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OCR for page 3388
Proc. Natl. Acad. Scz. USA
Vol. 96, pp. 3388-3395, March 1999
Colloquium Paper
This paper was presented at the National Academy of Sciences colloquium "Geology, Mineralogy, arid Human Welfare, "
held November 8-9, 1998 at the Arnold and Mabel Beckman Center in Irvine, CA.
Mineral surfaces and bioavailability of heavy metals:
A molecular-scale perspective
GORDON E. BROWN, JR.*~l, ANDREA L. FOSTER*, AND JOHN D. OSTERGREN~
*Department of Geological and Environmental Sciences, Stanford University, Stanford, CA 94305-2115, and Stanford Synchrotron Radiation Laboratory,
Stanford Linear Accelerator Center, P.O. Box 4349, Stanford, CA 94309
ABSTRACT There is a continual influx of heavy metal
contaminants and pollutants into the biosphere from both
natural and anthropogenic sources. A complex variety of abiotic
and biotic processes affects their speciation and distribution,
including adsorption onto and Resorption from mineral surfaces,
incorporation in precipitates or coprecipitates, release through
the dissolution of minerals, and interactions with plants and
microbes. Some of these processes can effectively isolate heavy
metals from the biosphere, whereas others cause their release or
transformation to different species that may be more (or less)
bioavailable and/or toxic to organisms. Here we focus on abiotic
adsorption and precipitation or coprecipitation processes involv-
ing the common heavy metal contaminant lead and the metal-
loids arsenic and selenium in mine tailings and contaminated
soils. We have used extremely intense x-rays from synchrotron
sources and a structure-sensitive method known as x-ray ab-
sorption fine structure (XAFS) spectroscopy to determine the
molecular-level speciation of these elements at concentrations of
50 to several thousand ppm in the contaminated environmental
samples as well as in synthetic sorption samples. Our XAFS
studies of As and Pb in the mine tailings show that up to 50% of
these contaminants in the samples studied may be present as
adsorbed species on mineral surfaces, which makes them poten-
tially more bioavailable than when present in sparingly soluble
solid phases. Our XAE7S studies of Se(VI) sorption on Fe2+-
containing sulfates show that this element undergoes redox
reactions that transform it into less bioavailable and less toxic
species. This type of information on molecular-level speciation of
heavy metal and metalloid contaminants in various env~ronmen-
tal settings is needed to prioritize remediation efforts and to
assess their potential hazard to humans and other organisms.
The earth's surface and near surface regions are dominated by
interfaces among solids, liquids, and gases. Such interfaces,
particularly those between natural solids and aqueous solutions,
play an enormously important role in a number of geological and
geochemical processes. For example, "water-rock" interactions
over geologic time have been major contributors to both the rock
cycle and the geochemical cycling of elements. In an environ-
mental context, interracial processes such as mineral dissolution,
mineral precipitation, and the sorption and Resorption of chem-
ical species are responsible for the release and/or sequestration of
heavy metals that may eventually become pollutants in soils and
groundwater. The importance of interracial processes is summed
up well in the following quotation from Werner Stumm (1~:
"Almost all of the problems associated with understanding the
processes that control the composition of our environment con-
cern interfaces, above all the interfaces of water with naturally
occurring solids."
Recognition of the importance of surface chemical reactions in
geochemical and environmental contexts can be traced back to
PNAS is available online at www.puas.org.
some of the pioneering work on soil-fluid interactions, particu-
larly the papers by H. S. Thompson (2) and J. Thomas Way (3),
who were early soil chemists. The paper by Way reported on the
filtration of "liquid manure" through a loamy soil, which resulted
in the manure being "deprived of color and smell." Almost 150
years later, we have a more fundamental understanding of the
sorption and cation exchange phenomena that account for Way's
observation. We are also beginning to develop a molecular-scale
understanding of the complexity and interplay of the chemical
and biological processes that control element cycling in soils and
sediments, some of which are shown in Fig. 1. Such processes
range from dissolution of mineral particles in soils, which can
release natural contaminants into pore waters, to the binding or
sorption of metals (M) and organic ligands (L) to mineral
surfaces, which can effectively immobilize contaminants and
reduce their bioavailability. Precipitation is another common
means of sequestering a heavy metal if the precipitated phase is
relatively insoluble. Some heavy metal contaminants such as lead
normally exist in minerals in one dominant oxidation state,
whereas others such as arsenic and selenium can exist in several
oxidation states and can undergo oxidation or reduction when
they interact with mineral surfaces or organic compounds, which
act as oxidants or reductants. Microorganisms and plants can
have a profound influence on chemical reactions involving con-
taminants. For example, microorganisms often play a major role
in the degradation of organic contaminants and in the oxidation
and reduction of heavy metals. In the case of plants, the root-soil
interracial region, referred to as the rhizosphere (circled area in
soil profile in Fig. 1) is an area of particularly intense chemical and
biological activity where organic acids, sugars, and other organic
compounds are exuded by live plant roots. The pH can be as much
as 2 units lower, and microbial counts can be 10- to 50-fold higher
at the root surface than in the bulk soil a few millimeters away.
Thus, mineral weathering rates and the solubility of mineral
elements and anthropogenic contaminants are generally greater
in surface soils, where plant and microbial activity are higher than
in deeper parts of the soil and geologic column.
Because of this complexity, experimental and theoretical stud-
ies that probe the nature of these processes at a fundamental level
are difficult to perform on natural samples, and the results are
sometimes difficult to interpret because of the large number of
interacting abiotic and biotic processes. In fact, it is often difficult
to distinguish between abiotic and biotic transformation mecha-
nisms involving heavy metals and metalloids, as both can affect
the pathways of reactions in a synergistic fashion (see, e.g., ref. 4~.
One approach to this problem involves analytical and experimen-
tal studies of simplified model systems in which variables can be
controlled to simulate processes in more complex natural systems.
Abbreviations: XAFS, x-ray absorption fine structure; EXAFS, ex-
tended x-ray absorption fine structure; MANES, x-ray absorption near
edge structure.
lTo whom reprint requests should be addressed. e-mail: gordon@
pangea.stanford.edu.
3388
OCR for page 3389
Colloquium Paper: Brown et al.
.
Processes
in
Environmental
Science Microorganism
~ ~ t i
-
-
-
-
-
-
-
Soil Profile
O honzon
. A horizon
I B horizon
l ~
| C horizon
| Bedrock
Solution Complex
1 ~ | Desorption Adsorption i, - ~Orqan/c
I? End Dissolution Precipitation t
Redox
~ egos ~ ~ r
(Degradation/Release M nix
O H O
u ~n
HO - C - CH2- C ~ CH2- C ~ OH
~g.fO
O OH a~2-cH-C-OH
H2 2
~C- C - H Organic Ligand (L)
AL
Mfred
Lied
Organic Reductant
It is also essential to carry out parallel studies of the natural
systems to place constraints on the variables and types of pro-
cesses that control contaminant speciation and distribution and to
develop testable hypotheses that can be addressed by appropri-
ately chosen model systems. We have adopted this combined
model system/natural system approach to help interpret analytical
results on natural systems containing heavy metal and metalloid
contaminants.
Ultimately, the impact of heavy metals and other environmen-
tal contaminants on humans and other organisms depends on
their concentration levels, toxicity, and bioavailability, i.e., the
extent to which they are absorbed by the blood or stored in
internal organs. The toxicity and bioavailability of a heavy metal,
in turn, depend in part on reactivity and solubility, which are
determined by the speciation or chemical form of the element.
