pdf file - ETH - UP - Environmental Physics - ETH Zürich

Impacts of Ocean Acidification on Shelled Pteropods in the
Southern Ocean
Term Paper in Biogeochemistry and Pollutant Dynamics
Author:
Tutor:
Nanina Blank * Prof. Dr. Nicolas Gruber **
Handed in: June 22 nd 2007
Reviewed: November 23 rd 2007
*
Opfikonstr. 49, 8050 Zürich, nblank@student.ethz.ch
** Environmental Physics (UP); Institute of Biogeochemistry and Pollutant Dynamics (IBP),
Departement of Environmental Sciences (D-UWIS) at the ETH Zurich,
nicolas.gruber@env.ethz.ch
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Abstract
Roughly a third of the anthropogenic carbon dioxide emitted into the atmosphere has been taken up
by the ocean (Sabine et al., 2004), where it changes both the biological and geochemical processes.
The carbon dioxide makes the sea water more acidic and shifts the chemical equilibrium towards
smaller carbonate ion concentrations. The Southern Ocean is naturally poor in carbonate ions due to
its position in the global ocean circulation system and is thus especially vulnerable to changes
induced through carbon dioxide uptake (Sarmiento and Gruber, 2006). In this ecosystem planktonic
shelled mollusc, such as pteropods, are abundant. They use carbonate ions and dissolved calcium to
form aragonite for their shells. For them to be able to make those shells the water has to be
supersaturated with respect to carbonate. In general, surface waters have higher carbonate ion
concentrations and are thus supersaturated with respect to aragonite, while the deep sea has lower
carbonate ion concentrations and is thus undersaturated. As the ocean takes up anthropogenic
carbon dioxide, the saturated part of the water column shrinks. According to simulations of the
IS92a scenario (IPCC, 2000) the whole water column of the Southern Ocean will become
undersaturated with regard to aragonite by the end of this century (Orr et al., 2005). That poses a
great threat to the pteropods, because their shells show extensive erosion when exposed to
undersaturated conditions (Orr et al., 2005; Feely et al., 2004) and it is not known, whether they can
maintain their calcification. The exact pH dependence of aragonite calcification in pteropods is not
known, but if they have a similar sensitivity as other aragonite calcifying organisms, ocean
acidification will have an adverse if not fatal impact. The disappearance of pteropods from the
Southern Ocean might have a severe impact on the ecosystem, especially on the organisms relying
heavily on pteropods as a food source. There is also a resulting loss in calcium carbonate flux to the
deep sea, but compared to the changes for the pelagic ecosystem this is likely of minor concern.
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Index
1 Introduction ....................................................................................................... 4
2 Ocean Acidification........................................................................................... 6
2.1.1 Changes in pH and carbonate.................................................................................6
2.1.2 Changes in saturation state of aragonite and calcite .............................................7
2.2 Ocean Acidification in the Southern Ocean....................................................................8
2.2.1 Changes in saturation depth .................................................................................10
3 Pteropods......................................................................................................... 12
3.1 Characteristics ..............................................................................................................12
3.2 Pteropods in the calcium carbonate cycle ....................................................................13
3.3 Pteropods in the Southern Ocean..................................................................................14
3.4 Sensitivity to changes in pH and carbonate ..................................................................14
4 Impacts of ocean acidification on pteropods: Discussion .......................... 16
4.1 Effects of elevated CO 2 on calcification........................................................................16
4.2 Effects on ocean carbon system ....................................................................................17
4.3 Cascading effects...........................................................................................................17
4.4 Further effects ...............................................................................................................17
4.5 Summary of effects on pteropods ..................................................................................18
5 Future Research .............................................................................................. 18
6 References....................................................................................................... 19
Figures
Figure 1: Concentrations of H 2 CO 3 *, HCO - 3 and CO 2- 3 as a function of pH ................................6
Figure 2: CO 2 dissolution and aragonite formation in the ocean ...................................................7
Figure 3: Carbonate ion concentration versus depth......................................................................8
Figure 4: Currents feeding the Southern Ocean.............................................................................9
Figure 5: Atmospheric CO 2 scenarios and surface [CO 2- 3 ] in the Southern Ocean. ....................10
Figure 6: The aragonite saturation state in the surface ocean in the years 1994 and 2100..........11
Figure 7: The euthecosomatour pteropod Cavolinia tridentate. ..................................................12
Figure 8: A pelagic pteropod........................................................................................................12
Figure 9: Vertical distribution of thecocomes in western Subantarctic and Antarctic waters. ....14
Figure 10: Shell dissolution in a live pteropod ............................................................................15
Figure 11: Calcification rate versus saturation state. ...................................................................16
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1 Introduction
The absorption of anthropogenic carbon dioxide (CO 2 ) alters the ocean’s biological and
geochemical processes. Global emissions from human activity, such as the combustion of fossil
fuel, have boosted the CO 2 concentration from its pre-industrial value of about 280 parts per million
(ppm) to presently 380 ppm (Sabine et al., 2004). Roughly a third of the emissions have been taken
up by the ocean, turning them more acidic. The threat of this continuing ocean acidification to
marine life has only recently been recognised and many questions as to how marine ecosystems will
react are not yet answered (Kleypas et al., 2006). A decrease in pH (increase in hydrogen ions) will
shift the chemical equilibrium towards smaller carbonate ion (CO 3 2- ) concentrations. This has been
shown to affect the calcification of marine organisms, which form their tests and shells out of
calcium carbonate (CaCO 3 ) (Orr et al., 2005; Feely et al., 2004). Biogenic CaCO 3 excreted in the
upper water layers and sinking to the deep sea triggers a global cycle of carbon fluxes. Calcification
is found from the lowest trophic level, the phytoplankton, up to larger invertebrates such as
echinoderms.
The Southern Ocean is especially vulnerable to changes induced through CO 2 uptake due to its
position in the global ocean circulation system (Sarmiento and Gruber, 2006). In this ecosystem,
planktonic shelled molluscs, the pteropods, form the dominant part of the calcifying community
(The Royal Society, 2005). As their habitat is very sensitive to the uptake of anthropogenic CO 2 ,
they are likely to be among the first to be affected by ocean acidification. It is the aim of this paper
to summarize and assess what is known about the changes these organisms are faced with and how
they might respond to it.
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2 Ocean Acidification
2.1.1 Changes in pH and carbonate
Anthropogenic emissions have increased the concentration of CO 2 from its pre-industrial value of
280 parts per million (ppm) to 380 ppm today (IPCC, 2001). It will continue to rise and may well
reach the 1000 ppm mark by 2100, depending on further emissions. CO 2 is soluble in water and is
therefore taken up by the ocean out of the atmosphere. About 30 % of the anthropogenic CO 2,
roughly 120 Petagrams (Pg = 10 15 g) of carbon (C), has already been absorbed by the sea at a rate of
about 2 Pg C per year during the last 20 years (Sabine et al., 2004; Sarmiento and Gruber, 2006).
Different scenarios for the future CO 2 concentration have been developed by the Intergovernmental
Panel on Climate Change (IPCC) (IPCC, 2001). In this paper I chose the scenario IS92a as a
reference for my considerations. That is to be coherent with the studies of Orr et al., 2005 and Feely
et al., 2004, which both used the IS92a for their experiments and contain the most important data on
pteropods and ocean acidification for this paper. The IS92a scenario was developed in 1992. Then it
was considered an intermediate scenario with mid-range CO 2 emissions. Current emissions are
higher than assumed in IS92a, though, and rather follow the newer scenario A1F in Figure 5 a.
When CO 2 dissolves in water, it leads to a reduction in both pH and carbonate ion (CO 3 2- )
concentration (Figure 2 a-c). The more CO 2 is taken up, the lower is the consequent pH – resulting
in an acidification – and the lower is the resulting carbonate ion concentration (Figure 1). The ocean
is able to buffer its pH through the reaction with dissolved carbonate:
H 2 CO 3 * + CO 3 2- ↔ 2 HCO 3
-
or with calcium carbonate (CaCO 3 ) from the sediments or from calcifying organisms:
H 2 CO 3 * + CaCO 3 ↔ Ca 2+ + 2 HCO 3
-
Due to its slow circulation, the ocean is a slow sink, though, taking thousands of years to equilibrate
with the atmosphere from the surface to the bottom. The sediments would have a huge capacity to
buffer the pH changes induced by dissolved CO 2 , but the ocean circulation is too slow compared
with the rapid emissions and the sediment dissolution itself is a slow process. The average pH of the
ocean has already decreased by 0.1 units compared to its preindustrial value, which corresponds to a
26 % increase in the hydrogen ion concentration. Models under the IS92a scenario predict a drop of
surface ocean pH of more than 0.3 by 2100 and up to 0.77 by 2300, a condition the world’s oceans
likely have not experienced in the last 300 million years (Caldeira and Wickett, 2003; The Royal
Society, 2005).
Figure 1: Plot of the concentrations of H 2 CO 3 *, HCO - 3 and CO 2- 3 as a function of pH
The concentrations are plotted for a dissolved inorganic carbon concentration [DIC] of 2000
µmol kg -1 . Note that the vertical axis is logarithmic as well. From Sarmiento and Gruber, 2006.
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2.1.2 Changes in saturation state of aragonite and calcite
Changes in the ocean carbonate chemistry inevitably affect marine life. The first ones to be affected
are most likely calcifying organisms, which need the carbonate to build up their tests and shells.
There are two forms of calcium carbonate (CaCO 3 ) secreted by marine calcifying organisms,
aragonite and calcite. They differ from each other in their crystalline structures: aragonite has an
orthorhombic configuration and calcite a trigonal one. The different structure of aragonite results in
a about 50 % higher solubility in seawater than calcite (Feely et al., 2004). To secrete either of
them, the water has to be supersaturated with respect to the respective form. The state of saturation
is given by:
Ω aragonite = [Ca 2+ ] * [CO 3 2- ] / K sp aragonite ,
respectively:
Ω calcite = [Ca 2+ ] * [CO 3 2- ] / K sp calcite .
with Ω >1 indicating supersaturation and Ω

