Metal Retention
and Transport on Colloidal
Particles in the Environment
Ruben Kretzschmar1 and Thorsten Schäfer2
M
any potentially toxic trace metals and radionuclides are strongly
adsorbed onto surfaces of mineral and organic compounds in soils
and sediments, limiting their mobility in the environment. However,
recent studies have shown that trace metals in soils, groundwater, rivers, and
lakes can be carried by mobile colloidal particles. Understanding the release,
transport, aggregation, and deposition of natural colloidal particles is therefore of utmost importance for developing quantitative models of contaminant transport and the biogeochemical cycling of trace metals.
protozoans). Frequently, the close
association of mineral and organic
components leads to a different
colloidal behavior (i.e. heteroaggregation between different colloid types) compared to the behavior of pure mineral and organic
particles (homo-aggregation within a colloid type).
KEYWORDS: colloids, groundwater, radionuclide, soil,
Colloidal particles originating in
soils are also found in rivers and
lakes. Especially during strong
rainfall events, surface runoff and
resulting soil erosion can transport large amounts of colloidal particles into adjacent water bodies. Soil colloids can
also be transported into rivers and lakes by water rapidly
infiltrating through soil drainage systems. Organic colloids
in rivers and lakes contain typically 40–80% fulvic compounds of primarily soil origin and a significant proportion
(varying seasonally between 10 and 30%) of structural, fibrillar polysaccharides (rigid biopolymers) released from
plankton (exudates) or cell wall remnants (Buffle et al.
1998). The stability of inorganic colloids in these systems
strongly depends on the relative proportions of stabilizing
small, fulvic compounds and destabilizing rigid, fibrillar
biopolymers (Buffle et al. 1998).
trace metal, transport, vadose zone
INTRODUCTION
Colloidal particles are ubiquitous in environmental systems
and can consist of inorganic mineral phases, organic biopolymers or microorganisms (i.e. biocolloids). In general, the
importance of colloid transport in natural systems depends
on the presence, stability, and mobility of colloids and their
interaction (sorption and reversibility) with the pollutant of
interest (McCarthy and Zachara 1989). Colloidal particles
are usually smaller than about 1 µm in at least one dimension and have a high specific surface area (10–800 m2 g-1).
They can have a significant impact on the transport of trace
metals, which are strongly sorbed to solid phases and therefore rather immobile in the absence of mobile particles
(Kersting et al. 1999). Biocolloid transport is relevant both
as a pathogen risk and in remediation strategies to degrade
or immobilize redox-sensitive contaminants. Also synthetic
colloids (e.g. zero-valent iron) have been used in a remediation strategy to clean up aquifer systems contaminated by
redox-sensitive pollutants (Kaplan et al. 1996). In this
paper, we provide an overview of recent developments in
the understanding of colloid-facilitated transport of metals,
and we identify some research challenges to be addressed.
COLLOIDS IN THE ENVIRONMENT
Colloidal particles are found in virtually all natural waters,
including surface waters (i.e. rivers and lakes), soil porewaters, and shallow and deep groundwater systems. In soils
and aquifers, mobile colloidal particles originate mostly
from in situ generation of submicron-sized mineral and
organic matter particles, which are naturally present. Such
particles can be composed of clay minerals, mineral precipitates (e.g. Fe, Al, Mn, or Si oxides and hydroxides, carbonates, phosphates), organic biopolymers (e.g. humic and
fulvic acids, polysaccharides, and exocellular biopolymeric
material), and biocolloids (including viruses, bacteria, and
1 Department of Environmental Sciences, S-ENETH, ETH-Zurich,
CHN F23.2, 9092 Zurich, Switzerland.
E-mail: [email protected]
2 Forschungszentrum Karlsruhe, Institut für Nukleare
Entsorgung (INE), Postfach 3640, D-76021, Karlsruhe, Germany.
E-mail: [email protected]
ELEMENTS, VOL. 1,
PP.
