&EPA
United States
Environmental Protection
Agency
Municipal Environmental Research EPA 600/2 78 157
Laboratory August 1978
Cincinnati OH 45268
Research and Development
Attenuation
of Pollutants
in Municipal Landfill
Leachate
by Clay Minerals
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RESEARCH REPORTING SERIES
Research reports of the Office of Research and Development, U.S. Environmental
Protection Agency, have been grouped into nine series. These nine broad cate-
gories were established to facilitate further development and application of en-
vironmental technology. Elimination of traditional grouping was consciously
planned to foster technology transfer and a maximum interface in related fields.
The nine series are:
1. Environmental Health Effects Research
2. Environmental Protection Technology
3. Ecological Research
4. Environmental Monitoring
5. Socioeconomic Environmental Studies
6. Scientific and Technical Assessment Reports (STAR)
7. Interagency Energy-Environment Research and Development
8. "Special" Reports
9. Miscellaneous Reports
This report has been assigned to the ENVIRONMENTAL PROTECTION TECH-
NOLOGY series. This series describes research performed to develop and dem-
onstrate instrumentation, equipment, and methodology to repair or prevent en-
vironmental degradation from point and non-point sources of pollution. This work
provides the new or improved technology required for the control and treatment
of pollution sources to meet environmental quality standards.
This document is available to the public through the National Technical Informa-
tion Service, Springfield, Virginia 22161.
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EPA-600/2-78-157
August 1978
ATTENUATION OF POLLUTANTS IN MUNICIPAL
LANDFILL LEACHATE BY CLAY MINERALS
by
R. A. Griffin and N. F. Shimp
Illinois State Geological Survey
Natural Resources Bldg.
Urbana, Illinois 61801
Contract No. 68-03-0211
Project Officer
Mike H. Roulier
Solid and Hazardous Waste Research Division
Municipal Environmental Research Laboratory
Cincinnati, Ohio 45268
MUNICIPAL ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
CINCINNATI, OHIO 45268
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DISCLAIMER
This report has been reviewed by the Municipal Environmental Research
Laboratory, U. S. Environmental Protection Agency, and approved for publi-
cation. Approval does not signify that the contents necessarily reflect the
views and policies of the U. S. Environmental Protection Agency, nor does
mention of trade names or commercial products constitute endorsement or
recommendation for use.
ii
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FOREWORD
The Environmental Protection Agency was created because of increasing
public and government 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 that environment and the interplay between its components
require a concentrated and integrated attack on the problem.
Research and development, that necessary first step in solving a prob-
lem, involves defining the problem, measuring its impact, and searching for
solutions. The Municipal Environmental Research Laboratory develops new and
improved technology and systems to prevent, treat, and manage wastewater and
solid and hazardous waste pollutant discharges from municipal and community
sources, to preserve and treat public drinking water supplies, and to min-
imize the adverse economic, social, health, and aesthetic effects of pol-
lution. This publication is one of the products of that research; it is a
most vital communications link between the researcher and the community.
This report presents results from laboratory investigation of the
capacity of clay minerals to remove pollutants from municipal landfill
leachates. These results are applicable to the design of clay mineral liners
for sanitary landfills and to the land disposal of municipal and hazardous
wastes.
Francis T. Mayo, Director
Municipal Environmental Research
Laboratory
iii
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ABSTRACT
The first part of this project was a laboratory column study to evaluate
the potential of mixtures of sand and calcium-saturated clay minerals for
attenuating and preventing pollution of water resources by pollutants in
municipal solid waste landfill leachate. Chloride, Na, and water-soluble
organic compounds (COD) were relatively unattenuated by passage through the
clay-sand columns; K, NH^, Mg, Si, and Fe were moderately attenuated; and
heavy metals — such as Pb» Cd, Hg, and Zn — were strongly attenuated by even
small amounts of clay. Concentrations of Ca, B, and Mn in the column effluents
increased markedly over the original leachate concentrations.
Montmorillonite was found to have the highest attenuation capability,
followed by illite and then kaolinite. Precipitation, with resultant accumula-
tion in the surface layers of the columns, was found to be the principal
attenuation mechanism for the heavy metals Pb, Cd, Hg, and Zn. The cation
exchange capacity of the clay minerals was concluded to be the dominant
attenuation mechanism responsible for removal of other substances from the
leachate.
The second part of the project involved batch studies of adsorption of Cr,
Cu, Pb, Cd, Hg, and Zn by montmorillonite and kaolinite from water solutions
and from landfill leachate. The adsorption in leachate proved to be 50 to 90%
lower in most cases than the clays' adsorption capacity for the metal ions in
pure aqueous solutions. The adsorption of the cations Cr(III), Cu, Pb, Cd^
Hg, and Zn increased with increasing pH while the anions Cr(VI), As, and Se
decreased with increasing pH. At pH values greater than about 5.3, precipi-
tation of the heavy metal cations was found to be an important attenuation
mechanism while adsorption was the principal mechanism for the anions over
the entire pH range studied.
Pollutant adsorption by clay minerals (and hence the mobility of pollut-
ants in clays and clay soils) was significantly affected by other, non-hazard-
ous solutes in the leachates. This effect was so pronounced that information
on movement of pollutants in one landfill leachate cannot be directly applied
to predicting the movement of the same pollutants present at the same con-
centrations in a different landfill leachate.
To evaluate the relative pollution hazard for municipal leachates, a
ranking system was developed. Results of the study are applicable to the use
of clay minerals as liners for sanitary landfills and to the land disposal of
municipal and hazardous wastes.
This report was submitted in fulfillment of Contract Number 68-03-0211,
by the Illinois State Geological Survey, under the partial sponsorship of the
Environmental Protection Agency. Work was completed as of August 1975.
iv
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CONTENTS
Abstract iv
List of Figures vl
List of Tables lx
Acknowledgments x
1. Introduction 1
2. Conclusions 4
3. Recommendations 7
4. Experimental 9
5. Column Leaching Study 16
6. Confirmation of Laboratory Column Studies by Comparison
with Field Data 34
7. Effect of pH on Exchange-Adsorprtion or Precipitation of
Lead from Municipal Leachates by Clay Minerals 44
8. Effect of pH on Copper, Zinc, and Cadmium Removal from
Deionized Water and Landfill Leachate by Clay Minerals .... 61
9. Effect of pH on Adsorption of Chromium from Landfill-
Leachate by Clay Minerals 83
10. Effect of pH on Adsorption of Arsenic and Selenium from
Landfill-Leachate by Clay Minerals 102
11. Mercury Removal from Municipal Landfill-Leachate by Clay
Minerals 113
12. Summary of Adsorption Studies 120
13. Application of the Results to the Problem of Landfill Design . . 123
References . . 128
Appendices
A. List of Publications . . 134
B. Procedures Used in Chemical and Physical Characterization of
Clay Minerals and Liquid Samples 136
Clay Mineral Characterization 136
Atomic Absorption Methods 136
Determination of Hg by Flameless A.A 139
Neutron Activation Analysis Methods for Hg and As 141
X-Ray Fluorescence Methods 143
Optical Emission Spectrometry Methods 143
Determination of Carbon 143
Methods for Ammonium, TDS, COD, and Chloride 144
Methods for Sulfate, Phosphate, and Boron 145
Determination of Base Exchange Capacity 146
Determination of Surface Area 146
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LIST OF FIGURES
No. Page
1. Diagram of column apparatus used in leachate study 13
2. Relative column effluent concentrations for several elements as
a function of pore fraction of leachate passed through columns
containing (a) 2% mohtmorillonite, (b) 8% montmorillonite, and
(c) 16% montmorillonite clay * 18
3. Relative column effluent concentrations for several elements as
a function of pore fraction of leachate passed through columns
containing (a) 2% montmorillonite, (b) 8% montmorillonite, and
(c) 16% montmorillonite clay 19
4. The attenuation number related to cation exchange capacity (a)
K, (b) NHi,, (c) Na, and (d) Mg 24
5. (a) Ca attenuation number related to cation exchange capacity.
(b) Mh attenuation number related to clay percentage 25
6. Manganese elution related to percentage of kaolinite leached
with natural and sterile leachate 30
7. Hardness of water expressed as CaCOa, in Silurian Dolomite
aquifer in DuPage County, northeastern Illinois 36
8. "Hardness halo" effect shown as a function of distance (m)
from the Winnetka and DuPage landfills in northeastern
Illinois 37
9. Hydraulic conductivity of kaolinite-sand columns as a
function of leaching time 39
10. Hydraulic conductivity of montmorillonite-sand columns as
a function of leaching time 40
11. Hydraulic conductivity of illite-sand columns as a function
of leaching time 41
12. The amount of Pb removed from DuPage leachate by kaolinite
at 25° C plotted as a function of pH 47
13. The amount of Pb removed from DuPage leachate by montmoril-
lonite at 25° C plotted as a function of pH 48
vi
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No. Page
14. The amount of Pb removed from Blackwell leachate by
kaolinite at 25° C plotted as a function of pH 49
15. The amount of Pb sorbed per gram of kaolinite at pH 5.0
and 25° C plotted as a function of the equilibrium Pb
concentration 52
16. Pb sorption data for kaolinite and montmorillonite at pH
5.0 and 25° C plotted according to the Langmuir isotherm
equation 53
17. Distribution of Pb (II) species in 4 x lO"1* M Pb(N03)2 and
uptake by 0.1 g kaolinite from 60 ml of solution 56
18. The amount of Cu, Zn, or Cd removed from solution per gram
of kaolinite at pH 5.0 and 25° C, plotted as a function of
the equilibrium concentration 67
19. The amount of Cu, Zn, or Cd removed from solution per gram
of montmorillonite at pH 5.0 and 25° C, plotted as a
function of the equilibrium concentration 68
20. Cu, Zn, and Cd removal data for kaolinite in deionized
water solutions at pH 5.0 and 25° C, plotted according
to the Langmuir equation (Eq. 8) 71
21. Cu, Zn, and Cd removal data for montmorillonite in
deionized water solutions at pH 5.0 and 25° C, plotted
according to the Langmuir equation. Numerals indicate
corresponding isotherms in Fig. 19 72
22. Cu, Zn, and Cd removal data for kaolinite in deionized
water solutions at pH 5.0 and 25° C, plotted according
to the competitive Langmuir equation (Eq. 12) 74
23. Cu, Zn, and Cd removal data for montmorillonite in
deionized water solutions at pH 5.0 and 25° C, plotted
according to the competitive Langmuir equation (Eq. 12) 75
24. The amount of Cu, Zn, or.Cd removed from DuPage leachate
solutions by kaolinite at 25° C, plotted as a function
of pH 78
25. The amount of Cu, Zn, or Cd removed from DuPage leachate
solutions by montmorillonite at 25° C, plotted as a
function of pH 79
26. Chromium (VI) adsorption-pH curves for montmorillonite
at 25° C 87
vii
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No. Page
27. Chromium (VI) adsorption-pH curves for kaolinite at
25° C .............................. 88
28. Distribution of Cr(VI) species for various Cr(VI)
concentrations .......................... 89
29. Adsorption isotherms for Cr(VI) at pH 4.0 and 25° C ........ 91
30. Langmuir plots of Cr(VI) adsorption data at pH 4.0 and 25° C. . . • 92
31. Removal of Cr(III) from solution by kaolinite ........... 94
32. Chromium (III) adsorption-pH curves at 25° C ............ 96
33. Chromium (III) adsorption isotherms at pH 4^.0 and 25° C ...... 97
34. Langmuir plots of Cr(III) adsorption data at pH 4.0 and 25° C . . . 98
35. The amount of As(V) removed from DuPage leachate solutions
by kaolinite and montmorillonite at 25° C plotted as a func-
tion of pH ............................ 105
36. Species distribution diagram for As(V) and Se(IV) ......... 106
37. The amount of As (III) removed from DuPage leachate solutions
by kaolinite and montmorillonite at 25° C plotted as a
- function of pH .......................... 107
38. The amount of As(V) or As (III) removed from DuPage leachate
solutions at pH 5.0 and 25° C per gram of clay plotted as
a function of the equilibrium arsenic concentration ....... 108
39. The amount of Se(IV) removed from DuPage leachate solutions
by kaolinite and montmorillonite at 25° C plotted as a
function of pH .......................... 110
40. The amount of Se(IV) removed from DuPage leachate solutions
at 25° C and several pH values per gram of clay plotted
as a function of the equilibrium Se concentration ........ Ill
41. Removal of Hg from DuPage landf ill-leachate and pure
solutions plotted as a function of pH at 25° C .......... 116
42. Removal of various forms of Hg from DuPage landf ill-leachate
solutions by kaolinite plotted as a function of pH at 25° C . . . 117
43. Effect of clay content on hydraulic conductivity and attenuation
of Pb, NH4, and Cl for a 40-cm thick liner ............ 124
44. X-ray diffraction patterns of the clay minerals used in
leachate pollutant attenuation studies .............. 137
viii
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LIST OF TABLES
No. Page
1. Chemical Characterization of the Clay Minerals Used in
Attenuation Studies of Leachate Pollutants ............ 10
2. Summary of Chemical Characteristics of Landfill Leachates
3. Design of Experiment and Some Physical and Chemical Properties
of the Column Contents
4. Rank of Some Chemical Constituents Found in Municipal Leachate
According to Their Relative Mobility Through Clay Mineral
Columns ............................. 21
5. Mean Attenuation Number (ATN) of Some Chemical Constituents
Found in Municipal Leachates for Three Clay Minerals ....... 23
6. Chemical Constituents in DuPage Leachate, Ranked According to
Pollution Hazard ....... • .................. 33
7. Pb Removal Parameters Used to Compute Sorption Isotherms from
52 ml Reaction Volumes ...................... 50
8. Maximum Removal of Pb from pH 5.0 and 25° C Solutions Computed
Using the Langmuir Equation ........ . .......... 54
9. Thickness (cm) of a Square Meter of a 30% Clay Liner Needed to
Remove Pb from 762 Liters (201 gal) of Solution Per Year ..... 59
10. Total Content of Cu, Zn, or Cd in 50 ml of Solution ........ 69
11. Comparison of Langmuir Adsorption Maximums in Deionized Water
with CEC Values ......................... 76
12. Adsorption Maxima for Cr(VI) by Montmorillonite and Kaolinite
at 25° C for Various pH Values .................. 93
13. Adsorption Maxima for Cr(III) by Montmorillonite and Kaolinite
at 25° C for Various pH Values .................. 99
14. Removal of Heavy Metals from Solutions by Kaolinite at pH 5.0 . . . 121
15. Estimated Landfill Liner Thickness Necessary for Attenuation
of Some Leachate Constituents Per Cubic Meter of Refuse
During a 20 Yr. Fill Life .................... 126
ix
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ACKNOWLEDGMENTS
In addition to the authors listed, the following were involved in writ-
ing parts of this report: A. K. Au, Keros Cartwright, R. R. Frost, R. H.
Gilkeson, G. M. Hughes, G. D. Robinson, R. R. Ruch, J. D. Steele, and W. A.
White. Their names are also listed in a footnote at the beginning of each
section to which they contributed.
The authors gratefully acknowledge the U.S. Environmental Protection
Agency for partial support of the work under Contract No. 68-03-0211, Cin- •
cinnati, Ohio; the American Colloid Company, Skokie, Illinois, for supplying
the montmorillonite clay; the Minerva Company, Elizabethtown, Illinois, for
supplying the illite clay; and the Ottawa Silica Company, Ottawa, Illinois,
for supplying the quartz sand used in this study.
The authors also wish to give thanks to those Geological Survey staff
members who gave freely of their time and talents. Among those are J. M.
Mellske, D. R. Dickerson, R. H. Shiley, W. J. Armon, J. K. Kuhn, J. A.
Schleicher, L. R. Camp, G. B. Dreher, J. K. Frost, L. R. Henderson, M. C.
Charles, S. K. Friesen, R. M. Trandel, and D. B. Heck.
Special thanks go to R. J. Helfinstine and W. E. Cooper for special
equipment and to P. L. Johnson for typing assistance.
x
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SECTION 1
INTRODUCTION
During the past 30 years, the landfill method for disposal of municipal
and industrial waste has been widely used in the United States. More than 90%
of our nation's wastes are now placed on the land for ultimate disposal at
the approximately 14,000 landfills throughout the nation (Garland and Mosher,
1975). These 14,000 landfills accept more than 360 million tons of household,
commercial, industrial, and municipal solid waste per day, at a cost of more
than 4.5 billion dollars annually (Black et al., 1968), As industry in the
United States complies with the Clean Air Act and the Federal Water Pollution
Control Act, the volume of industrial solid wastes, sludges, and liquid con-
centrates of pollutants is expected to double in the next 10 years. The dis-
posal of such huge volumes of solid waste by landfilling is not without its
environmental impact. When refuse buried in a landfill comes in contact with
water, then leachate, a mineralized liquid high in organic substances, is
produced and may move out of the fill and pollute the ground water.
Garland and Mosher (1975) have cited several examples of pollution by
leachates migrating from landfills. An example of severe economic damage in-
curred by pollution of a drinking water aquifer by leachate from a landfill
occurred in New Castle County, Delaware (Apgar and Satterthwaite, 1975).
Leachate from the landfill migrated more than 800 feet and contaminated the
Potomac drinking water aquifer 4 years after the landfill site had been
closed. The drinking water was contaminated with such high levels of organic
compounds and metal ions that it was no longer potable. To date, this problem
has cost the county $800,000 for interim solutions and, if the dump must be
moved to completely remedy the situation, the cost may reach as high as 20
million dollars. In addition to the monetary costs, the county estimates that
it will take 10 years to restore full use of the aquifer.
In another case, reported by Garland and Mosher (1975), contamination of
ground water by selenium was found to extend more than 2 miles from a dump
site in Long Island. In this case as elsewhere, heavy-metal contamination may
impart no odor or color to indicate that the water is contaminated.
The solid waste problem is most acute in the metropolitan areas, where
competition for the available land is intense. A city of 2 million inhabi-
tants generates 5000 tons of solid waste per day (Wirenius and Sloan, 1973),
which rapidly fills the conveniently situated landfill locations. The problem
of finding environmentally acceptable sites close to metropolitan areas is
compounded by urban sprawl and persistent opposition from citizen and en-
vironmental groups. The rapid increase in problems and the costs associated
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with transporting refuse for long distances now make it prudent to consider
sites that were previously unacceptable because they failed to meet certain
geologic and hydrologic critera. As of January 1975, Illinois had 2,040 waste
disposal sites recorded with the Illinois State Environmental Protection
Agency. Since 1965, the management of refuse disposal in Illinois has greatly
improved, owing to passage of the Illinois Refuse Disposal Act, which as-
signed the regulation of solid waste disposal to the Department of Public
Health. More comprehensive regulation was provided in 1970 with the creation
of the Environmental Protection Agency by passage of the Environmental Pro-
tection Act (HB3788). Regulations were passed that were designed to insure
that solid waste disposal facilities are located at such sites and operated
in such a manner as will protect the physical environment and public health.
These regulations have been enforced for the past 10 years. During this time,
the Illinois State Geological Survey has assisted the regulatory agencies by
evaluating the hydrogeologic conditions at proposed or operating waste dis-
posal sites. During the past 8 years, the Survey has appraised at about 100
sites annually the conditions relative to pollution hazard. Some sites were
not approved for geologic reasons, including locations in floodplains or
gravel pits, on fractured rock over aquifers, on steep grades, or in areas
of special environmental significance. Other sites were approved but were
never put into operation because of persistent opposition from citizen and
environmental groups or for other reasons.
The future of landfill disposal is clear. Acceptable disposal sites will
be difficult to find, their location will be approved only after certain
geologic and hydrologic criteria are met, and greater care will be required
in their operation. The relative unavailability of geologically acceptable
sites in close proximity to metropolitan areas and the rapidly escalating
costs associated with transportation of waste materials across long distances
now make it economically feasible to consider physically modifying geo-
logically unacceptable sites that may be ideal in other respects. The
Illinois State Geological Survey has conducted extensive studies of the move-
ments of pollutants through various geologic strata and in a variety of
hydrologic settings at several landfill sites in northeastern Illinois
(Hughes et al., 1971). Cartwright and McComas (1968) also conducted geo-
physical investigations at the same sites. The above studies, and others by
Apgar and Langmuir (1971) and Emrich (1972), have indicated that pollutants
in leachate can be detected at variable distances from a landfill, depending
on the clay content of the soil or the hydraulic conductivity of the soil
strata. It has therefore been suggested that a clay liner in the bottom of
previously unacceptable sites, such as gravel pits or old quarries, could
make them acceptable for disposal of municipal and/or industrial wastes. How-
ever, no sound evidence existed to indicate how thick such a layer must be or
what types of clay minerals were best suited for removal of toxic metals in
the presence of municipal leachate.
This paper reports the results of a study (1) to investigate and evalu-
ate the attenuating properties of several clay minerals for the pollutants
contained in leachates from municipal solid waste, and (2) to determine the,
capacity of the two major clay mineral types for removing the heavy metals —
Cr, Cu, Pb, Cd, Hg, As, Se, and Zn — from solution and the effect municipal
leachates have on this capacity at various pH values. The study was also de-
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signed to provide insight into the mechanisms responsible for attenuation of
heavy metals as well as to evaluate the potential use of clay minerals as
liners for waste disposal sites. Such liners would be used to prevent or
mitigate pollution of ground and surface waters by liquid effluents.
It is clear that, based on their intrinsic properties alone, municipal
leachates are noxious waste streams that pose a potential threat to public
health (Hanks, 1967; Peterson, 1974). To assess the magnitude of the pollu-
tion hazard of such a stream is a difficult problem and was the subject of
the U. S. Environmental Protection Agency (1973) decision model for screen-
ing, selecting, and ranking hazardous wastes streams. Use of the priority
ranking formulation requires evaluation of the "critical product" (pollution
hazard). At present, no actual waste-stream data for municipal leachates is
available. Therefore, one goal of this study was to provide waste-stream
data that could be used to determine the mobility index of several of the
major pollutants found in municipal leachates £hat had passed through simu-
lated clay mineral landfill liners. The mobility index thus derived could be
used to compute the pollution hazard for most of the chemical constituents
found in municipal leachate.
This report is presented in separate chapters that describe different
aspects of the work that was performed. For the convenience of the reader,
each chapter includes its own abstract. The column leaching studies (Sections
5 and 6) used "natural" landfill leachate as described in the Experimental
section (Section 4). Sections 7 through 11 report equilibrium adsorption
"batch" studies using small amounts of clays mixed with aqueous solutions of
metal salts and landfill leachates that had been "spiked" with metal salts.
The results of these investigations will also find application in the land
disposal of industrial and energy production wastes.
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SECTION 2
CONCLUSIONS
The results from the column leaching and adsorption studies have yielded
complementary data that have allowed us to make predictions as to the ex-
pected magnitude of the reduction in concentration or attenuation in soil for
most of the common chemical constituents found in landfill leachates. The
data also indicate that several mechanisms may be responsible for attenuation
of the various pollutants under differing environmental conditions.
Strong evidence is presented supporting the conclusion that the cation
exchange capacity is the principal chemical property of a clay mineral to
effect attenuation. Of the three clays used in this study, montmorillonite
was found to have the highest attenuation capability, followed by illite and
then kaolinite. This order was concluded to be a result of their respective
cation exchange capacities.
The pH of the leachate was found to significantly affect the amount of
attenuation. It was concluded that the heavy metal cations — Pb, Cd, Cu,
Cr(III), Hg, and Zn — were attenuated primarily by an exchange-adsorption
mechanism which was affected by pH and competition from other cations. How-
ever, at pH values between 5 and 6, a large increase in removal can be ex-
pected due to increased adsorption of metal complex ions and to formation of
insoluble heavy metal hydroxide and carbonate compounds. It was therefore
concluded that at high pH the primary mechanism of attenuation for these ions
was precipitation. The effect of pH on the attenuation of the heavy metal
anions Cr(VI), As, and Se was found to be the opposite of the cations and it
was concluded that precipitation was not an important attenuation mechanism.
Rather, the adsorption of the anions was found to correlate well with the
distribution of certain ionic species in solution. It was concluded that
HCrOIT was the species adsorbing in this study, since Cr(VI) adsorption be-
came zero as the pH was raised to 8.4, corresponding to the disappearance of
HCrO^ from solution in favor of the CrO^ ion. The adsorption of Cr(VI) was
also found to start to decrease as the pH was lowered past 2, corresponding
to the decrease in HCrO^ ion in favor of HaCrO^. Likewise, As and Se adsorp-
tion were also found to correspond to the distribution of HaAsO^ and HSeOJ
species in solution. These results led to the conclusion that the principal
attenuation mechanism for the heavy metal anions was adsorption of the mono-
valent species from solution. It was also concluded that at higher pH values
the heavy metal anions would be significantly more mobile than the cations.
The relative mobilities of the heavy metals, as determined from equi-
librium adsorption data from pure solutions of the metals at pH 5, were:
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Cr(III) < Cu < Pb < Zn < Cd < As(V) < As(III) < Se < Cr(VI)
The cationic heavy metals are generally adsorbed to a greater degree than are
the anionic heavy metals. However, this ranking is dependent on pH and ionic
competition and therefore changes somewhat in different leachates.
A significant point shown in the ranking is the importance of the
valence state of an element to the amount of that element removed from solu-
tion by clay minerals. Cr(III) species were removed to a much greater extent
than Cr(VI), and more As(V) was removed than As(III). These data suggested
that safer disposal of certain elements may be achieved if the element werejf
converted to the form most strongly attenuated prior to disposal.
The formation of Pb and Cd organic complexes in leachates was measured,
and their effect was determined to be of secondary importance to adsorption
and precipitation. This seemed to be due to competition from high concentra-
tions of other cations present in leachates. Due to the complex interactions
between inorganic, organic, and volatile forms of Hg, the mobility and rel-
ative importance of organic Hg complexes could not be accurately assessed.
The results of this study have led to the conclusion that passage of
leachate through a Ca-saturated clay material will result in high attenuation
of the heavy metals; in moderate attenuation of K, NHi*, Mg, and Si; and in
relatively low attenuation of Cl, Na, and water-soluble organic compounds
(COD). It was further concluded that the oxidation-reduction potential of the
leachate controlled the attenuation of Fe and Mn. Under strongly anaerobic
conditions, Fe and Ma will probably not be attenuated and may even elute in
substantial concentrations due to the dissolution of oxide coatings on the
clay surfaces. However, under mildly anaerobic conditions, substantial at-
tenuation can occur.
Substantial concentrations of Ca were eluted from the columns. Since a
very highly significant linear regression (r = 0.97) was obtained for the
amount of Ca eluted versus the cation exchange capacity of the clay, and
since the sum of the amount of K, NHi,, Na, and Mg removed from the leachate
agreed within about 3% with the amount of Ca eluted, it was concluded that
Ca elution was due to exchange with the other cations present in the leach-
ate. It was also concluded that this Ca elution observed in the laboratory
experiment corresponded to the "hardness halo" effect observed in field
monitoring wells around sanitary landfill sites.
Significant reductions in hydraulic conductivity were observed when
landfill leachate was passed through the columns. It was concluded that
microbial action is primarily responsible for the observed reductions and
that hydraulic activity of clay-sand liners placed in the bottom of a land-
fill will decrease after a period of leaching with municipal effluent. It was
further concluded that montmorillonite clays will decrease in permeability to
a greater extent than the other clays. This was assumed to be due to its
tendency to swell to a much greater extent than other types of clay.
A pollution hazard ranking system was developed. Using this system, it
was concluded that NHit was a 30 times greater pollution hazard than any other
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constituent found in the DuPage leachate. It was also concluded that in a
fresh, young leachate, COD or Fe most likely would present the highest pollu-
tion hazard and that the characteristics of each leachate and earth material
must be considered when evaluating their potential for pollution. The well
known variability in composition of municipal landfill leachates (Garland
and Mosher, 1975) and the difficulty in predicting leachate composition be-
fore landfilling is begun will make it difficult to apply this hazard-ranking
system. Nevertheless, the ranking system has the advantage of quantifying the
expected pollution hazard of a given leachate and allows comparisons of the
pollution hazards of one leachate with another. The ranking system also aids
regulators, landfill designers, and researchers by focusing attention on
those chemical constituents in the leachate with the highest potential for
serious pollution.
The mobility of a given leachate constituent, and hence the thickness of
a clay liner necessary to attenuate the constituent, depends to a large de-
gree on the element and the form of the element, the adsorption capacity of
the earth material, the cations present initially on the exchange complex,
the chemical composition of the leachate, and the pH of the leachate. The
adsorption capacity of the clay minerals and the reversible nature of ex-
change-adsorption reactions have important environmental consequences. Indus-
trial wastes containing heavy metals placed in sanitary landfills could alter
the ionic composition and/or pH of the leachate. A change in pH may release
large amounts of potentially toxic heavy metals into the aqueous phase,
especially in places where precipitates may have accumulated. Other ions in
the waste compete with the heavy metals and may exchange with them, thus al-
lowing metal ions to come into solution. These multiple interactions must be
considered when a disposal site is designed and when the environmental impact
of adding heavy-metal wastes to municipal landfills is assessed.
-------
SECTION 3
RECOMMENDATIONS
This study dealt specifically with attenuation of the inorganic con-
stituents of leachate, but, because leachates contain high concentrations of
organic compounds, also needed is a laboratory investigation dealing with the
attenuation of the specific organic compounds in leachate. The results of our
study, using COD as an indicator of organic-compound movement, showed poor
attenuation of organics. At the same time, the pollution hazard index in-
dicates that the organic fraction of leachate poses a serious pollution
hazard, especially for young leachates. It is therefore recommended that an
attenuation study be initiated, involving a comprehensive organic analysis to
determine removal of specific organics by passage through clay liners and/or
soils.
The results of this study also illustrated the important role that the
pH of the leachate or waste stream plays in pollutant removal by clay
minerals. It is therefore recommended that, in conjunction with the organic
removal studies, further studies be carried out to determine techniques which
will allow the pH to be manipulated to enhance and economically optimize
pollutant (organic and inorganic) removal by earth materials. Also, landfill
disposal of anionic forms of heavy metals such as Cr(VI), As, and Se should
be closely scrutinized because of their relatively high mobility and the fact
that manipulation of pH conditions to enhance removal of cationic heavy
metals such as Pb, Cd, Zn, Cu, and Cr(III) may actually increase the mobility
of the anionic metal ions.
