 EXPERIMENTAL CONSIDERATIONS RELEVANT TO VOLATILE ORGANIC SOLUTES
17.6.1 Stock solutions shall be prepared from pure standard materials
(either in the liquid or gaseous phases) or purchased as certi-
fied solutions. It is recommended that stock solutions be
prepared in methanol. The use of pipetts to transfer solutions
is not recommended but rather glass syringes should be used to
prevent losses due to volatilization. Because of the toxicity
of some volatile organic compounds, preparation and transfer of
solutions should be done in a fume hood, and a NIOSH/MESA
approved toxic gas respirator be used by the analyst.
17.6.2 PREPARATION OF STOCK VOLATILE SOLUTE SOLUTIONS
17.6.2.1 Place approximately 9 mL of methanol into a 10-mL
ground glass stoppered volumetric flask. Allow the
flask to stand unstoppered until all methanol wetted
surfaces have dried. Weigh the flask with the remain-
ing methanol to the nearest 0.01 mg, and immediately
add the test solute, using a syringe, 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 contacting the neck or sides of the
flask. Dilute to volume with methanol, stopper and mix
by inverting the flask several times.
17.6.2.2 Transfer the stock solution into a teflon-sealed
screw-cap vial. Store, with none or minimal head-
space, at approximately 4°C. All stock solutions
must be replaced after 1 month, or sooner if compari-
son with check standards indicate a loss of accuracy.
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17.6.2.3 Stabilize the temperature of the stock solution at
20°C before preparing secondary solutions.
17.6.2.4 Storage of all solutions must be done such that head-
space within the storage container is zero or
minimized.
17.6.3 PREPARATION OF SOLUTIONS FOR VOLATILE ORGANIC COMPOUND
ADSORPTION EXPERIMENTS
17.6.3.1 Place 990 mL of type IV water which has been boiled
and cooled to 20°C into each of a series of 1 L clean
amber glass bottles. (Generally eight solute
concentrations are required for completion of the
adsorption procedures.) Seal the bottles with open-
top screw caps fitted with teflon lined septa.
17.6.3.2 Inject known volumes of the stock solution prepared
in subsection 17.6.2 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.
17.6.3.3 Solutions stored in containers with headspace are not
stable and should be discarded 1 hour after prepar-
ation if not used in an experiment.
17.6.4 FILLING OF REACTION CONTAINERS
17.6.4.1 Upon immediate completion of subsection 17.6.3 pour
each solute solution carefully to minimize agitation
into pre-weighed reaction containers or those of
known volume (see 17.6.5) and containing known
amounts of adsorbent. Fill the containers such that
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no headspace is present. Gentle shaking of the
container to remove trapped air from the adsorbent
may be required. Place the teflon-faced septum and
aluminum seal on the container and invert to assure
no headspace is present. The volume of solution
added to the container is assumed to be that volume
determined in subsection I/.6.5.
17.6.5 DETERMINATION UF REACTION CONTAINER VOLUME
17.6.5.1 When transferring the solutions prepared in sub-
section 17.6.3 into the reaction containers, the
solutions should be poured quickly but gently into
containers of predetermined weight or volume.
17.6.5.2 Because the volume of solute solution is not measured
during transfer into the reaction containers, this
volume is determined indirectly.
17.6.5.3 The reaction containers to be used in 17.6.4 and each
containing the same mass or masses of adsorbent used
in 17.6.4 are used to determine the respective
container volume for each soil:solution ratio used.
Type IV water is pipetted into each container until
there is no headspace. The volumes of water added to
each of the containers is measured using a calibrated
syringe and the container volume is assumed to be
that of the volume of waters added.
17.6.5.4 Alternatively, the volume of solution added may be
determined by weighing the container containing .the
adsorbent before and after addition of the
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solution. The weight is converted to a volume from
knowledge of the density of the added solution.
17.6.6 Throughout all experiments the use of blanks is recommended to
determine effects of adsorption/desorption from containers as
well as losses due to volatilization. Refer to 17.8.5.11 for
further dicussion of the use of blanks.
17.6.7 For further information regarding the preparation of solutions
for volatile constituents or the analyses of these constituents
refer to U.S. EPA test methods 601 and 602 in Methods for
Organic Analysis of Municipal and Industrial Wastewater, EPA-
600/4-82-057.