205–210
The release of colloidal particles in soils and aquifers can be
strongly enhanced by fluctuations in water saturation, flow
velocity, and aqueous solution composition (i.e. pH, ionic
strength, redox potential). Such variations are insignificant
or at least gradual in most porous or fractured aquifer systems, but can be very pronounced and frequent in the
vadose zone (the unsaturated zone, including soils) and
fast-flow karst systems. For example, an intense rainfall
event typically results in rapid infiltration of water with low
ionic strength through soil macropores, causing flowinduced mobilization and chemical conditions favoring
colloid release and transport. Chemical and physical perturbations in aquifers can also be related to other factors
such as land use, artificial groundwater recharge, anoxic
and oxic groundwater mixing, or contaminant plumes
leaching from waste-disposal facilities. Such plumes can
lead to strong gradients in ionic strength, ionic composition, contaminant concentrations, and redox potential.
Colloids in chemically disturbed systems are commonly
formed by precipitation of small particles from aqueous
solutions. For example, Mavrocordatos et al. (2000) found
100–200 nm wide globular colloids, rich in Fe, Ca, and
organic C, in waters from a karst system in Switzerland. The
colloids presumably formed when Fe2+-rich water from an
acidic, anoxic peat drained into the neutral and oxic karst
system, leading to oxidation of Fe2+ to Fe3+ and precipitation of iron hydroxide colloids. Also, in acid mine drainage,
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colloids can be formed by precipitation. Zänker et al. (2002)
found that colloids containing Fe, Al, and trace metals
formed by the natural mixing of acidic, sulfate-rich acid
mine drainage water with near-neutral shaft or surface
waters. Mobile colloids can also form in highly alkaline
environments. Wan et al. (2004) reported the formation of
high concentrations of colloidal calcite in columns packed
with sediment from the US Department of Energy’s
Hanford site (Washington State, USA), after they were
leached with a highly saline and highly alkaline (pH 14)
solution. This experiment simulates on-site leakage of highlevel nuclear waste from underground storage tanks. These
examples show that mobile colloids can form from solution, depending on the local geochemical conditions.
Frequently, trace metals or pollutants are associated with
these colloids via sorption or coprecipitation.
A
B
CHARACTERIZING NATURAL COLLOIDS
Major advances in sampling and characterizing natural colloidal particles have been made during the past decade.
Some sampling procedures, especially at high pumping
rates, can introduce artifacts, including mobilization of colloids attached at sediment surfaces and creation of colloids
by changes in O2 and CO2 contents, temperature, pH, and
redox potential. New sampling protocols at low purging
and pumping rates while monitoring water quality parameters (especially turbidity) have been established to address
these concerns (e.g. Backhus et al. 1993). Important colloid
properties include (1) concentration (mass concentration,
number concentration); (2) size and shape (size distribution, particle morphology); (3) composition (mineralogy,
elemental composition, organic components); and (4) surface characteristics (specific surface area, surface charge,
sorption capacity). In this paper, we highlight a few selected
techniques. For further information, the reader is referred
to papers by McCarthy and Degueldre (1993), Buffle and
Leppard (1995), and Kretzschmar et al. (1999).
The mass concentration and elemental composition of colloids in water samples is routinely determined by inductively coupled plasma – atomic emission spectroscopy
(ICP–AES) or mass spectrometry (ICP–MS) using direct suspension nebulization. Typically, the water samples are analyzed with and without ultrafiltration or ultracentrifugation
in order to operationally differentiate among truly dissolved species, colloids, and suspended particles. Recent
developments include the coupling of size fractionation
techniques (e.g. asymmetrical flow field flow fractionation,
AsymFFFF) with ICP–MS to determine the size-dependent
colloid composition and trace metal–colloid association
(Plaschke et al. 2001). Size fractionation in AsymFFFF is
achieved in a thin ribbon-like channel in a laminar carrier
flow, applying a cross-flow perpendicular to the channel
flow. The sequence by which colloidal species are eluted is
determined by their diffusion coefficient and consequently
by their size.