Due to the relatively, high toxicity of Hg and the complex interactions
between inorganic, organic, and volatile forms of the element, much more
research is necessary to determine the environmental impact of Hg in landfill
environments. In conjunction with the study of Hg, Pt has also been reported
to be methylated in aquatic systems, thus establishing a previously unknown
mechanism of transfer. The major source of potential redistribution of com-
plexed forms of Pt is expected to come from the disposal into landfills of
spent catalysts from catalytic-converter-equipped automobiles. Thus, it is
recommended that Hg and Pt be studied to quantitatively determine the adsorp-
tion capacity of clays and soils for the inorganic and organic forms of Hg
and Pt. It is also important to determine the magnitude and role of micro-
organisms in the methylation of Hg and Pt to volatile forms under the con-
ditions found in municipal landfills.
A major task still before us is the utilization of quantitative soil
chemistry data to make predictions of pollutant migration beneath landfills.
-------
It is therefore recommended that further studies be carried out that involve
•a cooperative effort between soil chemists, groundwater hydrologists, and
modelers for the computer implementation of the prediction process and to
identify possible gaps in knowledge that may still bar the successful pre-
diction of long-term pollutant migration from disposal sites.
-------
SECTION 4
EXPERIMENTAL
The clays used in this study were the purified clay minerals kaolinite
(1:1 lattice type), montmorillonite (2:1 expanding lattice), and illite (2:1
nonexpanding lattice, mica type). These clay minerals were chosen for study
because they are the most common clay minerals in earth materials that would
be used individually or in combinations for landfill sites. Earth materials
containing one or more of these clay minerals can generally be obtained
locally for landfill liners.
The clays were brought to the laboratory, where they were crushed,
ground, and purified by sedimentation techniques to obtain the <2 ym particle
fraction that contained essentially pure clay minerals. Chemical analyses of
the three clays are given in Table 1. Details of the methods used and results
of X-ray diffraction analyses of the clays are given in the appendix. The
predominantly Ca-saturated, <2 ym fractions of the clays were then used in
the column leaching and heavy metal adsorption studies.
The municipal landfill-leachates used in this study were collected from
the DuPage County landfill (well MM 63) and from the monitoring well located
on the Blackwell Forest Preserve landfill. The details of the site descrip-
tions and well locations are given by Hughes et al. (1971). The leachate was
collected by using a tubing pump and large plastic containers that were
equipped with valves to allow continuous purging of high-purity C02 or argon
gas to maintain anaerobic conditions. The DuPage leachate collected initially
for use in the column leaching study was stored under argon. Both the DuPage
and Blackwell leachates collected at a later date for use in the heavy metal
adsorption studies were stored under C02. Purging with C02, a naturally oc-
curring landfill gas, was found to be a more satisfactory method than argon
purging since it permitted the leachates to be stored in the laboratory for
longer periods of time without significant changes in pH. The leachate used
in the column study was collected with a tubing pump and was split into two
53-gallon plastic closed-head drums. One drum was taken to the Argonne
National Laboratories and sterilized by gamma ray irradiation using a cobalt
source which gave a dose of 3.36 x 106 rad at the center of the drum. Both
the sterile and natural drums were stored under refrigeration with the
temperature maintained at 3 to 5° C. The drum containing the microbially
active leachate was stored with argon purged over the top of the leachate,
while the sterilized drum was stored over a 12% ethylene oxide/88% freon gas
mixture which maintained both sterility and an anaerobic environment. Mer-
curic nitrate salt was later added to the drum (Table 2) to maintain steril-
ity. Plate counts were performed on both leachate drums using potato dextrose
-------
TABLE 1. CHEMICAL CHARACTERIZATION OF THE CLAY MINERALS USED
IN ATTENUATION STUDIES OF LEACHATE POLLUTANTS
Element
Kaolin it e (Pike
County, Illinois)
(ppm)
Exch.*
Total
Monttnorillonite
(American Colloid
Co . southern
bentonite)
(ppm)
Exch.*
Total
Illite
(Minerva
Co. Mine)
(ppm)
Exch . *
Total
Ca
Mg
Na
K
NHn
Fe
Mn
Pb
Cd
Zn
B
Al
Si
Ti
Carbon (%)
Total
Organic
Inorganic
CEC
(meq/lOOg)
Surface
area (m2/g]
2,592. 3,700 13,120.
76.8 1.8QO 680.
43.2 929 24.0
87.2 8,200 240.
13.0 40 43.
<2.0 6,600 <2.0
0.06 29 0.02
<2.0 46 <2.0
<0.2 <3 <0.2
0.80 20 1.00
46 -
221,800
217,700
14,700
0.54
0.51
0.03
15.1
34.2
22,300
25,500
178
1,100
38
25,500
25
<15
<3
40
3
95,600
284,800
1,300
0.93
0.92
0.01
79.5
86.0
5,248 23,350.
800. 10,430.
115.2 1,050.
800. 56,270.
50. 62.5
<2.0 28,730.
0.37 <390.
<2.0 93.8
<0,3 18.8
2.5 37.5
43.8
130,100.
226,500.
4,010.
2.19
1.81
0.38
20.5
64.6
* Exch an ge ab1e
10
-------
TABLE 2. SUMMARY OF CHEMICAL CHARACTERISTICS OF LANDFILL LEACHATES
/
Component
COD
BOD
TOC
Organic acids
Carbonyls as
acetophenone
Carbohydrates as
dextrose
pH
Eh (m.v.)
TS
TDS
TSS
E.G. (mmhos/cm)
Alkalinity
(CaC03)
Hardness
(CaC03)
Total P
Ortho P
NHit-N '
N03+N02-N
Al
As
B
Ca
Cl
Na
K
Sulfate
Mn
Mg
Fe
Cr
Hg
Ni
Si
Zn
Cu
Cd
Pb
Range of all
values given
by Garland and
Mosher (1975)
(mg/D
40 -
9 -
256 -
-
-
-
4 -
-
0 -
0 -
6 -
^ —
0 -
0 -
0 -
6 -
0 -
0 -
-
-
-
5 -
34 -
0 -
3 -
1 -
0 -
16 -
0 -
-
-
-
-
0 -
0 -
0 -
0 -
89,520
54,610
28,000
9
59,200
42,276
2,685
17
20,850
22,800
154
85
1,106
1,300
4,080
2,800
7,700
3,770
1,826
1,400
15,600
5,500
1,000
10
17
5
DuPage leachate
DuPage used in column
Blackwell Forest leachate used study
Preserve leachate in sorption (mg/1)
(Hughes, 1971b) study
(mg/1) (mg/1) Natural Sterile
39,680.
54,610.
-
-
-
-
7.10
-180.
-
19,144.
-
10.90
3,255.
7,830.
6.
-
-
1.70
2.20
4.31
-
-
1,697.
900.
-
680.
1.66
-
5,500.
. 0.20
-
-
-
-
0.05
<0.05
"
1,362.
-
-
333.
57.6
12.
6.79
-155.
-
5,910.
-
7.20
4,220.
1,100.
<0.1
-
809.
-
<0.1
0.11
33.
49.
1,070.
822.
516.
<0.01
<0.1
204.
4.40
<0.1
0.0008
0.3
15.1
0.03
<0.1
<0.01
<0.1
1,340.
-
-
333.
57.6
12.
6.9
+7.
-
-
-
10.20
-
-
<0.1
-
862.
-
<0.1
0.11
29.9
46.8
3,484.
748.
501.
<0.01
<0.1
233.
4.2
<0.1
0.0008
0.3
14.9
18.8
<0.1
1.95
4.46
10,603.*
-
-
290.
90.1
11.
7.2
+75.
-
-
-
10.42
-
-
<0.1
-
773.
-
<0.1
0.14
28.5
43.2
3,311.
744.
491.
<0.01
<0.1
230.
3.0
<0.1
0.87*
0.3
15.0
16.3
<0.1
1.88
4.26
*Added as- a result of sterilization maintenance
11
-------
agar media. No growth occurred on plates inoculated with the sterilized
leachate while active growth of microbial colonies was observed on plates
inoculated with the natural leachate, thus indicating that sterility had been
achieved.
The results of chemical analysis of the leachates are presented in Table
2. Chemical procedures used are given in the appendix. For comparison, Table
2 also contains a summary of the range of leachate characteristics found for
more than 20 other leachates as given by Garland and Mosher (1975). It is
useful to note that the two leachates used in this study have widely differ-
ent chemical compositions. The DuPage leachate is approximately 15 years old
and has a lower total salt, phosphate, and sulfate content than the Black-
well. In addition, the organic matter of the DuPage leachate consists mainly
of microbially resistant compounds, which have been found to be more mobile
in soils than are biodegradable compounds (Hughes et al., 1971, Gowler,
1970). The Blackwell leachate, on the other hand, is younger and ranks among
the most concentrated ever reported, especially with regard to BOD and Fe.
The laboratory apparatus used in the leaching study consisted of labora-
tory columns containing mixtures of clay minerals and washed quartz sand
through which leachate was passed. A diagram of the column apparatus design
appears in Figure 1. The columns and apparatus were constructed to simulate
the slow (<2 pore volumes per month), saturated, anaerobic flow of leachate
as it is thought to occur at the bottom of a landfill.
~\
Pore Volume = Tl - ( Bulk Density J Volume of Column {l}
|_ Particle DensityJ
The entire column leaching system, was maintained under an argon atmosphere to
maintain anaerobic conditions. A tubing pump lifted the leachate to a 5-gal-
lon plastic carboy which acted as both a constant head device (Harriot
bottle) and a temperature equilibration reservoir. The leachate was then
passed through the columns and the effluents were collected in graduated
cylinders, which also allowed measurement of the flow rates. The outflow tube
was maintained above the top of the columns to insure saturated flow. The
level of the outflow tube was moved either up or down to maintain relatively
constant flow rates throughout the experiment.
The columns were constructed of 2-inch acrylic tubing, to which man-
ometer outlets were fitted vertically on the column at five locations. To
simulate field conditions, the leachate containers and columns were either
painted black or masked with black tape to prevent growth of organisms, such
as algae or photosynthetic bacteria, which are not indigenous to deep refuse
leachate.
The sand grains were coated with the clays, according to the methods
given by Grim and Cuthbert (1945) . The clay minerals and sand were then uni-
formly packed in the columns to a depth of 40 cm, except for the 32% and 64%
montmorillonite columns, which were 30 cm thick. The columns were packed to
bulk densities approximating those of naturally occurring glacial tills
(~1.8 g/cc; Manger, 1963). Table 3 gives some chemical and physical proper-
ties of the column contents. The hydraulic conductivities for each particular
12
-------
CVI
55 Gal. Water Refrigeration
Leachate Drum Bath Unit
Figure 1. Diagram of column apparatus used in leachate study.
13
-------
TABLE 3. DESIGN OF EXPERIMENT AND SOME PHYSICAL AND CHEMICAL PROPERTIES
OF THE COLTJMN CONTENTS
Natural Leachate
Treatment
Cation
Exchange
Capacity
(meq./lOOg)
Set A*Set Bt
Bulk density
(g/cc)
Initial hydraulic
conductivity
(cm/sec)
Set A Set B Set A
Set B
100 Sand
2 Mtt
4 M
8 M
16 M
32 M
64 M
2 Ktt
4 K
8 K
16 K
32 K
64 K
4 Itt
16 I
8K + 81
8M + 8K + 81
0.0
1.4
3.2
7.3
11.9
26.8
56.2
0.7
1.1
1.5
1.8
3.8
9.6
0.8
3.5
2.4
8.8
.1
2.3
4.3
7.2
12.1
24.0
55.5
0.4
0.8
1.4
2.5
3.4
8.5
0.9
3.2
2.3
8.5
1.71
1.71
1.77
1.79
1.87
1.55
1.23
1.68
1.76
1.80
1.87
1.66
1.22
1.80
1.83
1.90
1.64
1.71
1.72
1.74
1.78
1.86
1.52
1.11
1.70
1.74
1.77
1.90
1.55
1.32
1.81
1.91
1.98
1.69
1.27.10-3
9. 45. 10-"
4.34.10-"
4.70.10-"
1.22.10-5
1.27.10-6
3.05.10-7
7.44.10-"
4.78.10"5
9. 90.10-"
2.86.10~5
2.40.10-6
5.45.10-7
8. 17. 10-"
2.68.10"5
1.48.10-6
8.08.10-6
1.80.10-3
7. 93. 10-"
3. 47. 10-"
2. 61. 10-"
1.44.10-5
2.17.10-6
6.83.10-7
4. 53. 10-"
2.76.10'5
8. 25. 10-"
1.92.10-6
4.81.1Q-6
4.57.10'7
7. 16. 10-"
2.19-10-5
1.68.10'6
9.43.10-6
*Set A - Natural leachate
tSet B - Sterile leachate
ttM = Montmorillonite, I = Illite, K = Kaolinite
14
-------
clay content and bulk density agree with those given by Todd (1959) for
natural materials. The experimental design used in the study is also shown in
Table 3, which gives the percentages of clay mineral(s) in each column (to
which pure quartz sand was added to total 100%) . The experimental design in-
cludes a complete geometric progression of clay percentages from 2% to 64%
kaolinite and montmorillonite and two mixtures of clays as given in Table 3.
A 100% sand column was also included. Only 4 and 16 percentages of illite
were included because kaolinite and illite have very similar cation exchange
and lattice expansion properties, and thus a complete geometric array for
illite was not considered necessary.
After the leachate and the column contents were characterized, the
leachate was passed through the columns for periods of between 6 and 10
months, depending on the hydraulic conductivity of the individual column. The
hydraulic conductivity (K) was computed using the relationship:
K
AdH
where Q = flow rate in cm3 /sec
A = cross sectional area of column in cm2
dL = length of the column in cm
dH = head of water in cm.
During this time effluents from each column were collected periodically
and measurements were made for Na, K, Ca, Mg, Al, Zn, Pb, Cd, Hg, Fe, Mh,
NHi» , B, Si, Cl, chemical oxygen demand (COD), Eh, pH, and hydraulic conduc-
tivity. Finally, after approximately 15 pore volumes were leached, the clay
mineral columns were sectioned and the contents analyzed to determine the
vertical distribution of chemical constituents in each column.
Duplicate sets of columns were used in the experiment; one set of
columns was leached with natural effluent, another with sterilized effluent.
The sterilized treatment was used to determine how gross biological activity
might affect hydraulic conductivity of leachate through clay minerals used as
liners. The results of the experiment were statistically analyzed using the
paired t statistic to determine if there were significant differences in the
attenuation of each chemical constituent between sterile and natural leachate
and between clay mineral types. Linear regression and moving average analysis
were also performed on the column effluent data to determine relationships
between hydraulic conductivity, attenuation, and clay mineral properties.
In addition to the column leaching study > a series of separate equi-
librium studies on the capacity of clays to adsorb eight potentially hazard-
ous elements — Pb, Cd, Zn, Cu, Cr, As, Se, and Hg — were performed. From
these studies, adsorption isotherms for kaolinite and montmorillonite were
constructed to determine the adsorption capacities as a function of concen-
tration, pH, and ionic competition. The details of the experimental pro-
cedures used for each element are included in the section of the report deal-
ing with that particular element.
15
-------
SECTION 5
COLUMN LEACHING STUDY1
ABSTRACT
To evaluate the potential of clay minerals for attenuating the various
chemical constituents of landfill leachate, leachate was collected by an-
aerobic techniques from the 15-year old DuPage County sanitary landfill near
Chicago, Illinois, and passed through laboratory columns that contained vari-
ous mixtures of calcium-saturated clays and washed quartz sand. The columns
were constructed to simulate slow, saturated, anaerobic flow of leachate
through earth materials outside the landfill. Manometers were placed in each
column to measure any changes in permeability. Leachates were run through the
columns for periods of up to 10 months, during which time effluents were
periodically collected and analyzed for 16 chemical constituents. The column
contents were then cut into sections and analyzed to determine the vertical
distribution of chemical constituents in each column.
Chloride, Na, and water-soluble organic compounds (COD) were relatively
unattenuated by passage through the clay columns; K, NHi* , Mg, Si, and Fe were
moderately attenuated; and heavy metals, such as Pb, Cd, Hg, and Zn, were
strongly attenuated by even small amounts of clay. Concentrations of Ca, B,
and Mn in the column effluents increased markedly over the original leachate
concentrations. The increase in Ca was due to cation exchange with ions in
the leachate. The amount of Ca eluted from the columns was found, by mass
balances, to agree within 3% with the sum of Na, K, NIU , and Mg removed from
solution. The Mn increase probably resulted from a reduction of the oxidized
Mn on clay surfaces by the anaerobic leachate to more soluble ionic species.
Of the three clays used in the study, montmorillonite had the highest
attenuation capability, followed by illite and then kaolinite. This order
correlates well with the cation exchange capacities of the three clay min-
erals, which appear to be the dominant attenuation mechanism in these clays.
The principal attenuation mechanism for the heavy metals Pb, Cd, Hg, and Zn
was found to be precipitation, with resultant accumulation of the metals in
the surface layers of the columns.
A ranking system was developed for evaluating the relative pollution
hazard for municipal leachates. The new ranking method overcomes the problems
and objections found in evaluation of the critical product parameter for
JAuthors: R. A. Griffin, N. F. Shimp, J. D. Steele, R. R. Ruch, W. A. White,
and G. M. Hughes.
16
-------
municipal leachates by a method previously proposed to the U. S. EPA for
hazardous wastes.
Results of the study are applicable to the use of clay minerals as
liners for sanitary landfills and to the disposal of industrial and power
plant wastes in landfills and mines.
INTRODUCTION
The following results and discussion consider data from the columns
leached with normal leachate and do not include results from the sterilized
leachate. Details of the effect of sterile leachate on hydraulic conductivity
and chemical constituent attenuation will be the subject of Section 6. How-
ever, for those chemical constituents for which no significant difference in
attenuation between the normal and sterile leachate treatment was found, the
results were pooled to give added statistical significance to analysis of
clay-type effects. Those constituents for which no significant difference was
observed were Ca, Mg, Na, K, NHi,, Pb, Hg, Zn, and Cd.
COLUMN LEACHING STUDIES
Hydraulic conductivity and bulk density measurements of the column con-
tents are presented in Table 3. A wide range of hydraulic conductivities,
with values in agreement with those expected under field conditions from
similar materials (Todd, 1959), was observed. The relatively high bulk densi-
ties and slow flow rates used in this study closely simulate the conditions
observed in the field, which lends credence to the extrapolation of the re-
sults and conclusions presented here to field applications using clay liners
of similar composition.
Results of some column .effluent analyses are shown in Figures 2 and 3
plotted as relative concentration versus pore fraction. Relative concentra-
tion is the ratio of the column effluent concentration divided by the in-
fluent concentration. Thus the "breakthrough" point for a given element is
where the column effluent concentration equals the influent concentration and
has a value of one. A pore volume of effluent is defined as the volume neces-
sary to displace the volume of interstitial liquid in the pore spaces in the
column. The pore fraction is then given as the cumulative volume of column
effluent divided by the pore volume of the individual column.
Figures 2 and 3 illustrate the wide range of attenuation observed for
several elements contained in the leachate as it passed through columns con-
taining 2, 8, and 16% montmorillonite clay. The amount of reduction in con-
centration of a given element as it passes through the columns is reflected
by the shift of the curves toward higher pore volumes. The results shown in
Figure 2 for Cl, Na, NH4, and K are qualitatively in excellent agreement with
the results reported by Farquhar and Rovers (1975), who used soils in their
tests. The attenuation order also follows the general order of cation re-
placeability given by Grim (1968).
Figure 3 illustrates the negative attenuation or elution of Ca from the
columns. The relative concentrations greater than 1 indicate that Ca, and to
17
-------
1.20 i
0.00
0.0
4.0 6.0 8.0
Pore fraction
10.0
12.0
1.20-1
6.0 8.0
Pore fraction
1.20 i
0.00
2.0
4.0 6.0 8.0
Pore fraction
10.0
12.0
Figure 2. Relative column effluent concentrations for several elements as
a function of pore fraction of leachate passed through columns
containing (a) 2% montmorillonite, (b) 8% montmorillonite, and
(c) 16% montmorillonite clay.
18
-------
30,0-1
•£20.0-
«
o
o
o
> 10.0-1
a)
cr
0.0
2% Montmorillonite A
0.0
fr=T=-t — t—*—V
- Mrt
4.0 6.0 8.0
Pore fraction
10.0
12.0
8% moritmorillonite B
8.0 10.0 12.0
30.0 -i
0.0
0.0
16% Montmorillonite C
2-.0
4.0 6.0 8.0
Pore fraction
10.0
12.0
Figure 3. Relative column effluent concentrations for several elements as
a function of pore fraction of leachate passed through columns
containing (a) 2% montmorillonite, (b) 8% montmorillonite, and
(c) 16% montmorillonite clay.
19
-------
a lesser extent Fe and Mn, are eluting from the columns at much higher con-
centrations than the influent leachate at various pore fractions. The area
under the Ca curves in Figure 3 increases in proportion to the percentage of
clay in the column.
To quantify the observed attenuation, the area under each curve was in-
tegrated between pore fractions 1 and 11. The area between 0 and 1 pore frac-
tion was not included because it was merely the displacement of the deionized
water initially present in the column. The total area was that bounded by 10
pore volumes and relative concentrations between 0 and 1. The relative at-
tenuation number (ATN) was then obtained by subtracting the area under the
curve in each case from the total area and was expressed as a percentage. The
ATN numbers are unique for each element and each clay and express the rel-
ative mobilities of each element through each particular clay or clay mixture
column.
The attenuation number was computed for all columns and each chemical
constituent studied. The mean attenuation number for each chemical constit-
uent was used to rank the constituents according to their relative mobility
through the clay columns and is reported in Table 4. The constituents Al, Cu,
Ni, Cr, As, S, and POi* were found in such low concentrations in the DuPage
leachate that no attenuation order could be determined. The results showed
that greater amounts of the heavy metals Pb, Zn, Cd, and Hg than any other
element were removed from leachate, yielding an average of about 97% attenu-
ation for 10 pore volumes leached. The data for 2% montmorillonite (Fig. 2a)
show that Cd and Hg were the most mobile of the heavy metals studied.
Measurable amounts of Cd and Hg appeared in the effluents of the 2% clay
column after about 6 pore volumes were leached. Only traces of Cd and Hg ap-
peared in the effluents of the 8% and 16% clay treatments (Fig. 2b, 2c). Re-
movals of Pb and Zn were very high in all columns.
Results obtained from chemical analysis of sectioned columns revealed
large accumulations of all four heavy metals in the surface layers of each
column, including the 100% sand column. These removals could be attributed to
cation exchange replaceability, but a precipitation and/or filtration mechan-
ism appears a more plausible cause. Precipitation could involve formation of
heavy-metal hydroxides or carbonates, brought about by the relatively high pH
found in the column effluents. The average pH of the influent leachate was
6.9, while the average pH of the column effluents rose to values of 7.3-7.9.
The increase in pH could result in precipitation of the heavy metals from
solution in the columns. The precipitation mechanism is given further cre-
dence by the measurements of the effluent concentrations from the sand
columns, which have no cation exchange capacity and an average effluent pH of
7.8. No Pb and markedly reduced concentrations of Zn, Cd, and Hg eluted from
the sand columns. These data, along with the high accumulations of the metals
in the surface layers of the sand columns, indicate that precipitation of
heavy-metal hydroxides and carbonates can be an important attenuation mechan-
ism. This conclusion has been further verified by our equilibrium studies of
the effect of pH on heavy-metal removal by clay reported in Sections 7, 8, 9,
10, and 11 of this report.
Filtration of particulate material in the leachate by the earth
20
-------
TABLE 4. RANK OF CHEMICAL CONSTITUENTS IN MUNICIPAL LEACHATE
ACCORDING TO THEIR RELATIVE MOBILITY THROUGH CLAY
MINERAL COLUMNS
Chemical
constituent
Pb
Zn
Cd
Hg
Fe
Si
K
NH4
Mg
COD
Na
Cl
B
Mn
Ca
Mean
attenuation
number
99.8
97.2
97.0
96.8
58.4
54.7
38.2
37.1
29.3
21.3
15.4
10.7
-11.8
-95.4
-656.7
Principal
attenuation
mechanism
Precipitation/Exchange
Precipitation/Exchange
Precipitation/Exchange
Precipitation/Exchange
Oxidation-Reduction
—
Cation Exchange
Cation Exchange
Cation Exchange
Microbial Degradation
Cation Exchange
Dispersion
Artifact?
Elution from Clay
Exchanged from Clay
Relative
mobility
Low
Moderate
High
More Eluted
Than Applied
21
-------
materials is also a possible attenuation mechanism. Experiments in which the
DuPage leachate was filtered through 0.45 ym pore size membranes indicated
that only relatively small amounts of metals were retained on the membrane.
It was therefore concluded that filtration was not an important attenuation
mechanism for the heavy metals in this study but may be an important mechan-
ism for many other leachates.
Moderate attenuation was observed for the leachate constituents Fe, Si,
K, NHit, and Mg, which had values ranging between 58.4% and 29.3% attenuation.
Little attenuation was found for COD, Na, and Cl, values for which ranged be-
tween 21.3% and 10.7% attenuation. The elements Ca, Mn, and B were not at-
tenuated by the clays, but, rather, were found in substantially higher con-
centrations in the column effluents than in the influent leachate.
To determine what mechanisms were responsible for the observed differ-
ences in attenuation for each leachate constituent, the effect of clay type
was investigated. The mean attenuation obtained for each clay and chemical
constituent is tabulated in Table 5. The results show that no significant
difference in attenuation for the three clay minerals was observed for the
heavy metals (Pb, Cd, Hg, Fe, and Zn) or for B or Cl. Illite and kaolinite
attenuated Si significantly better than montmorillonite. Kaolinite was found
to elute Via in significantly higher amounts than either montmorillonite or
illite. No significant difference in COD attenuation between montmorillonite
and illite was found, and both attenuated COD significantly better than
kaolinite. Montmorillonite was found to attenuate the cations Na, K, NHi* , and
Mg and to elute Ca to a significantly greater degree than illite and kaolin-
ite.
A precipitation mechanism for the four heavy metals (Pb , Cd, Hg, and
Zn) , as discussed above, is consistent with the fact that no differences in
attenuation were found among the three clays. Because these four metals exist
in solution as cations, a significant clay-type effect would be expected if
cation exchange were the attenuation mechanism. No clay-type effect was ob-
served, which is taken as additional evidence that the primary attenuation
mechanism for these four heavy metals is precipitation.
To determine whether the higher attenuation of the cations Na, K,
and Mg and elution of Ca by montmorillonite was due to its higher cation ex-
change capacity, the attenuation numbers were plotted as a function of CEC
for the three clay minerals. The results are shown in Figures 4 and 5. The
very highly significant linear regression of attenuation numbers as a func-
tion of CEC led to the conclusion that the principal attenuation mechanism
for Na, K, NHij , and Mg was the cation exchange of these constituents for Ca.
Ca was the predominant exchangeable cation present on these clays at the
beginning of leaching (Table 1). To further confirm that this mechanism was
responsible, a mass balance for these five cations was computed for data
presented in Figure 2. The sums of the amounts of Na, K, NHi» , and Mg removed
from the leachate agreed within 3% with the amount of Ca eluted. Increases in
the concentration of alkaline earth metals in groundwater preceding a leach-
ate plume have been observed in the field during monitoring of the ground-
water chemistry around landfill sites. Such increases, termed the "hardness
22
-------
TABLE 5. MEAN ATTENUATION NUMBER (ATN) OF SOME CHEMICAL CONSTITUENTS
FOUND IN MUNICIPAL LEACHATES FOR THREE CLAI MINERALS
Mean ATN
Chemical
constituent
Pb
Zn
Cd
Hg
Fe
Si
K
NHt,
Mg
COD
Na
Cl
B
Mn
Ca
Montmorillonite
99.6*
97.7
96.7
98.4
34.8
39.2
58.9
54.8
48.2
24.6
20.6
9.3
-16.lt
-73.2
-885.5
Illite
100.0
98.6
100.0
98.1
82.8
81.6
31.0
31.0
19.7
23.2
16.4
13.5
-12.8
-6.4
-233.3
Kaolinite
99.9
98.1
97.5
95.2
67.6
71.2
23.2
25.1
18.1
16.2
9.7
14.3
-11.5
-266.2
-190.2
All
columns
99.8
97.2
97.0
96.8
58.4
54.7
38.2
37.1
29.3
21.3
15.4
10.7
-11.8
-95.4
-656.7
*Underlined means are not significantly different (0.05).
tMinus numbers indicate elution.
23
-------
K-ATN = I5.1 + 5.5 CEC
70-
10-
r=
' •
90
CEC
r = .
CEC
70-
20-
NH4+-ATN = I6.
r = .96
15 0
100
Mg-ATN=l0.4+5.l CEC
r = .95
80-
70-
60-
50-
0
15 0
40-
30-
20-
• Sand
A Illite
o Kaolinite
® Montmorillonite
CEC
a Clay mixtures
Figure 4. The attenuation number related to cation exchange capacity (a)
K, (b) NHlt, (c) Na, and (d) Mg.
24
-------
-2400
Ca-ATN = 48.8-i49.8 CEC
r=.97
200
-600
-500-
-400
o
o>
Hi
c
CEC
Mn-ATN = -50.6-l6.6
(% Kaolinite)
r = .95
Mn-ATN=-O.I-5.93
(% Montmorillonite)
r = ,95
• Sand
A Illite
20 30 40
Clay content (%)
O Kaolinite
® Montmorillonite
50
60
H Clay mixtures
Figure 5. (a) Ca attenuation number related to cation exchange capacity.
(b) Mn attenuation number related to clay percentage.
25
-------
halo," are likely due to the same mechanism responsible for the Ca elution
from the columns in this study. This is discussed further in Section 6.