17.7. PREPARATION OF MATERIALS TO BE USED AS ADSORBENTS
17.7.1 Samples of adsorbents such as soils, clays or sediments are
spread out on a flat surface in a layer, no more than 2 to 3 cm
deep. These samples will be allowed 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 (see Section 2). Anaerobic samples should be
processed 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 (i.e. N or Ar)
to prevent oxidation.
17.7.2 Weigh the entire sample after it has been air dried. Pass the
sample through a 2-mm screen sieve. Large aggregates shall be
crushed without grinding the sample using a clean mortar and a
rubber-tipped pestle. Those aggregates, such as pebbles and
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stones, that cannot be crushed shall be removed, composited,
and weighed.
17.7.3 Mix the sieved material until the sample is homogeneous. Use a
riffle splitter, or some other unbiased splitting procedure
(ASTM, Method C702-Reducing Field Samples of Aggregate to
Testing Size2, ASTM, Method 02013-72-Preparing Coal Samples for
Analysis3) to obtain subsamples of appropriate size.
17.7.4 The determination of the moisture content of the air-dried
sample shall be done using the ASTM-D2216, Laboratory
Determination of Moisture Content of Soils Method.1*
17.7.5 Determine the mass of the sample required for study corrected
for moisture content.
17.6.5.1 Determination of 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)
M = percent moisture
17.8 DETERMINATION OF SOIL:SOLUTION RATIOS FOR IONIC SOLUTES
17.8.1 A series of soilrsolution ratios ranging from 1:4 to 1:500
shall be tested and evaluated for the construction of
adsorption isotherms (see Sections 9 and 11).
2 Annual Book of ASTM Standards, Part 14.
3 Ibid., Part 26
•* Ibid., Part 19
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17.8.2 The soil isolation ratio is defined as the mass of adsorbent in
grams based on an oven-dry equivalent (subsection 17.7.5) per
volume in milliliters of solution.
17.8.3 It is recommended that the following soil:solution ratios be
used: 1:4, 1:10, 1:20, 1:40, 1:60, 1:100, 1:200, 1:500. The
need for soil isolation ratios greater than 1:500 is relatively
uncommon for most ionic solutes. In certain circumstances,
however, soil:solution ratios greater than 1:500 may be re-
quired to meet the criteria outlined in subsection 17.8.5.14;
in such cases, 1:1000, 1:2000, 1:5000 and 1:10,000 ratios are
suggested. The determination of a soil:solution ratio may be
an iterative process, whereby the eight ratios between 1:4 and
1:500 are tested before attempting the extremely "dilute"
systems (i.e., 1:1000, and higher). Using an iterative process
will reduce the amount of solute solution used, and will help
insure that enough solution will exist to complete the entire
procedure. Ratios less than 1:4 should not be used due to
limitations in mixing.
17.8.4 An example of how different soi1 :solution ratios are made is
given below for an air-dry moisture content of 3%:
Soil isolation Air-dry Oven-dry equivalent Volume of solution
ratio (g/mL) weight (grams) of adsorbent (grams) containing solute (mL)
200
200
200
200
200
200
200
200
1:4
1:10
1:20
1:40
1:60
1:100
1:200
1:500
51.5
20.6
10.3
5.15
3.43
2.06
1.03
0.412
50.0
20.0
10.0
5.00
3.33
2.00
1.00
0.400
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17.8.5 SOIL:SOLUTION PROCEDURE
17.8.5.1 Calculate the masses of adsorbent samples for the
various soil isolation ratios based on an oven-dry
equivalent weight (subsection 17.7.5) such that the
volume of adsorbent plus solution occupies 80% to 90%
of the container for nonvolatile solutes, and 100% of
the container for volatile solutes.
17.8.5.2 Weigh the samples of adsorbent to be used in the
soilrsolution series. If handling anaerobic
adsorbent-solute systems, steps 17.8.5.2 to 17.8.5.7
should be conducted in a glove box or bag before
placing the containers on the rotary extractor.
17.8.5.3 Place the weighed samples into clean, labeled
containers.
17.8.5.4 Pipet the solution containing the solutes (stock
solution) into each container containing the
adsorbent. The volume of solution should be
identical in all containers.
17.8.5.5 Pipet the stock solution into a container containing
no adsorbent. This sample will be the "blank." 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.