The term "speciation," as used here, refers to (i) the identity of
the element, (ii) its oxidation state, (iii) its physical state ti.e.,
phase association; presence in a liquid, gaseous, or solid phase
(amorphous or crystalline), colloidal particle, animal or plant cell,
or biofilm; presence as a surface coating or thin film on a solid,
as a sorption complex (monomeric or polymeric) on a solid,
colloidal particle, or an organic substance; etc.], (iv) its empirical
40
30
-
~ 20 ~
J
ID
_ . _
. ~ -
10
Broken Hill, Australia ( 1994)
~ -
.
-.-~ Minnesota (Urban, 1987)
Leadville, CO (1987)
Butte, MT ( 1990)
0 1000 2000 3000 coo 5000
Lead in Surface Soil (ppm)
FIG. 2. Blood lead levels in humans in several major lead mining
districts (modified after ref. 12~.
Proc. Natl. Acad. Sci. USA 96 (1999J 3389
1~ 3
~ 1 l
~ ~ l
_1 ~ ~Primary
_ Mineral
\Organic
Coating
Kaollnit+Polymer
Complexes
FIG. 1. Schematic illustration of a variety of
molecular environmental science processes af-
fecting contaminant elements in soils and
groundwater.
formula, and (v) its detailed molecular structure. For example,
chromium in the Cr(VI) form occurs as a tetrahedral oxoanion
in aqueous solution, is quite soluble and mobile in groundwater
and surface waters, and is toxic to organisms (5, 6~. In contrast,
when in the Cr(III) form, chromium is typically bonded to six
oxygens in an octahedral complex, is relatively insoluble and
immobile, and poses little risk to organisms (5 ). As shown by this
example, a fundamental understanding of the potential impact of
a particular environmental contaminant on humans must include
a molecular-level description of its speciation, as well as a number
of other interrelated factors (7~. This point is illustrated by the
blood lead levels measured in humans in several sites, where lead
levels in surface soils range from hundreds to thousands of ppm
(Fig. 2~. These data (8) show that significantly different levels of
blood lead are found in human populations at the different sites
and do not correlate directly with bulk Pb soil concentrations.
One possible explanation for these differences is the presence of
different species of lead at these sites, with lead at some sites being
in a more bioavailable form. When present in crystalline solids
such as galena or pyromorphite, lead should be less soluble, less
mobile, and less bioavailable than when present as sorbed species
on mineral surfaces, where it may be relatively easily removed into
solution by a reduction in pH from neutral to acidic, such as might
occur when a lead-contaminated soil sample is ingested by an
organism (see, e.g., ref. 9~. These differences are well illustrated
by the results of in viva swine studies using soil and mine waste
material from Leadville, CO, which have shown that Pb bioavail-
ability ranges from <5% for selected tailings materials, which are
dominated by galena (PbS), to 45% for surface soils where the
majority of Pb occurs as "Fe-Mn-Pb oxide" phases (10~.
Although the correlation between speciation of environmental
contaminants and their bioavailability has been shown in recent
studies (11-17), there is relatively limited use of information on
molecular-level speciation in setting maximum contaminant lev-
els (see, e.g., ref. 18~. For example, the current EPA limit for
arsenic in drinking water is 50 ppb total arsenic without desig-
nation of the species of As present, which can occur as inorganic
oxoanions tAs(III)O33- and As(V)O43-], as organic arsenicals,
or in other forms; the trivalent form of As, either in inorganic or
organic compounds, is generally more toxic than pentavalent
compounds of As (19~. In the past, cleanup efforts for contam-
inated soils or mine tailings were often driven by total contam-
inant element concentration without sufficient attention to spe-
ciation or the relative toxicities or bioavailabilities of individual
OCR for page 3390
3390 Colloquium Paper: Brown et al.
species of a contaminant element (18~. This practice is due in
large part to the difficulty in deriving molecular speciation
information for low concentration levels of a given contaminant
and in applying this information to field-scale situations. Potential
consequences of this approach are remediation efforts and major
expenditures that may not be warranted because of the presence
of a heavy metal or metalloid in chemical forms that are not
readily bioavailable.
In this paper, we summarize recent spectroscopic work on the
molecular-scale speciation of arsenic, selenium, and lead in
contaminated soils and mining wastes with the objective of
relating speciation to potential bioavailability. We also briefly
discuss some of the natural sources of these elements as well as
general concepts needed to understand variations in their uptake
or sorption on mineral surfaces in contact with bulk aqueous
solutions. These heavy metals (Pb) and metalloids (As, Se) are
among the most important environmental contaminants, affect-
ing millions of people. For example, arsenic pollution in drinking
water in Bangladesh is currently causing a health risk to over 70
million people (204. Arsenic also poses a potential risk in the
Mother Lode District of California, where tailings from gold
mining activities of the 1800s contain arsenic levels as high as
5,000 ppm (21, 22~. Similarly, selenium contamination in the
Central Valley of California threatens wildfowl populations as
well as agricultural production in this region where Se concen-
trations in topsoils and efflorescences are as high as 50 ppm (23~.
Lead is arguably the most widespread elemental contaminant
worldwide and has had a major impact on human welfare (24, 25~.
Lead has many diverse physiological and biochemical effects in
animals and humans, particularly children, where it has been
shown to affect intelligence quotients in children exposed to
lead-containing paints and soils (26, 27~.
In the contaminated soil and mine tailings samples considered
in this paper, As, Se, and Pb concentrations ranged from as low
as 50 ppm to as high as several weight percent. In the model
laboratory sorption samples considered here, the amount of
heavy metal present on particle surfaces, measured in percentage
of an ideal monolayer (ML), ranged from less than logo ML to
greater than 100% ML, or, in terms of concentration/unit area,
from as low as 0.1 ,umol/m2 to >20 ,umol/m2. At these very low
concentrations, sensitive element-specific probes capable of pro-
viding molecular-scale structure-composition information on
heavy metal speciation are essential. Synchrotron-based x-ray
absorption fine structure (XAFS) spectroscopy methods are well
suited for this type of work because of the extremely high intensity
of synchrotron x-rays and their energy tunability, which provide
element sensitivity and element specificity, respectively. XAFS
spectroscopy is capable of providing quantitative information on
the geometry, composition, and mode of attachment (inner- vs.
Outer-sphere surface complexes vs. three-dimensional precipi-
tates; monodentate vs. bidentate surface attachment, etc.) on
specific elements at surface coverages as low as ~0.05 ML by
using existing synchrotron radiation sources and x-ray detectors
(28-30).
The information derived from this work includes a molecular-
scale description of the dominant typos) of surface complexes or
precipitates that form when these metals or metalloids partition
from aqueous solution onto the mineral surface, a process
generally referred to as sorption when the details of the adsorp-
tion mechanism are not fully known. We also present similar
spectroscopic studies on the speciation of these heavy metals/
metalloids in several natural multiphase systems, including Se-
contaminated soils and sediments and As- and Pb-contaminated
mining wastes. We show that significant fractions (up to 50
atom%) of As and Pb in our samples of mine tailings from the
Mother Lode District of Central California and Leadville, CO,
respectively, occur as sorbed species that are potentially bioavail-
able. We also discuss spectroscopic results on Se in contaminated
soils that indicate redox transformations from more toxic to less
toxic species.