No significant linear regression of attenuation as a function of either
CEC or clay percentage was obtained for Fe, Si, COD, Cl, B, or Mn, except for
a very highly significant linear regression obtained for Mn elution and clay
percentage. The results of the regression of Mn elution as the percentage of
kaolinite arid montmorillonite increased are presented in Figure 5b. No sig-
nificant linear regression could be obtained with illitic clay because only
two illite columns were used. The data illustrated in Figure 5b show that
approximately three times as much Mn eluted from the kaolinite-containing
columns as from the montmorillonite columns. The data presented In Table 1
indicate that kaolinite contains only slightly more total Mn than mdntmoril-
lonite. However, surface Mn is three times more abundant on kaolinite than on
montmorillonite and corresponds to the increased elution from the kaolinite
columns. No correlation of Mn elution was observed with CEC^ and the amounts
of surface-extractable Mn correlate with the amounts eluted from the various
columns. These facts, along with the anaerobic conditions in the columns,
have led to the conclusion that Mn elution is due to the reduction of surface
coatings of Mn compounds on the clays to more soluble reduced ionic species.
The increase in Mn elution from kaolinite columns is proportional to, and
apparently due to, the larger amount of Mn on the surface, where it is readi-
ly available for reduction and solubilization by the anaerobic leachate.
The behavior of Fe is similar to that of Mn; however, it is more sensi-
tive than Mn to the oxidation potential in the leachate. During the early
stages of leaching, Fe was also solubilized and eluted from the columns, as
is shown in Figure 3. However, unlike Mn, Fe showed a net attenuation of
58.4% when the entire 10 pore volumes leached were considered. The results of
this study, therefore, indicate that Fe may be either eluted or not attenu-
ated by clay liners if the leachate is strongly anaerobic, or it may be
strongly attenuated under weakly anaerobic conditions.
Other data from this study indicate that attenuation of COD was rel-
atively low after passage of DuPage leachate through clay columns. This re-
sult is in agreement with those of Urioste (1971) , who reported poor removal
of COD when leachate was ponded on soils. The lack of a strong clay effect
arid of a significant correlation of attenuation of COD with either clay per-
centage or CEC indicate that the observed attenuation was probably the result
of microbial degradation of the organic compounds. The fact that the leachate
is relatively old probably accounts for the relatively low reduction in COD.
It has only a small percentage of readily degradable organics and a low nu-
trient status, owing to the lack of PO^ (Table 2).
Chloride attenuation was also relatively low, only 10.7%. This low at-
tenuation is not surprising, because Cl is considered as a mobile non-inter-
acting anion in soil systems. The low Cl attenuation was not a function of
the type or amount of clay mineral present and is attributed to physical
dispersion in the porous column media, with perhaps a small amount of inter-
action at anion exchange sites on the clay, or to other chemical reactions.
Since Cl is a negative ion, it would be attracted only by the positive
26
-------
charge at the edge of the clay minerals. Oxygen and OH are the negative ions
of the lattice, and any other negative ion would have to be nearly the same
size as the oxygen ion in order to coordinate with and continue or substitute
for the oxygen. Because the chloride ion is about two and one half times the
volume of the oxygen ion, it is too large to replace or coordinate with the
oxygen and hydroxyl ions, although fluorine can do so because it is about the
same size as the oxygen and hydroxyl ions.
No attenuation of B was observed in this study; rather, a small elution
of boron from the columns took place throughout the leaching experiment. This
may be interpreted as a slight solubilization of B from the clay minerals or
sand in the columns. However, no effect of type or percentage of clay and no
changes in elution with leaching time were found during the study. These
facts have led to the conclusion that the B results may be an artifact of the
experiment. It has been suggested that B may be dissolving from the borosili-
cate glass tubing used throughout the apparatus. More importantly, B may be
solubilized from the spun glass wool used as a filter to keep sand and clay
from migrating into each of the five manometers and the outflow tubing (Fig.
1). The high surface area of the spun glass and the neutral pH of the leach-
ate make it plausible that the boron elution may be due to contamination. In
any case, no attenuation of boron was observed in this study.
The results presented in Figure 4 and Table 5 permit the three clay
minerals to be ranked, according to their overall attenuating ability for the
chemical constituents found in municipal leachate, as follows:
montmorillonite > illite > kaolinite.
The montmorillonite used in this study has properties similar to those
of smectites produced by weathering of micas, chlorites, and other crystal-
line minerals. The cation exchange capacity is lower than that for montmoril-
lonite produced from weathering of volcanic ash and basaltic rocks.
The montmorillonite used produces clay material (e.g., Porter's Creek
Clay) that have much higher permeabilities than those from Wyoming-type mont-
morillonite. Such high permeability should make this montmorillonite more
useful than some other types of montmorillonite in making landfill liners in
humid climates. With its high cation exchange capacity, it would adsorb more
of the cations in the leachate than either illite or kaolinite. With its
greater permeability, it allows more water to pass through the liner than do
many other types of montmorillonite. The low permeabilities of other mont-
morillonite types would probably increase the hazard of lateral seepage from
the sides of the landfill in humid climates.
Some sodium montmorillonites tend to' shrink when sodium is exchanged for
divalent and trivalent cations or when salt concentrations are high. This
shrinkage sets up tension, which in turn produces cracks called syneresis
cracks. Syneresis cracks were common in irrigation ditches in Colorado that
were lined with sodium montmorillonite (personal communication, R. D.
Dirmeyer, Jr., 1961). Shrinkage could be reduced considerably by using cal-
cium montmorillonites and mixing them with other earth materials (e.g., 16 to
32% montmorillonite and 68 to 84% sand).
27
-------
The kaolinite used in this study is a fine-grained material in which the
crystallinity of the kaolinite crystals is poor. It has a high cation ex-
change capacity compared with other kaolinites. Well crystallized kaolinites
with large crystals have a cation exchange capacity of 1 to 5 meq/100 g. The
kaolinite used in this study would be better suited than most kaolinites for
landfill liners. The permeability of this kaolinite is lower than that of
well crystallized kaolinite with large crystals. The kaolinite in most sedi-
ments has a cation exchange capacity and permeability between those of the
kaolinite used and the well crystallized kaolinites.
The illite used is similar in cation exchange capacity and permeability
to the illite found in most sediments. The sediment from which the experi-
mental illite was taken contained only one clay mineral — the illite —
whereas most sediments that contain illite also contain other clay minerals.
It was concluded that the attenuation order was due principally to the
cation exchange capacity of each of the three clays.
Attenuation of Leachate
During the period of time from collection of the leachate through the
establishment of hydraulic equilibrium, the leachate was stored in a re-
frigerated (3 to 5° C) condition with either argon or sterilant gas being
purged slowly over the top of each respective drum. Chemical analyses were
performed weekly on the leachate to monitor possible changes in composition.
As a result of this monitoring, it was determined that the COD of the sterile
drum was steadily rising. It was determined that the active component of the
sterilant gas, ethylene oxide, was able to react with the chloride ion pre-
sent in the leachate which acted as a nucleophile to produce ethylene chloro-
hydrin (Rosenkranz and Wibdkowski, 1974). When it was discovered that the COD
was rapidly rising, the use of the sterilant gas was discontinued. Argon was
then used as the purge gas for both drums and mercuric nitrate salt was
added to the sterile drum to maintain sterility. Further monitoring of both
drums was continued at approximately 1 week intervals during the 10 month
period during which leaching of the columns occurred. The COD values were
found to remain constant at the value obtained when purging of sterilant gas
was discontinued. The value reported in Table 2 for COD, and for all the
other constituents, is the average of the 37 separate analyses performed
during the 10 month period of leaching.
Determination of the attenuation of chloride and the other major com-
ponents of DuPage leachate has been described in Section 5. The results of
this study indicated that there was an average 6% greater attenuation of
chloride in the columns leached with sterile leachate than those leached with
the natural leachate. This greater attenuation is attributed to the reaction
of chloride with the ethylene oxide to form ethylene chlorohydrin. Other than
the slight increase in chloride attenuation, no other significant difference
was apparently due to the increase in COD in the sterile leachate as compared
to the natural leachate.
There were, however, other significant differences between the sterile
and natural leachate treatments which were not attributed to the higher COD
28
-------
of the sterile leachate. Figure 6 illustrates the difference observed in Mn
elution from the colums. It should be noted that a negative attenuation num-
ber indicates that more Mn eluted from the column than was present in the
influent leachate. It can be seen that much higher levels of Mn were found in
effluents from the columns leached with natural leachate. It was concluded in
Section 5 that this elution was due to reduction of surface coatings of Mn
oxides on the clays by the anaerobic leachates. This conclusion is further
verified by the difference in Mn elution between the natural and sterile
leachate treatment. This difference is attributed to the stronger anaerobic
environment provided by the active microorganisms present in natural leach-
ate. Inspection of the data in Table 2 shows that the average Eh (oxidation
potential) reading of the natural leachate was 10 times lower than the
sterile leachate, even though both were well in the anaerobic range (Eh read-
ings less than 197 m.v. are considered to reflect anaerobic conditions). A
similar result was obtained for Fe in that significantly (.05 level) greater
mobility of Fe was found in columns leached with natural leachate as compared
to those leached with sterile leachate. A mechanism similar to that for Mn is
postulated as the reason for the observed differences between the natural and
sterile leachate.
Those chemical constituents for which no significant difference in at-
tenuation between the normal and sterile leachate was found were Ca, Mg, Na,
K, NHit, Pb, Hg, Zn, and Cd.
POLLUTION HAZARD OF LEACHATE
In addition to determining the relative mobilities of the various chemi-
cal constituents of leachate through clays (liner materials) , their relative
pollution hazard should be evaluated. Ranking wastes in terms of their exist-
ing or potential threats to public health and/or the environment has been the
subject of the "Priority Ranking System" suggested for development by the
U.S. -EPA (1973).
The priority ranking formula is:
R = Q/CP {3}
where R = ranking factor ,
Q = annual production quantity for the waste being ranked, and
CP = critical product for the waste being ranked.
A critical product is the lowest concentration at which any of the
hazards of concern become manifest in a given environment, multiplied by an
index representative of the waste's mobility into that environment. Thus, for
a municipal leachate that would be discharged through a clay liner or soil
into an aquifer used for drinking water, the toxicity factor could be the
public water supply limits for the given element, and the mobility index
could be the attenuation numbers derived above.
Evaluation of the critical product for municipal leachates moving
through soils or clay liners by the U.S. -EPA (1973) formula proved to be awk-
ward and unsatisfactory for several reasons. The first problem encountered
29
-------
-600
0
O 2 4
8
16
Koolinite (%)
Figure 6. Manganese elution related to percentage of kaolinite leached
with natural and sterile leachate.
30
-------
was that negative attenuation numbers, such as those obtained for calcium,
were not accounted for by the formula. A transformation of the data would 'be
necessary to express the negative numbers in a manner that could be used in
the CP formulation.
The second problem encountered was evaluation of the CP for the four
heavy metals. An upper boundary on the attenuation number, such as the 100%
removal, used to express the heavy metal removed from leachate is not allowed
conceptually by the formula. Instead, what is required is the actual attenu-
ation as determined by leaching until "breakthrough" of the particular ele-
ment is achieved. For Pb , an estimated 300 pore volumes would have to be
leached through an average column to achieve breakthrough. It was deemed im-
practical to actually measure the breakthrough of the heavy metals, and esti-
mating their breakthrough required additional analytical data that could
cause errors in estimating the pollution hazard.
A third problem with evaluating the pollution hazard by using the CP was
that it was not specific for the waste being evaluated in that the concen-
tration of the element of interest in the waste did not enter into the evalu-
ation of the hazard. That this is a serious fault can be illustrated by
simple examples: an element with a relatively high toxicity and mobility in-
dex could get a high hazard rating, even though only a trace was present in
the waste; conversely, an element with a relatively low toxicity and mobility
could receive a low hazard rating even though it was present in very high
concentrations .
A fourth criticism of the CP rating was that it was conceptually il-
logical because large CP values indicated a low pollution hazard and, con-
versely, a very small number represented a very high hazard. It was con-
sidered more logical to express high pollution hazards as large numbers when
a relative scale was used or when pollution potentials were evaluated.
To overcome these objections to the CP formulation of a pollution hazard
index for municipal leachates, the ranking equation was changed as follows:
R = (Q) (HI) {4}
where R and Q are as previously defined and
HI = the pollution hazard index for the waste.
The pollution hazard index (HI) is a toxicity index for the element
within a given leachate, multiplied by a mobility index for the element in a
particular leachate-clay system. The pollution hazard for the whole leachate
is that for the constituent with the highest hazard within the particular
leachate .
HI = () (100 - ATN) {5}
where C = the effective concentration of the chemical constituent,
DWS = the drinking water standard (U.S. -EPA, 1972), and
ATN = the attenuation number for the given element.
31
-------
The effective concentration is defined as the concentration of the
chemical constituent in the leachate plus the concentration of the constitu-
ent that may be leached from the soil or clay. When attenuation is occurring,
the effective concentration is merely the concentration of the constituent in
the influent leachate. When elution from the columns is occurring, as it did
for the three elements B, Ca, and Mn, the effective concentration is the
leachate concentration plus the concentration eluted from the column. Table 6
presents the 15 chemical constituents for which ATN values are available,
ranked according to their pollution hazard, as determined by equation 5.
The results in Table 6 give a reasonable ranking of the chemical con-
stituents in terms of what would be expected from a gross overview of the
data. The ranking system has the advantage of quantifying the expected pollu-
tion hazard of a given leachate and allows comparisons of the pollution
hazards of one leachate with another. The ranking system also focuses atten-
tion on the chemical constituent with the highest pollution potential. In the
case of DuPage leachate, NffiJ" was found to have the highest pollution hazard.
In a fresh leachate, COD might be expected to have the highest pollution
potential. However, in the DuPage leachate the hazard index clearly indicates
that NHtj is a pollution hazard about 30 times greater than any other con-
stituent found in this particular leachate.
We feel that the proposed pollution hazard ranking system for municipal
leachates (equation 5) overcomes the objections posed previously for the CP
component of the Priority Ranking System. The toxicity index can in most
cases be computed readily from a chemical analysis of the leachate.
The evaluation of the toxicity index is flexible in that drinking water
standards need not be the criteria. LDso values, or some other toxicity eval-
uation, can be used in place of drinking water standards. What we think is
important is the computation of the ratio of the actual waste concentration
relative to whichever toxicity evaluator is used. The mobility index, how-
ever, must be determined experimentally or estimated from the data presented
in this paper. The results of this study indicate that the mobility index
will be a function of the CEC of the earth material, the cations present
initially on the exchange complex, the chemical composition of the leachate,
and the pH of the leachate.
32
-------
TABLE 6. CHEMICAL CONSTITUENTS IN DUPAGE LEACHATE, RANKED ACCORDING TO
POLLUTION HAZARD
Chemical
constituent
NHi,
B
COD
Hg
Cl
Ca
Cd
Fe
Na
Mn
K
Mg
Pb
Zn
Si
Effective concentration
D. W. standard
862/0.5
(29-9 + 3.5)/1.0
1340/50
0.87/0.002
3484/250
(46.8 + 307.3)/250*
1.95/0.01
4.2/0.3
748/270
(0.02 + 0.02)/0.05
501/250*
233/250*
4.46/0.05
18.8/5.0
14.9/250*
Toxicity
index
1724.
33.4
26.8
435.
13.9
1.42
195.
14.0
2.77
0.78
2.00
0.93
89.2
3.76
0.06
Mobility
index
62.9
111.8
78.7
3.2
89.3
756.7
3.0
41.6
84.6
195.4
61.8
70.7
0.2
2.8
45.3
Hazard
index
108,440.
3,734.
2,109.
1,392.
1,241.
1,072.
585.
582.
234.
153.
123.
65.7
17.8
10.5
2.7
*Actual value not established by EPA; therefore it was assumed to be -the
same as chloride.
33
-------
SECTION 6
CONFIRMATION OF LABORATORY COLUMN STUDIES
BY COMPARISON WITH FIELD DATA1
ABSTRACT
The results of the laboratory column leaching experiments were checked
at the DuPage County sanitary landfill and at other existing landfills where
detailed field data are available. These field data clearly show a "hardness
halo" corresponding to the Ca release observed in the column experiments. The
relative attenuation rates of some of the ions are also confirmed by the
field data.
Laboratory results show that the leachate reduced the hydraulic conduc-
tivity of the columns during the experiment. Although similar change in field
hydraulic conductivity was not clearly demonstrated, the field data suggest
that it took place.
These results suggest that overall pollution from landfill leachate
would be reduced by designing earth material landfill liners for higher per-
meability. Properly designed liners would selectively attenuate the toxic
pollutants from the leachate and allow the ground water to dilute the non-
toxic components which can be tolerated at much higher concentrations without
harmful effects. The study raises some basic questions on monitoring systems
and landfill design which should be addressed by regulatory agencies re-
sponsible for environmental quality.
INTRODUCTION
This paper principally relates two phenomena noted in the laboratory to
field observations around sanitary landfills: the elution of large amounts of
the calcium ion from the study columns (Fig. 3) and the reductions in hy-
draulic conductivities which resulted from the introduction of leachate to
the clay-sand mixtures. It is felt that the testing of laboratory results in
field situations is necessary before the data may be used in sanitary land-
fill design.
RESULTS AND DISCUSSION
Hardness Halo
Figure 3 (Section 5) illustrates the negative attenuation or elution of
Authors: K. Cartwright, R. A. Griffin, and R. H. Gilkeson.
34
-------
Ca, Fe, and Mn from the columns containing 2,8, and 16% montmorillonite clay
in sand. The relative concentrations greater than 1 indicate that Ca and to a
lesser extent Fe and Mn are eluting from the column at much greater concen-
tration than the influent leachate at various pore fractions. The area under
the Ca curves can be seen to increase in proportion to the percentage of clay
in the column, was quantified in Section 5 by integrating between pore frac-
tion 1 and 11, and was assigned a relative attenuation number (ATN) as shown
in Table 4 (Section 5).
The elution of Ca from the columns was attributed to an ion-exchange
mechanism; the replacing of the Ca bonded to the clays at their cation ex-
change positions by other ions in the leachate.
Soils in much of Illinois are carbonate rich, with the clays generally
having Ca in the cation exchange position, and have free carbonates in all
except the leached zone of the soils. The presence of excessive hardness in
the vicinity of sources of pollution has appeared in a number of articles,
but is rarely discussed as to its origin. An example is found in a DuPage
County study (Fig. 7, from Zeizel et al.,, 1962). There are two areas of the
county where the hardness, as CaCOs, in the shallow carbonate aquifer exceeds
1000 parts per million (ppm). The eastern area is a fairly heavily developed
residential area and the glacial drift, which protects the aquifer from pol-
lution, is relatively thin. No specific source of the high hardness can be
established; however, it is most likely due to a high concentration of home
septic systems. The western area of high hardness near West Chicago is
thought to have resulted from the discharge of large volumes of waste chemi-
cal salts to surface ponds.
Other examples can be found in Anderson and Dornbush's (1967) study of a
sanitary landfill in South Dakota, and Walker's (1969) discussion of ground-
water pollution in Illinois. Most recently, Henning et al. (1975) showed high
calcium in monitoring wells very close to a landfill trench at Mentor, Ohio;
the Ca concentrations both decreased with distance from the fill and were
lower in the refuse than in the closest wells.
Hughes et al. (1971) published the results of studies of five landfills
in northeastern Illinois, including the Old DuPage County landfill from which
the leachate was taken for this study. Monitoring continued for three years
following the completion of that report. Figure 8 was drawn, using data from
the Winnetka and Old DuPage landfills.
The Winnetka data (top, Fig. 8) shows considerable scatter. This may be
partly due to a mixture of points, some in the fine-grained alluvium and some
in the glacial till which have somewhat different properties. However, these
data suggest that the hardness approaches background within 9 to 15 meters of
the refuse which is somewhat less than the distance that the chloride ion
travelled (Hughes et al., 1971). Note that the four data points from piez-
ometer nest LW3 follow this pattern.
The till under the Old DuPage landfill clearly illustrates the increase,
then decrease in hardness. The till is separated from the refuse by 1 to 1.5
meters of sand. The hardness returns to background within about 1.5 meters of
35
-------
R 9 E
R 10 E
R II E
BLOOMINGDALE
WOOD DALEJ
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NAPERVILLE
HARDNESS (as CoC03)
I More than 1000
500-1000
300-500
Less than 300
0
5 6 Miles
Figure 7. Hardness of water, expressed as CaC03 , in the Silurian Dolomite
aquifer in DuPage County, northeastern Illinois (from Zeizel et
al., 1962).
•
-------
1500-
1000-
i
500-
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Average value \
of leachote V
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Distance (meters)
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•I (oldest part of fill)
/I
/ \
!-\
f X
/ :x^
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value of leachate
6 12 18 24 30 36 42
Distance (meters)
Figure 8. "Hardness halo" effect shown as a function of distance (m) from
the Winnetka and DuPage landfills in northeastern Illinois.
37
-------
travel, about half the distance of the estimated travel of the chloride ion
(Hughes et al., 1971). Note, in particular, the values shown for piezometer
nests LW5 and LW6. These data points are all for the younger, northern part
of the fill; data from the older parts of the fill do not fit the same curve
(all hardness concentrations fall too low).
The surficial sand transmits leachate-contaminated water south from the
older parts of the DuPage landfill and shows a similar hardness distribution.
The hardness levels are much lower in the old refuse in this area, and all
the values reflect that lower concentration. The hardness returns to back-
ground levels within 9 to 15 meters; however, chloride ion has moved 240 to
300 meters in the permeable sand layer.
All these data clearly show the presence of the "hardness halo" result-
ing from the movement of leachate into the surrounding till and sands, and
that the rate of the hardness front is less than that of the chloride ion.
The chloride ion is probably the best tracer of this type of pollution in
this environment. The distance of travel varied from slightly less to ap-
proximately 10% of that of the chloride ion; this probably is controlled by
the nature of the materials, cation exchange reactions, concentrations in the
leachate, and ground-water flow rates. Nevertheless, all the data show an in-
crease in the hardness of the water in the sediments over that in the leach-
ate.
Hydraulic Conductivity
The results of initial hydraulic conductivity and bulk density measure-
ments of the column contents in the laboratory study are presented in Table
3. These data indicate that a wide range of hydraulic conductivities, with
values in agreement with those expected under field conditions from similar
materials (Todd, 1959) were observed. The relatively high bulk densities and
slow flow rates used in this study closely simulate the conditions observed
in the field. This lends credence to the extrapolation of the results and
conclusions of the laboratory studies to those obtained in the field.
During the initial stage of the experiment, the columns were leached
with deionized water until steady state conditions were achieved. The columns
containing low percentages of clay reached hydraulic equilibrium relatively
rapidly while the high percentage clay columns required leaching for well
over a month to achieve steady manometer readings. When hydraulic equilibrium
was achieved, as indicated by steady flow rates and relatively constant man-
ometer readings from the five manometers located over the entire length of
each column, the leachate was added to the columns.
Significant reductions in hydraulic conductivity and significant dif-
ferences in the reductions between natural and sterile leachate were ob-
served. The results of hydraulic conductivity changes observed in columns
containing montmorillonite, kaolinite and illite clays are presented in
Figures 9, 10, and 11. The data presented in these figures were statistically
smoothed using the five-member moving-average method to show more clearly the
trends in the data. The raw data were statistically analyzed to determine
whether significant differences in hydraulic conductivity occurred between
38
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Leoching time (months)
Figure 10. Hydraulic conductivity of montmorillonite-sand columns as a
function of leaching time.
40
-------
~ 0
0
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in ~"
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Leaching time (months)
Figure 11. Hydraulic conductivity of illite-sand columns as a function
of leaching time.
-------
sterile and natural leachate. The data from the columns containing 4% kaolin
ite were rejected from the analysis when they were found to deviate by more
than 3 standard deviations from the overall mean change in hydraulic conduc-
tivity observed for all other columns. The manometer readings indicated that
the out-flow tubes were plugged. The reason why only 4% kaolinite treatments
had this problem is not clear.
The results of the statistical analysis indicated that columns leached
with natural leachate had significantly (.05 level) greater reductions in
hydraulic conductivity than those leached with sterile leachate. This result
is illustrated clearly in Figures 9, 10, and 11. Furthermore, the statistical
analysis showed that columns containing montmorillonite had significantly
greater average reductions in permeability than kaolinite or illite and that
there was no significant difference between kaolinite and illite. This result
is consistent with the swelling nature of montmorillonite clay and is not
surprising.
Field data to support this conclusion are not as clear as for the "hard-
ness halo." Fewer field tests have been made for hydraulic conductivities
than chemical tests for water quality. In addition, field tests may be ac-
curate only to approximately a half-order of magnitude.
At the Winnetka landfill (Hughes et al., 1971), 23 hydraulic conduc-
tivity tests were conducted, 7 on refuse, 4 on alluvium, and 7 on till. These
data are too scattered to show any significant differences with distance from
the refuse.
At the Old DuPage County landfill, 34 field hydraulic conductivity tests
were made, 14 on refuse, 14 on sand, and 6 on till. The data on the hydraulic
conductivity of the sand (all south of the fill) suggest some reduction in
hydraulic conductivity. The 10 tests made on monitoring wells less than 6
meters from the fill have a mean hydraulic conductivity of 4.32 x ID"1* cm/sec
(range 1.9 x 10~3 to 1.9 x 10~7), and those monitoring wells greater than 12
meters from the refuse (4 tests) have a mean conductivity of 2.59 x 10~3
cm/sec (range 7.6 x 10~3 to 9.5 x 10"1*). The data are not statistically sig-
nificant, but they do suggest a hydraulic conductivity reduction similar to
that noted in the laboratory.
The reductions in hydraulic conductivity observed in the laboratory
study with the DuPage leachate seem particularly significant. This is due to
the fact that the DuPage leachate is approximately 15 years old (Hughes et
al., 1971) and contains a relatively low percentage of organic compounds
which are readily degradable by microorganisms (Table 2). In addition, this
leachate has a low nutrient status with both phosphate and sulfate being ab-
sent in detectable quantities. Much higher amounts of microbial growth and
plugging might be expected from a younger leachate.
The results have led to the conclusion that if clay liners, natural or
man-made, of similar composition to those used in this study are used in
municipal landfills, significant reductions in hydraulic conductivity can be
expected due to microbial growth. Further, slightly higher reductions in hy-
draulic conductivity can be expected from montmorillonite clays, apparently
42
-------
due to their swelling tendency. It was also concluded that Mn and/or Fe may
be leached from clay surfaces in substantial amounts under highly anaerobic
conditions but may not be leached under mildly anaerobic conditions.
43
-------
SECTION 7
EFFECT OF pH ON EXCHANGE-ADSORPTION OR
PRECIPITATION OF LEAD FROM MUNICIPAL
LEACHATES BY CLAY MINERALS1
ABSTRACT
The capacity of kaolinite and montmorillonite clay minerals to remove Pb
from municipal landfill leachates and the mechanisms by which removal is
achieved were studied to evaluate the potential usefulness of clay minerals
as liners for waste disposal sites under conditions of varying pH and ionic
competition.
»
Montmorillonite was found to remove as much as 5 times more Pb from
various solutions than did kaolinite. Results indicated that Pb removal was
reduced as much as 85% by leachate when compared to the amounts removed from
pure Pb(N03)2 solutions. A precipitate was found to form in leachates at pH
values above 5 and was identified as PbC03. The complexing capacity of DuPage
leachate for Pb was measured and the extent of complexation was found to be
11%. The higher-ionic-strength Blackwell leachate had no measurable com-
plexing capacity for Pb. Increased adsorption of Pb(NOa)2 was found to cor-
respond to the appearance of Pb-hydroxyl species in solution.
It was concluded that Pb removal from solution is primarily a cation ex-
change-adsorption reaction that is affected by pH and ionic competition. It
was also concluded that formation of Pb-organic complexes are of secondary
importance in landfill leachates due to competition from high concentrations
of other cations. At pH values above 6, a large increase in Pb removal from
solution by clay can be expected, due to either increased adsorption of Pb-
hydroxyl complexes or formation of PbCOa in landfill leachates.
The thickness of clay liners necessary to remove Pb from solutions of
PbCNOs), 0.1 M NaCl, and landfill leachates at concentrations ranging between
10 and 1000 ppm Pb and at pH values from 3 to 8 were computed. Some undesir-
able environmental consequences of the reversible Pb exchange-adsorption re-
action with clay may ensue where pH and ionic, competition are unfavorable.
INTRODUCTION
This section reports the results of an investigation, the purpose of
which was to determine the capacity of the two major clay mineral types for
removing Pb from solution and the effect municipal leachates have on this
capacity at various pH values. Another purpose of the investigation was to
JAuthors: R. A. Griffin and N. F. Shimp.
44
-------
gain insight into the mechanisms responsible for attenuation of Pb as well as
to evaluate the potential use of clay minerals as liners for waste disposal
sites.
Lead was chosen for study because documented evidence shows that low Pb
levels in drinking water can cause death to humans. In one case in Australia,
94 adults died of chronic lead poisoning because throughout childhood they
drank water collected from roofs of houses painted with lead-pigmented paint
(Henderson, 1955). There is also evidence (Broadbent and Ott, 1957) that or-
ganic chelates form with hydrolyzable metals and that they may make Pb more
mobile in soils or clay liners when municipal leachates are present than in
effluents that do not contain high concentrations of organic compounds.
EXPERIMENTAL
-j
The Pb removal studies were conducted by placing a known weight of clay,
between 0.100 and 1.000 g, into a 125 ml Erlenmeyer flask. The weight of clay
used was chosen to give an estimated 20 to 50% change in the Pb concentration
of the solution at equilibrium. A 50 ml aliquot of either deionized water,
0.1 M NaCl solution, or leachate and then a 2 ml aliquot of a Pb(NOa)2 solu-
tion were pipetted into the flask. The pH of the solutions were adjusted with
either HN03 or NaOH over the pH range of interest, and the volumes of acid or
base added were recorded. The volumes added were usually less than 1 ml. The
flasks were tightly stoppered, and as a result the COa liberated from the
leachate solutions caused a slight positive pressure in the flask, which
aided in maintaining anaerobic conditions during equilibration. (For a period
of time after addition of acid, the stoppers were removed to relieve exces-
sive pressure, and then the flasks were restoppered.) The results of rate
studies indicated that 4 hours were necessary for Pb in leachate to equili-
brate with kaolinite. This result is in agreement with Beevers (1966), who
found that Pb(NOa)2 solutions equilibrated in from 1 to 12 hours, depending
on the clay mineral. The samples in this study were shaken for at least 24
hours in a constant temperature bath at 25 ± 0.5° C to insure equilibration.