2.2 Ocean Acidification in the Southern Ocean
The calcifying organisms in the Southern Ocean are more vulnerable to decreases in pH and
carbonate ion concentration because the surface CO 3 2- concentration there is naturally low. The
boundaries of the Southern Ocean are not well defined since there are no continents bordering it,
but the definition adopted here is that it encompasses all regions south of 45° S. The higher
vulnerability to pH and carbonate decreases is due to the vertical distribution of [CO 3 2- ] caused by
(i) the biological carbon cycle and (ii) the predominant circulation. Here is how theses two
processes influence the pH and carbonate ion concentration in the water column:
i) In the upper water layers, photosynthetic organisms assimilate and calcifying organisms
form CaCO 3 . The assimilation of CO 2 influences the CO 3 2- concentration, it increases by 1 mol with
each mol of organic matter produced (simplified, for more detailed information see Sarmiento and
Gruber, 2006). The formation of CaCO 3 also has an impact on CO 3 2- concentration: it consumes 1
mol carbonate ions per mol biogenic CaCO 3 produced: Ca 2+ + CO 3 2- CaCO 3 . So, assimilation
produces CO 3 2- and calcium carbonate production reduces it. The organic matter, though, outweighs
the CaCO 3 by a factor of up to 5 (Sarmiento and Gruber, 2006), thus the carbonate ion
concentration is high in the sunlight penetrated part of the water column (Figure 3 a). When
organisms die, both organic matter and CaCO 3 skeletons and shells are eventually decomposed. The
remineralization of 1 mol organic matter uses up the 1 mol of CO 3 2- that was released during its
formation just like the dissolution of 1 mol CaCO 3 releases again the 1 mol of CO 3 2- it consumed
for its formation. But the ratio of the two decomposition processes varies with depth. While most of
the organic matter is already remineralized in shallow depths (Figure 3 b), the CaCO 3 sinks deeper
before it is dissolved. In the deep sea the organic matter-to-CaCO 3 ratio approximates one, but the
deep remains undersaturated (Figure 3 c). The resulting profile shows high [CO 3 2- ] in the upper and
low [CO 3 2- ] in the deep water layers. The depth at which waters are exactly saturated (Ω = 1) with
regard to a particular mineral phase is called the saturation horizon or saturation depth.
Figure 3: Plot of the carbonate ion
concentration versus depth. Also shown are the
carbonate saturation concentrations with regard
to calcite (solid line) and aragonite (dashed line).
The surface waters (a) are supersaturated with
respect to both aragonite and calcite. Further
down in the water column carbonate ions are
strongly decreased by remineralization of organic
matter (b). In the deep carbonate ions increase
slightly but stay undersaturated (c). The profile
was taken in the North Pacific at 42° N, 152° W.
Modified from Sarmiento and Gruber, 2006.
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ii) The Southern Ocean shows extensive upwelling, drawing up the deep waters from the
Atlantic, the Pacific and the Indian Ocean. This upwelling is fed by two major loops (Figure 4), one
conveying deep water from the North Atlantic and another running at the bottom of the Antarctic
northward and returning at mid-depth to the Southern Ocean. Hence the Southern Ocean is
constantly fed with deep water depleted of carbonate ions.
Figure 4: Simplified
graphic of the
currents feeding the
Southern Ocean.
The dashed circle
indicates the upwelling
region in the Southern
Ocean, fed by deep
North Atlantic water
(a) and deep water
from the Pacific,
Atlantic and Indian
Ocean basins (b).
After Sarmiento and
Gruber, 2006.
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2.2.1 Changes in saturation depth
Due to its upwelling carbonate ion-poor water the Southern Ocean is amongst all the seas the first to
suffer aragonite undersaturation at the surface. As seen in the example in Figure 3 the saturation
state is not uniformly distributed over the water column. In general surface waters are
supersaturated with respect to aragonite while the deep sea is undersaturated. According to the
model simulations of Orr et al.(2005), CO 2 concentrations in the atmosphere under the IS92a
scenario will reach 788 ppm. The resulting carbonate ion concentration at the surface of the
Southern Ocean will drop to 55 ± 5 µmol kg -1 , which is clearly below the aragonite saturation of 66
µmol kg -1 (Figure 5). The saturation horizon will shoal from its present depth of approximately 730
m all the way up to the surface (Orr et al., 2005), leaving the entire water column of the Southern
Ocean undersaturated (Figure 6).
Figure 5: Atmospheric CO 2
scenarios and
corresponding average
surface [CO 3 2- ] in the
Southern Ocean.
Time series of average
surface [CO 3 2- ] (b) in the
Southern Ocean for the IPCC
CO 2 scenarios (a) (IPCC,
2000).
Modified from Orr et al.,
2005.
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a: 1994
45°S
b: 2100
45°S 45°S
Figure 6: The aragonite saturation state in the surface ocean in the year 1994 (a) and 2100 (b),
respectively, as indicated by ∆[CO 3 2- ] A . The ∆[CO 3 2- ] A is the in situ [CO 3 2- ] minus that for aragoniteequilibrated
sea water at the same salinity, temperature and pressure. Shown are the median concentrations
of the Ocean Carbon-Cycle Model Intercomparison Project (OCMIP-2) on the surface in the year 2100 under
the IPCC scenario IS92a. Positive ∆[CO 3 2- ] A indicates supersaturation; negative ∆[CO 3 2- ] A indicates
undersaturation. The red dashed line (a) shows the approximate extent of the Southern Ocean, the black
dashed line (b) indicates where the aragonite saturation horizon has reached the surface. Modified from Orr
et al., 2005.
Since calcite is not as soluble in sea water as aragonite it will need a lower pH to reach
undersaturated conditions. However, calcite undersaturation in the Southern Ocean is thought to
follow the one of aragonite in only 50-100 years (Orr et al., 2005).
Second to follow the conditions of the Southern Ocean is the Arctic Sea, because it also has a very
low buffer capacity. In regions of the North Pacific and North Atlantic the aragonite saturation
horizon will shoal all the way to the surface by 2100 (Figure 6).
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3 Pteropods
3.1 Characteristics
Pteropods are shelled pelagic snails and belong to the zooplankton group. There are two Orders of
the Class of the Gastropoda, which are commonly referred to as pteropods. One being the
Thecosomata, the shelled pteropods considered in this paper, the other being the shell-less
Gymnosomata. The name ‘pteropod’ is derived from its most striking physiological feature, a winglike
modification of the molluscan foot (ptero- for wing, -poda for foot) (Figure 7 and Figure 8).
Thecosomes include 48 to 58 planktonic species. As they are very fragile, specimens get easily
damaged during collection and pose difficulties in their taxonomic assignment. The species differ
from each other in shell morphology, which is very multifaceted from spirally coiled or needle-like
to triangular. In order to minimise weight, the shells of thecosomes are very thin with a gauge of 6
µm to 100 µm. Being negatively buoyant, thecosomes use their wings for propulsion and
locomotion. The jerky flapping movement led to their common name ‘sea butterflies’. Many
thecosomes migrate diurnally, flapping towards the surface during the night and sinking back to
deeper water during the day. (Lalli and Gilmer, 1989)
Figure 7: The euthecosomatour pteropod Cavolinia
tridentate.
From Kleypas et al., 2006.
Figure 8: A pelagic pteropod.
Photo by Russ Hopcroft.
All thecosomes spend their premature life as males, changing sex as they age (= protandrous
hermaphrodites) (Lalli and Gilmer, 1989). Mating occurs between males that are about to mature to
females, so called reciprocal fertilisation. The sperm is kept until the transition is complete and the
females release fertilised eggs into the water. The lifespan of shelled pteropods is quite long,
ranging from months to 2.5 years (Fabry, 1989; Kleypas et al., 2006).
Being a larger zooplankton, the pteropods diet consists of phytoplankton and small zooplankton,
and it provides food for many predatory species of the higher trophic levels. Thecosomes secrete a
mucous web both for feeding and buoyancy (Lalli and Gilmer, 1989). The web is orbicular or
funnel-shaped and can exceed the thecosomes body size many times over. It is periodically drawn
in and the entangled prey is forwarded to the mouth. If disturbed, the animal will abandon its web.
Deserted feeding webs of thecosomes locally influence the ecosystem they live in. While drifting in
the water, organic particles entangle in the mucus and it thus serves as a breeding ground for
bacteria and as a food source for small grazers. The pteropod feces transport biogenic matter out of
the surface waters to the deep.
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Thecosomes are preyed on by larger carnivorous zooplankton, fishes, marine mammals such as
baleen whales and ringed seals, marine birds and gymnopods. Commercial fish like mackerel,
herring, North Pacific salmon or yellowfin tuna are also known to prey on shelled pteropods. To
some of the predators, thecosomes make up only an occasional part of their diet whenever available,
some are heavily dependent on them (Seibel and Dierssen, 2003) and some feed exclusively on the
marine snails. An example is the thecosome Clio antarctica in the Antarctic, whose only food
source is Limacina helicina, another thecosome (Seibel et al., 2003).
3.2 Pteropods in the calcium carbonate cycle
Pelagic snails form their shells out of aragonite, taking up carbonate ions (CO 3 2- ) dissolved in the
upper water layers and calcium (Ca 2+ ), which is abundant throughout the water column (Figure 2 e).
Shells of the dead molluscs sink, exporting the aragonite to deeper water where it is either deposited
forming sediments on the sea-floor or is dissolved if sinking below the aragonite saturation depth
(Figure 2 f, g).
There seems to be no agreement yet as to the exact contribution of pteropod aragonite shells to the
global export flux of carbonate from surface waters. This results from the lack of appropriate
measurement techniques and lack of observations. Pteropods have different sizes, generation times
and dissolution rates than foraminifera or coccolithophores, the major calcite producers, and are
therefore difficult to compare. Deep sediments as well as sediment traps show limited numbers of
pteropod shells because aragonite dissolves rapidly in undersaturated water both during sinking
through the water column and lying on the sea bed. So-called ‘pteropod ooze’, sediments with a
high proportion of calcium carbonate from pelagic snails, are found primarily in shallow waters
above the aragonite compensation depth. Studies suggest that aragonite contributes 12 % (Lalli and
Gilmer, 1989) to 50 % of the global carbonate export from surface waters to the deep (Berner in
Berger, 1977; Byrne et al., 1984). The estimated 50 % was based on a method that only used data
from above-average pteropod-rich sediments. 50 % is thought to be too high by Berger (1977) and
Lalli (1989). Berner and Hojo (1984) suggested 12 % of the calcium carbonate flux to be
contributed by aragonite (Lalli, 1989). Since they used sediment traps to determine the amount of
sinking thecosomes, the shells may already have been partly dissolved when the traps were
collected and the 12 % therefore been an underestimation. However there is little doubt that
aragonite plays a rather important role in the global oceanic carbon cycle. In the Ross Sea and the
Antarctic Polar Front pteropods have been found to dominate the export flux of carbonate (Orr et
al., 2005).
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3.3 Pteropods in the Southern Ocean
Some species are confined to very narrow environmental niches. In the Southern Ocean mainly four
species occur, namely Clio sulcata and Clio antarctica, Limacina helicina and Limacina retroversa
(Figure 9). The Limacina sp. are found in the shallower Subantarctic and Antarctic waters,
respectively, to a depth of about 250 m while the Clio sp. live between 200 and 800 m depth.
Thecosome population densities reach thousands of individuals per m 3 , at times even dominating
the zooplankton community (Cabal et al., 1999; Fabry, 1989).
Figure 9: Vertical distribution of
thecocomes in western
Subantarctic and Antarctic
waters.
From Lalli and Gilmer, 1989.
3.4 Sensitivity to changes in pH and carbonate
Very little is known about the pteropods reaction to elevated CO 2 concentrations and the resulting
effects. Up to today there are no long-term observations. However, there have been a few studies
looking at the reaction of the aragonite shells to increased carbon dioxide concentrations. Shells of
pteropods cannot persist in water undersaturated with respect to aragonite. Shells collected from
sediment traps below the aragonite saturation depth show extensive surface roughening and etching
(Byrne et al., 1984). In recent experiments live pteropods have been subjected to elevated CO 2
concentrations and consequent undersaturated conditions (Orr et al., 2005; Feely et al., 2004) as the
IPCC ‘business as usual’ scenario for CO 2 emissions (IS92a) would induce in the Southern Ocean.
The exposure resulted in pitting, peeling back of the exterior layer, and extensive erosion (Figure
10).
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Figure 10: Shell dissolution in a live pteropod. a-d, Shell from a live pteropod, Clio pyramidata,
collected from the subarctic Pacific and kept in water undersaturated with respect to aragonite for 48
h. The whole shell (a) has superimposed white rectangles that indicate three magnified areas: the shell
surface (b), which reveals etch pits from dissolution and resulting exposure of aragonite rods; the
prismatic layer (c), which has begun to peel back, increasing the surface area over which dissolution
occurs; and the aperture region (d), which reveals advanced shell dissolution when compared to a
typical C. pyramidata shell not exposed to undersaturated conditions (e). From Orr et al., 2005.
Conclusions are that the pteropods will not be able to produce their shells and thus to survive under
conditions undersatured with respect to aragonite and will be restricted to the shrinking latitudinal
and vertical expansion of saturated waters. If the Antarctic species will be able to migrate to and
survive in the warmer waters of lower-latitude oceans, which will stay supersaturated, is unknown.
Elevated carbon dioxide concentrations can further affect marine organisms directly in their
physiology. Acidification of body fluids through gills and body surface reduces the bloods ability to
transport oxygen. Some organisms react with decreased egg production or loss of sperm motility.
Effects on embryos and larvae have also been observed. However, there have so far been no studies
conducted assessing such impacts on pteropods.
There are no further studies assessing the response of pteropods to changes in parameters that they
will most likely face during the next century. The higher CO 2 concentrations will affect the
pteropods indirectly through the induced changes in the animal’s ecosystem. The pteropod’s main
food source, the phytoplankton, may be affected in various ways. Elevated CO 2 concentrations can
have an effect on the assimilation of inorganic carbon. And the lower pH affects the speciation and
bioavailability of nutrients and micronutrients. These two processes influence growth rates and
productivity of the phytoplankton and hence the food abundance for pteropods (The Royal Society,
2005). But none of these influences has been assessed in studies.
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4 Impacts of ocean acidification on pteropods: Discussion
4.1 Effects of elevated CO 2 on calcification
Since there are no studies examining the calcification rate of pteropods under elevated CO 2
concentrations, results from other calcifying organisms may help to make assumptions. There are
several patterns after which groups of organisms react. Kleypas et al. (2006), summarised results of
calcification rates of various organisms.
i) In the case of the examined corals it seems that calcification does not only rely on
supersaturated water, but also on the state of saturation (Ω). Calcification rates were reduced with
decreasing saturation state, although waters were still supersaturated with regard to aragonite
(Figure 11 a, dashed line). Calcification stopped completely in undersaturated water. If pteropods
reacted in the same manner, they would not be able to survive in undersaturated regions of the
oceans. However, there is the possibility that calcification can be maintained through cellular
mechanisms. Imaginable are ion pumps that can keep carbonate ion concentrations higher inside the
cells to enable calcification. The exact mechanisms how pteropods secrete their shells are not
known. If such techniques exist, the reaction of calcification might rather look like the solid line in
Figure 11 a. Depending on the state of undersaturation at which calcification stops, pteropods might
be able to compensate for a certain rate of shell dissolution as observed in Figure 10 through
continuing calcification. But even if they could – if they really respond according to this pattern -
they would already be affected now as the saturation state is still above 1 but decreasing.
ii) The examined foraminifera secrete calcite and show a different behaviour. Their
calcification rates remained rather unchanged over a wide range of Ω but declined as soon as the
water became undersaturated with respect to calcite (Figure 11 b). The slope of the decline is again
dependent on the presence of mechanisms that sustain calcification even in undersaturated water. If
pteropods reacted after this scenario, a decreasing saturation state would not affect them until it
sinks below 1.
Figure 11: Schematics of calcification rate versus saturation state.
After whatever pattern pteropods react to elevated CO 2 , there seems to be no doubt that their
calcification is reduced in waters undersaturated with respect to aragonite. Models and studies have
also convincingly shown that the Southern Ocean will be first to suffer undersaturation all the way
up to the surface. The deeper dwelling species Clio sulcata and Clio Antarctica will be first to meet
such conditions, followed by those further up in the water column. Whether they can keep on
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calcifying and compensate for the erosion of their shell is unknown, but is very doubtful if
carbonate ion concentrations in the Southern Ocean keep on dropping after 2100. Adaptation to
lower carbonate ion concentrations is possible but would have to happen within this century, which
is an extremely short time in evolutionary terms. And it is further aggravated by the pteropods
relatively long generation time. I have doubts that the pteropod species endemic to the Antarctic and
Subantarctic will be able to change over from their cold habitat to warmer lower-latitudinal waters.
Ocean currents would first have to allow their dispersal to the temperate zone. Their physiology
would have to be able to adapt to temperature increases of up to 10° C (Lalli and Gilmer, 1989).
Furthermore, they would have to find a niche in an alien ecosystem. Overall, it seems unlikely that
the Antarctic species will be able to evade extinction if the predicted CO 2 scenarios apply.
4.2 Effects on ocean carbon system
Due to the rather small contribution of pteropod shells to the global CaCO 3 flux, a change in
pteropod abundance will likely affect the global ocean carbon system less than it affects the
Antarctic ecosystem. The contribution of aragonite produced by pteropods is not huge but certainly
not negligible. A back-of-the-envelope calculation for the Southern Ocean can be made by the
following assupmtions: Aragonite secreting corals have shown a reduction in calcification within a
range of 13 to 83 % under [CO 2 ] of 840 ppm (Kleypas et al., 2006). If we assume that the average
decrease in calcification of these organisms (42 %) roughly applies to pteropods, that the [CO 2 ] will
rise according to the IS92a scenario, and that pteropods make up for half of the calcium carbonate,
then the decrease in export of CaCO 3 to the deep Southern Ocean by 2100 will be in the order of 20
%. The local ecosystem is far more affected by the decline of the pteropod population than the
global CaCO 3 flux.
4.3 Cascading effects
The decline or even disappearance of pteropod populations in the Southern Ocean is certainly going
to affect the ecosystem. Predators not entirely relying on pteropods will probably be able to
compensate for the lost food source with other species. But this results in increased predation
pressure on any alternative prey. Any predator feeding exclusively on pteropods will most probably
share the fate of its prey, either redistributing along with the pteropods or facing extinction (Royal
Society, 2005). Discriminating calcifying organisms will probably lead to a shift in the ecosystem
structure in favour of non-calcifying organisms. If there is less aragonite exported to deeper water,
the community of deep sea species may be affected. Deep-sea biodiversity and ecology has been
found to be dependent on calcifying planktonic organisms in near-surface waters (Kleypas et al.,
2006).
4.4 Further effects
• The pH decrease used by studies that observed direct mortality or decreased reproduction in
marine organisms is in most cases much higher than what is expected during the 21 st
century. I therefore assume that calcification in pteropods is inhibited long before these
other effects appear. Early life stages seem to be most vulnerable to pH changes. To draw a
final conclusion, effects of lower pH on larval stages of pteropods would therefore have to
be investigated.
• Phytoplankton will surely be affected by the ocean acidification and the global climate
change. But indirect impacts through the changes in phytoplankton have not been assessed
to an extent that would allow conclusions on the consequences for pteropods.
• Climate change increases sea temperatures. Direct effects on the pteropods’ physiology are
not known, but since the shoaling of the aragonite saturation depth already constrains the
- 17 -