17.8.5.6 Close the bottles, insuring a water-tight seal, and
place on rotary tumbler for mixing.
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17.8.5.7 Collect and preserve an aliquot of the stock solution
to determine the concentration of the solute(s)
before contact with reaction containers, adsorbent,
phase separation materials, etc. (initial solute
concentration). The volume and preservation tech-
niques of the aliquot will vary depending on the
solute and analytical method.
17.8.5.8 Continuously agitate samples at 29 ± 2 rpm for 24 ±
0.5 hours, at room temperature (22 + 3°C).
17.8.5.9 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, pH, and any changes in the adsorbent or
solution.
17.8.5.10 Separate the solid and liquid phases of each sample,
using either centrifugation or filtration (subsection'
17.5.2). Determine the electrical conductivity of an
aliquot of each supernate (see Section 6). Collect
and preserve aliquots of each supernate of sufficient
volume to determine the solute concentration.
17.8.5.11 After analysis of all the solutions generated by the
soil:solution procedure, a comparison of the initial
solute concentration(s) and blank samples is neces-
sary to determine if there was adsorption or
desorption of the solute onto or from surfaces other
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than the adsorbent. If the difference between the
blank and initial solute concentrations as calculated
using 17.8.5.11.1 is greater than 3%, a correction in
the adsorption data must be made.
17.8.5.11.1 Determine the percent difference between
the initial concentration and the blank
solute concentration:
% 0 = —- - — x 100
Lo
where % D = percent difference
C0 = initial solute concentration
(i.e., mg/L, U9/L)
Cg = solute concentration (i.e.,
rng/L, u9/U in blank
solution.
The difference in concentration shall be
subtracted from all adsorption data,
excluding the stock or initial concen-
tration value. If % 0 is a negative
value, the solute concentration in the
blank was greater than the initial
solute concentration. This would imply
that there is a contamination problem.
Laboratory technique and/or cleaning
procedures should be examined.
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If % D is a positive value, then the
blank solute concentration was less than
the initial solute concentration. The
difference in concentration shall be
added to all adsorption data, excluding
the initial concentration value.
17.8.5.12 Using the initial solute concentration and the final
solute concentration for the various soilrsolution
ratios tested, the percent of solute adsorbed can be
calculated:
% A = .". x 100
Co
where: % A = percent adsorbed,
Co = initial solute concentration (i.e.,
my/L, 9/L etc.), and
C = solute concentration after contact with
the adsorbent.
17.8.5.13 Select a soil:solution ratio in which between 10 to
30% of the highest solute concentration was adsorbed.
This soil:solution ratio will be used to determine
the equilibration time (subsection 17.10), and to
generate data for the construction of a constant
soil:solution isotherm (CSI). Often, several
soil:solution ratios will generate solute adsorp-
tion between 10 and 30%. The selection of a specific
soil to solution ratio is the investigators'
prerogative, with the limitation that it should be
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one of those listed in subsection 17.8.3 (see
Sections 9 and 11).
17.9 DETERMINATION OF SOIL:SOLUTION RATIOS FOR NONIONIC SOLUTES
17.9.1 While finding a suitable soil-.solution ratio for ionic solutes
must be done empirically, a useful soil:solution ratio for
nonionic solutes (hydrophobic organics) may be calculated if the
organic carbon content of the adsorbent and the water solubility
of the solute are known. The equations and their derivations
for determining the soilrsolution ratios for non-ionic solutes
are given in Section 10.
17.9.2 The soi1tsolution ratios listed in subsection 17.8.3 most
closelymatching the calculated soil :§olution ratio shall be
used throughout this procedure. If the calculated ratio is in
the middle of two ratios listed in subsection 17.8.3, the lower
ratio (greatest mass of absorbent per milliliter of solute) is
recommended to obtain the highest precision and accuracy.
17.10 DETERMINATION OF EQUILIBRATION TIME
17.10.1 Use the soil:solution ratio determined in subsection 17.8.5.13
for inorganic solutes, and subsection 17.9 for hydrophobic
organic solutes for the equilibration time determination(s).
17.10.2 A minimum of four agitation times is recommended to determine
the equilibration time. Recommended times are 1, 24, 48, and
72 hours, and represent the amount of time the solution and
adsorbent are in contact.