The equilibrium pH was recorded, the' samples were centrifuged, and the solu-
tions were analyzed for their Pb concentration by atomic absorption. The dif-
ference between the initial concentration and the equilibrium concentration
was used to compute the amount of Pb removed from the solution at the par-
ticular pH by a given clay mineral. This procedure was Carried out for a
range of initial Pb concentrations that varied between 10 and 1,000 ppm.
The resulting data were plotted as amount of Pb removed from solution
per gram of clay versus pH. The resulting family of curves allowed construc-
tion of adsorption isotherms for any individual pH value of interest through-
out the pH range 3 to 6. Representative adsorption isotherms and plots of the
data according to the Langmuir (1918) adsorption equation were constructed at
various pH values to determine the maximum amount of Pb that could be ad-
sorbed from the various solutions by the two clay minerals and to try to gain
insight into the mechanisms of adsorption. The Langmuir (1918) equation in
its linear form is:
-C_ = -L + ± {6}
x/m kb b
45
-------
where C is the equilibrium concentration of Pb, x/m is the amount of Pb ad-
sorbed per unit mass of clay, 1/b is the slope where b is the adsorption
maximum for Pb, and k equals the slope/intercept where k is a term relating
to the energy of adsorption.
The ability of the two municipal leachates to complex metal ions, in
particular Pb, was studied. The leachates were centrifuged at 1000 rpm for 10
minutes and then filtered through a 0.45 \m pore-size Millipore membrane held
in an anaerobic bacteria filter holder under argon pressure. The leachate ob-
tained after centrifugation and filtration was considered to contain only
soluble organics and was used in the complexation studies. Successive ali-
quots of Pb were added to the leachate and equilibrated for several hours.
The concentration of free and complexed lead ions in solutions were deter-
mined from pulse polarographic wave heights using the methods and equipment
described by Gadde and Laitinen (1973a).
Lead was removed from leachate solutions as a white precipitate at pH
values greater than 6. The precipitate was separated from solution on a 0.45
lam Millipore membrane, washed with deionized water and dried at room tempera-
ture. The chemical compound was then identified from its X-ray diffraction
pattern.
RESULTS
The results of Pb removal from 25° C solutions of DuPage leachate by
kaolinite and montmorillonite clay minerals were plotted as a function of pH
in Figures 12 and 13, respectively. Similar results obtained for Pb removal
from Blackwell leachate by kaolinite are shown in Figure 14.
The data presented indicate that Pb removal from landfill leachate in-
creases with increasing pH values and with increasing concentration of Pb in
solution. Increasing Pb concentration is indicated on the figures by increas-
ing alphabetical order, and the initial Pb concentration, weight of clay
used, micrograms of Pb added, and volume of solution that correspond to the
alphabetical designations are given in Table 7. A blank (no clay) solution of
leachate with Pb added was carried along through the experiment and the re-
sults of these also appear on the figures.
Data for Pb removal from leachate plotted as a family of curves of in-
creasing concentration have the advantage that sorption isotherms may be con-
structed from the plot, by using the information given in Table 7, for any
desired pH value from pH 3 to pH 6.
Pb sorption isotherms can be constructed from these plots by first
placing a vertical line across the family of curves at the pH of interest.
The amount of Pb removed from solution is found on the graph at the points
where the vertical pH line intersects each curve. The equilibrium Pb concen-
trations that correspond to the chosen pH value are then computed from the
amounts of Pb removed at each concentration, as determined from the graph and
the information for each Pb curve given in Table 7 by using the following re-
lation:
46
-------
7.500
7.000
Figure 12. The amount of Pb removed from DuPage leachate by kaolinite at
25° C plotted as a function of pH.
47
-------
6.000
Figure 13. The amount of Pb removed from DuPage leachate by montmoril^
lonite at 25° C plotted as a function of pH.
48
-------
5.000-
5.5 6.0 6.5 7.0
2.5
Figure 14. The amount of Pb removed from Blackwell leachate by kaolinite
at 25° C plotted as a function of pH.
49
-------
TABLE 7. Pb REMOVAL PARAMETERS USED TO COMPUTE SORPTION ISOTHERMS FROM
52 ml REACTION VOLUMES
Curve
DuPage leachate -
kaolinite (Fig. 2)
DuPage leachate -
montmorillonite
Blackwell leachate -
kaolinite (Fig. 4)
A
B
C
D
E
F
G
H
I
J
A
B
C
D
E
F
G
H
I
J
A
B
C
D
E
F
Initial Pb
concentration Pb added
(ppm) (micrograms)
9.62
19.23
38.46
57.69
76.92
96.15
115.38
153.85
192.31
384 . 62
9.62
19.23
38.46
57.69
76.92
96.15
115.38
153.85
192.31
384.62
19.23
38.46
76.92
96.15
192.31
288.46
500
1,000
2;000
3,000
4,000
5,000
6,000
8,000
10,000
20,000
500
1,000
2,000
3,000
4,000
5,000
6,000
8,000
10,000
20,000
1,000
2,000
4,000
5,000
10,000
15,000
Clay weight
(grains)
0.500
0.500
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
0.500
0.500
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
1.000
i.ooo
50
-------
Equil. Pb = Pb added (Ug) - (Pb removed (u/g) x wt. clav
sample volume
Pb sorption curves were constructed by this method for several pH values
from the plots given in Figures 12, 13, and 14. Representative curves are
presented in Figure 15 along with sorption isotherms obtained at pH 5.0 and
25° C for Pb sorption from Pb(N03)2 solutions and 0.1 M NaCl solutions, fol-
lowed by the DuPage and Blackwell leachates, respectively.
The sorption curves for the two leachates show a sharp upswing occurring
at equilibrium concentrations of approximately 200 ppm Pb. Qualitatively
identical curves, also with a sharp upswing at about 200 ppm Pb, were ob-
tained by using montmorillonite clay. The data presented in Figure 13 may be
used to verify this finding. Sorption isotherms computed from Figures 12, 13,
and 14 at pH 3.0 or 4.0 did not exhibit the sharp upswing. A sharp upswing in
a sorption isotherm at higher concentrations is generally viewed as initia-
tion of precipitation of an insoluble compound. The fact that the sharp rise
in Pb removal occurred at about 200 ppm Pb under the COa partial pressures in
the flasks, while at pH 4.0 no sharp increase was observed, is consistent
with solubility computation which assume a mechanism of PbCOs formation.
To predict the maximum amounts of Pb that could be sorbed by the two
clays from the various solutions, the kaolinite sorption data plotted in
Figure 15 and the sorption data obtained for montmorillonite were plotted ac-
cording to the Langmuir adsorption isotherm equation. The results are illus-
trated in Figure 16.
The Langmuir equation was found to describe the data obtained for
Pb(1103)2 sorption by both kaolinite and montmorillonite and for Pb sorption
from 0.1 M NaCl solutions by kaolinite over the entire concentration range
studied. The results obtained for the leachates, however, were somewhat dif-
ferent. The Langmuir equation was found to describe the sorption up to con-
centrations between 30 and 40 ppm, at which point a sharp change in slope
occurred, giving two distinct linear regions to the isotherms. The adsorption
maximums computed from the slopes of the lines shown in Figure 16 are given
in Table 8. From the values in Table 8 a quantitative estimate of the sorp-
tion differences noted in Figure 15 and a comparison of the sorption capaci-
ties of the two clays can be made. In Pb(NOs)2 solution, montmorillonite
sorbed approximately five times more Pb than kaolinite, while in DuPage
leachate it sorbed less than twice as much as kaolinite. This result indi-
cates that the competitive ions contained in landfill leachate affect the
relative sorption affinity of Pb for the montmorillonite; i.e., leachate re-
duced Pb sorption by montmorillonite proportionately more than it reduced Pb
sorption by kaolinite.
DISCUSSION
The data presented above suggest that several mechanisms are responsible
for removal of Pb from solutions of varying ionic composition and pH.
Precipitation was found to be an important mechanism in landfill leach-
ates, as is shown by the removal of Pb from the blank solutions, which con-
51
-------
16.0
0 25 50 100
200 300
Equilibrium Pb concentration (ppm)
400
500
Figure 15. The amount of Pb sorbed per gram of kaolinite at pH 5.0 and
25° C plotted as a function of the equilibrium Pb concentra-
tion.
52
-------
0.12
0.11-
0.10-
0.09-
0.08
0.07
Koolinite in Blackwell leachate
Kaolinite in
Du Page leachate
Kaolinite in
O.I M NaCI
Montmorillonite in
Du Page leachate
0 25 50
200 300 400
Equilibrium Pb concentration (ppm)
500
Figure 16- Pb sorption data for kaolinite and montmorillonite at pH 5.0
and 25° C plotted according to the Langmuir isotherm equation.
53
-------
TABLE 8. MAXIMUM REMOVAL OF Pb FROM pH 5.0 and 25 C SOLUTIONS
COMPUTED USING THE LANGMUIR EQUATION
KAOLINITE
Pb(N03)2
0.1 M NaCl
DuPage leachate
Blackwell leachate
MONTMORILLONITE
Pb(N03)a
DuPage leachate
Micrograms/g
Region 1 Region 2
15,914
10,240
1,680 8,530
986 2,401
82,428
1,811 11,133
Meq Pb+^/lOO g clay
Region 1 Region 2
15.36
9.88
1.62 8.23
0.95 2.32
79.56
1.75 10.75
54
-------
tained no clay (Figures 12, 13, and 14). Losses of Pb from the DuPage leach-
ate were observed at pH values greater than about 6 and in Blackwell leachate
at pH values above 5. A white precipitate was observed forming in the leach-
ate solutions at pH values greater than 6. It was filtered out, and the
chemical compound was identified by its X-ray diffraction pattern as a highly
crystalline PbC03. The peaks were sharply defined, and no peaks other than
those attributed to PbC03 were observed. This is offered as evidence that
PbC03 formation was the compound responsible for Pb removal from leachate
solutions at the higher pH values, and it is presumed to be the cause of the
apparent formation of a precipitate at concentrations of Pb greater than 200
ppm observed in the sorption isotherms (Figure 15).
Stumm and Morgan (1962) showed that the occurrence of metal hydroxyl
species can affect the sorption of hydrolizable metal ions. They found that
the pH at which metal hydroxyl species formed corresponded to the pH at which
metal ion sorption became significant. To check the role of hydrolysis of the
Pb ion on its sorption by clay at various pH values and in solutions where
precipitate formation did not effect Pb removal, the distribution of various
hydroxyl species in a 4 x 10"^ M Pb(N03)a solution in the pH range 3 to 8
was obtained from Gadde and Laitinen (1973b), who computed the species dis-
tribution by using the constants given by Olin (1960). The distribution of
Pb hydroxyl species, along with data obtained for sorption by kaolinite from
a solution with an initial concentration of 4 x 10""* M Pb as Pb(N03)2, is
illustrated in Figure 17.
It is evident from the plots in Figure 17 that species other than Pb4"4"
are relatively insignificant (<1%) at pH values less than 6. Sorption at pH
values below 6 are not related to the hydroxyl species of Pb, but rather to
Pb4"*" ion. The decrease in Pb sorption at low pH values is apparently due to
an increase in competition for sorption sites, with H4" and its related com-
petitive effects on Pb sorption caused by the dissolution of AT*"1 ' ions from
the clay crystal lattice (Grim, 1968). At pH values above 6, a sharp rise in
Pb sorption occurred coincident with the formation of hydroxyl Pb species.
It therefore seems likely that at least a portion of the observed increase
in Pb sorption with decreasing pH and the rapid increase in the amount of Pb
sorbed coincident to the formation of monovalent Pb-hydroxyl species are
consistent with a cation exchange mechanism for Pb removal from solution by
clay minerals. Table 8 exhibits further evidence that cation exchange is the
principal mechanism for Pb removal by clay minerals.
The sorption maximums for Pb(N03)2 solutions, computed from the slope
of the Langmuir plots, for kaolinite and montmorillonite are 15.36 and 79.56
meq Pb++/100 g clay, respectively. These values can be compared to the cation
exchange capacity (CEC) values of 15.1 for kaolinite and 79.5 meq/100 g for
montmorillonite that were determined by the ammonium acetate method and re-
ported in Table 1. The CEC values are within 2% of the Pb sorption maximums
computed from the Langmuir equation; i.e., Pb++ sorption is merely another
method of measuring the cation exchange capacity of a clay.
Further evidence of a cation exchange mechanism is the reduction in Pb
sorption in solutions containing 0.1 M NaCl and also as the total salt con-
tent of the two leachates studied increases. The decrease in Pb sorption is
55
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Figure 17. Distribution of Pb (II) species in 4 x 10 ^ M Pb(N03)2 and
uptake by 0.1 g kaolinite from 60 ml of solution.
56
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attributed to increasing competition for cation exchange sites by Na+ in the
0.1 M NaCl solutions and to an increase in the divalent cation competition in
the two leachates. For example, the Blackwell leachate contains much more Fe
in solution than the DuPage leachate (Table 2) . Such high levels of com-
peting ions could account for the large reductions in Pb sorption observed in
the presence of landfill leachate compared to the sorption in pure Pb(N03)2
solutions.
The results for Pb sorption from leachate, plotted according to the
Langmuir adsorption equation, show a distinct two-slope character. This shape
of curve has been attributed to adsorption at sites of distinctly different
energy (Griffin and Burau, 1974; Griffin and Jurinak, 1973). However, in the
present study competition from other cations in solution seems to be re-
sponsible for the change in slope, because it was observed only in the multi-
component cation systems. Solutions of Pb(N03)2 or NaCl did hot exhibit the
sharp change in slope for Pb sorption that was observed for the leachates.
One of the assumptions of the Langmuir equation is that the adsorbent
surface is homogeneous with respect to the energy of the adsorption sites.
However, in a multicomponent cation system the sites are occupied by cations
with various retention energies relative to Pb; i.e., Pb can displace certain
cations, such as Na+, much more easily than it can replace cations such as
Ca"1"1". This reaction is postulated to affect the shape of the adsorption iso-
therm by filling the lower energy sites preferentially; i.e., Pb first ex-
changes with a cation, or a group of cations, of similar exchange energy.
This phase of the sorption is attributed to the initial slope of the Lang-
muir plot. As the concentration of Pb in solution is increased, the chemical
potential gradient is increased until it is sufficient to initiate exchange
of the cation, or group of cations, with the next highest energy of reten-
tion relative to Pb. This second energy level of exchange is postulated to
produce the sharp change in slope of the Langmuir plots in the leachate solu-
tions.
An explanation, other than competition for cation exchange sites, for
the observed reductions in over-all Pb sorption is the tendency of Pb to
form metal-organic complexes with the organic compounds present in landfill
leachate. These organic-metal complexes can lower the activity of the Pb in
solution, thus reducing the chemical potential gradient for sorption. Gadde
and Laitinen (1973a) showed that Pb forms stable complexes with organic com-
pounds found in soils and that these compounds were able to solubilize Pb
present in different forms in the solid phase.
To determine the role of Pb-organic complexes in the observed Pb sorp-
tion reductions in leachates, the complexing capacity of Pb in the two
leachates was measured.
In the DuPage leachate, the extent of complexation (ppm complexed) was.
found to be 22 ppm at a Pb concentration of 200 ppm. It is clear from the
above data that only about 11% Pb is complexed, while it would take more than
50% complexation to explain the reduction in Pb sorption by an organic com-
plexing mechanism.
57
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The fact that formation of a Pb-organic complex cannot be used to ex-
plain more than a small fraction of the observed reduction in Pb sorption is
emphasized by the results obtained from using Blackwell leachate. In this
study it was found that addition of successive aliquots of Pb to the leachate
gave approximately the same incremental response (yA-current) in the wave for
free lead ion. Apparently up to 80 ppm total Pb, no complexation of Pb was
observed. It was noted that Fe+ or its weak complex with leachate is pre-
sumed to have produced the large increase in polarographic current observed at
a potential ~1.4 V. From the results of this study, it appears that Pb"*"1" is
either unable to compete with Fe"1"4" or other cations present in large amounts
in Blackwell leachate, or the leachate has little or no complexing capacity.
The former explanation appears to be the more plausible.
The results of the above studies have led to the conclusion that Pb re-
moval from solution is primarily an exchange-adsorption reaction that is af-
fected by pH and ionic competition. The formation of Pb-organic complexes was
concluded to be of secondary importance in landfill leachates due to competi-
tion from high concentrations of other cations. At pH values above 6, removal
of Pb from solution by clay can be expected to increase substantially, owing
either to increased adsorption of Pb-hydroxyl complexes or to formation of
PbC03 in landfill leachates.
Disposal Site Design Application
An example of how the data from the study can be used is its application
to the question posed at the beginning of the paper — how thick a proposed
clay liner must be to remove all Pb from landfill leachates, from industrial
waste streams of similar ionic strength to the leachates (0.1 M NaCI), or
Pb(NOs)2 solutions at various pH values and Pb concentrations. The results of
the computations are presented in Table 9.
The table gives the thickness of a square meter of a 30% clay liner,
packed to a bulk density of 1.60 glee., that contains enough clay to remove
all the Pb from 762 liters (201 gal) of solution. This particular volume is
the amount generated from a typical sanitary landfill containing municipal
solid waste placed 3 meters (10 feet) deep and having an annual net infil-
tration of 254 mm (10 inches) (U.S.-EPA, 1974). The thicknesses of the clay
liner given in the table, therefore, effect total removal of Pb for a year
by a square meter of liner at the given concentrations of Pb and pH values.
They are, of course, the minimum thickness possible since they represent an
idealized situation. The actual thickness necessary in a field application
will be somewhat greater to allow for non-equilibrium conditions, physical
dispersion, diffusion, and the normal engineering safety factors.
The information compiled in Table 9 indicates that only relatively thin
layers of clay, especially montmorillonite, are necessary for removal of Pb
unless the pH values are very acid and the Pb concentrations are high. The
high sorption capacity of clay minerals and the reversible nature of exchange
adsorption reactions have important environmental consequences. Soils and
surface waters may change in ionic composition or pH as environmental con-
ditions change. A sudden decrease in pH may release large amounts of poten-
tially toxic Pb into the aqueous phase, especially in places where PbC03 has
58
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TABLE 9. THICKNESS (cm) OF A SQUARE METER OF A 30% CLAY LINER
NEEDED TO REMOVE Pb FROM 762 LITERS (201 gallons) OF
SOLUTION PER YEAR
Pb Concentration
KAOLINITE
Pb(N03)2
0 . 1 M NaCl
DuPage
Blackwell
MONTMORILLONITE
Pb(N03)2
DuPage
10 ppm 100 ppm
at pH at pH
3 58358
<1 — 5.3 1.8 <1
<1 — — 2.3
15.9 2.1 * 28.9 6.4 *
19.8 4.0 * 49.6 11.3 *
<1 - — <1
9.9 1.8 * 13.2 3.7 *
1000 ppm
at pH
358
10.0
15.5
79.4 * *
264.6 * *
1.93 —
18.0 * *
*Precipitation as PbC03
59
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accumulated. Cations, especially di- and tri-valent, compete with Pb and may
exchange with it, thus allowing Pb to come into solution. These multiple
interactions must be considered when a disposal site is designed and the
environmental impact of Pb assessed.
60
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SECTION 8
EXCHANGE-ADSORPTION OF Cu, Zn, AM) Cd
FROM DEIONIZED WATER AND LEACHATE
SOLUTIONS BY CLAY MINERALS1
ABSTRACT
The effect of pH on the removal (exchange-adsorption by kaolinite and
montmorillonite clay minerals plus precipitation) of copper, zinc, and
cadmium in deionized water (pH range 4 to 6) and municipal leachate (pH range
2 to 8) solutions was studied. Solutions contained up to 1000 ppm Cu, Zn, or
Cd. Families of removal versus pH curves were obtained that can be used to
construct removal isotherms for specific pH values. With certain exceptions,
very significant increases in the amounts removed from both solutions were
observed as the pH rose in the pH ranges studied. Precipitation contributes
significantly to removal of Cu from leachate above pH 5, Zn above 7, and Cd
above 6.
Three different pH 5.0 removal isotherms are presented for deionized
water solutions. These isotherms were constructed from removal curves ob-
tained by different experimental methods which were shown to influence the
interpretation of the results. The differences in the three isotherms are re-
solved by use of a Langmuir-type isotherm equation that was derived to ex-
press the simultaneous competitive adsorption of two cations. The equation
reveals that, under certain conditions, the amount of exchange adsorption of
a cation should be independent of its solution concentration. This proved to
be true for the exchange adsorption of Cu, Zn, and Cd from deionized water
solutions by the clay minerals at pH 5.0.
Isotherms of leachate removal at pH 5.0 were constructed and compared
with the deionized water isotherms for the same pH. The amount of removal at
pH 5.0 from leachate is significantly lower than it is from deionized water
solutions because of competition from the other cations present in the leach-
ate. Competition prevents the amount of exchange-adsorption at low concentra-
tions from becoming independent of concentration. The leachate isotherms are
specific for the leachate used and may not approximate the exchange-adsorp-
tion from another leachate of different ionic strength and composition.
The mobility of Cu, Zn, and Cd in soils or clay minerals is dependent
upon the pH and ionic strength of the solution as well as on the CEC of the
soils or clay minerals. The CEC value is of little importance at sufficiently
1Authors: R. R. Frost and R. A. Griffin
61
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high (above 7) pH values because precipitation is then more important than
cation exchange in the removal of Cu, Zn, or Cd from solution.
INTRODUCTION
The results of chemical analyses of 20 municipal leachates have been re-
ported by the U. S. Environmental Protection Agency (EPA, 1974). These analy-
ses showed that leachate was similar to sewage sludge effluent with respect
to its high content of organic matter, nitrogen, phosphorus, and potassium
and also indicated levels as high as 10 ppm for Cu, 1000 ppm for Zn, and 17
ppm for Cd. The accumulation of these potentially toxic heavy metals in soils
makes long-term application of municipal leachates to the land hazardous be-
cause Cu, Zn, and Cd enter the human food chain by accumulating in plants.
However, a lack of basic data on the reactions of these metal ions with soil
colloids in the presence of a complex solution matrix, such as leachate,
hampers efforts to determine what levels of application can be tolerated
without degradation or loss of the soil resource to food crop production.
The original purpose of this study was to measure the adsorption of Cu,
Zn, and Cd by kaolinite and montmorillonite clay minerals from a landfill
leachate at pH 5.0. However, the study was later expanded to investigate the
adsorption of low concentrations of heavy metal ions by earth materials from
deionized water solutions and their removal (adsorption plus precipitation)
from a leachate in the pH range 2.0 to 8.0. The results of this study give
insights into the mobility of Cu, Zn, and Cd in soils irrigated with leach-
ates and can be applied to the design of clay liners for municipal and in-
dustrial waste disposal sites.
THEORETICAL
In any adsorption study, the amount adsorbed is usually measured as a
function of adsorbate concentration in the medium surrounding the adsorbent.
It is generally desirable to be able to fit the adsorption data to an adsorp-
tion isotherm equation so that "parameters" associated with the adsorption
isotherm can be calculated for comparisons and correlation with other data.
The Langmuir (1918) equation is given as:
/ KbC <•_•!
x/m = TTKC {7}
where x/m = amount of adsorbate adsorbed per gram of adsorbent, C = the equi-
librium ion concentration in solution, b = the adsorption maximum, and K = a
constant related to the bonding-energy of the adsorbate to the adsorbent.
This equation has been used extensively in studies of adsorption of ions from
solution by soils and clay minerals. Eq. {7} can be rearranged into a linear
form, where:
C/(x/m) = I/(Kb) + C/b. {8}
The application of Eq. {8} to experimental data for Zn adsorption by
soils (Shuman, 1975) has produced two linear portions of the plot. Following
62
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the lead of Syers et al. (1973), Shuman (1975) attributed this result to two
types of adsorption sites. However, the adsorption model upon which Eq. {7}
is based assumes that the surface of the adsorbent contains only one type of
adsorption site. Therefore, there is no justification for arguing the exis-
tence of two sites on an adsorbent surface when a one-site model equation has
been applied to the experimental data. Also, inspection of Eq. {7} shows
that, at low concentrations, it reduces to:
x/m = KbC. {9}
Therefore, Eq. {8} has no significance at low concentrations, even
though C/(x/m) values can be calculated and plotted. If a two-site adsorption
model is to be discussed, then the following equation should be applied to
the experimental data:
where « - fraction of sites with bonding-energy coefficient KI . In practice,
Eq. {10} would be almost impossible to apply to experimental data.
The linear form of the Langmuir equation (Eq. {8}) holds best for the
plateau region of the adsorption isotherm (high concentrations) . This is true
regardless of the equilibrium concentrations because the deviations from the
linear plot occur as the equilibrium concentration approaches zero. A close
look at Eq. {8} confirms this point. At, or near, the plateau region of the
adsorption isotherm, (x/m) is about constant and, hence, the real plot is
(C/constant) versus C, which must be a straight line. But at low concentra-
tions, where (x/m) is increasing as C is increasing, values of C/(x/m) can
change at a faster rate than they do at higher concentrations at which C is
the only variable changing to an appreciable extent. Therefore, all Langmuir
plots (Eq. {8}) will probably show two or more straightline segments if data
points are obtained at sufficiently low and high concentrations.
The most important point ignored by the application of Eq. {7} to ad-
sorption of cations from solution by soils and clay minerals is that the ad-
sorption process is primarily one of cation exchange, and for every cation
adsorbed one or more cations must be desorbed. The latter' can then compete
for adsorption sites. Boyd, Shubert , and Adamson (1947) developed an adsorp-
tion equation for the simultaneous competitive adsorption of two equally
charged cations, A and B. By formal analogy with the Langmuir (1918) equation
for adsorption from a binary gaseous mixture the following equation is ob-
tained:
b K /K^ (C./C )
f i \ , _ A D A r>
(x/m)A = ! +
The linear form of Eq. {ll} is
63
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CA/CB = _S_ 1\ {12}
(x/m)A KA b b CB
Where the concentration of A is small compared to the concentration of B, so
that adsorption of A does not produce a detectable change in the concentra-
tion of B (e.g., adsorption on a Ca-saturated clay from solutions with high
Ca content), a plot of C /(x/m) versus C should be linear. The plot should
also be linear when (x/mJA approaches the cation exchange capacity (CEC), be-
cause Cg would then become a constant value and would not change as CA
changes. From Eq. {ll} it can be seen that the amount adsorbed, (X/III)A» must
depend upon the ratio of the equilibrium concentrations of the exchanging
cations and not upon the actual concentrations in solution. Where CB»CA,
(x/m)A will become a linear function of CA, and when CA»CB , (x/m)A will be-
come a constant and independent of the concentration of A.
Equation {11} demonstrates that, under appropriate experimental condi-
tions, the amount of exchange-adsorption should be independent of the solu-
tion concentration of the cation adsorbed. If it is desirable to study the
migration of a cation through soils or clays by measuring its adsorption from
a pure solution, then a fixed weight of soil sample and a fixed solution
volume should be maintained throughout the concentration range being studied.
However, it is not sufficient to measure the exchange-adsorption from pure
cation solutions. For example, to determine how far the Cu, Zn, or Cd in 250
ml of 200 ppm deionized water solutions will migrate down a clay column the
following procedure should be followed. One gram of clay is placed in the
250 ml of solution (initial concentration, C); adsorption will occur to give
an equilibrium concentration, Cj. In the process, (200-Ci) x 0.25 mg of
cation is adsorbed and an equivalent number of moles of exchangeable cations
on the clay will be desorbed. Hence, the solution phase will now contain a
mixture of cations from which the amount of exchange-adsorption measured from
a pure solution with a concentration of Ci. This process should be carried
out stepwise, C to Ci, Ci to C2, .. . , Cn-l to Cn, until Cn -*- 0. Thus, the
amount of adsorption from solution concentration GI will be dependent upon
the initial concentration of the solution from which Ci was derived.
EXPERIMENTAL
After pH adjustment, all clay-leachate or clay-deionized water suspen-
sions in this study were shaken in a constant temperature bath at 25 ± 0.5°
for at least 24 hours to insure complete equilibration. The equilibrium pH
values of the clay suspensions were measured, the clay suspensions centri-
fuged, and the supernatant solutions were analyzed by atomic absorption
spectroscopy for their Cu, Zn, or Cd concentrations. Blanks (i.e., no clay)
of spiked leachate or deionized water solutions that had been prepared along
with the clay suspensions also were analyzed to determine the initial Cu, Zn,
or Cd concentrations present in the solutions. The amount of Cu, Zn, or Cd
removed from solution by a given clay at a particular pH was calculated as
the (initial equilibrium concentration) x (solution volume/sample weight) /
1000. The amount of Cu, Zn, or Cd removed from solution was plotted as a
function of pH.
64
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Three types of experiments were conducted on the removal of Cu, Zn and
Cd by the clay minerals from deionized water solutions of Cu, Zn or Cd'ni-
trates. In the first, 1.000 g of kaolinite or montmorillonite and 50-ml ali-
quot s of the appropriate nitrate solution were placed into 125-ml Erlenmeyer
flasks. The approximate concentrations chosen were: 25, 50, 100, and 200 ppm
Cu and 10, 40, 200, and 400 ppm Zn for both kaolinite and montmorillonite;
100, 200, and 500 ppm Cd for kaolinite; and 100, 500, and 1000 ppm Cd for
montmorillonite.
Three or four replicate clay suspensions were prepared for each concen-
tration of Cu, Zn, or Cd. The pH values of the replicate clay suspensions
were adjusted to various values in the pH range 4 to 6 with dilute HN03 or
dilute NaOH solutions. The use of NaOH was avoided if possible because Na
ions can compete for adsorption sites.
In the second type of experiment, the weight of the clay sample and/or
the volume of solution were chosen so that the total amount of Cu, Zn, or Cd
in solution per gram of clay would be a constant. The following quantities
were used: about 12.5 mg of Cu/g of kaolinite (i.e., 250 ml at 25 ppm Cu/
0.5g) at 10, 25, 125, 200, and 500 ppm Cu; about 62.5 mg of Cu/g of mont-
morillonite (i.e., 250 ml at 25 ppm Cu/O.lOg) at 25, 125, and 500 ppm Cu;
about 20.0 mg of Zn/g of kaolinite (i.e., 50 ml at 40 ppm Zn/0.10g) at 10,
40, 200, and 400 ppm Zn; about 100 mg of Zn/g of montmorillonite (i.e., 250
ml at 40 ppm Zn/O.lOg) at 10, 40, 200, and 400 ppm Zn; about 25.0 mg Cd/g of
kaolinite (i.e., 50 ml at 50 ppm Cd/O.lOg) at 20, 50, 200, and 500 ppm Cd;
about 125.0 mg Cd/g of montmorillonite (i.e., 250 ml at 50 ppm Cd/O.lOg) at
50, 100, 200, and 1000 ppm Cd. Three or four replicate suspensions were pre-
pared for each concentration of Cu, Zn, or Cd. The pH values of the replicate
clay suspensions were adjusted to various values in the pH range 4 to 6.