pteropods towards warmer lower-latitude waters an additional rise in temperature will
probably increase worsen their situation.
4.5 Summary of effects on pteropods
Pteropods will be most affected by decreased calcification. Other effects of pH decline and
temperature rise may add to the negative impact but are secondary. It is unlikely that pteropods in
the Southern Ocean will survive the 21 st century. Their disappearance will have a fatal impact on
the ecosystem, especially on the organisms relying heavily on pteropods as a food source. The loss
in calcium carbonate flux to the deep sea is not irrelevant but of minor concern.
5 Future Research
Only recently have scientists recognised the threat posed by ocean acidification to aragonite
excreting organisms in the high latitude oceans. There is a substantial need for more research:
• Intensive monitoring should deliver reliable data about the absolute abundance of pteropods
in the Southern Ocean and the latitudinal distribution pattern of aragonite export to the deep
sea and how that changes with time. This would allow estimating the consequences of
reduced aragonite production by pteropods.
• The response in calcification rate of pteropods to elevated CO 2 concentrations, as they are
expected until the end of this century, should be quantified. This would require observation
over longer periods. Additionally the calcification mechanisms of pteropods should be
determined.
• Long term studies could possibly answer the question of how reduced calcification affects
the survival of pteropods, its population dynamics, and how their ecology changes (Kleypas
et al., 2006).
• Given the experimental data up to date about their reaction to undersaturated water, one has
to contemplate the disappearance of pteropods altogether from the Southern Ocean.
Therefore, it should be examined how the ecosystem reacts to an absence of pelagic snails.
• Examination of pteropod fossils and reconstruction of their evolutionary pathway may allow
some predictions about their ability to adapt to the changes in their ecosystem.
- 18 -