17.10.3 Weigh the adsorbent on a oven-dry basis (subsection 17.7.5)
and place into clean, labeled containers. If handling
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anaerobic systems, steps 17.10.3 and 17.10.8 should be
conducted in a glove box or bag, before placing the containers
on the rotary extractor.
17.10.4 Pipet the solute solution into the various containers at the
times designated in 17.10.2. Immediately cap the container
and place on rotary extractor at 29 ± 2 rpm at room
temperature (22 ± 3°C).
17.10.5 Pipet the solute solution into a container containing no ad-
sorbent; this is the blank and should be agitated for 72
hours.
17.10.6 Collect and preserve an aliquot of the stock solute solution.
17.10.7 Remove the containers at the appropriate times from the rotary
extractor and record the solution temperature, pH and any
changes in the adsorbent or solution. If handling anaerobic
suspensions, return the containers to a glove box or glove bag
before opening the containers.
17.10.8 Separate the solid and liquid phases using either centrifuga-
tion or filtration (subsection 17.5.2). Determine the
electrical conductivity of an aliquot of each supernate.
Collect and preserve aliquots of each supernate of sufficient
volume for the solute concentration determinations.
17.10.9 Determine the rate of change in the solute concentrations at
the various times by
(C! - C2)
%AC = r x 100
4
where %AC = percent change,
C^ = concentration of the solute at time t, and
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C2 = concentration of the solute after 1, 24, 48,
or 72 hours.
17.10.10 The equilibrium time is defined as the minimum amount of time
needed to establish a rate of change of the solute concentra-
tion equal to or less than 5% per a 24-hour interval (see
Section 13).
17.11 CONSTRUCTION OF THE ENVIRONMENTALLY CONSERVATIVE ISOTHERM (ECI) FOR
IONIC AND NONIONIC SOLUTES
17.11.1 The construction of an ECI requires that the soil:solution
(subsection 17.8 and 17.9) and equilibrium (subsection 17.10)
procedures be completed.
17.11.2 If the equilibrium time as determined by 17.10.9 is equal to
or less than 24 hours, the data obtained from the
soil:solution procedure can be used in construction -of an
ECI. However, if the equilibrium time is greater than 24
hours, the soil:solution ratio determination procedure must
be redone at the equilibrium time determined by 17.10.9.
Refer to Section 12 for discussion of the advantages and
limitations of the ECI.
17.11.3 Since subsection 17.9 yields a single soi1rsolution ratio for
nonionic solutes, additional ratios should be selected which
bracket the calculated ratio. It is recommended that a
minimum of eight soil :solution ratios be used, and that these
ratios be selected from those listed in subsection 17.8.3.
These ratios will be evaluated as outlined in 17.8.5. When
volatile solutes are under study, refer to subsection 17.6
for experimental considerations.
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17.11.4 A minimum of five data points should be used to construct an
ECI. Soil rsolution ratios resulting in less than 10% of the
solute being adsorbed should not be used in construction of
the ECI (refer to Section 12 for justification). It is
recommended that as much data as possible generated by the
soil rsolution ratios prescribed in subsection 17.8 and
meeting the above criteria be used in construction of an
ECI. If less than five of the soil rsolution ratio data
generated in subsection 17.8 meet the criteria listed above,
variations in the recommended soil rsolution ratios can be
used in generating additional data.
17.11.5 Using the data generated by the soil -.solution procedure, the
amount of solute adsorbed per mass of adsorbent can be
calculated.
17.11.5.1 Determination of the amount of solute adsorbed per
mass of adsorbent:
where x/m = amount of solute adsorbed per unit mass of
adsorbent,
m = mass of adsorbent in grams added to reaction
container,
C0 = 'initial solute concentration before exposure
to adsorbent,
C = solute concentration after exposure to
adsorbent at equilibrium, and
V = volume of solute solution added to reaction
container.
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17.11.6 Construction of an ECI requires that: 1) the x/m value for
each soil:solution ratio meets the criteria in 17.11.2, and
2) the corresponding equilibrium concentration value (C) of
the solute.
17.11.7 CONSTRUCTION OF AN ECI
17.11.7.1 Using linear graph paper, plot the equilibrium
concentration (C) on the coordinate (x-axis) and
the corresponding x/m value as the dependent
variable (y-axis). Refer to Section 12 for an
example.