In the third type of experiment, stepwise removal of Cu, Zn, or Cd from
solution was studied. Some of the second type of experiment that used 250-ml
solutions were taken as the first step in the third type of experiment. For
the second step, the supernatant solutions from replicate clay suspensions in
the first step were sampled for analysis and then mixed together. The com-
bined solution was used with fresh clay samples to prepare three replicate
clay suspensions which were adjusted to various pH values in the pH range 4
to 6. The supernatant solutions from the second step were used in a third
step, etc.
Experiments on the removal of Cu, Zn, or Cd from leachate were carried
out by pipeting 50-ml aliquots of leachate into 125-ml Erlenmeyer flasks.
Either the leachate had been spiked with Cu, Zn, or Cd nitrates to give the
desired concentration prior to the previous step, or 2.0 ml of an appropriate
stock solution was pipeted into the flasks containing the 50-ml aliquots of
non-spiked leachate. Several replicates were prepared for each concentration
of Cu, Zn, or Cd used. The pH values of the replicate spiked leachate solu-
tions were adjusted to various values throughout the pH range 2 to 8. Appro-
priately sized samples of either kaolinite or montmorillonite,were then
weighed out and transferred to the flasks. The weight of clay used was chosen
so that the amount of Cu, Zn, or Cd removed from the leachate solutions could
be determined with some precision from the difference between the initial and
65
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final solution concentrations.
It was discovered during preliminary experiments that, when the pH of a
clay-leachate suspension was adjusted to a particular value, the pH of the
suspension would rise on shaking and that the removal of Cu, Zn, or Cd in-
creased with increasing pH. Thus, addition of the clay to the spiked leachate
solutions after pH adjustments have been made avoids such potential problems
as the possible irreversible removal of Cu, Zn, or Cd. The irreversible re-
moval could occur if the pH of the prepared clay-spiked leachate suspension
was high and the pH of the suspension was then adjusted to a much lower
value. The experimental procedure used should produce true equilibrium re-
moval data.
Several individual experiments also were carried out in which clay
samples were placed in a mixture of 50-ml aliquots of pH 5 leachate or de-
ionized water and 2-ml aliquots of different stock solutions. The pH of the
resulting suspensions was repeatedly adjusted to 5.0 until equilibrium was
obtained.
RESULTS AND DISCUSSION
The amount of Cu, Zn, or Cd removed from deionized water solutions was
plotted versus pH. Except in certain cases, very significant increases in the
amounts removed were observed as the pH rose in the range 4 to 6. The initial
concentration of Cu, Zn, or Cd in solution, the weight of clay used, the
final solution volume after pH adjustments, and the removal versus pH curves
themselves can be used to construct "adsorption" or "removal" (adsorption
plus precipitation) isotherms at different pH values by use of the following
equation:
•Equilibrium concentration (ppm) = initial concentration (ppm)
Amount removed (mg/g) x wt. clay (g) x 1000. r. •,
Final solution volume (ml)
Isotherms were calculated from the data recorded for the deionized water
solutions at pH 5.0, and the results are shown as isotherm types I, II, and
III in Figures 18 and 19.
The type I isotherms (Figs. 18 and 19) are those obtained when 1.00 g of
kaolinite or montmorillonite was placed in about 50 ml of Cu, Zn, or Cd solu-
tions. The total amounts of Cu, Zn, or Cd present in 50 ml of solution at
different concentrations are given in Table 10. For kaolinite (CEC 15.1 meq/
100 g), 75.5 ymoles of a divalent cation/g would be required for complete ex-
change; for montmorillonite (CEC 79-5 meq/100 g), 397.5 ymoles/g would be re-
quired for complete exchange. Thus, as indicated in Table 10, at low concen-
trations of Cu, Zn, or Cd, insufficient cations are present in 50 ml of solu-
tion for complete exchange of 1.00 g of clay (especially montmorillonite).
The amount of Cu, Zn, or Cd that can be adsorbed is therefore necessarily
limited by the number of Cu, Zn, or Cd ions initially present in solution.
The number of Cu, Zn, or Cd ions actually adsorbed will depend on how well
the desorbing Ca ions compete with the Cu, Zn, or Cd ions remaining in solu-
tion. The type I isotherms in Figures 18 and 19 specifically represent the
66
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KAOLINITE
200 300 400 500 600 700
Equilibrium concentration (ppm)
r
800
900
Figure 18. The amount of Cu, Zn, or Cd removed from solution per gram of
kaolinite at pH 5.0 and 25° C, plotted as a function of the
equilibrium concentration. Curve I - fixed weight clay/fixed
solution volume; Curve II - total amount of clay exchangeable
ions/total amount of Cu, Zn, or Cd ions in solution held about
constant; Curve III - stepwise isotherms; Curve IV - DuPage
leachate isotherms. Open symbol data points were obtained
from clay suspensions adjusted several times to pH 5.0 instead
of being calculated from removal curves at pH 5.0.
67
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MONTMORILLONITE
32-
28-
24-
20-
~ 16-
= 8-
o
in
4-
E -
o 18-
-o 16-
0)
8H
< 0-
Zn
100
200 300 400 500 600 700
Equilibrium concentration (ppm)
800
Figure 19. The amount of Cu, Zn, or Cd removed from solution per gram of
montmorillonite at pH 5.0 and 25° C, plotted as a function of
the equilibrium concentration. The labels I - IV and the open
symbols have same meaning as in Fig. 18.
68
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TABLE 10. TOTAL CONTENT OF Cu, Zn, OR Cd IN 50 ml OF SOLUTION
Initial concentration
in solution Cu Zn Cd
ppm
ymoles in 50 ml solution volumes
4
10
25
40
50
100
200
400
500
1000
3.15
7.87
19.7
31.5
39.4
78.7
157.4
314.8
393.4
786.9
3.06
7.15
17.9
30.6
38.2
76.5
153.0
306.0
382.5
765.0
1.78
4.45
11.1
17.8
22.2
44.5
89-0
178.0
222.5
445.0
69
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amount of Cu, Zn, or Cd adsorption by 1.00 g of clay from about 50 ml of
solution.
The type II isotherms in Figures 18 and 19 were obtained when the weight
of clay samples and/or solution volumes were adjusted so that the ratios
(total number of Cu, Zn, or Cd ions)/(number of exchangeable cations on the
clay samples) were held about constant (at least 2/1) and independent of the
concentrations in the initial solution. Although, at equilibrium, the equi-
librium concentration of Cu, Zn, or Cd (CA) could not be considered to be
much greater than the concentration of desorbing Ca ions (Cg) , the ratio of
CA/CB should be a constant and, hence, Eq. {11} says that (x/m)A should be
constant and independent of CA- This is shown by the type II isotherms in
Figures 18 and 19. The "scatter" in the experimental data points for the type
II isotherms is attributed to experimental errors due primarily to problems
in measuring accurately the small changes in concentration that are due to
adsorption. These errors must then be multiplied by large factors to compute
the amount of metal removed per gram of clay.
The type I and II isotherms plotted according to the linear form
(Eq. {8}) of the Langmuir equation (Eq. {?}) are shown in Figures 20 and 21.
The Langmuir plots for the type I isotherms (1.00 g clay/50-ml solution) show
a definite two-slope character (except the plot for Zn-kaolinite) ; in.fact,
three linear segments of the Langmuir plot appear to exist for the Cu-mont-
morillonite plot. The Langmuir plots (Fig. 21) of the type II isotherms for
montmorillonite (Fig. 19; the total amount of Cu, Zn, or Cd initially present
in solution /g clay was held about constant) show only a one-slope character
throughout the concentration range investigated. The C/(x/m) values defi-
nitely approach zero as the equilibrium concentration approaches zero. This
is precisely what should occur if the competitive Langmuir equation
(Eq. {ll}) is valid, because Eq. {11} says that, if the equilibrium ratio
CA/CB is constant, (X/IH)A is independent of the actual concentration, and,
therefore, as CA approaches zero, CA/(X/III)A must approach zero.
The Langmuir plots (Fig. 20) for the Cu and Cd type II isotherms with
kaolinite (Fig. 18) do show a two-slope character. It was observed during the
Cu adsorption experiments that the pH of the Cu-kaolinite suspensions in de-
ionized water decreased when the flask was shaken, thereby indicating that
hydrolysis of the Cu"1"2 ion was occurring in solution (i.e. , Cu+2 + H20 =
CuOH+ + H+, etc.). Data on hydrolysis of cations that were assembled by Mes-
mer and Baes (1974) show that no appreciable concentrations of CuOH4" or
Cua(OH)2+2 will exist in solutions of low Cu concentration below pH 7, but,
in the pH range 5 to 6, precipitation can occur from solutions of about 400
ppm Cu. Thus, precipitation is a reasonable explanation for the occurrence of
the second slope in the Langmuir plot for Cu-kaolinite. Although similar ar-
guments could be advanced for the Cd-kaolinite Langmuir plot, it appears
from Mesmer and Baes (1974) that, in the range of Cd-concentrations used,
hydrolysis and precipitation cannot be considered to be contributing factors
to the "adsorption" of Cd by kaolinite around pH 5.0.
Shuman (1975) obtained two-slope Langmuir plots when he plotted Zn ad-
sorption by Georgia soils data according to Eq. {8}. John (1972) found no
significant correlation between the CEC values and Cd Langmuir adsorption
70
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KAOLINITE
0 100200300 0 100200300 0 100200300400
Equilibrium concentration (ppm)
Figure 20. Cu, Zn, and Cd removal data for kaolinite in deionized water
solutions at pH 5-0 and 25° C, plotted according to the
Langmuir equation (Eq. 8). The roman numerals beside the
plots indicate the corresponding isotherm in Fig. 18.
71
-------
Montmorillonite
100 200 300 0 100 200 300 0 100 200 300 400 500 600 700
Equilibrium concentration (ppm)
Figure 21. Cu, Zn, and Cd removal data for montmorillonite in deionized
water solutions at pH 5.0 and 25° C, plotted according to the
Langmuir equation. Numerals indicate corresponding isotherms
in Fig. 19.
72
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maximums for 30 different soil samples. Both Shuman (1974) and John (1972)
used a fixed weight of soil and a fixed volume of solution in their adsorp-
tion measurements. The results of this study indicate that both Shuman (1974)
and John (1972) would have obtained somewhat different results and conclu-
sions if they had equilibrated their soil samples with sufficient solution
volumes of low-Zn or Cd concentration so that the total Zn or Cd content of
the solutions would exceed the CEC values of their respective soil samples.
The appropriate solution volumes will necessarily depend on the CEC values of
the soil samples (the larger the CEC, the larger the solution volume re-
quired) and what kind of exchangeable cations are present on the soil sam-
ples.
The type III isotherms in Figures 18 and 19 are "stepwise" isotherms.
These isotherms show the competitive effect of the desorbing exchangeable
cations initially present on the clay minerals on the removal of other heavy
metal cations from solution. Thus, the type I and II isotherms in Figures 18
and 19 do not provide sufficient information to predict the migration of Cu,
Zn, or Cd in pure solutions through soils or clays.
The adsorption data plotted by using the linear form (Eq. {12}) of the
competitive Langmuir equation (Eq. {11}) are shown in Figures 22 and 23. It
was assumed that the number of ions (presumed to be all Ca, which is more
reasonable for kaolinite than for montmorillonite) desorbed from the clays
into solution equals the amount of Cu, Zn, or Cd adsorbed from solution. It
can be shown that CA (ppm)/ACA (ppm) = CA(moles/1) /Cca(moles/1) , where CA =
equilibrium concentration of Cu, Zn, or Cd and ACA = ^initial - CA, so that
(CA/ACA)/(x/m)A versus (CA/ACA) was plotted in Figures 22 and 23. For step-
wise isotherms, the ACA used included the ACA values from previous steps. Al-
though there is appreciable scatter to the data, reasonably straight lines
(except for Cd-kaolinite) , can be drawn through most of the data points,
which include all data plotted for deionized-water solution isotherms I, II,
and III in Figures 18 and 19. Some data points lie above the line drawn
through the data points. At these data points, C/AC is generally small and
the number of ions adsorbed (x/m) is limited by the total number of ions
initially present in solution. Negative deviations from the main linear
region can occur when Eq. {8} is used at low concentrations. Positive devi-
ations from the main linear region can occur when Eq. {12} is used for small
values of C/AC.
The adsorption maximums calculated from the type II Langmuir plots in
Figures 20 and 21 and the competitive-Langmuir plots in Figures 22 and 23 are
given in Table 11.
The question arises as to why there is such poor correlation between the
adsorption maximums calculated from the CEC values and the experimental ad-
sorption maximums if cation exchange is the primary adsorption mechanism.
Bittel and Miller (1974) have determined the selectivity coefficients for
Cd/Ca on kaolinite and montmorillonite to be about one (i.e., Ca and Cd ions
compete on about an equal basis for adsorption sites) , but they do not
specify a pH for their experimental measurements. The data obtained in this
study indicate that the selectivity coefficients for Cu, Zn, and Cd on Ca-
saturated kaolinite and montmorillonite are less than one at pH 5 and ap-
73
-------
1.0-
0.5-
0.0
•g 2.0 H
x
~ 1.5-
3 1-0-
G 0.5-
^
o o.o
1.5-
1.0-
0.5-
0.0
Cd
Zn
Cu
KAOLINITE
o"
-*r~**
0.0
1.0
2.0 3.0 4.0
C/AC
5.0
6.0
Figure 22. Cu, Zn, and Cd removal data for kaolinite in deionized water
solutions at pH 5.0 and 25° C, plotted according to the com-
petitive Langmuir equation (Eq. 12). The data point symbols
correspond to those used in Fig. 18.
74
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MONTMORILLONITE
o.o
Figure 23. Cu, Zn, and Cd removal data for montmorillonite in deionized
water solutions at pH 5.0 and 25° C, plotted according to the
competitive Langmuir equation (Eq. 12). The data point symbols
correspond to those used in Fig. 19-
75
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TABLE 11. COMPARISON OF LANGMUIR ADSORPTION MAXIMUMS IN DEIONIZED
WATER WITH CEC VALUES
Cu Zn Cd
Source K* Mt KM KM
mg/g
II - Langmuir plots 3.08 18.5 2.86 19.2 5.0 31.2
(Figs. 21 and 22)
Langmuir plots 3.33 21.0 3.00 23.6 - 50.0
(Figs. 23 and 24)
Calc. from CEC 4.80 25.3 4.94 26.0 8.45 44.8
values - Table 1
*Kaolinite
tMontmorillonite
76
-------
proach one at hzgher pH. Therefore, the experimental adsorption maximums for
Cu, Zn, or Cd would not approach the CEC values until the equilibrium ratio
C/CCa was larger than was the case in any of our experiments. Because the ad-
sorptxon of Cu, Zn, or Cd is pH-dependent , the adsorption maximums will of
course, show definite pH-dependence. Prior to the onset of precipitation
which is dependent on both concentration and PH, decreasing competition from
IT" and increasing concentration of hydrolyzed ions (e.g., Cu2(OHO+2 ZnOH+
and CdOlT) are possible reasons for the "adsorption" to increase as pH in- '
creases.
Curves for pH versus the removal of Cu, Zn, or Cd from leachate by kaol-
inite and montmorillonite are shown in Figures 24 and 25, respectively. The
data listed beside each curve and in the figure captions were used in Eq. {13}
to calculate the type IV isotherms at pH 5.0, which are shown in Figures 18
and 19. This was done so that direct comparisons could be made with the de-
ionized-water isotherms. Figures 18 and 19 show that the removal from leach-
ate is appreciably lower than the removal from deionized-water solutions.
Boyd, Shubert, and Adamson (1947) gave a general expression for the exchange
adsorption of a cation, A, from a mixture of equally charged cations in solu-
tion, which is as follows:
b K. C.
\ A A
Equation {14} says that the exchange adsorption of any one cationic
species at constant concentration will decrease as the concentrations of
other cationic species increase. Thus, from solutions of constant ionic
strength, in which the total ionic content is large compared to the total
number of exchangeable cation on any soil or clay mineral sample placed in
that solution, the amount of Cu, Zn, or Cd removal cannot approach the CEC of
the soil or clay mineral at low concentrations, as is true in the case of ex-
change-adsorption from deionized water (i.e., increasing solution volume will
usually not affect the equilibrium concentration ratio of the cation of
interest to the other cations in solution to any appreciable extent). Only
when the concentrations of Cu, Zn, or Cd in solution exceed the combined con-
centrations of all other cationic species in solution will the amount of ex-
change-adsorption of Cu, Zn, or Cd approach the CEC of the soil or clay min-
eral sample being tested. In practice, except at very low pH, precipitation
of Cu, Zn, Cd and other heavy metal ions as hydroxides and/or carbonates will
occur at moderate concentrations so that precipitation rather than cation ex-
change adsorption can become the principal mechanism for the removal of heavy
metal ions from solution.
The sharply rising portions of the removal curves shown in Figures 24
and 25 can easily be interpreted, with the aid of information assembled by
Mesmer and Baes (1974), as being caused primarily by precipitation of Cu, Zn,
or Cd carbonates, hydroxides, or hydroxide-carbonates. Removal of Zn and Cd
from the leachate by both kaolinite and montmorillonite is .greatly reduced
compared to their removal from deionized water solutions at pH 5.0. The
amount of Zn and Cd removed is reduced proportionally about the same for both
77
-------
KAOLINITE
25.0-
21.0-
17.0-
13.0-
t
n
» 9.0-
»
; 5.0-
•
•
i i.o^
•
' 0.9-
2 0.8 H
o> 0.7 H
o
E 0.6 H
0)
v_
*- 0.5 H
c
0.4 H
0.3-
0.2-
0.1-
0.0
Cu
1000
192
52.5ml
887
/
52.5ml
I ' I ' I ' I
2468
PH
Cd
8
Figure 24. The amount of Cu, Zn, or Cd removed from DuPage leachate
solutions by kaolinite at 25° C, plotted as a function of
pH. The plots are labeled with the initial solution con-
centration (ppm) of Cu, Zn, or Cd from which each plot was
obtained. Unless otherwise indicated, 2.0 g of clay and a
total leachate solution volume of 50.5 ml were used to
obtain each data point.
78
-------
MONTMORILLONITE
Figure 25. The amount of Cu, Zn, or Cd removed from DuPage leachate
solutions by montmorillonite at 25° C, plotted as a function
of pH. The plots are labeled with the initial solution con-
centration (ppm) of Cu, Zn, or Cd from which each plot was
obtained. Unless otherwise indicated, 0.5 g of clay and a
total leachate solution volume of 50.5 ml were used to obtain
each data point.
79
-------
kaolinite and montmorillonite, but the reduction in the amount of Cu removed
from the leachate is appreciably greater for montmorillonite than for kaolin-
ite. This can readily be explained on the basis that the Cu-leachate iso-
therms represents a combination of exchange-adsorption and precipitation. The
amount of Cu removed by precipitation, for example 1 mg/g, will be about the
same for leachate solutions at pH 5, whether kaolinite or montmorillonite is
present in solution. However, the 1 mg/g represents about 20% of the CEC for
kaolinite, while it represents only about 4% of the CEC for montmorillonite.
The decrease in removal from leachate, therefore, appears much greater for
montmorillonite than for kaolinite. The actual decrease in Cu exchange-ad-
sorption is probably proportionally the same for both clays. The removal of
Cu from leachate reaches a maximum about pH 7 (Figs. 24 and 25) and then de-
creases for pH values above 7. This behavior is due to the amphoteric charac-
ter of Cu(OH)2 precipitates, which redissolve in basic solutions by forming
Cu(OH)^, etc. ions.
The amount of Cu, Zn, or Cd removed from leachate has no significant ap-
parent dependence on leachate volume, as can be seen from the leachate iso-
therms (Figs. 18 and 19), on which data points obtained for different clay
sample weights but constant leachate volumes have been plotted on one iso-
therm curve. However, the amount removed could become dependent on the leach-
ate volume if the clay sample is very large.
Equation {14} implies that at some given ionic strength, where
all. j Cj
is about constant and large compared to one, (x/m)^ = Constant x C^. The iso-
therms for Cd (Figs. 18 and 19) are linear to quite high concentrations.
The isotherms for Cu, Zn, or Cd removal from leachate were not plotted
according to the linear form of a Langmuir adsorption isotherm equation, al-
though, most assuredly, straight line plots would have been obtained and
"adsorption" maximums could be calculated. The reasons for not plotting them
were, first, none of the leachate isotherm plots have really reached the
plateau region in the concentration range studied (up to 1000 ppm) . Second,
any adsorption maximums calculated from the Langmuir plots would be somewhat
meaningless because the amount of exchange-adsorption from leachate is
limited by competition, owing to the high ionic strength of the leachate and
not because all the adsorption sites have been occupied by Cu, Zn, or Cd ions,
as is implied with a Langmuir adsorption maximum. Third, if an adsorption
maximum is calculated from the rising part of an isotherm, it represents the
amount adsorbed at some higher hypothetical concentration that may or may not
lie on the real adsorption isotherm. Finally, our purpose was to determine
the maximum amount of Cu, Zn, or Cd that can be removed by kaolinite or mont-
morillonite from leachate at any particular concentration up to approximately
1000 ppm. As we have seen for removal from deionized water, with appropriate
experimental conditions, the amount removed is independent of concentration
and is proportional to the CEC values of the clays. It would presumably ap-
proach the CEC at higher (volume)/(weight of clay) ratios. But, with leach-
80
-------
ate, the ionic strength is not a variable to be adjusted and the conditions
cannot be created where at low concentrations of Cu, Zn, or Cd the ratio of
the (equilibrium concentration of Cu, Zn, or Cd)/(concentration of other
cations) is large. The maximum amount removed must then be a function of the
ionic strength of the leachate, the CEC of, the clay sample, and pH of the
leachate. Therefore, the maximum amount removed at any concentration and pH
from leachate (or for that matter any solution) is simply the value read from
the removal isotherm itself at the concentration of interest.
SUMMARY AND CONCLUSIONS
Under appropriate experimental conditions, the amount of Cu, Zn, or Cd
exchange-adsorbed from deionized water by purified kaolinite and montmoril-
lonite clay minerals is independent of the equilibrium concentration of Cu,
Zn, or Cd. However, the maximum amount of Cu, Zn, or Cd adsorbed in our ex-
periments was related to, but not equal to, the CEC values of the clay min-
erals, probably because the desorbing Ca ions effectively competed with the
Cu, Zn, or Cd ions present in solution. If cation-exchange adsorption experi-
ments are carried out at constant (solution volume)/(sample weight) ratios
for Cu, Zn, or Cd in deionized water, the amount of adsorption is necessarily
limited by the total amount of Cu, Zn, or Cd that was initially present in
solution at low concentrations. Therefore, the isotherm obtained is really a
plot of the amount of Cu, Zn, or Cd removed by a fixed weight of sample from
a fixed volume of solution versus concentration. At low concentrations a dif-
ferent isotherm can be obtained simply by using a different constant (solu-
tion volume)/(sample weight) ratio in the experiments. To properly simulate
field conditions, soil samples must be equilibrated with sufficient solution
volumes so that the total metal ion content of the solutions exceeds the ad-
sorption capacity (CEC values) of the respective soil samples. The appropri-
ate solution volumes will necessarily depend on the CEC of the soils and the
kind of exchangeable cations present on the soil samples.
If it is desirable to study the migration of Cu, Zn, or Cd from de-
ionized water through soils or clays by .means of batch experiments, the ex-
periments can be carried out stepwise (i.e., repeated treatments of a solu-
tion with new soil or clay samples) for each initial solution concentration
tested. That is necessary because of the increasing concentration of ex-
changeable cations in solution as exchange-adsorption of an individual cation
from solution occurs stepwise towards zero. This is not a particular problem
in high ionic strength solutions such as landfill leachate as discussed be-
low.
The three different types of isotherms obtained for the same range of
Cu, Zn, or Cd concentration are easily interpreted in terms of a cation ex-
change-adsorption mechanism. Also, the correct "Langmuir" isotherm equation
to apply to exchange adsorption data is Eq. {11}, which covers the simultane-
ous competitive adsorption of two cations, and not Eq. {7>, which is for a
single cation. Although strict application of Eq. {11} requires the exchange-
able cations on the adsorbent to be homoionic, the principal conclusion that
"under appropriate conditions" the amount of exchange-adsorption is indepen-
dent of concentration will apply, regardless of the number of exchangeable
cations on the adsorbent.
81
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The adsorption of Cu, Zn, or Cd from leachate by kaolinite and mont-
morillonite clay minerals is presumed to occur via a cation exchange mechan-
ism, but, because the high ionic strength of the leachate is relatively un-
variable in the adsorption experiments, the amount adsorbed does not become
independent of the concentration of Cu, Zn, or Cd in the leachate. Thus,
there is no way to predict the maximum amount of Cu, Zn, or Cd that will be
adsorbed from a given leachate by a given soil or clay mineral, at a given pH
and at some concentration without actually measuring it. At pH 5.0, precipi-
tation of Cu as a hydroxide-carbonate makes a very significant contribution
to the total amount of Cu removed by kaolinite but, because mohtmorillonite
has a higher amount of exchange-adsorption, the contribution of precipitation
to the total amount of Cu removed by montmorillonite is less significant. If
the pH of the leachate is lower than about 6.5, precipitation does not make a
significant contribution to the total amount of Zn and Cd removed by kaolin-
ite and montmorillonite.
The mobility of Cu, Zn, and Cd in soils or clay minerals is similar to
Pb and is dependent upon solution pH and ionic strength as well as on the CEC
of the soils or clay minerals. This is of little importance at sufficiently
high (above 7) pH values, because precipitation is then more important than
cation exchange in the removal of Cu, Zn, or Cd from solution. However, the
mobility of Cu would apparently reach a minimum at pH 7 and would subse-
quently increase at pH values above 7. At pH 5, Cu, Zn, and Cd in leachate
will be quite mobile in soils or clay minerals with low CEC values, especial-
ly when the ionic strength of the leachate is high. Thus, if adsorption data
at pH 5.0 is obtained either from a leachate very high in ionic strength or
from a low CEC soil or clay mineral, and the mobility of Cu, Zn, or Cd is
calculated, it can be stated that for higher pH values, for higher CEC val-
ues, and/or for leachates low in ionic strength, the mobility would be less
than it is at pH 5.0. These facts add a built-in safety factor to estimations
of adsorption and/or mobility.
Treatment of a soil with a waste stream or leachate will alter the ex-
changeable cation distribution of the soil; for example, a high Ca soil will
become a high Na soil if treated with a high Na-content waste stream or
leachate. Thus, other potential problems may be created that must be evalu-
ated in addition to the problem of heavy metal toxicity to plants or heavy
metal accumulation in the food chain when waste streams or leachates are dis-
posed of on agricultural land.
82
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SECTION 9
EFFECT OF pH ON CHROMIUM ADSORPTION FROM
LANDFILL LEACHATE BY CLAY MINERALS1
ABSTRACT
The adsorption of Cr(VI) and Cr(HI) species by kaolinite and montmoril-
lonite clay minerals was found to be highly dependent upon the pH of the clay
suspensions and the physical-chemical properties of the clay minerals. Solu-
tion ionic strength was found to be of secondary importance to estimations of
Cr(VI) and Cr(III) adsorption.
No precipitation of Cr(VI) was detected in the pH range 1.0 to 9.0. Pre-
cipitation of Cr(III) as an amorphous hydrated hydroxide starts to occur
above pH 4.5.
The adsorption of Cr(VI) from a given solution decreased as pH in-
creased. The Cr(VI) species distribution indicated that the HCrOiT ion was the
Cr(VI) species predominantly adsorbed. Montmorillonite adsorbed about four
times more Cr(VI) than kaolinite under similar conditions of pH and ionic
competition* Contrary to expectations, less Cr(VI) was adsorbed from pure
solutions than from leachate solutions.
The adsorption of Cr(III) increased as the pH of the suspensions in-
creased. At pH 2.5, the amounts of Cr(III) adsorbed were consistent with a
cation exchange mechanism involving Cjr3+ ions. As the pH is raised to 4.0,
the amounts adsorbed correspond to cation exchange adsorption of the hydro-
lized Cr(III) species, Cr(OH)t and Cr6(OH)it. The adsorption of Cr(III) is 3%
to 14% lower in leachate than in pure Cr(N03>3 solutions.
For a given type of clay, about 30 to 300 times more Cr(III) than
Cr(VI) is adsorbed depending upon pH and the ionic competition in solution.
The results of the study suggest that landfill disposal of Cr(VI) wastes re-
presents a potential pollution hazard due to its high mobility in earth ma-
terials and that safe disposal may require conversion of Cr(VI) wastes to
Cr(III) before disposal in landfills.
INTRODUCTION •>
Efforts to effectively dispose of most industrial heavy metal wastes
without polluting the environment have thus far proved unfruitful. Tradition-
ally, rivers or lakes have been used for the disposal of these potentially
Authors: R. A. Griffin, Anna K. Au, and R. R. Frost
83
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hazardous discharges. Industrial plants have also disposed of their wastes
through recharge basins or diffusion wells (Welsch, 1955), and into sewer
systems (Nassau County, N. Y., Department of Public Works Sewer Regulations,
1955). All of these disposal methods can contribute to contamination of sur-
face and ground water (Davids and Lieber, 1951). To minimize the problems
caused by heavy metal wastes in sewage treatment, laws have been enacted in
some northern Illinois counties forbidding the disposal of such wastes into
sanitary sewers. That prohibition has increased the pressure for permission
to dispose of these wastes in the available sanitary landfill sites. However,
unless specially designed, sanitary landfills also are a potential source of
surface and ground-water pollution (Walker, 1969). For that reason a demand
has arisen for information about the capacity of earth materials to adsorb
heavy metals from landfill leachates (Fuller, 1975).