(A) Scheme of the laser-induced breakdown detection
setup. The plasma created by a pulsed laser beam causes
an acoustic wave, which is detected by a piezoelectric receiver. In addition, the emitted light can be detected and spatially resolved by a microscope lens and a charge-coupled device (CCD) camera. (B) The figure
shows the relative number of breakdown events (BD probability) as a
function of laser pulse energy. The particle size distribution of a synthetic
three-modal colloid sample of polystyrene spheres is reproduced within
20% error. (C) From the fitted curve the particle size distribution is
appraised as shown in the histogram (data in Figures 2B and 2C reprinted
with permission from Walther et al. (2004); copyright 2004, American
Institute of Physics).
FIGURE 1
The composition of colloids can be further characterized by
a variety of established techniques such as X-ray diffraction
(XRD), X-ray fluorescence (XRF), Fourier-transform infrared
spectroscopy (FTIR), and total organic carbon analysis
(TOC). Compared to the mass concentration, which can be
obtained from elemental or gravimetric analysis, it is much
more difficult to determine in situ the number and concentration of colloidal particles. A very sensitive analytical
approach is the laser-induced breakdown detection (LIBD),
which is based on the formation of plasma due to a dielectrical breakdown in a focused pulsed laser beam when a colloidal particle is present (Kitamori et al. 1988). The plasma
is observed either by its light emission or by detection of
the acoustic shock wave generated by its rapid expansion
ELEMENTS
C
(FIG. 1). Because each plasma event corresponds to a single
colloid, the relative number of events per number of laser
shots, also called breakdown probability, provides a direct
measure of the colloid number concentration. The photon
flux required for breakdown decreases with increasing colloid size. Using this dependence, the colloid size and distribution can be ascertained by varying the energy of the laser
pulse (Walther et al. 2004).
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The shape of colloidal particles is studied mostly by transmission electron microscopy (TEM), high-resolution scanning electron microscopy (HR-SEM), or atomic force
microscopy (AFM). Perret et al. (1994) developed specialized
ultracentrifugation and staining methods for visualizing
the morphology of aquatic colloidal particles by TEM. More
recently, synchrotron-based scanning transmission X-ray
microscopy (STXM) combined with near-edge X-ray absorption fine-structure (NEXAFS) microspectroscopy was used to
characterize the size, shape, composition, and heterogeneity of mineral and organic colloids in soils, groundwater,
and surface-water samples. FIGURE 2 shows
STXM images of smectite colloids associated
A
with humic substances (Schäfer et al. 2003) and
a water-dispersible colloidal aggregate isolated
from a forested surface soil. Further characterization of the chemical heterogeneity of the particles can be achieved by cluster analysis
(Lerotic et al. 2004), in which pixels with spectral similarity are grouped. FIGURE 2 demonstrates that STXM is a very powerful tool to
resolve organic/inorganic colloid association
and chemical differences under nanoscale resolution.
Perhaps the most important properties of colloids are their surface area and surface charge.
The specific surface area determines, together
with the surface reactivity, the adsorption
capacity of the colloids for trace metals and
other environmental pollutants. While the
smallest particles do not contribute significantly
to the total mass of colloids in a water sample,
they usually contribute most of the surface
area. Therefore, they are of greatest importance
for colloid-facilitated transport of adsorbed
metals. The surface charge of colloids is
extremely important because colloidal stability
and transport are generally described in terms
of colloid and aquifer surface charge. Most natural colloidal particles have net negative surface
charge. Most clay mineral colloids have pHindependent negative surface charge, due to
the crystal structure of clay minerals. The
charge of natural organic matter (NOM) is usually dominated by carboxylic and phenolic
functional groups, creating pH-dependent negative charge. Only the hydroxides and oxides of
iron and aluminum have positive surface
charge at pH values below the pH at which the
surface charge is zero (pHPZC), which is near pH
8–9. However, in natural waters, adsorbed NOM
renders the surface more negatively charged.