17.11.7.2 Fit an adsorption equation, using either the
Freundlich or Langmuir-type equations to the data
plotted in 17.11.5.1.
17.11.7.3 The linear expression of the Freundlich equation is:
log (x/m) = log Kf + i/n log C
where x/m = amount of solute adsorbed per unit
mass of adsorbent,
Kf = a constant,
i/n = a constant (sometimes written as N), 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 17.11.5.1, where the
intercept equals log Kf and the slope equals
i/n. An example using the Freundlich equation is
given in Section 14.
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17.11.7.4 A linear expression of the Langmuir-type equation is:
1 . C
xm
where x/m = amount of solute adsorbed per unit
mass of adsorbent,
KL = a constant,
M = a constant, and
C = equilibrium concentration of the
solute after exposure to adsorbent.
A linear regression can be used to fit a curve
through the data plotted in 17.11.7.1, where the
intercept equals i/KLM and the slope equals i/M.
Examples using Langmuir-type equations are given
in Section 14.
17.11.7.5 From the application of the Freundlich or Langmui-
type equations, a coefficient of determination
(r2) can be determined that will indicate the
statistical accuracy of the regression used to
describe the adsorption data. Examples are given
in Section 15.
17.11.7.6 The equation resulting in the coefficient of
determination value closest to 1.0 is usually used
to generate a curve through the data plotted in.
17.11.7.1.
17.11.7.7 The data plotted in 17.11.7.1 and the curve of
"best fit," 17.11.7.6, represent an ECI.
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17.11.7.8 The following information should be reported with
the ECI: 1) temperature at which the tests were
conducted, 2) pH and EC of all solute solutions,
3) concentrations of stock (CQ) and blank (Cg)
solute solutions and the factor, if any, used to
correct data, 4) the soil .-solution ratios, their
corresponding solute solution volume and adsorbent
mass, the initial (C0) and final (C) solute
concentration and the percent of solute adsorbed,
5) the %AC for each equilibration time, 6) the
equation for the line of "best fit" and the
corresponding r2 value, and 7) a complete
description of the adsorbent.
17.12 CONSTRUCTION OF THE CONSTANT SOILrSOLUTION RATIO ISOTHERM (CSI) FOR
IONIC SOLUTES.
17.12.1 Unlike the ECI, where the initial concentration of the solute
is constant and the mass of adsorbent varies in each contain-
er, the CSI requires that the initial solute concentration
varies, and that the mass of adsorbent remains constant.
Refer to Section 12 for advantages and limitations of both
techniques.
17.12.2 The soil:solution ratio (% A between 10 to 30%) and the
equilibrium time (%AC <5% per a 24-hour interval), determined
in subsections 17.8 and 17.10 respectively, are recommended
to be used in the construction of a CSI.
17.12.3 Weigh the adsorbent (mass prescribed by the soil:solution
ratio) into clean, labeled containers. If handling anaerobic
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-159-
adsorbent-solute systems, steps 17.12.3 to 17.12.5 should be
conducted in a glove box or glove bags.
17.12.4 Make a series of approximately eight dilutions of the stock
solute solution (albeit a laboratory extract or field
leachate sample) such that there is a progressive decrease in
solute concentration. The most dilute solution should
contain the solute in a sufficient concentration so that the
amount of solute remaining in solution is above analytical
detection limits after contact 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.
17.12.4.1 The dilution of complex solutions may cause
changes in pH, redox potentials, etc. with the
subsequent precipitation of the solute(s) (see
Section 11). Effort should be made to limit such
reactions, and where such actions are not possible
or are felt inappropriate, the procedures in
subsection 17.11 may be used for determination of
adsorption isotherms.
17.12.5 Immediately after the dilutions of the stock solute solution,
pi pet the diluted solutions into containers containing the
adsorbent. Each solution should have a corresponding con-
tainer with the volume of solution in all containers being
equal.
17.12.6 Place the containers on the rotary extractor at 29 ± 2 rpm at
room temperature (22 + 3°C). Agitate for the time determined
-------
-160-
in subsection 17.10. 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 waste waters).
17.12.7 After the agitation period, remove the containers from the
rotary extractor and open. If the suspensions are anaerobic,
return the containers to a glove box or bag, then open the
containers. Observe and record the solution temperature, pH,
and any changes in the adsorbent or solution.