Chromium compounds are widely used in the leather, textile, chemical
manufacturing, metal finishing, and other industries. Approximately 30,000
tons of chromium-bearing wastes are discharged annually from the metal
finishing industries alone (U.S.-EPA, 1973), and problems of environmental
pollution have arisen. Chromium (VI) contamination of the village wells in
Douglas, Michigan, was reported in 1947 (Davids and Lieber, 1951), and as
early as 1952 chromium was found in high concentration in the ground water of
Nassau County, New York (Welsch, 1955).
In trace amounts, chromium is an essential element in the diet of some
animals and, presumably, human beings. However, at sufficiently high concen-
trations, all compounds of chromium are toxic (Smith, 1972). The valence
state of chromium has a considerable influence on its toxicity. It is well
established that Cr(VI) compounds are the most toxic and are usually irri-
tating and poisonous to all tissues (Baetjer, 1956).
Thus far, the distribution and impact of chromium on aquatic ecological
systems have not received extensive study, so that relatively little is known
about the transfer of the metal from waste streams to earth materials and
then to living systems. Because knowledge of the chromium-leachate system is
scant, the present study was conducted to investigate the effect pH has on
Cr(VI) and Cr(III) adsorption by clay minerals plus precipitation in deion-
ized water and municipal leachate solutions. It was also desired to gain in-
sight into the factors that affect the mobility of chromium as it passes
through soils or clay-mineral layers. These soil or clay layers may be po-
tentially useful as liners for waste disposal sites.
EXPERIMENTAL
Various concentrations of Cr(VI) in deionized water and DuPage leachate
were prepared using potassium chromate (KaCrOO , and 50-ml aliquots were
pipetted into Erlenmeyer flasks containing either 3 g of montmorillonite or
5 g of kaolinite. The weight of clay used was chosen so that the amount of Cr
removed from the solutions could be determined with some precision from the
difference between the initial and final solution concentrations. Several
replicate suspensions for each concentration were prepared, and their pH
values were adjusted with either HN03 or NaOH to various values in the pH
range 1.0 to 9.0.
84
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In preliminary experiments, the amount of Cr(VI) adsorption was found to
decrease as pH increased, and pH values of the Cr(VI)-clay suspensions rose
when the flasks were shaken. Also, apparent irreversible Cr(VI) adsorption
occurred if the Cr(VI)-clay suspensions had been equilibrated at one pH value
and again equilibrated at a higher pH value. Therefore, the flasks were
shaken for about 2 weeks in a constant temperature bath at 25 ± 0.5° C; the
pH values of the clay suspensions were measured each day, and acid or base
was added when necessary to maintain the initial pH value. At least one day
after the final pH adjustment, the pH values were recorded and the suspen-
sions were centrifuged. The supernatant solutions were then decanted into
plastic bottles and their pH adjusted to 2.0 to prevent any Cr adsorption by
the container. The equilibrium Cr concentrations C£q (ppm) in the supernatant
solutions were determined by atomic absorption spectroscopy. The initial Cr
concentration (Cj) in ppm was determined by analyses of blank sample, i.e.,
sample without clay, prepared at the same time as the clay suspensions. In
our experiments, Cj concentrations ranging from 5 to 300 ppm Cr as Cr(VI)
were used.
The amount of Cr(VI) adsorbed (x/m) in ing/g clay at a given pH was cal-
culated as (Cj - CEq) x VE/1000 where VE = final solution volume (ml)/weight
of clay sample (g). The amount of Cr(VI) adsorbed by a fixed amount of clay
from a given Cj solution at various pH values was plotted against the pH
values to obtain an adsorption-pH curve.
The experimental procedures for the Cr(III) adsorption studies were
similar to those used in the Cr(VI) experiments. It was determined that 0.100
g of the clay minerals would give the desired precision in determining the
change in Cr(III) concentrations at equilibrium. Chromium (III) nitrate
(Cr(N03)3 • 9H20) was used as a source of Cr(III). The study of Cr(III) ad-
sorption was generally limited to the pH range of 1.5 to 4.5 because of
Cr(III) precipitation around pH 5. Because of the precipitation, the pH of
the leachate was adjusted to about 4 prior to "spiking" with Cr(III). An
initial Cr(III) concentration range of 30 to 800 ppm was chosen for the ex-
periments.
RESULTS AND DISCUSSION
Adsorption isotherms can be calculated from a family of adsorption-pH
curves at different pH values. The amount of Cr adsorbed (x/m) in mg/g clay
at a particular pH value is read from an adsorption-pH curve for a particu-
lar Cx and the equilibrium concentration, CEq, in ppm is calculated from the
following equation:
r, e ^ (x/m) • 1000
CEq (ppm) = CX (ppm) - ^
where all the parameters are as previously defined.
Interpretation of the adsorption data was aided by application of the
Langmuir equation (1918). In the derivation of the Langmuir equation it is
assumed that (a) the surface is energetically homogeneous (b) the adsorbate
adsorbate interaction on the surface is negligible, and (c) the adsorbed
85
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molecules do not influence neighboring sites. The Langmuir adsorption equa-
tion in its linear form is:
1 • C {16}
_ i ,^,
x/m kb b
where C is the equilibrium concentration of the ion, x is the amount ad-
sorbed, m is the mass of adsorbent, b is the adsorption maximum, and k is a
constant that relates to the energy of adsorption.
Chromium (VI) Adsorption
No precipitation of Cr(VI) was observed in the pH range 1.0 to 9.0.
Families of Cr(VI) adsorption-pH curves are presented in Figures 26 and 27.
Several characteristics of Cr(VI) adsorption by the clays appear. First, the
adsorption of Cr(VI) decreases as pH increases. Second, Cr(VI) is not ad-
sorbed by the clays near pH 8.5 and above. Third, the amounts of Cr(VI) ad-
sorbed are small compared to the amounts of exchangeable cations on the clay
samples.
Diagrams showing the distribution of Cr(VI) species covering the experi-
mental concentration range are presented in Figure 28 and were calculated by
using the constants given in Butler (1964). Noteworthy is the rapid decrease
in the fractions of HCrOiT and CraO?" species above pH 5 and the corresponding
increase in the fraction of CrOff" species, which become the principal species
present in solution at about pH 8.5 (Fig. 28). Below pH 2 the fraction of
ions decrease rapidly as the fraction of HaCrOi* species increases.
The behavior of the Cr(VI) adsorption-pH curves in Figures 26 and 27
implies that the HCrO^ ion is the principal ion being adsorbed by the clay
minerals. Conversely, the lack of adsorption at pH values above 8.5 indicates
that the CrO*" ion is not adsorbed at all by either of the two clays. The
mechanism of Cr(Vl) adsorption by these clay minerals apparently cannot
neutralize the two negative charges present on the Cro£~ ion. On the other
hand, Cr207~ ions may be adsorbed at low pH values because of its more open
structure, CrO^ - 0 - CrOl, which places the two negative charges an ap-
preciable distance apart, as opposed to charges on adjacent oxygens that
occur in the CrO£~ structure. The charge separation on the CrzQ? ion may al-
low it to act essentially as two monovalent ions, with each negative charge
•fulfilling an adsorption site or one negative charge fulfilling an adsorption
site and the other negative charge being neutralized by a cation in solution.
If the distribution of Cr(VI) species in solution is the only factor govern-
ing adsorption of Cr(VI) over the pH range 2 to 5, a plateau should be ob-
served in the Cr(VI) adsorption-pH curves; however, the adsorption of Cr(VI)
continues to rise with decreasing pH to a pH value around 2. Therefore, in
the low pH range the pH probably modifies the structures of the clay minerals
to permit increased Cr(VI) adsorption to occur. Dissolution of clay minerals
is known to occur at low pH levels (Hofmann et al. , 1956). This can alter the
surface structure and surface area of the clays, resulting in changes in
their adsorption characteristics as the pH is lowered.
Adsorption isotherms were constructed from the Cr(VI) adsorption-pH
86
-------
0.70-
0.60- 93)
ol
e 0.50 H
-2 0.40-
E
2 0.30 H
-------
I I I
J u
-3,0.200-
\
o>
§ 0.150-
o
E 0.100-
o
•o
-e 0.050-
o
o 0.000-
-•185
A.K2Cr04
SOLUTION
B. DU PAGE
LEACHATE
2.0 4.0 6.0 8.0 2.0 4.0 6.0 8.0
PH
Figure 27. Chromium (VI) adsorption-pH curves for kaolinite at 25° C.
Initial Cr(VI) concentrations (ppm) are indicated beside
each curve and the equivalence volume for each curve is
10.Q ml/g kaolinite.
88
-------
1.0-
^—"^1
N
0 0.6-
«4—
O
C
O
I 0.4-
0.2-
0:0-
a. I.OxlO"4M
b. I.OxiO"3M
c. 6.0 x IO"3M
1.0
I. HCr04~
2. Cr04=
3. H2Cr04
4. Cr207=
3.0
5.0
PH
7.0
9.0
Figure 28. Distribution of Cr(VI) species for various Cr(VI) concentra-
tions.
89
-------
curves in Figures 26 and 27 using Equation {16} at pH values 3.0, 4.0, 5>Q,
and 7.0. Sample adsorption isotherms, constructed at pH 4.0, are shown in
Figure 29. For both clay minerals, more Cr(VI) was adsorbed from DuPage
leachate solutions than from pure KaCrOi* solutions throughout the pH range
3.0 to 7.0. This result is contrary to that which was expected. Evidently,
anions (e.g., Cl~ and HGO^) in the leachate do not compete favorably with
HCrOij ions, or adsorption would have decreased.
The effect of the Cl~ ion on the adsorption of Cr(VI) by clay was deter-
mined by adsorption experiments carried out with 20 ppm Cr(VI) in deionized
water with and without 1000 ppm of Cl~ added as NaCl. No appreciable change
in the adsorption of Cr(VI) was caused by the Cl~ ion. That more Cr(VI) was
adsorbed from DuPage leachate than from pure KaCrOi* solutions may be the re-
sult of formation of polynuclear complexes in the leachate solutions, organic
or inorganic in nature, which can be adsorbed by the clay. The high ionic
strength of the leachate may also contribute to higher adsorption of Cr(VI)
species by a depression of the diffuse double-layer surrounding the clay
particles (van Olphen, 1963), which allows more ions to approach the clay
surface and be adsorbed.
The adsorption isotherms for Cr(VI) at pH values of 3.0, 4.0, 5.0, and
7.0 were plotted according to the linear form of the Langmuir equation (Equa-
tion 16). All the Langmuir plots gave linear regression r2 values 0.99. The
Langmuir plots at pH 4.0 are shown in Figure 30. From the slopes of the Lang-
muir plots, adsorption maxima were calculated and are presented in Table 12.
The difference in the calculated adsorption maxima of montmorillonite and
kaolinite reflects the difference in the probable number of available adsorp-
tion sites, based on comparison of the structural differences and the surface
areas of the two clay minerals (Table 1). The precise mechanism for anion ad-
sorption by clay minerals is uncertain, but we assume, as have others, that
anion exchange plays an important role in the adsorption process. However,
the adsorption maxima presented in Table 12 represent the maximum amount of
Cr(VI) ions adsorbed at some sufficiently high concentration of Cr(Vl) ions
in solution, whereas the adsorption isotherms in Figure 29 represent the
maximum amount of Cr(VI) ions that can be adsorbed at any given concentration
of Cr(VI) ions in solution.
Chromium (III) Adsorption
During preliminary experiments on Cr(IIl) adsorption by kaolinite, the
removal curves shown in Figure 31 were obtained. The curve labeled "blank"
in Figure 31, representing removal of Cr(III) from a solution containing no
clay, shows that precipitation becomes a very important mechanism of Cr(IIl)
removal near pH 5.0. The precipitate formed was blue-gray, and X-ray diffrac-
tion patterns of the precipitate showed no definite crystalline structure.
Because only HN03 and NaOH were used to adjust the pH of the Cr(1*03)3 solu-
tion, it is reasonable to believe that the precipitate formed is a chromic
hydroxide. Murray (1956) stated that chromic hydroxide is a hydrous oxide
(Cr20s • nHaO) of indefinite composition that is blue-gray when its water
content is high. It was therefore concluded that chromic hydroxide was the
precipitate of our experiments.
90
-------
0.60-
0.50-
o. 0.40H
o>
E
^ 0.30-
o
I 0.20H
o
0.10-
0-
Montmorillonite in
Du Page leachate
-X -X
Montmorillonite in K2Cr04 solution
Kaolinite in Du Page leachate
/—Kaolinite in K
*o -o
solution
0
100 200 300
Equilibrium Cr concentration (ppm)
Figure 29. Adsorption isotherms for Cr(VI) at pH 4.0 and 25° C.
91
-------
400-
300H
O
x
o
100-
0-
Kaolinite in
solution
Koolinite in DuPage leachate
Montmorillonite in
0
Montmorillonite in
Du Page leachate
100 200
Equilibrium Cr concentration (ppm)
Figure 30. Langmuir plots of Cr(VI) adsorption data at pH 4.0 and 25° C.
92
-------
TABLE 12. ADSORPTION MAXIMA FOR Cr(VI) BY MONTMORILLONITE AND
-O
KAOLINITE AT 25 C FOR VARIOUS pH VALUES
mg/g
DuPage
pH Pure Solution Leachate Solution
Montmorillonite
3.0 0.400 0.667
4.0 0.256 0.526
5.0 0.147 0.417
7.0 0.052 0.169
Kao Unite
3.0 0.093 0.189
4.0 0.044 0.130
5.0 0.032 0.115
7.0 0.015 0.051
93
-------
100-
90-
80-
70-
0>
I 60H
o
>»
•° 50H
O)
I 40H
-------
Families of Cr(III) adsorption-pH curves for the PH range 1.5 to 4.5 are
presented in Figure 32. Cation exchange is generally accepted as the princi-
pal mechanism for cation adsorption by soils and clay minerals. Chromium
(III) is known (Rollinson, 1956) to be extensively hydrolyzed in acid solu-
ni0^l+SP^ieS SUCh SS Cr(°H) ' Cr(°H)i °r Cr2(OH)*+ or Cr2(OH12)6+, and
Cr6(OH)15. Therefore, the increasing adsorption of Cr(III) as pH increases
can, in part, be attributed to exchange-adsorption of hydrolyzed Cr(III)
species other than Cr ions.
Adsorption isotherms were constructed from the adsorption-pH curves
(Fig. 32) by using Equation {15} at pH values 2.5, 3.0, and 4.0. The iso-
therms constructed for pH 4.0 are shown in Figure 33. As was expected, the
adsorption of Cr(III) by the clay minerals is lower in DuPage leachate than
in pure Cr(N03)3 solutions. However, significantly less reduction in adsorp-
tion of Cr(lII) from DuPage leachate (3% to 14%) took place than observed in
tests of Pb, Cu, Zn, and Cd adsorption from DuPage leachate by the same clay
minerals. The Cr(III) species existing in solution are so strongly adsorbed
that the cations present in the DuPage leachate do not effectively compete
with the Cr(lII) species for exchange-adsorption sites.
The adsorption isotherm data for Cr(III) at pH values 2.5, 3.0, and 4.0
were plotted according to the Langmuir equation (Equation {16}) and the Lang-
muir plots for the pH 4.0 adsorption isotherms are shown in Figure 34. Ad-
sorption maxima calculated from the slopes of the Langmuir plots in Figure 34
are presented in Table 13. The amounts of the various hydrolyzed Cr(III)
species that could be adsorbed via a cation exchange mechanism by the mont-
morillonite and kaolinite clay minerals that had CEC values of 79.5 meq/100 g
and 15.1 meq/100 g, respectively, are presented in Table 13, because, if
cation exchange is the principal adsorption mechanism, the ratio of the ad-
sorption maxima should be about equal to the ratio of the CEC values for the
two clay minerals.
At pH 2.5 the adsorption maximum ratio is close to the CEC ratio (Table
13) , but the adsorption maxima themselves are higher than those based on ex-
change adsorption of Cr3+ ions. This implies that some hydrolyzed Cr(III)
species are being adsorbed even at pH 2.5. The question arises, however, as
to the validity of using CEC values obtained from Ntft ion exchange at pH 7.0
to calculate the expected adsorption maxima at pH 2.5. However, as there is 'a
general relative correlation between the adsorption maxima and CEC values, it
is reasonable to use adsorption maxima calculated from CEC values for com-
parison purposes. At pH 4.0, the ratio of the adsorption maxima increases to
about 13, and the amount of Cr(III) adsorbed by montmorillonite is even
larger than would be expected from the exchange adsorption of the Cr6(OH)i5
ion.
The high adsorption maximum ratio at pH 4.0 is the result of the large
adsorption maximum obtained for montmorillonite, which appears to be in-
directly caused by our experimental procedure. The rates of some of the
Cr(III) hydrolysis reactions are very slow, and, in one reported case (Las-
wick and Plane, 1959), about 107 days were required for Cr(IlI) solutions,
even at elevated temperatures, to reach equilibrium. In our Cr(III) experi-
ments, the clay mineral suspensions were adjusted to a particular PH value,
95
-------
150.0-
100.0-
o>
c
50.0H
o
o
0.0
.O
O
100.0-
o
50.0-
0.0-
A.Cr(N03)3
SOLUTION
KAOLINITE
C. Cr(N03)3
SOLUTION
MONTMORILLONITE^
765
;388
i . i i I 1 L
B.DUPAGE LEACH ATE
KAOLINITE
D. DU PAGE
LEACHATE
MONTMORILLONITE
1.5
2.5
3.5
4.5 1.5
pH
2.5 3.5
4.5
Figure 32. Chromium (III) adsorption-pH curves at 25° C. Initial Cr(III)
concentrations (ppm) are indicated beside each removal curve,
and the equivalence volume for each curve is 500.0 ml/g kaoli-
nite.
96
-------
100-
o>
E
•o
0)
T3
o
i_
o
0-
Montmorillonite
in Du Page leachate
Montmorillonite in
/Cr(N03)3 solution
-X
Kaolinite in Cr(NO,), solution
Kaolinite in Du Page leachate
100 200 300 400 500 600 700
Equilibrium Cr concentration (ppm)
Figure 33. Chromium (III) adsorption isotherms at pH 4.0 and 25° C.
The plots shown in dotted lines were obtained from the
"corrected" adsorption-pH curves.
97
-------
30
E 20
V.
X
-------
TABLE 13. ADSORPTION MAXIMA FOR Cr(III) BY MONTMORILLONITE AND KAOLINITE AT 25° C FOR VARIOUS pH VALUES
(mg/g)
pH 2.5
Pa Lb
17.9 --
3.3 —
5.4 —
pH 3.0 pH 4.0 Values
PL PL Cr+3
Montmorillonite
33.7 32.8 139.6 136.1 13.8
Kaolinite
5.0 5.0 10.7 14.7 2.6
Ratio of Adsorption Maxima
6.5 6.6 13.0 9.3 5.3
Based on CEC for Different Species
Cr (OH) 2+ Cr (OH)t Cr 6 (OH) ft
20.7 41.4 82.7
3.9 7.8 15.6
5.3 5.3 5.3
a b
Pure solutions, DuPage leachate solutions
-------
such as 4.5. After a few hours, hydrolysis caused the pH of the suspensions
to drop. The pH of the suspensions were readjusted to 4.5, but the pH again
dropped because of hydrolysis. The pH of the suspensions were all readjusted
to the desired pH value several times during a period of two weeks. They were
then shaken for 2 days before a final pH value was measured, and the suspen-
sions were centrifuged. What appears to occur in the pH range 3.5 to 4.5 is
as follows. Adsorption plus possible precipitation of Cr(III) takes place at
the higher initial pH values to which the suspensions had been adjusted. But,
when the pH of the suspensions drops due to hydrolysis, desorption of Cr(III)
species from the clay mineral or dissolution of any precipitate formed ap-
parently does not occur at rates fast enough to achieve true equilibrium in
the clay suspension. The pH values of the montmorillonite suspensions showed
larger decreases in pH than the kaolinite suspensions; this apparently pro-
duces much larger differences between the calculated and true equilibrium ad-
sorption maxima for montmorillonite than for kaolinite.
The adsorption-pH curves shown in Figure 32 were replotted as the amount
of Cr(III) adsorbed versus the highest pH values to which the clay suspen-
sions were adjusted. The adsorption isotherms were calculated from the "cor-
rected" adsorption-pH curves at pH 4.0 and are shown as the dotted isotherms
for montmorillonite in Figure 33. The "corrected" adsorption isotherms for
kaolinite are almost superimposable on the isotherms shown in Figure 33, and,
therefore, were omitted. Adsorption maxima from pure Cr(NOa)3 solutions, cal-
culated by Langmuir plots of the "corrected" isotherms at pH 4.0, are 72.2
mg/g for kaolinite. Thus, the "corrected" adsorption maxima agree with an ex-
change-adsorption mechanism involving hydrolyzed Cr(III) species. We have
learned, as Laswick and Plane (1959) pointed out, that the changes that
Cr(III) species undergo are generally quite slow, which makes the interpre-
tation of the experiments difficult.
CONCLUSIONS
The results of this study indicated that Cr(III) yielded the strongest
attenuation of all the heavy metals studied. This suggests that landfill dis-
posal of Cr(III) wastes would exhibit low mobility and probably would initi-
ate fewer pollution problems than would the other heavy metals. Above pH 6,
Cr(III) should be immobile because of precipitation. Below pH 4, Cr(III)
species were strongly adsorbed by both kaolinite and montmorillonite and
would have a relatively low mobility through soils or clay minerals used as
landfill liners. Between pH 4 and 6, the combination of adsorption and pre-
cipitation should render Cr(III) quite immobile.
In the pH range 1.5 to 4.0, 30 to 300 times more Cr(III) than Cr(VI) was
adsorbed by the clay minerals; and at higher pH values the ratio (Cr(III) re-
moved) /(Cr(IV) removed) became even larger because of increased Cr(III) re-
moval and decreasing adsorption of Cr(VI). The adsorption of Cr(VI) was low
relative to Cr(III), even at very low pH values where Cr(VI) adsorption was
strongest. Cr(VI) adsorption was the weakest of the heavy metals studied and
was markedly reduced as the pH was raised into the alkaline range. Thus, it
would become very mobile at high pH. Since Cr(VI) is the most toxic and mo-
bile form of Cr, landfill disposal of Cr(VI) wastes can potentially cause
serious pollution problems even if the landfill has a thick clay liner. The
100
-------
results of this study suggest that a conversion of Cr(VI) wastes to Cr(III)
by a process such as that devised by Shiga (1975) would greatly reduce the
hazard to water resources from Cr(VI) wastes disposed of in landfills.
101
-------
SECTION 10
EFFECT OF pH ON ADSORPTION OF As AND Se
FROM LANDFILL-LEACHATE BY CLAY MINERALS1
ABSTRACT
The effect of pH and ionic competition on arsenate (As(V)), arsenite
(As(III)), and selenite (Se(IV)) adsorption by kaolinite and montmorillonite
clay minerals from municipal landfill-leachate solutions were determined.
The results showed that pH had a strong influence on the amounts ad-
sorbed of all three of the elemental forms studied. Montmorillonite clay was
found to adsorb about twice as much As or Se as kaolinite. Leachate was found
to have little effect on the adsorption of As(V) or Se(IV), while As(III) ad-
sorption was reduced 30 to 50%. It was concluded that the principal adsorp-
tion mechanism was anion exchange of the monovalent species of each elemental
form.
The results of the study suggest that land disposal of As and Se wastes
under alkaline conditions represents a potentially high pollution hazard.
INTRODUCTION
As and Se are quite toxic (U.S.-EPA, 1972) and, therefore, they have a
high potential to produce pollution problems. The U. S. Environmental Protec-
tion Agency (EPA) cites several examples of instances where land disposal of
arsenic wastes have resulted in the poisoning of drinking water wells (U.S.-
EPA, 1973) and selenium has been reported to have polluted ground water as
far as two miles from a dump on Long Island (Garland and Mosher, 1975). As
and Se waste streams may be liquids, suspensions, or sludges. Sample waste
streams from copper, lead, and zinc smelting, from duplicating and photo
equipment manufacturing, and from pharmaceutical industries have been
measured and reported to contain from 1,000 to 30,000 ppm As and from 3,000
to 50,000 ppm Se (Lehman, 1973). These waste streams may be disposed of on
land in lagoons, in landfills, or by spreading. Land spreading of municipal
wastewater effluents (U.S.-EPA, 1975) and sewage sludges which may contain
low concentrations of As and Se is now being considered as a viable alterna-
tive to treatment.
The purpose of this study was to provide some basic data regarding the
effect of pH on the removal of As and Se from landfill-leachate by kaolinite
and montmorillonite, which are common clay minerals found in soils. The re-
1Authors: R. R. Frost and R. A. Griffin
102
-------
suits of the present study give insights into whether or not As and Se waste
streams can be safely disposed of in properly designed landfills for munici-
pal solid waste. The results of the study also provide basic information on
the mobilities of As and Se through soils. This can be of aid in the design
of land disposal systems for As and Se waste streams in general.
EXPERIMENTAL
Appropriate stock solutions of As(V) and As(III) were prepared by dis-
solving reagent grade Na2HAs0lf or NaAs02 in deionized water. Se(IV) stock
solutions were prepared by dissolving pure selenium metal in a minimum amount
of 1:1 HNOs. Some heating of the solution was necessary to speed up the dis-
solution. All stock solutions were adjusted to about pH 5 before use.
Fifty ml aliquots of leachate were pipeted into 125-ml Erlenmeyer flasks
containing from 1.00 to 5.00 g of either kaolinite or montmorillonite. The
weight of clay used was chosen so that the amount of arsenate, arsenite, or
selenite removed from the leachate solutions could be determined with some
precision from the difference between the initial and final solution concen-
trations. Several replicates of each clay suspension were prepared. The pH
values of the replicate clay suspensions were adjusted with HNOa or NaOH to
different values over the pH range 1 to 9. The clay suspensions were shaken
overnight and then 2.0 ml of an appropriate pH 5 stock solution of either As
or Se was pipeted into the flasks. Of the several experimental procedures
possible, the procedure used gave the most satisfactory results. The As or Se
clay suspensions were then shaken in a constant temperature bath at 25. ± 0.5°
C for at least 24 hours to insure complete equilibration. The equilibrium pH
values of the clay suspensions were measured, the clay suspensions centri-
fuged, and the supernatant solutions analyzed for their As or Se concentra-
tion by atomic adsorption spectroscopy. A NO-CaHz flame was used for As
analyses, and air-CaHa flame was used for Se analyses. Blanks (i.e., no clay)
were prepared along with the clay suspensions and were also analyzed to de-
termine the initial As or Se concentration. The amount of As or Se removed
from solution by a given clay at a particular pH was determined as:
. total sol'n vol. after pH adjustments
(initial - equilibrium cone.) x sample weight.
The amount of As or Se removed from solution was then plotted as a function
of pH. From the initial concentration of As or Se in solution, the weight of
clay used, the final solution volume after pH adjustments, and the removal
versus pH curves themselves, "adsorption" or "removal" isotherms can be con-
structed at different pH values by use of the following equation:
Equilibrium C (ppm) = Initial C (ppm) -
Am't. removed (ug/g) x Wt. Clay (g).
Final solution volume (ml)
RESULTS AND DISCUSSION
Arsenate (As(V)) and Arsenite (As(III)) Adsorption
103
-------
Curves for the removal of As(V) from leachate by kaolinite and mont-
morillonite versus pH are shown in Figure 35. It is interesting to note that
the amount of As(V) removed from solution goes through a maximum at about pH
5. The distribution of arsenate species in solution as a function of pH is
shown in Figure 36a. Comparison of Figure 35 with Figure 36a shows that the
As(V) removal curves follow the monovalent HaAsO^ species curve reasonably
well. Thus, it was concluded that the HaAsOIT ion is the principal As(V) ion
being adsorbed by the clay minerals. The non-adsorption or depressed adsorp-
tion of the HAsOij: ion is apparently due to the occurrence of negative charges
on adjacent oxygen atoms in the tetrahedral HAsO™ ion which results in repul-
sion of the ion from the clay surface. Some As(V) precipitation was observed
to occur above approximately pH 9 and can be seen on the removal curves in
Figure 35 as the curves turn upward around pH 9.
Curves for the removal of As(III) from leachate by kaolinite and mont-
morillonite versus pH are shown in Figure 37. In general, an increase in ad-
sorption of As(III) is observed as pH increases in the range 3 to 9. The
montmorillonite removal curves in Figure 37 show an interesting peak about pH
7. There is some question as to what As(III) species are present in the
leachate solutions and which As(III) species is actually being adsorbed. From
a study of As(III) adsorption by an anion exchange resin (Everest and Popiel,
1957), the AssOs on As3(OH)T0 species was proposed as the species being ad-
sorbed in the pH range 5 to 6, but the fraction of the total As(III) in solu-
tion existing as the AsaOs (As3(OH)Io) species is small and the amount of ex-
change adsorption was low. Other As(III) species apparently existing in solu-
tion are As2C% (As2(OH)f), As(OH)3, As(OH)^, and As020H= — with the latter
two species becoming important above pH 9. Whatever As(III) species are
present in solution, the fraction of the total As(III) in solution present as
monovalent As(III) species will increase as pH increases. Thus, the amount of
As (III) adsorption should increase as pH increases as observed in Figure 37.
There is no apparent explanation for the observed peaks about pH 7 on the
montmorillonite removal curves in Figure 37. Although no peaks are observed
o.n the kaolinite removal curves in Figure 37, they might have appeared if
more data points were obtained in the pH range 6.5 to 7.
Comparison of removal curves (not shown) for As(V) and As(III) from de-
ionized water solutions obtained at a single concentration for each clay
mineral with the corresponding leachate removal curves showed that the anion
competition present in the leachate surpresses the adsorption by the clay
minerals of As(V) slightly and of As(III) by some 30 to 50 percent.
Isotherms calculated from the removal curves in Figures 35 and 37 at pH
5.0 are shown in Figure 38. The higher adsorption of As(V) and As(III) by
montmorillonite compared to kaolinite simply reflect the structural and sur-
face area differences between the two clay minerals. Anion exchange sites are
thought to exist primarily at the broken edges of clay minerals (Grim, 1968).