Proc. Natl. Acad. Sci. USA 96 (1999)
Natural Sources of Heavy Metal Contaminants and Pollut-
ants. One common public misconception is that inorganic con-
taminants or pollutants (contaminants at sufficiently high con-
centrations to pose a hazard to exposed organisms) are intro-
duced into soils and natural waters only through anthropogenic
emissions. Because of their presence in minerals and the fact that
minerals undergo dissolution caused by chemical weathering,
heavy metals and metalloids are continually released into the
environment at various rates and concentrations from natural
materials. When concentrated in certain types of mineral deposits
exposed at the earth's surface, they can be released at sufficient
concentrations to be considered pollutants (e.g., arsenic released
from the weathering of arsenical pyrite and arsenopyrite). When
such deposits are mined, the release is typically enhanced because
of the crushing of ore and gangue (waste) materials and the
resulting increase in surface area of material exposed to weath-
ering processes. In these cases, pollution can become a serious
problem, especially if the oxidation of sulfide minerals, which are
often present in heavy metal ore deposits, causes acid mine
drainage, which can enhance the mobility of heavy metal cations
(see ref. 31 for a case study). For example, the oxidation of pyrite
(FeS2) produces ferric hydroxide solids, sulfate, and hydrogen
ions by means of a reaction of the following type:
FeS2(s) + 15/4 02(g) + 7/2 H2O(l) > Fe(OH)3(s)
+ 2 S04(aq) + 4 H+(aq)
[1]
which results in decreased pH, enhanced dissolution of minerals,
and enhanced release of cations that are adsorbed on mineral
surfaces in the vicinity of the dissolving pyrite. Desorption of
surface-bound Pb(II) and other cations from mineral surfaces
under acidic solution conditions results in an increase in the
concentration level of potentially bioavailable forms of heavy
metal cations in surface waters and groundwater.
The following four sections are intended to provide a brief
overview of some of the minerals that contain arsenic, selenium,
and lead as major, minor, or trace components. Some of the
phases listed are potential candidates for secondary minerals in
various types of base-metal deposits or tailings (e.g., Pb-Zn-Cu
sulfide deposits); others are primary minerals. A relatively com-
prehensive list of primary and secondary minerals containing
these elements can be found on the Internet at the address:
http://un2sg4.unige.ch/athena/mineral/minppcl.html. Although
some of the minerals listed below are rare and may seem
unimportant, they could be present at relatively high concentra-
tions in certain contaminated or polluted environments contain-
ing one or more of these heavy metals or metalloids.
Natural Forms of Arsenic in the Environment. Arsenic is the
51st most abundant element in crustal rocks (average concen-
tration = 1.8 ppm) and can occur in the 5 +, 3 +, O, 1-, and 2-
oxidation states in different geological environments (32~. In
general, reduced inorganic As found in sulfide minerals is
relatively low in toxicity, but oxidized inorganic As(III) and
As(V) compounds are significantly more toxic than many
organoarsenicals. Moreover, As(III) compounds are two to
three times more toxic than As(V) compounds (20~. Both
As(III) and As(V) form anionic species when in aqueous
solution, adsorbed to mineral surfaces, or incorporated into
precipitates. As(V) adsorbs more strongly to mineral surfaces
than does As(III), thus is generally less mobile and potentially
less bioavailable (33, 34~.
Arsenic occurs naturally as a major component in arsenates
"containing As(V)04 units] (e.g., adelite tCaMgAsO4~0H)],
chalcophyllite [Cu~xAl2(Aso4~3(so4~3(oH)27 33H2O] ~ clinoclase
tCu3(As044~0H)3i, duftite (CuPb(As04~0H)], hoernesite
tMg3(As0442 8H2O], scorodite tFeAsO4 2H2O]~; arsenites tcon-
taining As(III)03 units] (e.g., armangite tMn26(As~ sOso)
(OH)4CO3], ecdemite tPb6As2O7Cl4], finnemanite tPbs
(As03~3Cli, paulmooreite tPb2As2Os]' trigonite [Pb3Mn
OCR for page 3391
Colloquium Paper: Brown et al.
(As03~2(As02~0H)], trippkeite iCuAs2O44~; elemental arsenic
tAs(O)], sulfides and arsenides (e.g., arsenopyrite tFeAsS], lautite
[CuAsSi, skutterudite t(Cu, Ni)As3], smaltite t(Co, Ni)As3_x]~;
sulfosalts (e.g., orpiment (As2S3], realgar tAsS], tennantite [(Cu.
Fess; or as a minor or trace component in minerals (e.g.,
in arsenical pyrite, where As can occur in arsenopyrite inclusions
or as part of a solid solution, in jarosite ~KFe3(SO4~2~0H)6],
where As(V) substitutes for S(VI), or in hydrotalcite-like
anionic structures, which have the general formula
tM(II)~_xM(III)x(OH)2iX+(An~ n) mH2O, where M(II) and
M(III) are divalent and trivalent cations, respectively, and An-
represents anions such as CO32-, SO42-, CrO42-, AsO33-, and
AsO43-, which could occupy inner-layer positions to neutralize
the positive layer charge (35~.
Natural Forms of Selenium in the Env~ronment. Selenium is
the 66th most abundant crustal element (average concentra-
tion = 0.05 ppm) and occurs in the 6+, 4+, O. and 2- oxidation
states in different geologic settings depending on the Eh and pH
values (32~. In reduced forms, selenium is relatively insoluble and
immobile, thus poses little danger to organisms. However, in
oxidized forms, particularly Se(VI), selenium is mobile in aque-
ous solutions and poses a significant risk to organisms. Selenium
is present as a major component in selenates [containing
Se(VI)04 units] (e.g., olsacherite tPb2(SeO4~(SO4~], schmied
erite tPb2Cu(II)2(Se(IV)03)Se(VI)04~0H)4~; selenites tcon-
taining Se(IV)03 units] (e.g., chalcomenite tCuSeO3 2H2O],
hannebachite [Ca2(SeO332 H2O], molybdomenite tPbSeO3i, sofi-
ite ~Zn2SeO3C12~; elemental selenium tSe(O)~; selenides (e.g.,
berzelianite tCu2Se], umangite tCu3Se2~; or as a minor to trace
component in others (e.g., substituting for S in sulfides such as
pyrite and in sulfates such as barite and jarosite).
Natural Forms of Lead in the Environment. Lead is the 36th
most abundant element in the earth's crust (average concentra-
tion = 13 ppm) and is generally present in the 2+ oxidation state
in inorganic compounds and rarely present as elemental lead
(32~; it is not sensitive to oxidation or reduction over the normal
range of Eh and pH values encountered in various geologic
settings at or near the earth's surface. Lead can occur as a major
element in a wide variety of minerals (e.g., anglesite tPbSO4i,
cerrusite [PbCO3], galena tPbSi, pyromorphite tPbs(PO4~3(Cl,
OH, F)~), and as a trace or minor element, especially by substi-
tution for Ca2+ and K+ (e.g., Pb-bearing apatites and jarosites).