17.12.8 Separate the solid and liquid phases using either centrifuga-
tion or filtration (subsection 17.5.2). Determine the
electrical conductivity of an aliquot of each supernate.
Collect and preserve aliquots of each supernate of sufficient
volume for the solute concentration determinations.
17.12.9 Determine the solute concentration in the stock solution, the
dilute solutions before (CQ in equation 17.11.5.1) and after
(C in equation 17.11.5.1) exposure to the adsorbent. If
significant differences in the blank solutions (subsection
17.8.5.11) were ascertained, the adsorption data must be
corrected.
17.12.10 Using the data generated where the various solute concentra-
tions were exposed to the same mass of adsorbent, the amount
of solute adsorbed per mass of adsorbent (x/m) can be cal-
culated. Refer to equation 17.11.5.1 for calculation of x/m.
5 American Public Health Association. 1975 (14th edition) American Public
Health Association, Washington, D.C., p. 38-45.
-------
-161-
17.12.11 Construction of the CSI requires: 1) a x/m value for each
solute concentration, and 2) the corresponding equilibrium
concentration value (C) of the solute.
17.12.12 Construction of the CSI shall follow the same procedure and
reporting requirements as the ECI. Refer to subsection
17.11.5 for directions on construction of the ECI/CSI.
-------
-162-
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-------
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1980. Sorption of benzidine by sediments and soils: Soil Science, v.
129, p. 277-281.
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-173-
APPENOIX 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. A summary of the adsorbents
and their sample locations is given in Table A-l, and a summary of relevant
physicochemical characteristics is given in Table A-2, which also includes
mineralogical descriptions. The chemical composition of the materials
including major elements (Table A-3) and trace constituents (Table A-4) has
also been characterized.
The eleven clays and soils represented a wide range in physicochemical
properties and characteristics. The Catlin soil is a dark prairie soil
(Mollisol), containing a relatively high organic matter content in the surface
horizon. It is an important soil agriculturally, and the clay-size fraction
is dominated by illite. Mollisols dominate the Great Plains states. 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 is an Illinoian-age deposit and is fairly representative of
midwestern glacial tills. It is a sandy till, gray in color, calcareous where
unweathered, and the dominate clay is illite. At the sampling site (Table
A-l), the Sangamon Paleosol was a buried soil that had formed in the Vandalia
till, and was overlain by glacial loess. The Sangamon Paleosol, Vandalia
(ablation phase), altered (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 Wilsonville,
Illinois.
-------
-174-
The soil sample designated as EPA-14 was used by Hassett et al. (1980a,
1980b, 1981) and Zierath et al. (1980) in studies concerned with the
adsorption of hydrophobic solutes. The Cecil clay sample from South Carolina
was used by Roy et al. (1986) in a study concerned with the adsorption of
anionic mixtures. The kaolinite and illite clay samples have also been used
in previous studies (see Griffin et al., 1976; Griffin and Shimp, 1976, 1978,
and Frost and Griffin, 1977).
REFERENCES
Frost, R. R., and R. A. Griffin. 1977. Effect of pH on adsorption of copper,
zinc, and cadmium from landfill leachate by clay minerals: Journal of
Environmental Science and Health, v. A12 (4 and 5), p. 139-156.
Griffin, R. A., and N. F. Shimp. 1976. Effect of pH on exchange-adsorption
or precipitation of lead from landfill leachates by clay minerals:
Environmental Science and Technology, v. 10, p. 1256-1261.
Griffin, R. A., et al. 1976. Attenuation of pollutants in municipal landfill
leachate by clay minerals: Part 1 - Column leaching and field
verification: Environmental Geology Notes, no. 78, Illinois State
Geological Survey, 34 p.
Griffin, R. A., and N. F. Shimp. 1978. Attenuation of pollutants in munici-
pal landfill leachate by clay minerals: U.S. Environmental Protection
Agency, Cincinnati, Ohio, EPA-600/2-78-157 (NTIS: PB 287-140).
Hassett J. J., J. C. Means, W. L. Banwart, and S. G. Wood. 1980a. Sorption
properties of sediments and energy-related pollutants: U.S. Environ-
mental Protection Agency, EPA-600/3-80-041 (NTIS: PB 80-189-574).