Montmorillonite clay was found to adsorb approximately twice as much As
as kaolinite. Examination of the surface area data, as measured by N2 gas ad-
sorption given in Table 1, suggests that the montmorillonite has roughly
twice as much edge surface area as the kaolinite. The tetrahedral H2AsO£ ion
can align itself with the silica tetrahedral of the clay lattice and can form
104
-------
700
600-
400-
300-
o
E
0}
zoo-
100-
Kaolinite
190
I O.I
Montmorillonite
128.
59.O
21.3
pH
-i 1 r
Figure 35. The amount of As(V) removed from DuPage leachate solutions
by kaolinite and montmorillonite at 25° C plotted as a func-
tion of pH. Labels are the initial solution concentration
of As(V) in ppm. Each datum point was obtained by using
either 4 g of kaolinite of 1 g of montmorillonite in a total
solution volume of 52.2 ml.
105
-------
A. As (32)
B. Se(I2)
I i I I I L
Figure 36. Species distribution diagram for As(V) and Se(IV).
-------
400-
O)
o
o
300-
E
O 200
o
E
o>
V)
100-
J L_
Koolinite
68.0
Montmorillonite
024
I
8
0
pH
8 10
Figure 37. The amount of As(III) removed from DuPage leachate solutions
by kaolinite and montmorillonite at 25° C plotted as a func-
tion of pH. Labels are the initial solution concentration of
As(III) in ppm. Each datum point was obtained by using 4 g
of clay in a total solution volume of 52.5 ml.
107
-------
jcn
As(3Z) by montmorillonite
As(3Z:)by kaolinite
As(m) by montmorillonite
kaolinite
60
80
100
120
140
160-
180
Equilibrium concentration (ppm)
Figure 38. The amount of As(V) or As(III) removed from DuPage leachate
solutions at pH 5.0 and 25° C per gram of clay plotted as a
function of the equilibrium arsenic concentration.
108
-------
an extension of the crystal lattice which has a relatively high bond strength.
The arsenate ion can be compared to the large As3 (OH) To ion which is not
tetrahedral and thus is unable to align itself as effectively on the clay
edges. It was therefore concluded that the principal adsorption mechanism was
anion exchange of the monovalent species of each elemental form.
Selenite (Se(IV)) Adsorption
Curves for the removal of Se(IV) from leachate by kaolinite and mont-
morillonite versus pH are shown in Figure 39. The removal of Se(IV) goes
through a maximum in the pH range 2 to 3 and then decreases as the pH in-
creases. The distribution of Se(IV) species in solution as a function of pH
is shown in the Figure 36b. Although it is not as apparent as for As(V), the
monovalent HSeOJ ion appears to be the species predominantly adsorbed by the
clay minerals. It is evident from the data in Figure 39 that the adsorption
of Se from solution is rapidly converging on zero removal at approximately pH
10, which is the value at which the HSeO'a species disappears from solution.
This is evidenced by examining the species distribution diagram pre-
sented in Figure 36b. The HSeOJ ion has a trigonal-pyramidal configuration
which may account for its reduced adsorption; by contrast, the HaAsO^ ion has
a tetrahedral configuration. Thus, the configuration of the HSeO^ ion must be
an inhibiting factor in its adsorption by the clay minerals. On the other
hand, pH must play a significant role in modifying the clay mineral surface
structure so that increasing adsorption occurs with decreasing pH until the
point where a significant fraction of the total Se(IV) in solution is present
as the HaSeOs specie. At this point, adsorption starts to decline sharply as
the pH is lowered below a value of 2.
Comparison of removal curves (not shown) for Se(IV) from deionized water
solutions obtained at a single concentration for each clay mineral with the
corresponding leachate removal curves showed that the leachate supresses the
adsorption of Se(IV) by the clay minerals slightly. This is presumed to be
due to competition for anion exchange sites by the high concentrations of
anions present in leachate (Table 2) with the HSeO?.
Isotherms calculated from the removal curves in Figure 39 at pH values
3.0, 5.0, and 7.0 are shown in Figure 40. The higher adsorption of Se(IV) by
montmorillonite compared to kaolinite simply reflects the structural and sur-
face area differences between the two clay minerals as previously discussed.
CONCLUSIONS
The results of this study indicate that pH has a pronounced effect on
the amounts of As(V), As(III), and Se(IV) adsorbed from leachate by clay
minerals. It was concluded that the principal adsorption mechanism was anion
exchange and, from species distribution diagrams, that the adsorption was
due principally to the monovalent species of each element studied, thus lead-
ing to the strong pH dependency of the adsorption process.
The results of the study suggest that optimum As removal by soil ma-
terials from waste streams would be achieved by conversion of any As (.III J to
109
-------
600-
500-
0>
400-
o
o
£
o
,£ 300-
-a
o
E
o>
t.
-------
250-
I 1 1 I I
Montmorillonite
10 20 30 40 50 60 70 80 90 100 110 120
Equilibrium concentration (ppm)
Figure 40. The amount of Se(IV) removed from DuPage leachate solutions
at 25° C and several pH values per gram of clay plotted as
a function of the equilibrium Se concentration.
Ill
-------
As(V), and disposal of the waste in montmorillonitic soils at pH 5. Optimum
Se(IV) removal would result from disposal in a montmorillonitic soil at a pH
of 2 to 3. A high mobility and hence potential pollution hazard would be ex-
pected from land disposal of As(V) or Se(IV) wastes under alkaline condi-
tions.
112
-------
SECTION 11
MERCURY REMOVAL FROM LEACHATE
BY CLAY MINERALS1
ABSTRACT
Preliminary investigations regarding the effect of pH on the amounts of
Hg removed from DuPage leachate solutions by the two clay minerals, kaolinite
and montmorillonite, are reported.
Large amounts of Hg were removed from leachate, both in the presence and
absence of clay. This removal of Hg increased as the pH was raised. It was
estimated that 70 to 80% of the Hg was removed from leachate by precipitation
and/or volatilization, thus indicating that these were the predominant at-
tenuation mechanisms. Removal by clay accounted for 20 to 30% of the observed
Hg removal. About two thirds of the Hg removed by the clay was found to be
organic-Hg, and about one third was inorganic-Hg.
The results indicated that maximum Hg removal from leachate would be
achieved by disposal in montmorillonite clays or soils under alkaline pH con-
ditions.
INTRODUCTION
Because Hg has been the subject of many publications, its occurrence and
toxicity will be discussed only briefly here. There are two basic sources of
Hg in the environment: as it occurs in nature, and as it is redistributed in
nature by man's actions. Hg is widely distributed in the air, soil, and water
in low concentrations (Klein, 1972). Man in his utilization of Hg-related
technology has at times redistributed this Hg in nature. It is this source of
Hg in the environment that we must concern ourselves with since it often re-
sults in the release of dangerously high Hg concentrations (Goldwater, 1971).
The main source of environmental Hg contamination has been in the in-
dustrial sector (D'ltri, 1972a) , and much of this can be attributed to waste-
water discharge into rivers and streams (Turney, 1972; Derryberry, 1972). A
great deal of the mercurial waste, both organic and inorganic, disposed of in
this manner can be directly or indirectly converted by anaerobic microbes in-
to mono- or di-methyl-mercury (D'ltri, 1972a; Greeson, 1970), both of which
have been found to be extremely toxic.
As a result of studies on Hg contamination and poisoning, the disposal
Authors: R. A. Griffin and G. D. Robinson
113
-------
of industrial mercurial wastes is carefully controlled. However, the po-
tential for contamination still exists where gaps in knowledge preclude en-
lightened disposal practices. This potential also exists at municipal land-
fill sites accepting mercurial wastes. The leachates generated from these
sites are anaerobic and therefore have the ability to convert mercurial
wastes into their toxic forms. Since Hg is incorporated into many industrial
or consumer products such as paints, Pharmaceuticals, paper products, fluor-
escent lamps, mercury batteries, etc., the indiscriminate disposal of these
products in landfills by an uninformed population represents an important
potential environmental Hg contamination route (D'ltri, 1972b).
There has been very little research done on the chemical behavior of Hg
in municipal leachates. The information that is available to date on Hg in
municipal leachates is insufficient to evaluate the migration of Hg in land-
fills. It is therefore our purpose here to report the results of studies on
the adsorption of Hg from leachate by clay minerals. These studies were
initiated to determine whether or not clay minerals have potential for use as
landfill liners for attenuation of Hg which may be present in leachates.
EXPERIMENTAL
In order to determine the amount of Hg removed from solution, a series
of samples were prepared using HgCla in DuPage landfill-leachate, and HgClg
in DuPage landfill-leachate with clay (kaolinite or montmorillonite) added.
The latter two treatments were also used to determine the amounts of organic
and inorganic Hg removed from solution.
The amounts of organically-bound Hg and inorganic Hg in solution were
determined by the flameless Atomic Absorption Spectroscopy (AA) technique
described in the appendix.
Eight samples were prepared for each solution in each experiment. Before
addition of Hg, 50 ml aliquots of either deionized-H^Oj DuPage leachate, or
DuPage leachate and 1.000 g kaolinite were added to 125 ml Erlenmeyer flasks.
The pHs of the solutions were then adjusted using 0.1N, IN, 3N, and/or cone.
HNOs to obtain pH values between 1 and 9 (each set of samples spanned this
range). The samples were placed in a shaking waterbath at 25 ± 0.5° C, and
allowed to equilibrate overnight. The pH values were checked after 20-24
hours and 2 ml aliquots of 25 ppm Hg stock solution were added to each flask
giving a total volume of 52 ml and a Hg concentration of approximately 0.962
ppm. The samples were again placed in the shaking waterbaths and allowed to
equilibrate overnight. After 20 to 24 hours the final pH values of the
samples were recorded, and each sample was transferred to a 50 ml centrifuge
tube and centrifuged at 20,000 rpm for 5 minutes. The supernatant solutions
were decanted, acidified, and analyzed for their Hg content using flameless
AA.
The amount of Hg removed from solution by a given clay at a particular
pH was determined as follows:
114
-------
r (CI " CEo) * v?
CR = T E (19}
where CR = amount of Hg removed in yg/g clay.
Cj = initial Hg concentration in ppm,
CEq = equilibrium Hg concentration in ppm,
VF = total solution volume after pH adjustments in ml, and
W = weight of clay in grams.
The amount of Hg removed from solution was then plotted as a function of pH.
RESULTS AND DISCUSSION
Some examples of results obtained for Hg removal from various solutions
plotted as a function of pH are presented in Figure 41. It can be seen from
this figure that Hg is removed from solution, even from a presumably sterile
and pure HgCl2 solution (curve A). In the short time available for this study
we were not able to satisfactorily isolate all the reasons for this removal
of Hg. Three possible removal mechanisms are adsorption onto the walls of the
glassware and plastic bottles, precipitation, or volatilization. At the lower
pH values the Hg removal is thought to have been by adsorption onto the
glassware, although standard solutions may be satisfactorily stored at pH 2.
It is presumed that the adsorption in the sample solutions occurred at the
higher initial pH values and that the Hg did not have sufficient time to
totally desorb after the pH was adjusted to lower values. The increase in re-
moval in the higher pH range (5-8) is apparently the result of both precipi-
tation and adsorption. Volatilization was not considered as an important
mechanism for the Hg losses observed from the HgCl2 solutions since they were
presumably sterile and stoppered.
An increasing amount of Hg removal is displayed in curve B (leachate
blank) which illustrates the loss of Hg from a leachate solution. The removal
of Hg from solution here is again probably due to the same mechanisms; that
is, the increase in Hg removal (greater amount of Hg removed by leachate than
by deionized-H20) is probably the result of either increased adsorption or
increased precipitation. However, volatilization losses of organic Hg com-
pounds due to microbial transformations in the leachate are also possible.
Curve C in Figure 41 shows the total amount of Hg removed from leachate
with 1 g of kaolinite present. The amount of Hg removed by 1 g of kaolinite
is taken as the difference between curve C and curve B in Figure 41. This is
plotted as curve B in Figure 42.
Figure 42 shows a breakdown of various forms of Hg removal from leachate
solutions. The amount of organic- and inorganic-Hg removed by clay was deter-
mined by placing an aliquot from each sample into each of two BOD bottles.
One set was digested following the general procedure while the second set was
analyzed undigested. A set of leachate blanks was treated in the same manner.
There were then four separate sets in all, which included a digested blank
(HgCl2 in leachate), an undigested blank, a digested sample (HgCl2 in leach-
ate with 1 g kaolinite), and an undigested sample.
The total organic-Hg in solution was determined by taking the difference
115
-------
0.7
0.6-
0.5-
c
o
1 0.4H
o
o
E
£ 0.2H
o>
x
O.I -
0.0
HgCI2 in leochate
81 Ig kaolinite
HgCI2 solution
2.0
3.0
4.0
5.0
PH
6.0
7.0
8.0
Figure 41. Removal of Hg from DuPage landfill-leachate and pure
solutions plotted as a function of pH at 25° C. The initial
Hg concentration is 0.96 ppm and the final volume is 52 ml.
116
-------
0.7
0.6-
3 Q5~
0)
o
.c
o 0.4-
Q>
e
o
I
d>
0.3H
0.2-
0.1-
0.0
Total Hg removed
from leachate
Total Hg removed by clay
2.0
3.0
4.0
5.0
PH
6.0
7
7.0
8.0
Figure 42. Removal of various forms of Hg from DuPage landfill-leachate
solutions by kaolinite plotted as a function of pH at 25° C.
117
-------
between the digested and undigested blanks. The amount of organic-Hg in solu-
tion with 1 g kaolinite present was always less than the organic-Hg deter-
mined from the blanks. This difference was then taken as the amount of or-
ganic-Hg removed due to the presence of the clay. Figure 42 illustrates the
relative amounts of organic-Hg (curve C) and inorganic-Hg (curve D) removed
by 1 g kaolinite plotted as a function of pH. These results can be compared
to the total Hg removed by 1 g kaolinite in curve B and with the total Hg re-
moved from solution by all mechanisms in curve A.
The results given in Figure 42 illustrate that the Hg removal is pH de-
pendent, with increasing amounts of removal as the pH is raised within the
range 2 to 8.
From the data, it can be stated that approximately two thirds of the
total Hg removed by 1 g kaolinite was organic Hg (curve C) , while about one
third of the Hg removed by the clay was inorganic Hg (curve D). The total Hg
removed from solution (curve A) is apparently the result of several mechan-
isms operating simultaneously, i.e. adsorption by clay, volatilization, and
precipitation. The amount of Hg removed as a result of the presence of clay
was evaluated, as previously described, and accounted for approximately 20 to
30% of the total Hg removed from solution. It therefore appears that precipi-
tation and volatilization account for the largest amounts of Hg removed in
this study.
At this time, no data is available as to the relative amounts of Hg lost
by volatilization from the leachate as opposed to precipitation. In any
event, large amounts of Hg are lost from leachate solutions both in the
presence and absence of clay.
In an unreplicated experiment, the removal of Hg from solution by mont-
morillonite was found to be approximately 5 times greater than the removal of
Hg from solution by kaolinite. Since this experiment has not been repeated,
the results are tentative. However, this ratio is similar to the ratio of the
Cation Exchange Capacities (CEC) of these two clay minerals. Since this re-
sult is consistent with those for the previous heavy metal cations studied,
it was concluded that the difference in adsorption between montmorillonite
and kaolinite was due to the cation exchange of the various ionic forms of Hg
which may be present in leachate solutions of variable pH.
CONCLUSIONS
The results of this study indicate that removal of Hg from leachate
solutions is enhanced by clay minerals and is pH dependent. Substantial
amounts of Hg were removed from leachate by the clays, and these amounts were
concluded to be in proportion to the respective CEC values of the clays. Of
the amount of Hg removed by the clays, it was concluded that about two thirds
was organic Hg and one third inorganic Hg.
Large amounts of Hg were removed from solution, both in the presence and
absence of clay. These results lead to the conclusion that adsorption by
clays was not the major mechanism responsible for removal of Hg. Rather, pre-
cipitation and/or volatilization accounted for between 70 and 80% of the Hg
118
-------
removed from the leachate solutions and were concluded to be the predominant
attenuation mechanisms in these experiments.
The results of this study suggest that maximum removal of Hg from leach-
ate would be achieved by disposal under alkaline conditions. It is also sug-
gested that montmorillonitic clays or soils will remove substantially more Hg
than kaolinitic clays or soils.
119
-------
SECTION 12
SUMMARY OF ADSORPTION STUDIES
The results of these studies with heavy metals indicate that pH has a
pronounced effect on the amounts adsorbed from landfill leachates by clay
minerals. It was concluded that the principal adsorption mechanism is cation
and anion exchange, a mechanism that led to the strong pH dependency of the
adsorption process. In addition, species distribution diagrams led to the
conclusion that it is the monovalent species of each element that is
principally adsorbed by anion exchange.
A comparison of the relative amounts of heavy metals removed at pH 5.0
from 100 ppm equilibrium concentration solutions of the metals studied, both
cationic and anionic, is presented in Table 14. The table indicates that the
cationic heavy metals are generally adsorbed to a greater degree than the
anionic forms. However, this ranking is somewhat pH-dependent, because the
greatest anion adsorption occurs in acid solutions and the greatest cation
adsorption in alkaline solutions. The ranking therefore changes somewhat at
different pH values.
A significant point shown in Table 14 is the importance of the valence
state of an element to the amount of that element removed from solution by
clay minerals. Cr(III) species are removed to a much greater extent than
Cr(VI) species. The clay minerals removed 30 to 300 times more Cr(III) from
solution than Cr(VI). The table also shows more extensive removal of As(V)
than of As(III). These results indicate that safer disposal of certain ele-
ments may be achieved if, prior to deposition at the landfill or disposal
site, the element is converted to the form that would be most strongly at-
tenuated.
The information derived from the studies of the various elements indi-
cates that the amounts of heavy metals removed from leachate by clay minerals
depends to a large degree on the element and the form of the element in-
volved, the pH of the leachate, the adsorption capacity of the particular
clay mineral in the liner, and the ionic strength of the leachate.
The adsorption capacity of the clay minerals and the reversible nature
of exchange-adsorption reactions have important environmental consequences.
For example, if industrial wastes containing heavy metals are placed in a
landfill, changes in the ionic composition or pH of the leachate can occur.
A change in pH may release large amounts of potentially toxic heavy metals
into the aqueous phase, especially in places where precipitates may have ac-
cumulated. Other ions in the waste compete with the heavy metals and may
120
-------
TABLE 14. REMOVAL OF HEAVY METALS FROM SOLUTIONS BY KAOLINITE AT pH 5.0
Amount removed at
equilibrium concentration
of 100 ppm
(pmoles/g)
Element
Cr(III)
Cu
Pb
As(V)
Zn
As (III)
Cd
Se
Cr(VI)
Pure
solutions
769*
55-1
42.3
t
33.6
*
26.7
t
0.62
Du Page
leachate
576*
15.7
12.1
5.3
3.8
2.0
1.9
1.9
1.9
*Precipitation contributes to removal at pH 5.0.
tRemovals from 40 ppm solutions were approxi-
mately the same as removals from leachate.
^Removals from 40 ppm solutions were 30 percent
greater than removals from leachate.
121
-------
exchange with them, thus allowing metal ions to come into solution. These
multiple interactions must be considered when a disposal site is designed
and when the environmental impact of adding heavy-metal wastes to municipal
landfills is assessed.
122
-------
SECTION 13
APPLICATION OF THE RESULTS TO THE
PROBLEM OF LANDFILL DESIGN
LANDFILL DESIGN
Regarding pollutant control, the design of a landfill should take into
account three factors: the hydrologic system governing direction of pollutant
travel, the geochemistry of the water-sediment system, and the release rate
of unattenuated pollutants to surface or ground waters. The first item has
been the subject of a number of papers and will not be addressed here.
Current landfill design and,engineering practice are to construct clay
liners, either natural or artificial, very thick and containing high clay
percentages. The motive is to create relatively impermeable liners that will
contain the leachate and therefore protect the groundwater resources. This
approach can create difficulties in humid climates where infiltration exceeds
the capacity of the liner to dissipate the leachate. This causes what is re-
ferred to as the "bathtub" effect wherein the relatively impermeable clay
liner acts as a bathtub which fills up with leachate and then overflows. If
the leachate is not collected and treated, the overflow manifests itself in
the form of leachate springs on the surface and results in surface water
pollution and surface environmental degradation instead of ground water pollu-
tion. Neither form of pollution is desirable; both might be prevented by
using proper design features in the construction of the sanitary landfill.
If one does not choose to collect and treat the leachate and wishes to avoid
the "bathtub," the results of this research suggest an alternative landfill
design.
The chemical attenuation study reported in Section 5 has indicated that
most of the toxic constituents found in municipal leachates are moderately-
to-highly attenuated by passage through laboratory columns containing rela-
tively low percentages of clay minerals. If the assumption is granted that
the "bathtub" effect is an undesirable feature of clay liners -, then it
follows that it is desirable to determine the point of "optimal" attenuation,
i.e., the percentage of clay in a liner material which gives the maximum
attenuation balanced with a maximized hydraulic conductivity. Figure 43
represents a dual scaled graph with the initial hydraulic conductivity of
the montmorillonite columns as given in Table 3 plotted as a function of per-
centage of montmorillonite. The opposite scale is the attenuation number for
the chemical constituent of interest also plotted as a function of the per-
centage of montmorilldhite. The attenuation numbers (Table 5), as reported in
Section 5, are the percentage of removal of the element from the leachate upon
passage through 10 pore volumes of the clay-sand mixture. The attenuation
123
-------
Hydraulic conductivity
024
8 12 16 20 24 28 32 36 40 44 48 52 56 60 64 68 72
Montmorillonite (%)
Figure 43. Effect of clay content on hydraulic conductivity and
attenuation of Pb, NH^, and Cl for a 40-cm thick liner.
The bulk densities vary; densities for each clay content
are shown in Table 3.
124
-------
number scale is given as 0 at the point of minimum hydraulic conductivity and
100 at the point of maximum hydraulic conductivity.
For the heavy metals (Pb is used here as an example), even small amounts
of clay gave almost total removal. The heavy metals, even though toxic,
represent a minimal pollution hazard in municipal leachates because they are
attenuated very strongly and are usually present at low levels in leachates.
Therefore they can usually be ignored as far as determining the optimal clay
liner for a given leachate. At the other extreme are the relatively non-
interacting constituents represented by Cl. Chloride as shown here is rela-
tively unattenuated by even large amounts of clay. Figure 43 suggests that in
order to prevent chloride migration, relatively impermeable clay liners would
be necessary. Due to the non-toxic nature of the chloride ion, it also ranks
low along with the heavy metals in the pollution hazard index (Table 6). In
view of the problems associated with the "bathtub" effect, it seems unwise
to design clay liners to optimize chloride attenuation. Rather, it seems
prudent to design clay liners for optimum attenuation of the most hazardous
constituents found in a particular leachate. In the case of the DuPage leach-
ate used in this study, the pollution hazard index ranks NHj as 30 times more
of a pollution hazard than any other constituent found in this leachate. It
therefore seems reasonable to design a clay liner for DuPage leachate that
gives optimal attenuation of NH^ and all the other constituents should also
be attenuated to relatively safe levels for minimal pollution of the ground-
waters adjacent to the landfill site.
For the case of NH, (shown in Figure 43) if one extrapolates the curve,
it is apparent that 18-20% montmorillonite would give nearly total removal
from the leachate. If the 10% liner is doubled in thickness from the 40 cm
used in this study to 80 cm, it will contain enough montmorillonite to give
nearly total removal of the Nflt in 10 pore volumes of leachate and still
retain the relatively high hydraulic conductivity of 6 x 10~^ cm/sec. This
illustrates that, for a given desired hydraulic conductivity, the removal
capacity of the liner may be adjusted by changing its thickness (at the same
clay content) without greatly affecting transmission of water through it.
LINER THICKNESS
The thickness of various mixtures of sand and clay, as represented by
the total cation exchange capacity (CEC), to achieve total attenuation of
selected relatively mobile ions was calculated (Table 15). The removal ef-
ficiency will differ in leachates, depending upon relative ion strength. The
efficiencies used in Table 15 are based on the DuPage leachate used in this
study. The concentrations are given in parentheses under the average concen-
tration value. Increasing cation exchange capacity generally reflects in-
creasing clay contents. Thus a thicker liner with greater hydraulic con-
ductivity and lower CEC may be the optimal liner for attenuation.
Determining the release rate of nonattenuated or poorly attenuated con-
taminants from the clay liner (natural or man-made) to surface waters or
ground-water aquifers is necessary for good landfill design. In designing a
landfill, a decision must be made as to which ions should be totally attenuated
125
-------
TABLE 15. ESTIMATED LANDFILL LINER THICKNESS NECESSARY FOR ATTENUATION OF SOME LEACHATE
CONSTITUENTS PER CUBIC METER OF REFUSE DURING A 20 YEAR FILL LIFE*
NJ
ON
Constituent Initial**
Concentrat ion
NHi,
Na
K
Mg
Ave.
(ppm)
379
(830)
755
(740)
763
(530)
1,609
(240)
Max.
(ppm)
1,106
7,700
3,770
15,600
Ave.
(cm)
32
118
51
226
10
Max.
(cm)
92
1,208
252
2,191
CEC
Ave.
(cm)
16
59
26
113
20
Max.
(cm)
46
604
126
1,096
30
Ave.
(cm)
11
39
17
75
Max.
(cm)
31
403
84
730
*Assumption: Bulk density = 1.8 g/cc; 100 liters of leachate generated per m3 of refuse/yr.;
initial concentration decreases linearly to zero at 20 yr.; removal efficiencies for each
constituent were estimated using the average values given by Griffin, et al. (1975), NHi* = 37.1%,
Na = 15.4%, K = 38.2%, Mg = 29.3%.
**Concentrations taken from the twenty leachate analyses reported by EPA (1974) ; those in
parenthesis, (830), are the values of Old DuPage landfill leachate used in this study.
-------
and which should be eventually released to the environment. The chloride ion,
which moves essentially unattenuated, is the most obvious example of the latter
type.
CONCLUSIONS
The results of this study suggest an alternative to landfill design. The
use of hydraulic conductivity information and the pollution hazard rating for
a given leachate or waste stream can allow a different approach than the
prevalent "containment" type liner systems. These data suggest that overall
pollution would be lessened by designing landfill liners for higher permea-
bility and by selectively attenuating the most toxic pollutants from
leachate and allowing the groundwaters to dilute the nontoxic components
which can be tolerated at much higher concentrations than the toxic without
deleterious effects. Thus landfill stabilization and use for other productive
purposes could be achieved at much faster rates than can presently be
achieved by containment type liners.
127
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SECTION 14
REFERENCES
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Apgar, M. A., and D. Langmuir, Ground Water v. 9, p. 6 (1971).
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Brunauer, S. , P. H. Emmett, and E. Teller, Adsorption of Gases in Multi-
molecular Layers: J. Am. Chem. Soc. v. 60, p. 309-319 (1938).
Butler, J. N., Ionic Equilibrium - A Mathematical Approach, p. 360-363
Addison-Wesley, Reading, Mass. (1964). '
Cartwright, Keros, and M. R. McComas, Ground Water v. 6. no. 5 t>. 23-30
(1968).
Davids, H. W., and M. Lieber, Water and Sewage Works v. 98, no. 12 p. 528-
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APPENDIX A
LIST OF PUBLICATIONS
Griffin, R. A., and N. F- Shimp. 1975. Leachate migration through selected
clays. In "Gas and Leachate from Landfills: Formation, Collection, and
Treatment," E. J. Benetelli and John Cirello, Eds., Ecological Research
Series, EPA-600-9-76-004, U. S. Environmental Protection Agency,
Cincinnati, Ohio 45268. p. 92-95.
Griffin, R. A., and N. F. Shimp. 1975. Interaction of clay minerals and
pollutants in municipal leachate: WateReuse v. 2, p. 801-811.
Griffin, R. A., R. R. Frost, and N. F. Shimp. 1976. "Effect of pH on removal
of heavy metals from leachates by clay minerals," in Residual Management
Land Disposal, W. H. Fuller {Ed.}, Ecological Research Series,
EPA-600/9-76-015, U. S. Environmental Protection Agency, Cincinnati,
Ohio 45268. p. 259-268.
Griffin, R. A. , and N. F. Shimp. 1976. Effect of pH on exchange-adsorption or
precipitation of lead from landfill leachates by clay minerals: Environ.
Sci. and Technology v. 10, p. 1256-1261.
Griffin, R. A., Keros Cartwright, N. F. Shimp, J. D. Steele, R. R. Ruch,
W. A. White, G. M. Hughes, and R. H. Gilkeson. 1976. Attenuation of
pollutants in municipal landfill leachates by clay minerals, Part 1 —
Column leaching and field verification: Illinois State Geological Sur-
vey, Environmental Geology Notes 78, Urbana, Illinois, 34 p.
Griffin, R. A., R. R. Frost, A. K. Au, G. D. Robinson, and N. F. Shimp. 1977.
Attenuation of pollutants in municipal landfill-leachates by clay
minerals, Part 2 — Heavy metal adsorption studies: Illinois State
Geological Survey, Environmental Geology Notes 79, Urbana, Illinois,
47 p.
Frost, R. R., and R. A. Griffin. 1977. Effect of pH on adsorption of arsenic
and selenium from landfill leachate by clay minerals: Soil Sci. Soc. Am.
Jour., v. 41, p. 53-57.
Frost, R. R., and R. A. Griffin. 1977. Effect of pH on removal of copper,
zinc, and cadmium from landfill leachate by clay minerals: Jour.
Environ. Sci. and Health, Part A, v. 12, p. 139-156.
Cartwright, Keros, R. A. Griffin, and R. H. Gilkeson. 1977. Migration of
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landfill leachate through unconsolidated porous media: Advances in
Groundwater Hydrology, v. 12:(in press).
Griffin, R. A., A. K. Au, and R. R. Frost. 1977. Effect of pH on adsorption
of chromium from landfill-leachate by clay minerals: Jour. Env. Sci. and
Health, Part A, v. 12, p. 431-449.
Cartwright, Keros, R. A. Griffin, and R. H. Gilkeson. 1977. Migration of
landfill leachate through glacial tills: Ground Water, v. 15,
p. 294-305.