Specific adsorption of phosphate and other
anions can also add negative charge to the surfaces of Fe- and Al-oxide colloids, which is very
important in fertilized soils and nutrient-rich
rivers and lakes (Stumm and Morgan 1996).
colloid surface charge, the role of natural organic matter in
stabilizing colloidal particles in suspension, and the formation of colloids in supersaturated solutions. Biological factors include the growth of biocolloids and the release of surfactants and polysaccharides by microorganisms, which
strongly alter mineral surface properties and colloid aggregation. In this article, we address only selected topics of
current research. Recent detailed reviews are provided by
Ryan and Elimelech (1996), Kretzschmar et al. (1999), and
McCarthy and McKay (2004).
B
C
D
E
F
G
H
PROCESSES CONTROLLING
COLLOID MOBILITY
Understanding the key processes controlling the mobility
of colloids in soils, groundwater aquifers, and aquatic systems is a major challenge. The mobility of colloids is governed by a combination of physical, chemical, and in many
cases biological factors. Physical factors include the transport of water and gases, the mobilization of colloids by
shear and hydration forces, and the filtration of particles in
porous media with heterogeneous pore systems. Chemical
factors comprise the influence of solution composition on
ELEMENTS
Scanning transmission X-ray microscopy (STXM) images
of (A) smectite colloids associated with natural groundwater humic substances and (B) colloidal particles from the surface horizon of a forested soil. Images A and B were taken at an incident X-ray
energy of 280 eV [i.e. below the C(1s) absorption edge], and image
contrast is produced mainly by the mineral components. Images C and
D were taken at 290 eV [ i.e. above the C(1s) absorption edge], and carbon-rich areas now also appear dark. In both cases, the cluster analysis
(E, F) indicated two chemically distinct regions within the particles (yellow and orange: mineral–organic regions; red: organic regions; blue:
background). The averaged NEXAFS spectra of each of these regions are
shown in G and H.
207
FIGURE 2
S EPTEMBER 2005
The transport of colloidal particles in the environment is
dramatically influenced by the tendency of the colloids to
form larger flocks (aggregation) or to attach to stationary
solid–liquid or gas–liquid interfaces (deposition). The influence of aggregation and deposition on colloid transport in
porous media is schematically depicted in FIGURE 3, along
with other important processes influencing colloid transport. Finely dispersed, submicron-sized colloidal particles
can be transported over large distances in groundwater systems (Schäfer et al. 2005) because their settling velocities
are extremely low and filtration by straining is rather ineffective. The rates of colloid aggregation and deposition are
controlled primarily by the balance between electrostatic
interaction forces (attractive or repulsive, depending on surface charge) and attractive van der Waals forces acting upon
colloid–colloid and colloid–surface collisions, respectively.
Colloidal particles with similar surface charge must overcome a repulsive electrostatic energy barrier in order to
aggregate. Screening of surface charge by dissolved or
adsorbed electrolyte ions reduces the electrostatic repulsive
force and thereby enhances colloid aggregation. Thus,
increases in ionic strength, presence of multivalent counter
ions, and the adsorption of oppositely charged ions to the
surface of colloidal particles can lead to enhanced colloid
aggregation. Likewise, colloid deposition onto similarly
charged stationary surfaces in porous media is hindered by
repulsive electrostatic forces, but screening of surface
charge again results in strongly increased colloid deposition
rates. In contrast, colloid deposition onto oppositely
charged surfaces is rapid and in many cases irreversible, as
long as the surface charge of the colloids and surfaces
remains unchanged. This is of great importance in chemically heterogeneous porous media, where even small patches
of positively charged surfaces can lead to effective removal
of negatively charged colloids from the flowing solution
(Elimelech et al. 2000).
Colloid mobilization and transport in the unsaturated zone
from soils to the groundwater table (vadose zone) is an
important research topic, which has been addressed by
DeNovio et al. 2004. Understanding colloid transport in
unsaturated porous media is complicated by the presence of
dynamic air–water interfaces and the increasing discontinuity of water-filled pores with decreasing water content.
Additional mechanisms of colloid retention have been suggested, such as film-straining and deposition of colloidal
particles on the air–water interface. The colloids are trapped
in a thin stationary film of water between the air–water
interface and the water–solid interface (Wan and Tokunaga
1997). However, these colloidal particles may be remobilized when the water content increases and the air–water
interfaces disappear due to increasing pore space saturation.