Hassett, J. J., J. C. Means, W. L. Banwart, S. G. Woods, S. A. Khan, and A.
Khan. 1980b. Sorption of dibenzothiophene by soils and sediments:
Journal of Environmental Quality, v. 9, p. 184-186.
Hassett, J. J., W. L. Banwart, S. G. Wood, and J. C. Means. 1981. Sorption
of a-Naphthol: Implications concerning the limits of hydrophobic
sorption: Soil Science Society of America Journal, v. 45, p. 38-42.
Roy, W. R., J. J. Hassett, and R. A. Griffin. 1986. Competitive interactions
of phosphate and molybdate on arsenate adsorption: Soil Science (in
press).
Zierath, 0. L., J. J. Hassett, W. L. Banwart, S. G. Wood, and J. C. Means.
1980. Sorption of benzidine by sediments and soils: Soil Science, v.
129, p. 277-281.
-------
-175-
Table A-l. 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 Til 1 Member
ablation
altered
unaltered-
Champaign, Illinois A
Spartanburg, South Carolina B t
Cecil, Georgia Ap
Ceredo, West Virginia A
Elizabeth, Illinois
Pike County, Illinois
Macoupin County, Illinois Bt
(near Sawyerville)
Tifton, Georgia Ap
Glasford Formation
Macoupin County, Illinois B
(near Sawyerville)
(near Sawyerville) C
(near Eagerville) C
Typic Argiudoll
Typic Hapludult
Typic Hapludult
unknown
unknown
Plinthic Paleudult
-------
Table A-2. Summary of selected physicochemical characteristics of clays and soils used in the development of TRD.
adsorbent
Catlin silt loam
Cecil clay
Cecil clay loam
EPA-14
Illite clay
Kaolinite clay
Sangamon paleosol
Tifton loamy sand
Vandal i a Till
altered
unaltered
ablation phase
pH(l:l)a
6.1
4.5
4.6 '
4.5
7.9
8.1
6.1
4.7
7.4
7.5
6.4
sand
i
11
31
32
2
0
0
45
85
45
45
56
silt
% •
69
12
17
63
0
0
25
9
38
40
21
clay
21
58
51
34
100
100
30
5
17
15
23
organic
carbon
4.04
0.34
NDd
0.48
1.81
0.51
0.10
NO
0.18
0.34
0.10
CECb surface are
meq/lOOg (m^/g)
18.1
3.7
3.8
18.9
20.5
15.1
16.7
1.9
6.6
4.9
10.5
14.8
36.9
29.7
145e
34.2
22.9
1.7
7.3
5.6
10.6
ac illite
55-67
<5
5-6
13
70
8
33-36
0
71-77
75-82
32-58
Clay analysis
kaolinite expandable
-----
5-15
68-92
79-92
37
0
87
7-14
73-96
3-10
4-19
2-6
24-30
3-32
2-16
14
0
5
50-60
4-27
18-19
6-9
32-39
Other
:s clay-sized
minerals
chlorite
gibbsite, goethite
hematite
goethite, hematite
gibbsite
30% mixed layer
quartz
_
goethite
_
_
goethite
en
i
a pH of a 1:1 soil:water suspension
b Catlin exchange capacity
c Surface area by N adsorption using BET method
d no data available
e surface area by ethylene glycol (from Hassett et al., 1981)
-------
Table A-3. Summary of major element composition (in oxide form) of clay and soils used
in the development of the TRD (percent).
adsorbent
Catlin silt loarn
Cecil clay
Cecil clay loam
EPA-14a
Illite clayb
Kaolinite clay
Sangamon paleosol
Tifton loamy sand
Vandalia
altered
unaltered
ablation phase
sio2
72.5
44.8
66.2
ND
48.5
46.6
82.7
96.4
61.3
59.1
83.5
Ti02
0.73
1.15
0.94
ND
0.67
2.45
0.43
0.27
0.33
0.33
0.35
Al 0
2 3
10.8
30.0
20.4
ND
24.6
41.9
10.2
1.3
6.7
6.5
7.9
Fe
2
4.
10.
6.
6.
4.
0.
2.
0.
2.
2.
2.
°3
0
4
8
99
11
94
9
5
1
4
5
CaO
0
<0
<0
0
3
0
0
<0
9
9
0
.9
.1 .