6 References
Berger, W.H., 1977. Deep-sea carbonate: pteropod distribution and the aragonite compensation
depth. Deep-sea research, 25, p.447-452.
Byrne, R.H., Acker, J.G., Betzer, P.R., Feely, R.A., Cates, M.H., 1984. Water column dissolution
of aragonite in the Pacific ocean. Nature, 312, p.321-326.
Cabal J.A., Alvarez-Marques F., Acuna J.L., Quevedo M., Gonzalez-Quiros R., Huskin
I., Fernandez D., del Valle C.R., Anadon R., 2000. Mesozooplankton distribution and grazing
during the productive season in the Northwest Antarctic Peninsula (FRUELA cruises). Dee-sea
research, 49, p.869-882.
Caldeira and Wickett, 2003. Anthropogenic carbon and ocean pH. Nature, 425, p.365.
Cao, L., Caldeira, K., Jain, A.K., 2007. Effects of carbon dioxide and climate change on ocean
acidification and carbonate mineral saturation. Geophysical research letters, 34.
Fabry, V.J., 1989. Aragonite production by pteropod molluscs in the subarctic Pacific. Deep-sea
research, 36, p.1735-1751.
Feely, R.A., Sabine, C.L., Lee, K., Berelson, W., Kleypas, J., Fabry, V.J., Millero F.J., 2004.
Impact of anthropogenic CO 2 on the CaCO 3 system in the oceans. Science, 305, p.362-366.
IPCC, 2000. Emission scenarios. Special report, found at
http://www.grida.no/climate/ipcc/spmpdf/sres-e.pdf.
IPCC, 2001. Climate change, 2001: The scientific basis. Contribution of working group I. Found at
http://www.grida.no/climate/ipcc_tar/wg1/index.htm.
Kleypas, J.A., Feely, R.A., Fabry, V.J., Langdon, C., Sabine, C.L., Robbins, L.L., 2006. Impact of
ocean acidification on coral reefs and other marine calcifiers. Workshop report.
Lalli and Gilmer, 1989. Pelagic snails. Stanford university press, California.
Orr, J.C., Fabry, V.J., Aumont, O., Bopp, L., Doney, S.C., Feely, R.A., Gnanadesikan, A., Gruber,
N., Ishida, A., Joos, F., Key, R.M., Lindsay, K., Maier-Reimer, E., Matear, R., Monfray, P.,
Mouchet, A., Najjar, R.G., Plattner, G.K., Rodgers, K.B., Sabine, C.L., Sarmiento, J.L., Schlitzer,
R., Slater, R.D., Totterdell, I.J., Weirig, M.F., Yamanaka, Y. and Yool, A., 2005. Anthropogenic
ocean acidification over the twenty-first century and its impact on calcifying organisms. Nature,
437, p.681-686.
Riebesell, U., Zondervan, I., Rost, B., Tortell, P.D., Zeebe, R.E., Morel, F.M.M., 2000. Reduced
calcification of marine plankton in response to increased atmospheric CO 2 . Nature, 407, p.364-367.
Sabine, C.L., Feely, R.A., Gruber, N., Key, R.M., Lee, K., Bullister, J.L., Wanninkhof, R., Wong,
C.S., Wallace, D.W.R., Tilbrook, B., Millero, F.J,; Peng, T., Kozyr, A., Ono, T., Rios, A.F., 2004.
The oceanic sink for anthropogenic CO 2 . Science, 305, p.367-371.
Sarmiento and Gruber, 2006. Ocean biogeochemical dynamics. Princeton university press,
Princeton.
- 19 -