Heavy Metal Uptake on Mineral Surfaces. Sorption on mineral
surfaces is an important process that can bind and sequester heavy
metals and other aqueous contaminant ions. Sorption can dra-
matically reduce the mobility of contaminants in groundwater
and, in the case of redox-sensitive elements, result in their
transformation into a less (or more) toxic species through reduc-
tion or oxidation reactions. The effectiveness of sorption reac-
tions in binding an ion is determined by a number of variables,
including (i) pH, (ii) the charge on the mineral surface as a
function of pH, (iii) the type of sorption complex formed, (iv)
competition between different ions for the same types of reactive
surface sites, (v) the presence of organic and/or inorganic ligands
that can inhibit or enhance sorption of a metal ion, and (vi) the
presence of surface coatings such as biofilms that may block
reactive sites and/or create new sorption sites. A discussion of
each of these variables is beyond the scope of this paper. In this
section we review the general modes of sorption of aqueous
cations and anions on mineral surfaces and discuss the specific
types of surface complexes formed by As, Se, and Pb, as revealed
by x-ray absorption spectroscopy.
Mineral surfaces in contact with aqueous solutions have a point
of zero charge (pHpzc)' which is the pH value (or small range of
pH values) at which the surface is electrically neutral. Points of
zero charge of silicate and oxide surfaces have been critically
evaluated in several studies [e.g., (36, 37~] and range from 2-3
(SiO2) to ~12 (MgO). Below the pHpZc value, the charge of a
mineral surface is positive, indicating an excess of protons bonded
to the surface. Above the pHpzc' the surface is negatively charged
Proc. Natl. Acad. Sci. USA 96 (1999) 3391
indicating an excess of OH- groups. At very high pH values, oxo
ions (o2-) may occur on the surface of an oxide or silicate in
contact with an aqueous solution. The pH of uptake or release of
a heavy metal depends to a significant extent on whether the
heavy metal occurs as a cation or anion in solution. For example,
Pb(II) in aqueous solutions exists as a cation over a wide pH
range, thus its affinity for a given mineral surface generally
increases with increasing pH. In contrast, As(III), As(V), Se(IV),
and Se(VI) behave as anions in aqueous solutions, thus are not
strongly sorbed at high pH values where the mineral surface is
negatively charged.
The type of surface complex formed has an important effect on
the mobility of a metal ion. Some metal ions bond directly to the
mineral surface, losing waters of hydration and forming an
inner-sphere complex (see Fig. 1~. Such ions are relatively difficult
to desorb except for large pH changes, thus are relatively immo-
bile. However, if the metal ion forms a weakly bound outer-
sphere complex in which the ion is surrounded by waters of
hydration and no direct chemical bonds to the surface are formed,
the metal can be easily desorbed when pH changes. An indication
of the strength of binding of an aqueous cation or anion to a
mineral surface can be obtained from the macroscopic uptake
behavior as a function of ionic strength. For example, when an
increase in the ionic strength of the background electrolyte (e.g.,
NaNO3) reduces the degree of uptake of an ion, the ion is less
strongly bound to a mineral surface than when adsorbate uptake
shows no ionic strength dependence. In the former case, the ion
may form a weakly bound outer-sphere complex, whereas in the
latter case, a strongly bound inner-sphere complex may be
indicated. However, direct spectroscopic verification of the mode
of sorption is required to verify conclusions drawn from macro-
scopic measurements alone.
Among the ions examined here, Pb(II) forms strongly bound
inner-sphere complexes on many mineral and metal oxide sur-
faces (38-40), as do the oxoanions AsO43- (41-45), AsO33- (46),
and SeO32- (474. In contrast, the oxoanion SeO42- (47) forms
more weakly bound complexes that are thought to be predomi-
nantly of the outer-sphere type, although there is still some
controversy about the dominance of inner-sphere vs. outer-
sphere complexes in the case of selenate sorption on iron
(oxy~hydroxides (48~. The mode of sorption of these species was
determined by using XAFS spectroscopy.
The strength of binding and stability of an inner-sphere com-
plex will depend in part on how many bonds it forms with surface
functional groups. For example, bidentate or tridentate surface
complexes, attached by two and three bonds, respectively, will
typically be more difficult to desorb than monodentate complexes
(attached by a single bond), all other factors being equal. Another
important factor in the effectiveness of sorption reactions in
removing contaminant metal ions from solution is the possibility
of multinuclear complex formation (49~. In this case, more than
one metal ion is involved in the surface complex. The progression
from mononuclear to multinuclear surface complexes generally
occurs with increasing sorbate metal concentration (29, 50), with
precipitation occurring at a point where the solution becomes
supersaturated with respect to the hydroxide of the metal-ion
sorbate (504. When other ions are present, such as the anions
carbonate, sulfate, or phosphate, additional complications can
arise, including competition for surface sites, formation of solu-
tion complexes, or precipitation of solids made up of the sorbate
ion and the anion.
In addition to the types of adsorption described above, another
potentially important sorption mechanism involves incorporation
of certain types of metal ions into a solid coprecipitate made up
of the sorbate ion and metal ions derived from dissolution of the
mineral surface. Studies by several groups over the past few years
have shown that this type of sorption product can form in aqueous
systems containing sorbate ions like Co(II) (refs. 51-53) and
Ni(II) (54, 55) and sorbents like or-Al2O3, kaolinite
tAl2Si2Os(OH)4], and pyrophyllite tAl2Si4O~o(OH)2~. The result
OCR for page 3392
OCR for page 3394
OCR for page 3395
Representative terms from entire chapter:
mineral surfaces
3392 Colloquium Paper: Brown et al.
Ruth Mine (Trona~ CA) Argonaut Mire (Jackson CA)
· no sulfides
·Total As = 2000 ppm
·As(V) only
_ ~
_ ·~ 2% sulfides
_ _ ·Total As = 200 ppm
~_ ·As(V) and reduced As
_ -
As-A1 3.16 A ~
. As-Fe = 3.25 A
~ ~3
ing solid phase can have a hydrotalcite-type structure (35) and
may be relatively insoluble, and therefore decreases the metal ion
bioavailability. A different mechanism of heavy metal sequestra-
tion by means of a coprecipitation mechanism is produced by
amending Pb-contaminated soils with phosphate, resulting in the
growth of the Pb-phosphate pyromorphite (56~.
EXPERIMENTAL APPROACH
Details of the procedures we used in preparing model heavy
metal sorption samples for XAFS spectroscopy studies can be
found in Bargar et al. (38, 39~. All sorbents were synthetic
high-surface area powders (~50 - 100 m2/gJ and are available
commercially or can be synthesized. The identity and phase purity
of the powdered sorbents were determined by powder x-ray
diffraction, and in selected cases by transmission electron micros-
copy coupled with energy dispersive x-ray analysis. Surface area
Apache Tailings:
· Sulfide-rich
· acidic pHf Id. ~=i i. ;\ `,
· total Pb = 8330ppm ~a.
~,.
~- ~ s [~l,,I,{ ~i vi}N,[~.
~, ~ ; ! be, l't,l' `,~,~,4;~) )~/,,l:
_. it's! I~llrI7111~-rit~
,,, N No,,,! i, W I~l\~it r Hi,
_
;~
X
!5~
e_"-
Pb LIII EXAFS model spectra:
1
(~.5 ~
(~.4 '
_ 1
0 0.3
0(~.2 '
~ '
().1 '
van`edi ni (c plumboternte pl u nib<,j:l rat .c ite
Pro c. Natl. Acad. Sci. USA 96 (1999)
FIG. 3. AS K-edge EXAFS data
and fluorescent x-ray images for AS-
contaminated mine tailings from the
Mother Lode District of California
(derived from data reported in ref. 59~.