.1
.71
.27
.57
.50
.1
.4
.7
.6
MyO
0.71
0.19
0.19
ND
1.73
0.30
0.65
0.03
4.66
4.95
0.54
Na 0
2
0.84
0.07
0.04
0.21
0.14
0.13
0.45
0.01
0.52
0.49
0.56
V P2°5
2
0
0
2
10
1
1
0
2
2
1
.14 0.1
.54 0.1
.65 <0.1
.94 ND
.23 ND
.49 ND
.49 <0.1
.05 <0.1
.03 <0.1
.08 <0.1
.98 <0.1
a (data from Hassett et al., 1981)
b (data from Griffin and Shimp, 1978)
-------
Table A-4. Summary of trace element concentrations in the clays and soils used in the development of the TRO (mg/kg).
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.l.
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.l.
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-14a
10
_
450
-
-
-
87
.
11
.
8
1
23
8
46
_
-
216
_
-
200
6
16
2
-
<80
1
1
15
_
-
3
-
lllite c.b Kaolinite c.b Sangamon
_c
44
-
-
-
19
-
-
-
-
-
-
-
-
-
.
-
<390
-
94
-
-
-
.
-
-
-
.
-
.
-
-
38
6
46 230
500
2
<1
<3 " <1.3
38
52
15
11
3
1
12
-
28
27
0.5
29 970
<9
46 19
79
0.6
8
<1
5
55
-
-
5
<3
-
2
20 71
Tifton l.s.
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
Vandal i a Till
altered
6
172
359
2
<7
<1.3
25
39
8
15
3
1
9
-
19
23
0.3
388
<9
<14
68
0.4
6
<2
3
75
-
-
4
<3
-
2
42
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
a data from Hassett et al., 1981
b data from Griffin and Shimp, 1978
c no data available
-------
-179-
APPENDIX B. COMPOSITION OF THE METALLIC HASTE EXTRACT
USED IN THIS STUDY AND ASSOCIATED ADSORPTION ISOTHERMS
In order to test and refine the basic batch adsorption procedure for
ionic solutes, a metallic waste sample was collected on Nov. 1, 1984 from the
Sandoval Zinc Co., near Sandoval, Illinois. Grab samples were taken from a
dry slurry lagoon that was used to store metallic scrubber sludges (refer to
Gibb and Cartwright, 1982 for additional information). Samples were taken
from the surface and from a depth of about one meter. The samples were
composited, then 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 B-l) 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 since these two
soils represented widely different physicochemical materials.
The adsorption of barium, lead, and zinc from the extract by the two
soils was investigated and the results were incorporated into the TRD. The
adsorption isotherms are shown in Figs. B-l through B-3, and were generated
using the procedures described in the text.
REFERENCES
Gibb, J. P. and K. Cartwright. 1982. Retention of zinc, cadmium, copper, and
lead by geologic materials: Cooperative Groundwater Report 9, Illinois
State Water Survey - State Geological Survey, Champaign, IL 61820, 113 p.
American Society for Testing and Materials. 1979. Proposed methods for
leaching of waste materials: Annual Book of ASTM-A Standards, Part 31,
Water, Philadelphia, PA, p. 1258-1261.
-------
-ISO-
Table B-l. Chemical constituent concentrations obtained by the ASTM-A (water
shake extraction) performed on the Sandoval zinc slurry
(concentrations in mg/L)
pH
EC(dS/m)
Al
As
8
Ba
Be
Ca
Cd
Co
Cr
Cu
Fe
' K
Mg
Mn
Mo
Na
Ni
P
Pb
Sb
Se
Si
Sn
V
In
6.27
0.17
<0.05
<0.08
<0.08
2.25
-------
-181-
30-
20-
1
o
<
./
10-
I
0.5
I
1.0
I
1 5
I
2.0
Equilibrium barium concentration (mg/L)
Figure B-l.
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.
-------
-182-
Equilibnum lead concentration Img/L)
Figure 8-2. 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.
-------
-183-
4 -
100
T
200 300
Equilibrium zinc concentration (mg/L)
400
T
500
Figure B-3.
Zinc adsorption isotherms at 24°C of two soils using the metallic
waste extract. The average ,pH of the Sangamon Paleosol
suspensions was 5.9, and pH 4.3 for the Cecil clay.
------- |