Seibel and Dierssen, 2003. Cascading trophic impacts of reduced biomass in the Ross Sea,
Antarctica: Just the tip of the iceberg? Biological bulletin, 205, p.93-97.
The Royal Society, 2005. Ocean acidification due to increasing atmospheric carbon dioxide. Policy
document, found at www.royalsoc.ac.uk.
- 20 -

Review of Zero Valent Iron and Apatite as reactive materials for
Permeable Reactive Barrier
Author: Luca Geranio
Tutor: Dr. Evert Elzinga
Term Paper SS 07/08, major in Biogeochemistry and Pollutant Dynamics
Department of Environmental Sciences
ETH Zürich
June 2007
Abstract
Permeable reactive barrier (PRB) is a technology developed recently in the last years. It has obtained
promising results in the removal of several contaminants present in the groundwater. This Term paper
focuses the attention on two reactive materials, Zero valent iron and Apatite, employed in the PRB system,
giving an overview of the reactions and types of pollutants treated to date. The pollutants removal from
groundwater by these two reactive materials is based on different processes, that is, reduction of organic or
inorganic contaminants in the case of zero valent iron and immobilization of inorganic pollutants in the case
of Apatite. Zero valent iron acts as reductant either reducing directly the contaminants in a less harmful form
or also reducing pollutants which can subsequently precipitate. At the moment, Zero valent iron is the
material most frequently used in the field installations and it is particularly effective in the chemical
degradation of persistent chlorinated compounds into non-toxic and harmless by-products. In general
perhalogenated hydrocarbons tend to be reduced faster than hydrocarbons less halogenated and also the
dechlorination is more rapid at saturated carbons (e.g. CCl 4 ) centers than at unsaturated carbons (e.g.
TCE)[15]. Like anions, inorganic cations are reduced by Zero valent iron and precipitate as meagrely soluble
solids. Zero valent iron seems to have effect in the remediation of heavy metals such as chromium, arsenic,
technetium, selenium, copper, mercury and uranium.
Apatite II is a particular form of apatite (natural waste of fish industry) that can immobilize and sequester a
broad range of metals into new phosphate minerals and other low-solubility phases for a geologic period of
time. The apatite II provides a low but sufficient concentration of PO4 3- in solution in order to exceed the
solubility of the metal-apatite that rapidly precipitate, but only in the presence of an existing apatite structure
which acts as nucleating site or seed crystal. Apatite acts also as a material for non-specific adsorption of
most cationic metals from solution and is also an excellent buffer for neutralizing acidity through PO 4
3-
, OH - ,
and substituted CO 3
2-
, exerting control over chemical activities of other species leading to the precipitation of
oxihydroxide- and carbonate-metal phases. Metals in solution are immobilized on the apatite mineral by
surface sorption (main mechanism for most metals), precipitation (main mechanism for U, Pu, Pb,
lanthanides) or co-precipitation (transition metals). The relative contribution of adsorption and precipitation
to metal removal depends upon the environmental conditions, the mineral phases present, and the metal
concentration in solution [16] [14].

Table of contents
1. INTRODUCTION..........................................................................................................................ii
1.1 Permeable reactive barriers: the basic idea................................................................................ ii
1.2 Site characterization ................................................................................................................. iv
1.3 Choice of reactive media (RM) and laboratory tests.................................................................iv
2. GEOCHEMISTRY OF BARRIER MATERIALS..................................................................... v
2.1 Zero Valent Iron......................................................................................................................... v
2.1.1 Reactions and types of pollutants treated ........................................................................... v
2.1.2 Treatment of halogenated organic compounds with Zero Valent Iron.............................. vi
2.1.3.1 Chromium ................................................................................................................... viii
2.1.3.2 Technetium, Selenium, Arsenic ................................................................................... viii
2.1.4 Zero valent iron and inorganic cations ............................................................................viii
2.1.5 Precipitation problems........................................................................................................ix
2.2 Apatite ...................................................................................................................................... ix
2.2.2 Apatite II..............................................................................................................................x
2.2.2.1.1 Heterogeneous nucleation............................................................................................ xi
2.2.2.1.2 pH buffering................................................................................................................ xii
2.2.2.1.3 Surface chemi-adsorption............................................................................................xii
2.2.2.1.4 Biological stimulation................................................................................................. xii
2.2.2.2 Case study: use of apatite(II) in groundwater remediation contaminated with heavy
metals by mine waste in Idaho,USA ......................................................................................... xii
2.2.2.3 Benefits of use of Apatite II ......................................................................................... xiii
3. CONCLUSIONS AND REMAINING RESEARCH QUESTIONS....................................... xiv
4. AKNOWLEDGEMENT .............................................................................................................xv
5. REFERENCES............................................................................................................................. xv
ii