COLLOID-FACILITATED TRANSPORT
OF METALS AND RADIONUCLIDES
Under certain conditions, colloid-facilitated transport can
become the major transport mechanism of strongly sorbing
trace metals in soils and aquifers. This was demonstrated for
Pb transport in water-saturated soil columns (Grolimund et
al. 1996). The soil was first leached with a Pb-containing
NaCl solution of high ionic strength to simulate a contaminant plume. Subsequently, the soil porewater was
exchanged with low ionic strength CaCl2 solution to simulate fresh water infiltration. This leaching sequence resulted
in two transport fronts, first, an unretarded front in which
the salt concentration decreased, and second, a retarded
Ca2+–Na+ exchange front. Between these two fronts, the
low ionic strength and high Na+ saturation of the soil led to
colloid release and colloid-facilitated transport of Pb.
Kretzschmar and Sticher (1997) showed in column experi-
ELEMENTS
ments that Fe-oxide colloids that were stabilized by
adsorbed humic substances can also act as carriers for Pb.
Denaix et al. (2001) studied mobile colloids in a soil heavily
contaminated by former emissions from Zn and Pb smelters
in France. They found that 50% of the total Pb was bound
to colloids (mostly biocolloids), while 95% of the total Zn
was present as dissolved species. The remaining Zn was also
associated with biocolloids or smectites.
Colloid-facilitated transport can also play a very important
role in the radionuclide mobility in geological formations.
It has therefore to be considered in safety assessments of
geological nuclear waste disposal sites and risk assessments
of highly contaminated sites. For example, at the US
Department of Energy’s former plutonium-production
Hanford site (Washington State, USA), high-salinity, highly
alkaline radioactive waste containing 137Cs and 90Sr has
leaked from underground storage tanks into the underlying
sediments. Laboratory experiments on columns of Hanford
sediment showed that leaching with 1 M NaNO3 solution
and a subsequent decrease in ionic strength causes the
mobilization of colloids and illite-associated cesium (Flury
et al. 2002). Both findings suggest that colloid-facilitated
transport of 137Cs in these extreme environments is an
important transport mechanism. Numerous studies furthermore demonstrated that organic and inorganic colloids can
increase the mobility of Am, Np, Pu, and U on a laboratory
scale. Column experiments by, for example, Artinger et al.
(2003) on the influence of radionuclide–humic substances
contact time and Dardenne et al. (2002) on the effect of
metastable ferrihydrite mineral transformation showed
changes in metal sorption reversibility and therefore
colloid-enhanced radionuclide migration. Molecular understanding of surface complexation and structural incorporation of radionuclides is therefore a prerequisite for interpreting and reliably predicting the long-term influence of
colloids on radionuclide mobility. Another aspect is the
upscaling of colloid migration from laboratory to field
scale. A few field investigations (i.e. Vilks et al. 1997;
Kersting et al. 1999; Geckeis et al. 2004) have demonstrated
colloid-associated radionuclide mobility on a larger scale.
For example, the international Colloid and Radionuclide
Retention (CRR) project at the Grimsel Test Site (Switzerland) demonstrated that smectite colloids and colloid-associated 244Pu can be mobile in a natural rock fracture under
certain hydrodynamic conditions (flow velocity va ~ 39 m/d).
The alkaline (pH 9.6), low ionic strength (I = 2.2 mM)
groundwater at the Grimsel Test Site generally favors colloid stability and transport (FIG. 4).
Mobile colloids affect the transport, bio-availability, and
natural cycling of trace metals in lakes, rivers, and estuaries.