Fits to the EXAFS data are shown in
red and molecular models of the
AS(V) environments in the model
compounds used in the fits are also
shown.
was determined for powdered sorbents by N2-Brunauer-Emmet-
Teller (BET) measurements. Powdered sorbents were placed in
0.01M NaNO3 solutions and pH was allowed to equilibrate. In
separate experiments, the heavy metal was added to each solution
in contact with the sorbent at initial aqueous concentrations
ranging from ~0.1 to 0.01 M. pH was adjusted by addition of
NaOH to achieve a desired percentage uptake of the heavy metal
onto each sorbent, which was verified by graphite furnace atomic
absorption (GFAA) spectrophotometry analysis of the superna-
tant. The resulting sorption densities ranged from ~0.1 to >20
,umol/m2. Each sample was allowed to "equilibrate" at the final
pH for 24-36 hr. then before GFAA analysis the sample was
centrifuged and about 95% of the supernatant was removed.
Except at the highest metal concentrations, the solutions were
undersaturated with respect to the stable precipitate; thus, for
example, no Pb(OH)2 or PbO precipitate phase was expected to
Hamms Tailings:
· Carbonate-buffered
· near-neutral pH
· total Pb = 8520ppm
= -~-At, \~r~.s~s~1~
,~ my--- f\-~,~f hit, .'r~,<~r~.~ ;~.
pi':; a., >-
Colloquium Paper: Brown et al.
form in the case of the model lead sorption systems. In each case,
the sorption sample in the form of a wet paste was loaded into a
Teflon sample holder sealed with Mylar windows, and XAFS
spectra were collected at the Stanford Synchrotron Radiation
Laboratory (SSRL) on wiggler-magnet beamlines within 2 days
of sample preparation. XAFS spectra for the model sorption
samples were collected in the fluorescence-yield mode by using
either a Stem-Heald type gas-filled ionization chamber detector
(57) or a 13-element Ge detector under ambient conditions. The
XAFS spectra of crystalline model compounds were collected in
transmission mode under ambient conditions. In none of these
experiments was evidence found for oxidation or reduction of the
sorbate ion during or after XAFS data collection, as indicated by
the lack of energy shifts in the edge positions, extended x-ray
absorption fine structure (EXAFS)-derived metal-oxygen bond
lengths, and x-ray photoelectron spectroscopy measurements.
XAFS data analysis procedures used for the powdered sorption
samples are described in Bargar et al. (38, 39) and O'Day et al.
(58~.
XAFS spectroscopic analysis of the natural soil and mine waste
samples containing As, Se, and Pb was carried out by using the
experimental and data analysis procedures described in Foster et
al. (59, 60), Pickering et al. (61), and Ostergren et al. (62), for As,
Se, and Pb, respectively. These natural samples were not signif-
icantly modified from their conditions in the field, except for size
separation in selected cases, in an attempt to preserve the same
speciation of heavy metals present in the original samples. About
100 mg of each sample was placed in a Teflon holder and covered
with thin Mylar tape, and XAFS data were collected by using
fluorescence-yield detection (either a Stem-Heald detector or a
multielement solid state detector) at ambient temperature (298
K) and pressure [1 aim (101.3 khan. In addition to the normal
EXAFS spectral fitting, which yields information on the identity
and arrangement of first and second neighbors around the central
absorber, linear least-squares fitting of the x-ray absorption near
edge structure (XANES) spectra was conducted for the As- and
Se-contaminated samples, and linear least-squares fitting of the
XAFS spectra was conducted for the Pb-contaminated samples.
By using this approach (59, 61, 62), the quantitative speciation of
the heavy metal was determined by fitting the spectrum of the
contaminated sample with spectra of model compounds, includ-
ing both crystalline phases and model sorption samples with the
heavy element sorbed at low surface coverage on different
mineral or organic substrates. This approach works well when the
spectral signatures of the different models are significantly dif-
ferent, which is often the case. Examples of this type of fitting are
given in the sections below. The amount of surface-bound heavy
metals in the contaminated soil and mine waste samples was
determined in this manner.
In addition to XAFS analysis of the contaminated samples,
each sample was examined by powder x-ray diffraction, optical
microscopy, and electron microprobe to determine the types and
amounts of crystalline phases present (59, 62, 63~. In addition,
selected samples were examined by surface-sensitive x-ray pho-
toelectron spectroscopy (62~.
RESULTS AND DISCUSSION
As in Mine Tailings from the Mother Lode District, California.
Almost 150 years of gold mining in the Mother Lode of Califomia
has resulted in significant concentrations of arsenic in mine
tailings (up to 5,000 ppm), some of which have been used for
housing developments in the Sierra Nevada Foothills of Central
Califomia. Because of the toxic effects of high concentrations of
arsenic on humans and other organisms, there is concern about
these tailings, and a number of studies are underway to determine
the potential health hazard of these tailings to humans.
We have used XAFS spectroscopy to determine the oxidation
state, local coordination environment (to a radius of ~7 A around
As), and the relative proportion of different As species in model
Proc. Natl. Acad. Sci. USA 96 (1999) 3393
compounds and three Califomia mine wastes: a fully oxidized
tailings (Ruth Mine), a partially oxidized tailings (Argonaut
Mine), and a roasted sulfide ore (Spenceville Mine) (59~. Analysis
of the XANES spectra of these contaminated samples indicates
that As(V) is the predominant oxidation state in the Ruth and
Spenceville mine samples, but mixed oxidation states were ob-
served in the Argonaut mine-waste. We obtained qualitative
information about As(V) chemical speciation by fitting the
XANES spectra of mine samples with a linear combination of
component (model compound) spectra (Fig. 3~. These analyses
suggest the presence of As(V) species similar to those found in
scorodite (FeAsO4 2H2O) and As(V) adsorbed on goethite (~-
FeOOH) and gibbsite [y-Al(OH)3~. Nonlinear least-squares fits
of mine waste EXAFS spectra indicate variable As speciation in
each of the three mine wastes. We conclude that ferric oxyhy-
droxides and aluminosilicates (probably clay) bind roughly equal
portions of As(V) in the Ruth Mine sample. Our analysis suggests
that tailings from the Argonaut Mine contain ~20% reduced As
bound in arsenopyrite (FeAsS) and arsenical pyrite (Fes2-xAsx)
and ~80% As(V) in a ferric arsenate precipitate such as
scorodite. Roasted sulfide ore of the Spenceville Mine contains
As(V) substituted for sulfate in the crystal structure of jarosite
[KFe3(SO4~2~0H)6], and sorbed to hematite surfaces. Determi-
nation of solid-phase As speciation by means of EXAFS spec-
troscopy is a valuable first step in the evaluation of As bioavail-
ability, because the mobility and toxicity of As compounds vary
with As oxidation state. As bound in crystalline or x-ray amor-
phous precipitates is generally considered to be less available for
uptake by organisms than when sorbed to mineral surfaces.
Se-Contaminated Soils in the Central Valley of California.