1. INTRODUCTION
1.1 Permeable reactive barriers: the basic idea
Permeable reactive barrier (PRB) is a technology developed beginning from the 90s as a remediation for
polluted groundwater. It is an implementation consisting of a permeable zone which cleans up a
contaminated plume through immobilization or transformation of the pollutants in a less harmful form. In the
subsurface the flow of the water is intercepted by a perpendicular “wall” of reactive materials that can
degrade, precipitate, sorb or exchange contaminants which can so reach the innocuous or legal concentration
downgradient the barrier [1][2][8]. The principle of applying PRB's in remediating contaminated ground
water is illustrated in Figure 1. This treatment technology is so called “in situ passive method” because
typically exploits the natural gradient of the flow allowing the water but not the contaminants to pass across
the barrier [2][8][23]. It is a new valid alternative to traditional Pump and Treat systems due to its low cost
and mantenance once installed and its potential long duration and effectiveness.
A benefit of its application is the large number of pollutants that can be treated often bringing their
concentration below the detection limit. Moreover the site on the surface above the treated aquifer is
available for other activities and can be economically re-used once the installation is concluded and during
the remediation process. There is the possibility to treat waste plumes that are heterogeneous in composition
and concentration and a reduction of costs (~50%) and energy compared to Pump and treat (P&T)
methodology could be achieve. In addition PRB has lower impact to the rate of groundwater flow compared
to P&T. Finally it could be a strong potential application in urban areas. [2][8]
One of the mains drawbacks for the utilization of this technology is that only the part of plume moving
downstream from the typically immobile source across the reactive materials can be treated.
Another drawback is that the barrier is permanent and fixed and can not be shifted if there is a deviation in
the movement of plume. PRB technology needs careful study and characterization of the site prior the
installation in order to avoid this drawback. All the possible subsurface changes must be taken into account
and the configuration and the size of the barrier must be able to respond to them.
Other requirements are the extensive knowledge of hydrology in the aquifer and the precise localization and
description of the polluted plume. Furthermore PRB is applicable only for shallow plumes ( not deeper than
50 feet down the ground surface). In addition to these hassles a long process of remediation is needed if the
aquifer has a low hydraulic conductivity. Finally a possible lessening of permeability of the PRB due to
corrosion, clogging and fouling of the reactive media could happen after a certain time. Unfortunately few
field data concerning the longevity of the reactivity of the PRB are available. [2][8][23]
The PRB is emplaced digging a trench in the ground and subsequently filled with adapted reactive media in
accordance with the hydrogeology characteristics of the site and the pollutant to treat.
Figure 1. Principle of groundwater remediation by using PRB [8]
Basically there are two main types of configuration for the installation of the PRB that are used for the field
application: the continuous PRB and the Funnel-and-Gate Design.
iii

In the continuous scheme the plume encounters a wall of reactive media for all its width and height. The
reactive zone such as in Funnel and Gate configuration has permeability equal or bigger than that of the
aquifer and so the groundwater changes very little or does not change its flux velocity and natural gradient. It
also does not deviate around the reactive media. It is good norm to place the inferior part of the barrier into
impermeable strata in order to prevent the contaminant underflow phenomena.
In the Funnel and Gate design the groundwater is directed by means of a impermeable funnel (commonly
sheet pilings or slurry walls) towards the permeable zone or “gate” constituted to the media. The cross
sectional area of the gate is usually rectangular and much smaller than that of the plume in the aquifer and so
the flow velocity within the reactive site will be higher [23].
1.2 Site characterization
A complete understanding of the site destined to PRB implementation is compulsory. You have to
characterize the stratigraphy, its variation in fracturing and permeability, and the local position and extent of
plume. Naturally the aquifer location, the groundwater flow direction, flow velocity and contaminant
concentrations are also fundamental in the appraisal and characterization of the site.
The reactive zone of PRB must be large enough that the entire plume will pass through, moving under the
natural ground-water gradient, also considering the recharge and seasonal variation.
It is also necessary to place PRB in order to avoid that the contaminant plume migrates partially beyond site
boundaries. Microbial activity can have beneficial or negative effects. Microbes can degrade contaminants
but can also biofoul and diminish the permeability of the barrier [23].
1.3 Choice of reactive media (RM) and laboratory tests
The chemical or physical processes, involved in the treatment with PRB, comprise reduction, sorption,
precipitation and biochemical degradation (aerobic or anaerobic) of contaminants.
Sorption includes adsorption, absorption and ion exchange to reactive media. In general, adsorption is the
adhesion of vapour or dissolved matter to the surface of a solid by physical and/or chemical forces,
absorption is the incorporation of one substance into or through another of a different state (e.g., liquids in
solids, gases in liquids) [31] [32] [8]. Remediation based on sorption phenomena usually use media like
activated carbon, zeolite, peat for organic compounds and heavy metal removal.
In the precipitation the contaminants pass from the soluble forms into insoluble solid states which are
detained. In the degradation the pollutants are transformed in less harmful compounds by chemical or
biological reactions. The type or types of processes ,mentioned above, that actually occur in PRBs and
remove pollutants from solution, depend on both the pollutant and the barrier material itself [8][23].
Once obtained the information from the site characterization it has to be chosen a suitable reactant.
This information is extracted based on laboratory experiments.
The reactive materials may be mixed with sand to facilitate the passage of the water through the barrier and
its amount vary proportionally to the mass flux of contaminants requiring remediation [2][8]. Among all the
factors that influence the choice of an appropriate media, the chemical composition of the contaminant is the
main and most important. You have to consider if the reactive medium reacting with the detrimental
compounds in the water promotes the formation of toxic by-products. The medium has to be known in depth
and the ideal one, beyond to be not a source of pollutant itself, has to be inexpensive and durable (i.e.
reactive over a long time scale). It has to have a size of the particles that consents the passage of the flow
without constrains and it has not to be much heterogeneous, avoiding clogging [2][8][23].
In addition the environmentally compatibility of reaction products and by-products has to be considered (e.g.
Fe 2+ , Fe 3+ , oxides, carbonates) . In order to evaluate the suitable of the reactive materials laboratory tests
concerning the rate and mechanism (including the formation of by-products) of pollutant removal are
performed. These laboratory tests coupled with site characterization information are the base for the
designing and the implementation of the PRB. Sometimes when there are a huge well-known data
concerning the removal rate of the contaminant, the laboratory tests can be eliminated. This is not applicable
when there is a mixture of pollutants.
The tests should be carried out with the ground water coming from the plume.
Batch studies: it is the test more appropriated for the selection of the reactive materials for the barrier. The
rate in the remediation of pollutants and the longevity of different materials can be evaluated under
iv

controlled condition. Typically, different samples are prepared with in each one a mixture of the reactive
material to test and an aqueous solution containing dissolved contaminants. It is possible to test the reactivity
of different materials simultaneously [23]. The mixtures react for a given period of time and the
concentrations of the contaminants at the beginning and at the end of the contact time are measured [13].
Column studies: the conditions of this test, like flow velocity, are more similar to those of the field. Based on
this studies, you can obtain the residence time of the contaminant in the reactive zone that can be used , with
the flow rate, to determine the thickness of the media [23]. Column tests are less cheap than batch tests but
approximate better the reality and can be useful for the estimation of long term performance. In this method
the reactive materials are packed in granular form into a columns fed with contaminated water [13].
In synthesis laboratory tests serve to assess the effectiveness and rate of pollutants removal of potential
barrier materials. In addition lab experiments evaluate the reaction products that are formed in the
remediation process and their eventually toxicity.
2. GEOCHEMISTRY OF BARRIER MATERIALS
In the following sections, two often used barrier materials will be discussed in terms of the (geo)chemical
processes that remove pollutants from solution. These materials are zero valent iron and apatite.
The choice to focus the attention on these two medium is due to the fact that they engage in different types of
geochemical processes leading to contaminant removal. Zero valent iron is the most used reactive medium in
PRB remediation, whereas Apatite is a very promising and cheap media which can remove several metals
from groundwater. Zero valent iron has the ability to degrade diverse organic and inorganic toxic
compounds. Zero valent iron acts as reductant either reducing directly the contaminants in a less harmful
form or also reducing pollutants which can subsequently precipitate.
Apatite is an effective reactive material in PRB for the removal of toxic metal forms from solution [16][14].
The apatite II provides a low but sufficient concentration of PO4 3- in solution (about 100 ppb PO 4
3-
or less
resulting in no phosphate loading or eutrophication, particularly important in ecosystem restoration and
maintenance) in order to exceed the solubility of the metal-apatite. The resulting and rapid precipitation of
phases such as Pb-pyromorphite or U-autunite happens only in the presence of an existing apatite structure
which acts as nucleating site or seed crystal. Apatite also acts as an excellent material for specific adsorption
of most cationic metals from solution. Apatite is an excellent buffer for neutralizing acidity through PO 4
3-
,
OH - , and substituted CO 3
2-
, exerting control over chemical activities of other species leading to the
precipitation of oxihydroxide- and carbonate-metal phases.
Environmental conditions, mineral phases present and metal concentration in solution determine the relative
contribution of adsorption and precipitation to metal removal. The evaluation of which mechanism dominate
at any particular site is possible with simple feasibility studies on the contaminated groundwater and soil
under site conditions. [16][14]
2.1 Zero Valent Iron
2.1.1 Reactions and types of pollutants treated
Zero valent iron (ZVI) has a demonstrated effectiveness against a broad variety of contaminants, especially
towards the chlorinated aliphatic hydrocarbons (CAHs) [2][8][23].
Zero valent Iron is the most common reactive medium used in PRB remediation. ZVI is instable under
natural conditions and it has to be fabricated at high temperature. In the normal atmospheric oxygen
condition and at low temperatures, the oxidation of ZVI is negligible by means of the formation of oxide
films that act as inhibitors and do not enable the surface exposure. The standard potential of Fe 0 /Fe 2+ couple
is -0.440V and this negative potential permits ZVI to act as an electron donor and reduce several redoxlabile
compounds (oxidized species) [12]. The favourable thermodynamical electrochemical corrosion of
ZVI, which happens in the aqueous system, is necessary for the remediation of contaminants using this
material [30]. There is a passage of electrons from the iron surface to the oxidized organic pollutant, which
becomes reduced and completely or less harmful. At the same time there is a production of soluble cations of
the metal. As long as electron acceptors are present, corrosion processes and electron transfer can
continue[2][8][23].
v