Griffin, R. A., and A. K. Au. 1977. Lead adsorption by montmorillonite using
a competitive-Langmuir equation: Soil Sci. Soc. Am. Jour., v. 41,
p. 880-882.
Griffin, R. A., N. F. Shimp, J. D. Steele, R. R. Ruch, W. A. White, and
G. M. Hughes. 1976. Attenuation of pollutants in municipal landfill-
leachate by passage through clay: Environ. Sci. and Technology v. 10,
p. 1262-1268.
135
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APPENDIX B
PROCEDURES USED IN CHEMICAL AND PHYSICAL CHARACTERIZATION
OF CLAY MINERALS AND LIQUID SAMPLES
Clay Mineral Characterization
(W. A. White)
The kaolinite used in this study was collected from materials of Penn-
sylvanian age in Pike County, Illinois. The site description and location are
given by White (1959) , sample 996N. The montmorillonite used in the study was
southern bentonite from the American Colloid Company. The illite was from the
Minerva Company Mine^of the Allied Chemical Company.
The clays were brought to the laboratory, where they were crushed,
ground, and purified by sedimentation techniques to obtain the <2 ym particle
fraction that contained essentially pure clay minerals. The purified clay was
then floculated with 1 M CaCla, centrifuged, and spray dried. The clay
minerals present in the <2 ym fraction were identified by X-ray diffraction
techniques (Parham, 1962), and the results are given in Figure 44. The
positions of the diffraction peaks identify which clay minerals are present
in the sample. In addition, the areas under the peaks of the diffraction
patterns give a crude estimate of the relative amounts of each clay mineral
in the sample and were used to prepare the following summary:
(a) The montmorillonite sample was found to be almost completely
monomineralic, containing approximately 95% montmorillonite
and 5% mixed-layer materials close to montmorillonite in
composition.
(b) The illite sample was found to be monomineralic, with mica
minerals predominant. The sample contained approximately 70%
illite and 30% mixed-layer materials close to illite composition.
(c) The kaolinite sample could be characterized as a moderately
pure kaolinite, comparable to the Georgia hard kaolins. The
sample contained approximately 87% kaolinite, 8% illite, 5%
mixed-structure material, and a trace of quartz.
Atomic Absorption Methods for Cd, Fe, Mn, Pb, Si, Zn, Na, K, Ca, and Mg
(John Steele and David Heck)
136
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Illite, ;
ethylene- ;
glycol
Montmorillonite,
.ethylene-glycol
Kciolinite M
36
25 15
Degrees 29
Figure 44. X-ray diffraction patterns of the clay minerals used in
leachate pollutant attenuation studies.
137
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Atomic absorption methods were used for the following determinations:
filtered and acidified leachate (dissolved metals) — Cd, Fe, Mn, Pb, Si,
and Zn; membrane filters (suspended metals) — Cd, Fe, Mn, Pb, and Zn; am-
monium acetate extracts (exchangeable cations) — Cd, Fe, Mn, Pb, Zn, Na, K,
Ca, and Mg; and column sections and original clays (total cations) — Cd, Mn,
Na, Pb, and Zn.
Atomic absorption measurements were made using a Perkin-Elmer Model 306
Atomic Absorption Spectrophotometer. In the case of the dissolved, suspended,
and exchangeable metal determinations, measurements were recorded directly
from the electronic digital readout in the auto-concentration mode. For the
total cation determinations, absorbance measurements were recorded on a
Perkin-Elmer Model 056 Recorder. The following Perkin-Elmer burner heads were
used: a 4-inch-long slot, flat-head burner using an air-acetylene flame for
all of the dissolved and suspended cation determinations except Si; a 2-inch-
long slot nitrous oxide burner head using a nitrous oxide-acetylene flame for
Si determinations; and a three slot high solids burner head with an air-
acetylene flame for all of the exchangeable and total cation determinations.
Standard single element hollow cathode lamps were used. Corrections for back-
ground absorption were made simultaneously for total cation determinations
using a Perkin-Elmer Deuterium Arc Background Corrector.
All reagents used are ACS certified reagent grade chemicals, and stan-
dard stock solutions were prepared from high purity metals or compounds.
Calibration standards prepared from diluted stock solutions contained the
following matrices: dissolved metal standards — 1% v/v distilled HN03; sus-
pended metal standards — 4% v/v distilled HC1; exchangeable cation stan-
dards — IN ammonium acetate and 7.5% v/v distilled HNOs; and total cation
standards -- 1.2% v/v 48% HF, 1.5% v/v aqua regia (1:3:1 HN03-HC1-H20) , and
0.54% w/v HsBOa. For total Na determinations, the standards contained, in
addition to the above mentioned matrix for total cations, 1000 ppm Cs as
CsCl. For determination of Ca and Mg in the leachate fraction, the standards
were prepared to contain 1% w/v La20a and 2.5% v/v HC1. For determination of
Na and K, the standards were prepared to contain 1000 ppm Cs as CsCl.
Sample preparation
The filtered-and-acidified leachate and the ammonium acetate extract
samples required no sample preparation and were analyzed directly except that
for the analysis of Ca and Mg, 1% L^Os w/v and 1% v/v HC1 were added. Also
for the determination of K and Na, the final sample contained 1000 ppm Cs as
CsCl.
Column section dissolution
The methods described below are modifications of those of Bernas (1968)
and French and Adams (1973). Approximately 0.5 g of the clay-sand mixture,
previously dried at 110° C for several hours, is transferred to a 125 ml
linear polyethylene screw cap bottle. The sample is wetted with 1.5 ml aqua
regia, and 1.2 ml of 48% HF is added. The cap is screwed tight, and the
bottle is placed in a steam bath for 2 hours. The sample and acids are mixed
by an occasional swirling of the bottle during this period. The bottle is
138
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removed from the steam bath, and 9 ml of a H3B03 solution (0.06 g/ml) is
added to the sample to complex the fluoride. The bottle is again sealed and
placed on the steam bath for 0.5 hours, removed, and allowed to cool. The
dissolved sample is transferred to a 100 ml polyethylene volumetric flask,
the bottle is'washed repeatedly with deionized H20, and the washings are
added to the flask. The sample is diluted to volume with deionized H20 and
returned to the polyethylene bottle for storage.
Membrane filter dissolution
The membrane filters were treated according to methods modified from
Parker (1972). The membrane is placed in a 250 ml Pyrex beaker, and 3 ml of
distilled HN03 is added. The beaker is covered with a watch glass and heated
on a hot plate where the acid is evaporated to dryness. The additions of acid
and the heating steps are continued until a light residue is left. The resi-
due is taken up with 2 ml of distilled 1:1 HC1 and warmed gently. The sample
is then transferred to a 50 ml Pyrex volumetric flask, and the beaker is
washed repeatedly with small portions of deionized H20 and the washings added
to the flask. The sample is diluted to volume and transferred to a poly-
ethylene bottle for storage.
Atomic Absorption Spectrophotometric Procedures
The following analytical wavelengths were used: 228.8 nm (Cd), 248.3 nm
(Fe), 279.5 nm (Mn), 589.0 nm (Na), 283.3 nm (Pb) , 251.6 nm (Si), 213.9 nm
(Zn) , 422 nm (Ca), 285 nm (Mg) , and 766.5 nm (K). Where necessary, samples
are diluted to bring the metal concentration within the linear portion of the
calibration curve. For total Na determinations, samples are diluted 1 to 2
with the addition of 1000 ppm Cs as CsCl. In the case of the suspended, dis-
solved, and exchangeable metal determinations, the metal ion concentrations
are determined directly from the electronic digital readout in the auto-con-
centration mode, after appropriate calibration. For the total cation deter-
minations, the metal ion concentrations are calculated by solving for con-
centration in a least squares constructed calibration equation of absorbance
vs. concentration. A new calibration curve equation is calculated for each
set of analyses.
An estimate of the average relative standard deviation for the elements
determined by atomic absorption spectrophotometry is 10 percent or better.
Determination of Hg by Flameless Atomic Absorption
(R. A. Griffin and G. D. Robinson)
The Hg samples obtained in the investigation of the removal character-
istics of Hg from leachate solutions by clay minerals as a function of pH
were analyzed by flameless Atomic Absorption Spectroscopy (A.A.) using a
slight modification of the procedure described by the U. S. Environmental
Protection Agency (EPA) (U.S.-EPA, 1971).
Variations of the standard procedure described by U.S.-EPA (1971) were
implemented in order to gain maximum sensitivity from the procedure and the
available instrumentation. The apparatus used was a Perkin-Elmer Model 360
139
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A. A. equipped with a Perkin-Elmer Flameless Mercury Analysis System. The
absorption cell supplied by Perkin-Elmer was modified by replacing the plas-
tic end windows with quartz windows which were glued in place with epoxy
cement. This allowed greater light transmission and higher sensitivity and
stability to be achieved for the Hg analyses.
Reagents used:
Sulfuric Acid reagent grade cone.
Nitric Acid reagent grade cone.
Potassium Permanganate 5% w/v 5g KMnCK in 100 ml DI H20
Potassium Persulfate 5% w/v 5g K2S208 in 100 ml DI H20
Hydroxylanine Hydrochloride 5% w/v 5g NH2OH • HCl in 100 ml DI H20
Stannous Chloride - Dissolve 50g SnCla in 100 ml cone. HCl and 400 ml DI H20
Stock Mercury Solution 1000 ppm - Dissolve l.OSOg HgO in min. vol. 1:1 HCl -
dilute to 1 liter with DI H20
Working Mercury Solutions - Make successive dilutions of the stock Hg solu-
tion to obtain a standard containing 0.1 ppm. This should be made fresh
daily and acidified to .15% HNOa. Note: It was found that 10 and 25 ppm
working Hg standards could be stored in plastic bottles several weeks
if acidified to .15% HN03 with no apparent loss or adsorption of Hg.
This greatly reduced the steps involved in making fresh working stan-
dards of .1 ppm.
Waste Hg Absorbing Media - . 1 M KMnOi» and 10% H2SOif.
A calibration curve was obtained following the procedure outlined in
U. S.-EPA (1971) substituting a hydroxylamine-HCl solution for the hydroxyla-
mine sulfate solution, and a stannous chloride solution for the stannous
sulfate solution. It was found that this calibration curve had to be repeated
with each set of samples.
The samples for Hg analysis were placed in 60 ml plastic bottles and
acidified with cone. HNOs to pH ~2. A 1 ml aliquot was then taken from each
sample and placed in a 300 ml BOD bottle (these were run in triplicates) ;
the volume in the bottle was then brought up to 100 ml with DI H20. The
samples were then digested to oxidize the organo-mercury compounds which
would not otherwise respond to the flameless A. A. technique (U. S.-EPA,
1971).
The samples were digested by addition of cone. H2SOij and 2.5 ml of cone.
HN03 to each bottle with mixing after each addition. One ml of 5% w/v KMnOi,
solution was then added to each bottle, swirled, and allowed to stand for at
least 15 minutes. Two ml of K2S2Os were then added and allowed to stand for
at least 30 minutes. Two ml of NH2OH'HC1 were then added to each bottle to
reduce the excess permanganate. Immediately prior to aerating through the
absorption cell, 2 to 3 ml of SnCl2 solution were added to each bottle. The
samples were read on the absorbance mode and then converted to concentration
mathematically using linear regression analysis (calibration standards were
prepared fresh daily). The final Hg concentrations were then subtracted from
the initial concentration (0.962 ppm) to obtain the amount of Hg removed
from solution.
140
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Neutron Activation Analysis Methods for Hg and As
(Joyce Frost and Larry Camp)
Irradiations are tarried out in the University of Illinois' Advanced
TRIGA reactor, with a thermal neutron flux of 1.4 x 1012 neutrons cnT2 sec'1
(500 KW). The samples are contained in a rotary specimen assembly. Rotation
at 1 rpm during irradiation ensures an equivalent integrated flux for each
sample.
Samples and standards are irradiated in heat-sealed polyethylene snap-
cap vials which were previously cleaned with deionized water and acetone.
The gamma-ray spectrometer system consists of a 3" x 3" Nal(Tl) crystal
connected to a 400-channel analyzer. The analyses are comparative — e.g. the
activity of an element in the sample is compared to that of the element in a
known standard to determine concentration of the element in the sample.
Determination of Hg by Instrumental Neutron Activation Analysis in "Spiked"
Liquid Samples
This procedure was used for the column effluent samples (filtered and
acidified leachate samples), solubilized membrane filter samples, and am-
monium extract solutions of the clay column samples that originated after the
original sterile leachate was spiked with mercury to a resulting concentra-
tion of 4.0 ppm.
A two-hour irradiation was carried out on 5 ml portions of the samples
and two acidified mercuric nitrate standards, one containing 5.02ygHg/ml and
the other 0.502ygHg/ml. The samples were left to decay one week to decrease
interference from short-lived radioisotopes, especially 15 hour 2 Na and 36
hour 82Br which gives a high background in the low energy region of the gamma
ray spectrum. The activity in the samples and standards due to the 0.077 MeV
gamma ray of 65 hours 197Hg was then counted. At the same time uranium was
determined in those column effluent samples that unexpectedly had a high con-
centration of uranium, up to 0.20 ppm; these were the first several sets of
effluents collected from the columns that contained montmorillonite. The
uranium activity counted was the 106.1 keV gamma ray of 56.4 hour 239Np; the
standard was 5 ml of a solution of U02(N03)2 containing 1.02 yg U/ml.
Determination of Hg in Liquid Samples by Neutron Activation Analysis with
Radiochemical Separation
This procedure was used for the original leachate and for the precise
determination of mercury in some of the samples having a low mercury content.
The method follows closely that of Weiss, Koide and Goldberg (1971).' The
samples (1 to 3 irradiation vials, each containing 10 ml per sample) and
standard (10 ml of 5.02ugHg/ml) were irradiated for 2 hours. After a day's
decay, each sample was transferred to a beaker, and 10 mg of Hg-^ "carrier was
added. The pH of the solution was adjusted to 1 and elemental mercury pre-
cipitated by the addition of stannous chloride solution. The mercury was
separated on a Millipore filter and dissolved in aqua regia. This solution
141
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was neutralized with ammonium hydroxide, and thioacetamide solution was
added. The precipitate of mercuric sulphide was separated on a Millipore
filter; dissolved in aqua regia; transferred to a tared, glass, stoppered
vial; and made up with deionized water to 25 ml. The activity of the solution
due to the 0.077 MeV gamma ray of 197Hg was counted. The radiochemical
separation for the standard (250 yl) plus carrier was begun with the mercuric
sulphide precipitation step. Radiochemical yields (85-98%) were determined by
re-irradiation of a 1 ml aliquot of each sample and the standard.
Determination of Hg in Column Sections and Clay
One gram of sample was accurately weighed and irradiated for two hours
as previously described. The irradiated sample was then mixed with ~1 gram of
Alundum and transferred to a combustion boat which has had Hg carrier added
as Hg(NOs)a. The boat was then placed in a Vicor tube which in turn is placed
in a Lindberg "Heavy Duty" furnace, such that Oa can flow through and the
combustion gasses trapped in a dry ice cold trap. The furnace is operated for
15 minutes at 1000° C and allowed to cool.
The boat was removed from the cooled tube (at most, 400° C) and the cold
trap section (dry ice removed) placed end down into a polyethene 100 ml cen-
trifuge tube.
The foresection of the tube was washed with 10 ml of HNOa, and this
solution via transfer pipplets was used to wash down the cold trap section of
the tube. This was repeated twice more, but HaO substituted for the HNOs,
giving 30 ml of total solution. The centrifuge tube was placed in HaO bath
and heated. When hot (90° C or +), the solution was stirred and solid NH^Br
(AR) added, red Bra was evolved, and the solution was stirred. Solid Ag(NOs)a
was added to remove the excess Br (much of the radioactive Br had isotopical-
ly exchanged) and the solution was centrifuged. The liquid was transferred to
a 4 oz. wide mouth jar and labeled.
Ten ml of warm (90+° C) HN03-H20 (1-3) was added to the ppt in tube and
again centrifuged. This solution was added to the original 30 ml, and the
solution counted for 77.6 KEV y ray of Hg197 (65.5 hour t^) on a Nal detec-
tor. One ml of this solution was removed and put in 1 dram vial, sealed, and
labeled and irradiated at 15 minutes at 500 KW- This was compared with a
"carrier vial" and a "yield" factor found for correction.
Determination of Arsenic in Liquid Samples by Neutron Activation Analysis
with Radiochemical Separation
Ten ml of sample and 10 ml of an arsenic trioxide standard were irradi-
ated for 1 hour, then left to decay for a day. The sample with 30 mg of As"1"1"1"
carrier added, in a distillation flask,^was oxidized by boiling it for 5 min.
with 3 ml of 30% HaOa, to decompose organic arsenic compounds present (EPA,
1971). Then, after the addition of hydrochloric and hydrobromic acids,
arsenic was distilled from the solution (NAS-NS #3002, 1965). Elemental
arsenic was precipitated from the distillate by the addition of sodium hypo-
phosphite. The precipitate was separated on a weighed filter paper, re-
weighed, and the activity due to the 0.56 MeV gamma ray of 26.5 hour 76As was
142
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counted. The standard was carried through the same radiochemical separation.
Determination of Arsenic in Clay Samples by Neutron Activation Analysis with
a Radiochemical Separation
The procedure followed was that used by Ruch, Kennedy and Shimp (1970)
for the determination of arsenic in sediments. The irradiated clay samples,
with added As"1"*"1" carrier, are dissolved with hydrofluoric and perchloric
acids and hydrogen peroxide, and the silicon tetrafluoride evaporated off.
The residue was taken up in hydrochloric acid and arsenic was separated by a
hydrogen bromide distillation; the procedure henceforth was the same as that
used for liquid samples.
X-Ray Fluorescence Method for Mg. Al. Si, Ca. K. Ti. Fe, and P
(John Kuhn and Ray Henderson)
The column section samples and clays were air dried then ground to ~200
mesh and oven dried at 105° C for 6 hours. From the dried sample, 125 mg was
weighed into a graphite crucible containing 1.000 g of lithium tetraborate. A
depression made in the tetraborate prior to addition of the sample prevented
its contact with the crucible wall. Next, 125 mg of lanthanum oxide was added
as a heavy-element absorber, and the contents of the crucible' were mixed,
with a glass rod, as thoroughly as possible without scraping the crucible
wall or bottom. The mixture was fused in a furnace for 15 minutes at 1000° C,
removed, covered with a second crucible and allowed to cool to room tempera-
ture. The resulting fused pellet was weighed alone to determine fusion loss
and placed in the grinding vial of a No. 6 Wig-L-Bug with 2 percent by weight
of Somar Mix (a commercial mixture used as a grinding and plasticizing
agent). The sample was ground for 3 minutes, transferred to a die, and
pressed at 40,000 psi. Samples were then exposed to X-rays and the data com-
pared to values acquired from standard rock pellets prepared in the same
manner. The resulting concentrations determined on the samples are quantita-
tive for those major and minor elements listed (Mg, Al, Si, Ca, K, Ti, Fe,
and P).
Optical Emission Spectrometry Method for Be. V. Cr, Co. Ni, Cu, and Mo
(Gary Dreher)
Direct reading optical emission Spectrometry was used to determine trace
element concentrations in the original clay materials. The procedure used was
to prepare a mixture of 1 part Ba(N03)2 (5 mg) , 8 parts sample (40 mg) and 31
parts SPEX graphite powder (Spex Industries, Inc.) (155 mg) in a 1/2 inch di-
ameter, 1 inch long polystyrene vial containing two 1/8 inch diameter poly-
styrene beads. The sample mixture was shaken in a Wig-L-Bug for 60 seconds.
Fifteen milligrams of this mixture was loaded into each of 4 electrodes, 1/8
inch in diameter, having thin wall craters. The sample electrodes were arced
at 15 A. short-circuit current for 65 seconds with an arc gap of 4 mm, and
surrounded by a 10 SCFH flow of a gas mixture 80% in argon and 20% in oxygen.
Determination of Total Carbon (TC). Inorganic Carbon (1C), and Organic
143
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Carbon (PC)
(David Heck and Larry Kohlenberger)
Total Carbon (TC) and Inorganic Carbon (1C) were determined in the
column sections and clay samples. Organic Carbon (OC) was determined by dif-
ference of TC and 1C. The procedures are as follows:
Total Carbon
A zirconium silicate combustion boat containing a layer of dry alumina
powder (to prevent sample fusion to boat) is accurately weighed. A
layer of sample (1.5 to .30 g) is added, and the boat is reweighed in
order to obtain the sample weight. A Nesbitt absorption bulb containing
a bottom layer of magnesium perchlorate and topped with C02 absorbent
(lithasorb or ascarite) is also weighed. The weighed absorption bulb is
placed on the exit end of a zirconium silicate combustion tube which is
heated to 1350° C in a high temperature furnace. With dry oxygen flow-
ing through the tube continuously, the sample boat is slowly pushed
into the hot zone at a rate which will neither crack the boat nor blow
out the sample. After the sample has been in the hot zone for 4 minutes,
the sample boat is removed, the absorption bulb closed off, and the
sample allowed to cool. It is reweighed, the C02 absorbed determined by
difference, and % total carbon calculated.
Inorganic Carbon
A modified ASTM (1968) procedure, D 1756, is used whereby C02 is formed
by decomposing ~0.5 g sample with acid and absorbing the gas on Litho-
sorb.
Methods for NHt. TDS. COD, and Chloride
(Don Dickerson)
r
Ammonium ion (Nttf) was determined in the filtered (but not acidified)
leachate samples as follows. A suitable volume of leachate (1 ml 0.01 N HC1
is equivalent to 180 mg Nflt/liter) is placed in a Kjeldahl flask and diluted
to -20 ml with NHt free water. The flask is connected to the Kjeldahl dis-
tillation apparatus. The receiver contains 30 ml of 4% boric acid solution.
Fifty percent KOH solution is slowly added into the flask port followed by
5 ml of NaOH-Na2S203 solution (25 g NaOH, 5 g Na2S203, and 75 ml H20). The
solution is steam distilled for 4.minutes after NHt starts coming over. Three
drops of methyl red-bromocresol green indicator is added and the solution is
immediately titrated to the orange-pink end point withf>Q.01 N HC1 solution.
Determination of Exchangeable
Exchangeable NH£ was determined on the column sections according to a
modified procedure by Jackson (1958). One sample (containing approximately
.8 - 1.0 mg Nfft) is analyzed by the normal Kjeldahl procedure. A second
sample is combined with 20 ml of 1.0 N, acidified NaCl and distilled simi-
larily following the Kjeldohl procedure.
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Determination of Chloride
Chloride was determined in the filtered (but not acidified) leachate
samples as follows. A sample containing not more than 4 mg of Cl~ is diluted
to 25 ml with Cl-f ree water and then combined with 25 ml of isopropyl al-
cohol. While stirring, 2 drops of bromophenol blue is added and the pH ad-
justed to slightly acid by the addition of 1.0 N NaOH solutions or 1.0 N HN03
solutions as needed. One half ml of a saturated solution of diphenyl carba-
zone in isopropyl alcohol is added and the solution titrated to the endpoint
with 0.1 N Hg4* (as Hg(N03)2 in dilute HN03).
Determination of Chemical Oxygen Demand (COD)
The following procedure, adopted from Leithe, (1947) was used to analyze
filtered (but not acidified) leachate samples. A 1-10 ml leachate sample di-
luted to 50 ml with distilled water and 25 ml 0.10 N KaCraO? solution was
added to 0.5 g HgSOij in a 300-ml ground-joint flask with a Friedrich con-
denser. Concentrated sulfuric acid (75 ml) containing 1 g AgaSOij was added in
small portions with thorough swirling as well as two boiling stones as anti-
bumping aids, and refluxing was carried out for 2 hours. The mixture was
cooled, washed into a 500-ml Conical flask, diluted to 350 ml with distilled
water, treated with 3 drops ferroin indicator and titrated with 0.1 N Fe
(NHif) 2(80^)2 solution to a stable color change from blue-green to reddish-
brown. A blank value was determined using 50 ml of distilled water under the
same conditions.
Determination of Total Dissolved Solids (TDS)
TDS was determined in the filtered (but not acidified) leachate samples
as follows. Ten ml was placed in a weighed beaker, covered with a watch-
glass, and placed in an oven preheated to 250° F until evaporation was com-
pleted. The temperature was then raised to 350-375° C for an hour. The resi-
due was finally cooled in a vacuum desicator and weighed.
Methods for Sulfate, Phosphate, and Boron
(David Heck)
Sulfate was determined in the original leachate solutions as follows.
Fifty ml of sample was neutralized with concentrated HCl and then 0.5 ml
added in excess and the sulfate precipitated in almost boiling solution with
10 ml of 10% BaCl2 added drop by drop with constant stirring. After digesting
one hour until the BaSOi, settled, another drop of barium chloride was added
to make sure the precipitation was complete, and the BaSOi, filtered off,
washed with hot water, ignited in the muffle furnace one hour at 1000° C and
weighed as BaSOi*.
Determination of Phosphate
Phosphate ion was determined in the original leachate solutions as fol-
lows. Fifty ml of sample was brought to boil with 15 cc HN03 and 40 cc H20,
digested for an hour, cooled, filtered. The filtered material was washed with
a dilute HN03-NH^N03 solution. The residue was discarded, the filtrates were
145
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combined and 11 ml ammonium citrimolybdate reagent added, brought cautiously
to a boil, and then let to stand overnight. The resulting precipitate was
filtered through a fritted crucible, washed with the above solution and
water, dried at 100-105° C, cooled, and weighed.
Determination of Boron
Boron was determined on the column sections and clay samples according
to the following procedure.
A 25-50 mg solid sample was weighted into a clean 4 oz. plastic bottle,
and 10 ml of 2.5 NH2SOn and 4 ml 5% HF added. The bottle was fitted with a
tight cap, swirled, and allowed to stand overnight. The sample was diluted to
a total volume of 20 ml, and 10 ml of .001 M methylene Blue (.374 g/1) and
25 ml of purified ethylene dichloride added. The bottle was shaken for at
least 30 minutes, removed from shaker, and layers allowed to separate. A 5 ml
portion of the lower layer was pipetted into a 25 ml volumetric flask and
diluted to volume with ethylene dichloride. Unknown sample was then compared
to known standards (1-6 ppm range) on a Beckman DBG Spectrophotometer at 660
Vim. A reagent blank was required for standards and samples. Boron was deter-
mined in the leachate fractions according to the same procedure, however, two
ml were used as the sample.
Determination of Base Exchange Capacity
(Bill Armon)
Base Exchange Capacity was determined on the column sections according
to a procedure by Peech (1947), with slight modification. The 10 g sample was
leached with ammonium acetate to remove exchangeable cations and to saturate
the exchange complex with ammonia. The excess ammonia was removed by leaching
with ethyl alcohol and the remaining, exchangeable ammonia removed and deter-
mined by distillation in a Kjeldahl apparatus.
Determination of Surface Area
(Joe Thomas)
The surface areas of the clays were determined by the B.E.T. method
(Brunauer, Emmett, and Teller, 1938) using nitrogen as the adsorbate in a
continuous flow system (Nelsen and Eggertsen, 1958),
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
. REPORT NO.
EPA-600/2-78-157
4. TITLE AND SUBTITLE
ATTENUATION OF POLLUTANTS IN MUNICIPAL LANDFILL LEACHATE
BY CLAY MINERALS
3. RECIPIENT'S ACCESSION NO.
5. REPORT DATE
August 1978 (Issuing Date)
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
R. A. Griffin and N. F. Shimp
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
Illinois State Geological Survey
Urbana, Illinois 61801
10. PROGRAM ELEMENT NO.
1DC618
11. CONTRACT/8R£NX NO.
68-03-0211
12. SPONSORING AGENCY NAME AND ADDRESS
Municipal Environmental Research Laboratory--Cin., OH
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
13. TYPE OF REPORT AND PERIOD COVERED
Final Report 12/72 to 8/75
14. SPONSORING AGENCY CODE
EPA/600/14
15. SUPPLEMENTARY NOTES
Project Officer: Mike H. Roulier (513) 684-7871
16. ABSTRACT
The first part of this project was a laboratory column study of attenuation of
pollutants in municipal solid waste landfill leachate by mixtures of sand and calcium-
saturated clays. Chloride, Na, and COD were relatively unattenuated by passage through
the clay columns; K, NHd, Mg, Si, and Fe were moderately attenuated; and the heavy
metals Pb, Cd, Hg, and zn were strongly attenuated even in columns with small amounts
of clay. Calcium, B, and Mn were higher in column effluents than in the applied
leachate. Precipitation was the principal attenuation mechanism for the heavy metals;
cation exchange was responsible for any attenuation of the other elements. The clays,
in order of increasing attenuation capacity, were Kaolinite, Illite, Montmorillonite.
The second part of the project involved batch studies of adsorption of Cr, Cu, Pb,
Cd, Hg, and Zn by Montmorillonite and Kaolinite from water solutions and from landfill
leachate. Adsorption of the cations Cr(III), Cu, Pb, Cd, Hg, and Zn increased with in-
creasing pH; adsorption of the anions Cr(VI), As, and Se decreased with increasing pH.
Above pH = 5.3 precipitation of the cations was an important mechanism while adsorption
was the principal mechanism for the anions over the pH range studied. Because adsorp-
tion/mobility of any element was affected by other solutes in leachate, adsorption
information on one leachate may not be directly applied to predicting adsorption of the
same element at the same concentration in another leachate.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b. IDENTIFIERS/OPEN ENDED TERMS
COS AT I Field/Group
Attenuation
*Transport Properties
*Soil Chemistry
Contaminants
Arsenic
Cadmium
Copper
Industrial Wastes
Chromium
Iron
Lead (Metal)
Mercury (Metal)
Selenium
Zinc
Pollution Hazard Rating
Chemical Oxygen Demand
Groundwater Pollution
Clay Liners
Municipal Solid Waste
Leachate
Ammonium (ion)
13B
18. DISTRIBUTION STATEMENT
RELEASE TO PUBLIC
19. SECURITY CLASS (This Report)
UNCLASSIFIED
21. NO. OF PAGES
157
20. SECURITY CLASS (This page)
UNCLASSIFIED
22. PRICE
EPA Form 2220-1 (Rev. 4-77)
147 -fa U. S. GOVERNMENT PRINTING OFFICE: 1978-757-140/1445 Region No. 5-11
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