Selenium occurs naturally in sediments and soils in many parts of
the Western U.S. and is assumed to be incorporated in pyrites in
marine sedimentary rocks such as shales. When such soils are
irrigated for agricultural purposes, this indigenous element be-
comes soluble and is transported in agricultural drainage waters
to ponds and reservoirs where it becomes concentrated in water-
borne plants and animals (up to 3,000 ppm). One result of this
concentration process was discovered in the early 1980s by
government scientists at the Kesterson National Wildlife Refuge
in Merced County, California. Wildlife, particularly waterfowl,
died or were born deformed from consumption of high levels of
selenium (64~. Similar problems have since been documented at
nine sites in eight Western states comprising some 1.5 million
acres of farmland. This problem has major financial and health
implications in the San Joaquin Valley of California, a vast area
that produces a significant portion of the nation's vegetables,
fruits, and other crops. If farmers are prevented from draining
irrigation waters in this region, the rapid buildup of salts in the soil
will quickly make production of these crops impossible. About
500,000 acres of farmland in the San Joaquin Valley one
quarter of the Valley's agricultural acreage are at stake, with an
annual crop production worth about $500 million (65~.
To provide molecular-level information on the chemical forms
of selenium present in Se-contaminated soils from the Kesterson
Reservoir area, we have carried out XAFS spectroscopy studies
that showed that selenate and selenite are present in the top few
cm of soil adjacent to the drainage ponds but are reduced to
elemental selenium at lower soil levels (61, 66~. In carefully
controlled laboratory studies of soil columns to which selenate-
containing solutions were added, Tokunaga et al. (67, 68) found
that selenate is rapidly converted into elemental selenium. The
reduction can occur by means of both biotic (4, 69) and abiotic
(70) pathways. However, when reoxidized to the selenate form
during irrigation, selenium becomes highly mobile and is trans-
ported as an aqueous complex in drainage waters. Various
solutions to this problem have been proposed, including bacterial
reduction, immobilization, and removal of selenium from drain-
age waters or the use of drainage waters to irrigate land on which
salt-resistant plants such as cotton, Eucalyptus trees, and atriplex
are grown. None have been adopted and some skeptics doubt that
3394 Colloquium Paper: Brown et al.
a viable solution, which satisfies environmental, financial, and
political constraints, will be found. An eventual solution to this
problem will require a detailed knowledge of the redox chemistry
of selenium, its speciation in soils and groundwaters, and the
effect of microbial organisms and inorganic reductants on its
speciation.
Pb in Mine Tailings from Leadville, CO. Pb is a ubiquitous
environmental contaminant in soils because of the intensive
use of Pb in batteries, paints, alloys, and solder, ammunition,
gasoline additives, and other commercial products and the
production of lead by mining and smelting activities. A recent
study of the history of atmospheric lead deposition over the
past 12,370 years, as measured in a peat bog in the Jura
Mountains of Switzerland (71), has shown that the greatest
lead flux (15.7 mg/m2/yr) occurred in 1979. This level is 1,570
times the natural background value of 0.01 mg/m2/yr. However,
since the elimination of tetraethyl lead as a gasoline additive,
beginning in the 1970's in the U.S. and some other countries,
lead contamination levels have dropped significantly (25~.
They still exceed natural background levels by orders of
magnitude in soils in many nonurban localities where soils have
become polluted as a result of mining and smelting activities
and in urban localities where paints and other anthropogenic
sources of lead contaminate soils. The bioavailability of Pb is
known to vary widely among different Pb species, and this fact
is often cited to explain apparently dramatic variations in Pb
bioavailability from site to site (e.g., ref. 104. Understanding
the detailed relationship between speciation and bioavailabil-
ity necessarily begins with a complete, accurate, and direct
identification of Pb species in environmental media, such as
soils and mine waste. Working toward this goal, researchers
have recently begun applying synchrotron radiation-based
x-ray absorption spectroscopic (XAS) techniques to determine
the molecular-scale details of Pb speciation at contaminated
sites (62, 63, 72, 73~. Here we summarize the results of our
work on Pb-bearing mine tailings from Leadville, CO (62~.
Using the unique advantages of XAS techniques, we empha-
size the identification and characterization of poorly crystal-
line and/or fine-grained species, such as sorption complexes
and poorly crystalline coprecipitates, which are likely to con-
trol Pb bioavailability and mobility in natural systems.
Bulk Pb concentrations range from 6,000 to 10,000 ppm in the
two tailing piles we sampled at the Leadville site. These concen-
trations necessarily raise human health and environmental con-
cerns, but bioavailability and chemical lability of Pb in these
materials vary dramatically and show little correlation with bulk
concentrations (10~. Because these samples are heterogeneous
multiphase mixtures (Fig. 4), a variety of complementary ana-
lytical methods were used, including powder x-ray diffraction,
scanning electron microscopy, electron probe microanalysis, x-ray
photoelectron spectroscopy, and synchrotron radiation-based x-
ray absorption. By using this suite of techniques, and XAS
techniques in particular, in conjunction with physical and chem-
ical separation techniques, we were able to identify and charac-
terize a number of species not amenable to detection by conven-
tional microanalytical techniques. In particular, we found direct
spectroscopic evidence for Pb adsorbed to mineral surfaces and
variations in this surface-bound component with pH as would be
predicted on the basis of simplified model system studies of
adsorption processes. Least-squares fitting of EXAFS spectra
shows that 50% of the total Pb in selected samples of the
carbonate-buffered tailings with near-neutral pH occurs as ad-
sorption complexes on iron (hydr~oxides, whereas Pb speciation
in sulfide-rich low pH samples is dominated by Pb-bearing
jarosites; we find no evidence for adsorbed Pb in these latter
samples. Importantly, the dominant Pb species in each of the
tailing piles could not be definitively identified without the
molecular-scale information provided by EXAFS analysis. Be-
cause these species likely control Pb transport and bioavailability
in these environments, our results clearly illustrate the need for
Proc. Natl. Acad. Sci. USA 96 (1999J
molecular-scale characterization as basis for understanding the
behavior of and health risks posed by Pb in natural environments.
SUMMARY AND CONCLUSIONS
The heavy metal Pb and the metalloids As and Se are among the
most common environmental contaminants resulting from an-
thropogenic activities and the weathering of natural mineral
deposits. These elements occur in a variety of chemical forms or
species that can vary widely in solubility, mobility, toxicity, and
bioavailability, depending on their speciation. In contaminated
soils and mine tailings, for example, they can occur in primary
minerals, secondary minerals formed by weathering of primary
m~nerals in situ, solid precipitates formed by reactions of con-
taminant ions in groundwater with other aqueous ions, and
adsorbed species. Although it is relatively easy to determine the
types of solid phases present in a contaminated soil or mine
tailings sample and the concentration levels of heavy metals and
metalloids they contain by using a combination of x-ray diffrac-
tion and analytical methods, it is considerably more difficult to
assess the importance of adsorbed heavy metal/metalloid species,
particularly at very low surface coverages. Adsorbed species may
comprise a significant fraction of the heavy metal or metalloid
present, and they are often the most bioavailable fraction.
By using a combination of synchrotron-based XAFS spectros-
copy and other analytical methods, the molecular-level speciation
of As, Se, and Pb, including the types of sorbed species, was
determined for selected mine tailings and contaminated soils.
Sorbed As and Pb species were found to be major components
with potentially high bioavailability in mine tailings from the
Mother Lode District of California and Leadville, CO, respec-
tively. Similar studies have shown that the most toxic and poten-
tially bioavailable forms of selenium, Se(VI), are transformed
rapidly to environmentally benign forms in contaminated soils
through a combination of biotic and abiotic processes.