Fe 0 Fe 2+ + 2e -
2H + + 2e - H 2(g)
Fe 0 + 2H + Fe 2+ + H 2(g) Net reaction
The surface area of Iron per unit volume of pore water, so called specific surface area, plays a determining
role in the rate of degradation of the contaminants. It influences directly the number of active surface sites
exposed to the groundwater plume that are very important for the initialization, mediation and degradation
reactions, independently from the nature of the pollutant. However an elevated specific surface can lead to a
lessening of the permeability of the barrier which should be much greater than that of the surrounding aquifer
layer. Other intrinsic factors, for ZVI, like the grain size and shape, the manufacturing process, the content in
alloying element (P, Ni, S, C, Cr), have an additional role. The chemistry of the groundwater and the
consequent influence in the corrosion process is important too: for example many dissolved species like
chloride, carbonate, sulphate enhance the corrosion processes and can lead to the formation of unstable
minerals[23]. The mineral precipitation is a factor of great importance in the appraisal of the performances of
a reactive permeable barrier in the long term [3]. If water contains a high amount of carbonates, there is an
increase in calcite (CaCO 3 ) and siderite (FeCO 3 ) precipitates. It can moreover be observed the precipitation
of iron oxides and hydroxides, among which the ferric hydroxide ( Fe(OH) 3 ), the ferrous hydroxide
(Fe(OH) 2 ), the magnetite (Fe 3 O 4 ) and maghemite ( Fe 2 O 3 ); these are responsible of the ” effect
coating”[15][23][12]. In field applications of PRBs iron is always used in mixture with sand to avoid a
complete clogging[11].
Table 1. Contaminant treatable by reactive material in PRBs[23]
vi

2.1.2 Treatment of halogenated organic compounds with Zero Valent Iron
ZVI is one of the reactive medium more used for the treatment of numerous organic compounds and some
inorganic compounds present as contaminants in the groundwater (see Table 1). In the last years several
researches have been focused above all to the degradation of chlorinated compounds like TCE and PCE, by
means of reactions occurring on the Fe 0 surface. In general perhalogenated hydrocarbons tend to be reduced
faster than hydrocarbons less halogenated and also the dechlorination is more rapid at saturated carbons (e.g.
CCl 4 ) centers than at unsaturated carbons(e.g. TCE)[15].
The degradation process of the chloroether derivates can involve three steps[15]:
1. the contaminant adsorption on the reactive sites of the barrier;
2. reaction on the ZVI surface reaction;
3. final products desorption;
Nowadays the reaction between Iron and Chlorinated solvents is considered as an abiotic reductive
dehalogenation [23] which can take place based on three degradation mechanisms, according to the
conditions of the groundwater ,such as the amount of oxygen. This reaction occurs at the surfaces of Fe(0)
that, through their corrosion and donation of electrons, permits the chlorinated hydrocarbon’s reduction and
dehalogenation to hydrocarbon (non-toxic product). The resulting chloride is then dispersed in the water
phase[23].
The three mechanisms are[15]:
A)
Fe 0 Fe 2+ +2e - Anodic reaction
RCl + 2e - + H + RH + Cl - Cathodic reaction
Fe 0 + RCl + H + Fe 2+ + RH + Cl - Net reaction
The reduction happens for transferring electrons from surface metal to the chlorurated molecule adsorbed on
it.
B) The ferrous ions (Fe 2+ ) yielding through the ZVI corrosion by water are further oxidated to ferric
ions(Fe 3+ ) and than the chlorinated compounds are reduced. The net reaction is:
2Fe 2+ +RCl+H + 2Fe 3+ + RH + Cl -
2Fe 0 +O 2 +2H 2 O 2Fe 2+ + 4OH -
4Fe 2+ +4H + +O 2 4Fe 3+ + 2H 2 O
C) H 2 produced through the iron corrosion by the water shifts its electron to the chlorurated substance.
H 2 +RCl RH +H + + Cl -
It is most likely that the dehalogenation proceeds via sequence A).
When the compounds to treat, like the chlorinated aliphatic compounds, and the oxygen have similar oxiding
potential, they can compete for being the favoured electrons acceptor. If the oxygen is sufficiently present,
Fe 2+ is further oxidized to Fe 3+ and can precipitate as hydroxide or (oxy)hydroxides at the elevated pH in the
reactive zone of the ZVI[12].
Aerobic system[2][12]:
2Fe 0 +O 2 +2H 2 O 2Fe 2+ +4OH -
4Fe 2+ +4H + +O 2 4Fe 3+ +2H 2 O
Fe 3+ +3OH - Fe(OH) 3(s)
Under anaerobic conditions the corrosion reaction proceeds more slowly, yields an increase in the pH (due to
the consumption of H + like in the aerobic condition) and leads to the formation of ferrous (oxy)hydroxides
instead of ferric (oxy)hydroxides[2][12].
Fe 0 +2H 2 O Fe 2+ + H 2 + 4OH -
Fe 2+ +2OH - Fe(OH) 2(s)
vii

2.1.3 Zero valent iron and inorganic anions or oxyanions
Elements which can be in the oxyanion form under environmental conditions include selenium, arsenic,
chromium, technetium, antimony, nitrate, phosphate, sulphate. The anionic species are very soluble in water
and they are not attracted to the common negative surface of the minerals in aquifer [28]. This fact implies
that these oxyanions can be potentially persistent in high concentration in ground water. These elements have
in common that the reduced forms (which can be generated by the ZVI) are the immobile ones. The ZVI is
then used in the barrier to reduce the oxidized species in the groundwater to the reduced forms which then
precipitate out of solution and are thus removed from the groundwater[23].
2.1.3.1 Chromium
Cr can appear commonly in environment as Cr(VI) or Cr(III). Cr(III) is relatively a micronutrient, non–toxic
is adsorbed by some minerals and forms scarcely soluble hydroxide precipitates . Hexavalent Cr, instead is
carcinogenic, more soluble and then more persistent and mobile[23].
Several researches have been done on Cr interactions with Fe(0), and actually there are PRBs that have been
specifically designed to treat Cr-contaminated groundwater with Fe(0).
2-
In the ground water Cr(VI) usually appears in the CrO 4 form and it is not adsorbed by aquifer materials. In
the ZVI system, the reduction of Cr(VI) to Cr(III) coupled with the oxidation of Fe(0) to Fe(II)and Fe(III)
and the subsequent precipitation of Fe(III)-Cr(III) oxyhydroxides or hydroxides seems to be the main
mechanism. Laboratory tests have investigated some materials containing reduced iron to see which was the
most effective has been seen being in the Cr(VI) removal through reduction and precipitation: the most
suitable medium was found to be Fe(0). Also column tests and field applications have shown a rate of Cr(VI)
reduction and precipitation by Fe(0) suitable for ground water remediation system.
Sequestration of Cr(VI) provokes a steep increase in pH from initial neutral condition(6,5100mV to Eh