Colloid-bound transport of trace metals is of great importance in contaminated rivers, for example in the Vistula
River in Poland, which has been polluted in Pb and Cd by
zinc and lead mines in Upper Silesia (Guéguen and
Dominik 2003). In the Sacramento River in California, USA,
Hg was transported mainly by organic colloids and in a
“residual fraction” consisting of Hg mineral phases (Roth et
al. 2001). In a study of leachates from tailings at the New
Idria and Sulphur Bank mercury mines, in California, Lowry
et al. (2004) found that up to 95% of the total Hg was
present as inorganic colloidal phases, predominantly as
HgS. Powell et al. (1996) reported that Fe-rich colloids are
very important carriers for the trace metals Ni, Cu, and Cd
in an estuary in northwest Florida, USA. Similarly, colloids
rich in iron and organic matter carry about 90% of the total
uranium load in the Kalix River in Sweden, and uranium
transport through the estuary into the Baltic Sea is strongly
influenced by colloid aggregation as the salinity of the
water increases (Andersson et al. 2001). In several streams
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A
B
(A) Processes controlling the mobility of colloidal particles in water-saturated porous media: (1) release of colloids, (2) aggregation of colloids, (3) immobilization by gravitational
settling of flocks, (4) immobilization by straining of flocks or single particles, (5) immobilization by colloid deposition on solid–water interfaces,
(6) transport of dispersed colloidal particles over long distances by flowing water.
FIGURE 3
(B) In unsaturated porous media, additional mechanisms of colloid
retardation may occur: (7) immobilization by film straining and (8)
immobilization by deposition on air–water interfaces.
in the Pineland area of New Jersey, USA, an area characterized by sandy soils, peats, and acidic, organic-rich surface
waters, trace metals are bound largely to organic colloidal
materials (Ross and Sherrel 2004). The studies cited above,
along with many others, document the major role of colloidal particles in the transport and cycling of trace metals
in the environment.
RESEARCH CHALLENGES
Future studies need to address key challenges concerning
colloidal transport in saturated and unsaturated (vadose
zone) systems. First, representative colloid sampling in seasonally variable environments and the use of appropriate
techniques for sampling the vadose zone or weakly permeable formations remain critical challenges. Second, a quantitative understanding of colloid formation, release, stability,
transport, and deposition in heterogeneous porous media
remains to be developed. While a theoretical framework
exists for describing colloid transport and deposition in
defined model systems, important effects of charge heterogeneity, variable water saturation, preferential flow, surface
roughness, and variable redox conditions are still poorly
understood. Third, the molecular mechanisms, thermodynamics, and kinetics of sorption–desorption reactions of
trace metals, metalloids, and radionuclides on colloid surfaces must be further investigated because colloid-facilitated
transport critically depends on these processes. With the
rapid development of computational methods and new
analytical techniques, including synchrotron-based spectroscopy and microscopy, there will be exciting new opportunities for addressing these issues. .
ELEMENTS
Colloid migration experiment in a 2.23 m natural fracture
(dipole)—part of the international joint Colloid and
Radionuclide Retention (CRR) project at the Grimsel Test Site,
Switzerland. The colloid injection concentration of smectite colloid with
a size of 109 ± 10 nm, which was determined by laser-induced breakdown detection (LIBD), was adjusted to 20 mg/L. Breakthrough curves
are shown for the conservative tracer 131I (blue solid line), tetravalent
244Pu (red stars), and smectite colloids (black balls; detected in-line by a
mobile LIBD system). Radionuclide and colloid concentrations are normalized to the injected mass or activity m0. A vertical line marks the peak
arrival time of colloids and colloid-associated 244Pu, which typically
appear ahead of the conservative tracer (second vertical line) due to size
or charge exclusion effects. Background of graph gives a schematic
overview of the Grimsel Test Site tunnel and boreholes. Green, filled circles mark the injection (10 mL/min) and extraction boreholes (150
mL/min) of the asymmetrical dipole used for this study. For details see
Geckeis et al. (2004). Data of Figure 4 have been published in Geckeis
et al. (2004). COPYRIGHT 2004, OLDENBOURG WISSENSCHAFTSVERLAG.
209
FIGURE 4
S EPTEMBER 2005
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REVIEWS IN MINERALOGY AND GEOCHEMISTRY
Mineralogical Society of America and The Geochemical Society
MICRO- AND MESOPOROUS MINERAL PHASES
Volume 57, 2005 Giovanni Ferraris and Stefano Merlino (eds)
ISBN 093995069-3. US$40
For more description and table of contents of this book, and
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