A detailed knowledge of molecular-level speciation of heavy
metal/metalloid contaminants and the bioavailability of the dif-
ferent species of a contaminant element is necessary for setting
max~mum contaminant limits. Knowledge of speciation is also
required for efficient and cost-effective remediation efforts. This
point is well illustrated by cleanup efforts at the former uranium
processing plant at Fernald, OH, which serves as the host for the
Uranium in Soils Integrated Demonstration. This demonstration
project used carbonate soil-washing procedures to remove
hexavalent uranium. However, this conventional remediation
procedure was not totally effective because of the presence of
secondary phases containing U(IV) and U(VI), the latter be~ng
in the form of insoluble phosphates. These uranium species were
detected by using a combination of XAFS, optical luminescence,
Raman spectroscopy, scanning electron microscopy, and powder
x-ray diffraction (74) and micro-XAlFS spectroscopy (75~.
Changes in remediation procedures based on this type of specia-
tion information could result in significant cost savings and more
efficient cleanup of environmental contaminants, both of which
should be major societal goals.
The studies by our group presented in this paper were supported by
the Department of Energy, Basic Energy Sciences, and the National
Science Foundation. All of the synchrotron-based spectroscopic work
was carried out at the Stanford Synchrotron Radiation Laboratory
(SSRL), which is supported by the Department of Energy (Basic
Energy Sciences and Office of Biology and Environmental Research)
and the National Institutes of Health. We are grateful to the staff of
SSRL for their technical support of this work. We also acknowledge
the collaborations of T. Tokunaga and S. Myneni (Lawrence Berkeley
National Laboratory) and I. Pickering (SSRL) in the studies of
Se-contaminated soils; G. Morin, F. Juillot, and G. Calas (University
of Paris 7) in the studies of Pb-contaminated soils from northeastern
France, and G. A. Parks and the late T. N. Tingle (Stanford University)
in the studies of As- and Pb-contaminated mine tailings.
Colloquium Paper: Brown et al.
2.
3.
4.
31.
Stumm, W., Werhli, B. & Wieland, E. (1987) Croat. Chem. Acta 60,
429-456.
Thompson, H. S. (1850) R. Agric. Soc. Engl. J. 11, 68-74.
Way, J. T. (1850) R. Agric. Soc. Engl. J. 11, 313-379.
Oremland, R. S., Hollibaugh, J. T., Maest, A. S., Presser, T. S.,
Miller, L. G. & Culbertson, C. W. (1989)Appl. Environ. Microbiol.
55, 2333-2343.
; Gauglhofer, J. & Bianchi, V. (1991) in Metals and Their Compounds
in the Environment, Merian, E., ed. (VCH, Weinheim, Germany),
pp. 853-878.
6. Babich, H. & Stotzky, G. (1983) in Aquatic Toxicology, Nriagu, J. O.,
ed. (Wiley, New York) pp. 1-46.
7. Chaney, R. L., Ryan, J. A. & Brown, S. L. (1997) Proc. Workshop
on Environmentally Acceptable Endpoints: Chlorinated Organics,
Energetics, and Heavy Metals, (American Academy of Environmen-
tal Engineers).
i. Gulson, B. L., Mizon, K. J., Law, A. J., Korsch, M. J., Davis, J. J.
& Howarth, D. (1994) Econ. Geol. 89, 889-908.
9. Ruby, M. V., Davis, A., Schoof, R., Eberle, S. & Sellstone, C. M.
(1996) Environ. Sci. Technol. 30, 422-430.
10. Casteel, S. W., Brown, L. D., Dunsmore, M. E., Weis, C. P.,
Henningsen, G. M., Hoffman, E., Brattin, W. J. & Hammon, T. L.
(1997) Draft Report Bioavailability of Lead in Soil and Mine Waste
from the California Gulch NPL Site, Leadville, Colorado, (United
States Environmental Protection Agency).
11. Davis, A., Drexler, J. W., Ruby, M. V. & Nicholson, A. (1993)
Environ. Sci. Technol. 27, 1415-1425.
12. Davis, A., Ruby, M. V. & Bergstrom, P. D. (1992) Environ. Sci.
Technol. 26, 461-468.
13. Davis, A., Ruby, M. V., Goad, P., Eberle, S. & Chryssoulis, S. (1997)
Environ. Sci. Technol. 31, 37-44.
14. Dieter, M. P., Matthews, H. B., Jeffcoat, R. A. & Moseman, R. F.
(1993) J. Toxicol. Environ. Health 39, 79-93.
15. Freeman, G. B., Johnson, J. D., Liao, S. C., Feder, P. I., Davis,
A. O., Ruby, M. V., Schoof, R. A., Chaney, R. L. & Bergstrom, P. D.
(1994) Toxicology 91, 151-163.
16. Gasser, U. G., Walker, W. J., Dahlgren, R. A., Borch, R. S. &
Burau, R. G. (1996) Environ. Sci. Technol. 30, 761-769.
17. Ng, J. C., Kratzmann, S. M., Qi, L., Crawley, H., Chiswell, B. &
Moore, M. R. (1998) Analyst 123, 889-892.
18. Davis, A., Ruby, M. V., Bloom, M., Schoof, R., Freeman, G. &
Bergstrom, P. D. (1996) Environ. Sci. Technol. 30, 392-399.
19. Yamauchi, H. & Fowler, B. A. (1994) inArsenic in the Environment,
Nriagu, J. O., ed. (Wiley Interscience, New York), pp. 35-53.
20. ScienceScope (1998) Science 281, 1261.
21. Greenwald, J. (1995) in Time Magazine, p. 1.
22. Vogel, N. (1995) in San Jose Mercury News, p. 1.
23. Presser, T. S. (1994) in Selenium in the Environment, Frankenberger,
W. T., Jr. & Benson, S. L., eds. (Dekker, New York), pp. 139-155.
24. Nriagu, J. O. (1998) Science 281, 1622-1623.
25. Nriagu, J. O., ed. (1978) The Biogeochemistry of Lead in the
Environment (Elsevier, New York).
26. Ewers, U. & Schilpkoter, H.-W. (1991) in Metals and Their Com-
pounds in the Environment, Merian, E., ed. (VCH, Weinheim,
Germany), pp. 971-1014.
27. Agency for Toxic Substances and Disease Registry (1988) The
Nature and Extent of Lead Poisoning in Children in the United States:
A Report to Congress. DHHS Doc. No. 99-2966 (U. S. Department
of Health Human Services, Public Health Service, Atlanta, GA).
28. Brown, G. E., Jr. (1990) in Mineral-Water Inte~ace Geochemistry,
Reviews in Mineralogy, Vol.23, Hochella, M. F., Jr. & White, A. F., eds.
(Mineralogical Society of America, Washington, DC), pp. 309-363.
_~. Brown, G. E., Jr., Parks, G. A. & O'Day, P. A. (1995) in Mineral
Surfaces, Vaughan, D. J. & Pattrick, R. A. D., eds. (Chapman &
Hall, London), pp. 129-183.
30. Brown, G. E., Jr., Parks, G. A., Bargar, J. R. & Towle, S. N. (1998)
in Proceedings of the American Chemical Society Symposium on
Kinetics and Mechanisms of Reactions at the Mineral/WaterInterface,
Sparks, D. L. & Grundl, T., eds. (Am. Chem. Soc., Washington,
DC), in press.
Moore, J. N. & Luoma, S. N. (1990) Environ. Sci. Technol. 24,
1278-1285.
32. Greenwood, N. N. & Earnshaw, A. (1984) Chemistry of the Elements
(Pergamon, Oxford).
33. Frost, R. R. & Griffin, R. A. (1977) Soil Sci. Soc. Am. J. 41, 53-57.
34. Korte, N. C. & Fernando, Q. (1991) Crit. Rev. Environ. Control 21,
1-39.
Cavani, F., Trifiro, F. & Vaccari, A. (1991) Catal. Today 11, 173-301.
Parks, G. A. (1965) Chem. Rev. (Washington, DC) 65,177-198.
Proc. Natl. Acad. Sci. USA 96 (1999' 3395
37. Sverjensky, D. A. (1994) Geochim. Cosmochim. Acta 58, 3123-3129.
38. Bargar, J. R., Brown, G. E., Jr. & Parks, G. A. (1997) Geochim.
Cosmochim. Acta 61, 2617-2637.
39. Bargar, J. R., Brown, G. E., Jr. & Parks, G. A. (1997) Geochim.
Cosmochim. Acta 61, 2639-2652.
40. Strawn, D. G., Scheidegger, A. M. & Sparks, D. L. (1998) Environ.
Sci. Technol. 32, 2596-2601.
41. Waychunas, G. A., Rea, B. A., Fuller, C. C. & Davis, J. A. (1993)
Geochim. Cosmochim. Acta 57, 2251-2269.
42. Manceau, A. (1995) Geochim. Cosmochim. Acta 59, 3647-3653.
43. Waychunas, G. A., Davis, J. A. & Fuller, C. C. (1995) Geochim.
Cosmochim. Acta 59, 3655-3661.
44. Fendorf, S. E., Eick, M. J., Grossl, P. & Sparks, D. L. (1997)
Environ. Sci. Technol. 31, 315-328.
45. Foster, A. L., Brown, G. E., Jr., Tingle, T. N. & Parks, G. A. (1998)
Am. Mineral. 83, 553-568.
46. Manning, B. A., Fendorf, S. E. & Goldberg, S. (1998) Environ. Sci.
Technol. 32, 2383-2388.
47. Hayes, K. F., Roe, A. L., Brown, G. E., Jr., Hodgson, K. O., Leckie,
J. O. & Parks, G. A. (1987) Science 238, 783-786.
48. Manceau, A. & Charlet, L. (1994) J. Colloid Interface Sci. 168, 87-96.
49 Chisholm-Brause, C. J., O'Day, P. A., Brown, G. E., Jr. & Parks,
G. A. (1990) Nature (London) 348, 528-531.
50. Chisholm-Brause, C. J., Brown, G. E., Jr. & Parks, G. A. (1991) in
XAFS VI, Sixth International Conference on X-Ray Absorption Fine
Structure, Hasnain, S. S., ed. (Ellis Horwood, New York), pp.263-265.
51. d'Espinose de la Caillerie, J.-B., Kermarec, M. & Clause, O. (1995)
J. Am. Chem. Soc. 117, 11471-11481.
Towle, S. N., Bargar, J. R., Brown, G. E., Jr. & Parks, G. A. (1997)
J. Colloid Interface Sci. 187, 62-82.
53. Thompson, H. A., Parks, G. A. & Brown, G. E., Jr. (1999) Geochim.
Cosmochim. Acta, in press.
54. Scheidegger, A. M., Lamble, G. M. & Sparks, D. L. (1997) J. Colloid
Interface Sci. 186, 118-128.
55. Scheidegger, A. M., Strawn, D. G., Lamble, G. M. & Sparks, D. L.
(1998) Geochim. Cosmochim. Acta 62, 2233-2245.
56. Traina, S. J. & Laperche, V. (1999) Proc. Natl. Acad. Sci. USA 96
3365-3371.
57. Lytle, F. W., Greegor, R. B., Sandstrom, D. R., Marques, E. C.,
Wong, J., Spiro, C. L., Huffman, G. P. & Huggins, F. E. (1984) Nucl.
Instrum. Methods 226, 542-548.
58. O'Day, P. A., Chisholm-Brause, C. J., Towle, S. N., Parks, G. A. &
Brown, G. E., Jr. (1996) Geochim. Cosmochim. Acta 60, 2515-2532.
59. Foster, A. L., Brown, G. E., Jr., Tingle, T. N. & Parks, G. A. (1998)
Am. Mineral. 83, 553-568.
60. Foster, A. L., Brown, G. E., Jr. & Parks, G. A. (1998) Environ. Sci.
Technol. 32, 1444-1452.
61. Pickering, I. J., Brown, G. E., Jr. & Tokunaga, T. (1995) Environ.
Sci. Technol. 29, 2456-2459.
62. Ostergren, J. D., Brown, G. E., Jr., Parks, G. A. & Tingle, T. N.
(1999) Environ. Sci. Technol., in press.
63. Morin, G., Juillot, F., Ostergren, J. D., Ildefonse, P., Calas, G. &
Brown, G. E., Jr. (1999) Am. Mineral. 84, 420-434.
64. Ohlendorf, H. M. & Santolo, G. M. (1994) in Selenium in the
Environment, Frankenberger, W. T., Jr. & Benson, S. L., eds.
(Dekker, New York), pp. 69-117.
65. Benson, M. (1993) San Jose Mercury News, July 4, p. 1A.
66. Tokunaga, T. K., Brown, G. E., Jr., Pickering, I. J., Sutton, S. R. &
Bajt, S. (1997) Environ. Sci. Technol. 31, 1419-1425.
67. Tokunaga, T. K., Pickering, I. J. & Brown, G. E., Jr. (1996) Soil Sci.
Soc. Am. J. 60, 781-790.
68. Tokunaga, T. K., Sutton, S. R., Bajt, S., Nuessle, P. & Shea-
McCarthy, G. (1998) Environ. Sci. Technol. 32, 1092-1098.
69. Losi, M. E. & Frankenberger, W. T., Jr. (1998)J. Environ. Qual. 27,
836-843.
70. Myneni, S. C. B., Tokunaga, T. K. & Brown, G. E., Jr. (1997) Science
278, 1106-1109.
71. Shotyk, W., Weiss, D., Appleby, P. G., Cheburkin, A. K., Frei, R.,
Gloor, M., Kramers, J. D., Reese, S. & Van Der Knapp, W. O.
(1998) Science 281, 1635-1640.
72. Cotter-Howells, J. D., Champness, P. E., Charnock, J. M. &
Pattrick, R. A. D. (1994) Eur. J. Soil Sci. 45, 393-402.
73. Manceau, A., Boisset, M.-C., Sarret, G., Hazemann, J.-L., Mench, M.,
Cambier, P. & Prost, R. (1996) Environ. Sci. Technol. 30, 1540-1552.
74. Morris, D. E., Allen, P. G., Berg, J. M., Chisholm-Brause, C. J.,
Conradson, S. D., Donohoe, R. J., Hess, N. J., Musgrave, J. A. &
Tait, C. D. (1996) Environ. Sci. Technol. 30, 2322-2331.
Bertsch, P. M., Hunter, D. B., Sutton, S. R., Bajt, S. & Rivers, M. L.
(1994) Environ. Sci. Technol. 28, 980-984.