2.1.4 Zero valent iron and inorganic cations
Like anions, inorganic cations are reduced and precipitate as meagrely soluble solids. Industrial, mine and
nuclear sites are the principal sources of Hg, Cu, Tc and complex such as UO 2
2
, responsible of their high
concentration. Laboratory experiments have shown for mercury, technetium, uranium and copper treated
with ZVI a reduction and a subsequent coprecipitation within secondary precipitates.
For U(VI) the reaction proposed is the following[23]:
Fe 0 +UO 2+ 2(aq)
Fe 2+ +UO 2 (s)
UO 2 (s) (uraninite) can be amorphous or crystalline precipitate and its solubility is in the range of 10 -8 mol/l in
a pH range between 4 and 14. Its solubility can enhance under oxidising conditions and below pH 4.
UO(VI) is reduced to U(IV) spontaneously by elemental iron[23].
2.1.5 Precipitation problems
Geochemical changes such as pH increases and oxygen elimination, occur in water passing a barrier of Fe
(0). These changes can lead to precipitation of secondary minerals onto the reactive surface which can
influence the reactivity and permeability of the ZVI system over the time[22]. These redox conditions
permits precipitation of secondary minerals from ions typically present in ground water as well as some
ground water contaminants [29]. The typically secondary minerals formed in PRBs are magnetite (Fe 3 O 4 ),
hematite(α-Fe 2 O 3 ), goethite (α-Fe 3+ O(OH)), lepidocrocite ( ɤFeOOH) , calcite (CaCO3), aragonite (CaCO 3 ),
siderite (FeCO 3),
2+ 3+
green rust ([Fe (1–x) Fe x (OH) 2 ] x+ [x/n A n–·m H 2 O] x– , where x is the ratio Fe 3+ /Fe tot ), ferrous
hydroxide Fe(OH) 2 , ferrous sulfide (FeS 2 ), and marcasite (FeS 2 ). In general calcium carbonates and siderite
are found close to the entrance of a PRB, while magnetite, ferrous hydroxide, green rust and iron
oxyhydroxides generate throughout a PRB. Accumulation of secondary minerals is responsible of loss of
pore space and reactive surface area of the medium in PRBs, which can alter flow paths, residence times, and
effectiveness of a PRB treatment. Due to site-specific geochemical and hydrogeological conditions, and the
relatively long period over which mineral deposition occurs, it is complicated developing a general
assessment of mineral precipitation in PRBs using field and/or laboratory data [29].
The effects that an implementation of a ZVI barrier produces intra-wall and down gradient are a loss of
dissolved oxygen, decreasing Eh, reduction in carbonate alkalinity increasing in pH value up to 9 or 10 and
in Fe 2+ that can also precipitate as oxyhydroxide colloids. Loss of cementation between the grains and
precipitate formation can generate mobile colloidal particles that can transport toxic substances. Deeper
explorations have to be carried out to know which type of geochemical characteristic of the plume can
influence the liberation and mobilization of immobilized metals and the transportation of the colloidal
materials. Iron hydroxides, with the proceed of the corrosion, can form a more and more thick passivating
layer on the surface of iron grains. This phenomenon gradually occludes the Iron surface and reduces its
reactivity [23]. Since Ferrous (oxy) hydroxide is thermodynamically unstable, it might be further oxidized to
magnetite or goethite that is non-passivating and seems to allow sufficient contaminant degradation
instalments over the years [23]. Under conditions of pH neutral, it can be formed a mixed valence
(Fe 2+ /Fe 3+ )compound, said “green rust”, as intermediate product of magnetite (Fe 3 O 4 ) formation from iron
hydroxide. The “green rust” is stable only at low grades of oxide reduction and its oxidation leads usually to
the formation of maghemite (Fe 2 O 3 ) or lepidocrocite ( ɤ FeOOH), with a prevalence in the formation of the
first compound compared to the second[7]. The maghemite, differently from the magnetite, is responsible for
iron passivation. If ”effect coating” prevents the Ionian ferrous passage in solution, the surface of the iron is
polarized by work of Fe 2+ ion and diminishes therefore the tendency of the iron to corrode itself. It has also
been seen that, if the concentration of organic chlorinated compounds increases, the half life of the polluting
agents increases too. That depends on the increasing of the passivated iron surface and on the emphasized
saturation of the reaction site. The passivation phenomenon is influenced moreover from the composition of
the iron [10]. The mineral precipitation can involve, at the same time of reactivity reduction, also various
negative impacts on the hydraulic conditions of the barrier [3]. The filling of the pores can raise the
heterogeneity inside the barrier, leading therefore to the formation of preferential paths [9], increasing of
fluid speed and meaningful lessening of the time of residence inside of the PRB. Today there are studies
describing the performances in the long term of the PRB containing zerovalent iron.
ix

2.2 Apatite
2.2.1 Structure
Apatite is a microporous mineral described with general formula Ca 5 (PO 4 ) 3 (OH,F,Cl), or more accurately by
the unit cell contents [Ca 4 ][Ca 6 ][(PO 4 ) 6 ][F] 2 . [Ca 4 ][Ca 6 ][(PO 4 ) 6 ][F] 2 is composed by CaO 6 columns linked
together with PO 4 groups in order to form “a hexagonal network like a honeycomb with channels extending
right through the structure in c direction “ [25].
This one-dimensional tunnel structure is very stable and strong, and is chiefly due to the arrangement of
calcium and phosphate. The structure can be considered a tunnel structure with walls composed of cornerconnected
CaO 6 and PO 4 polyhedra as relatively invariant units. Compared to fluorapatite, in the hydroxyl
apatites the presence of OH - , that is a little larger than F - , is responsible for the resulting expanded structure.
The microporus structure with tunnels filled to Calcium and anions (OH,F) leads to an adjustment that best
satisfies bond-length requirements [25]. Little changes in the ionic radii of the tunnel atoms cause the
expansion or contraction of the channel. The channels can accommodate different contents, accepting large
cations of different valence through the introduction of framework counter ions. Apatites are reversible ion
exchangers for some anions and cations, are stable in the subsurface and do not induce microbial
blooms[25].
2.2.2 Apatite II
2.2.2.1 Reactions of apatite and metals
Apatite II has the general formula [Ca 10-x Na(PO 4 ) 6-x (CO 3 ) x (OH) 2 where x10 -20 [16] .
Metals in solution are immobilized on the apatite mineral by surface sorption (main mechanism for most
metals), precipitation (main mechanism for U, Pu, Pb, lanthanides) or co-precipitation (transition metals).
It has particular characteristics that render it fit for optimal performance in the field, namely: its substituted
2-
CO 3 and sodium that destabilize the structure and make the material more soluble, no substituted F, low
trace metal concentration, poor crystallinity (>90% amorphous) and high microporosity [6]. These properties
influence the kinetics and solubility. F as minor constituent in the structure raises lattice stability and
decreases solubility and dissolution rate. Instead carbonate has the opposite effect[16]. Differences in the
performance among various apatite phases result from differences in those properties which influence the
kinetics and solubility, e.g., crystallinity and minor element chemistry. A higher degree of crystallinity
decreases solubility and dissolution rate, making the apatite less reactive and less effective, whereas lower
crystallinity increases solubility. Thus, the amorphous form of the solid is the most reactive. The presence of
F as a minor constituent in the apatite structure increases lattice stability, decreasing solubility and
dissolution rate. The presence of carbonate as a minor constituent in the apatite structure decreases lattice
stability, increasing solubility and dissolution rate[16].
Phosphate-Induced Metal Stabilization (PIMS) is a remediation technology that stabilizes a broad range of
radionuclides and metals like Pb, U, Cd, Zn, Cu, and Al which are bound into new phosphate minerals and
low solubility phases. The metal–phosphate phase is very stable over geological time because of its low
solubility products ( K sp ), as shown in table 2 [6][14]. These metal-phosphate products are
thermodynamically stable over a broad range of environmental conditions, and precipitate rapidly, thus
ensuring a stable and long term immobilization of metal pollutants. They have great durability, low solubility
and are stable over a wide range of conditions [14].
x

Table2-Solubility products of metal phosphates [27]
Metals and apatite react very quickly ([17], [20], [26], [4], [5], [21])on molecular scale, but at a macroscopic
dimension, limiting parameters like grain size, flow rate, barrier design, influence the efficiency with which
dissolved metals come into contact with the surface of the reactive media. Apatite II has treated successfully
contaminated soils, groundwater and wastewater for Pb, U, Cd, Zn, Al stabilizing between 5% and 50 % of
its weight in metals depending upon the metal and the environmental conditions[16] . Pb,Cd, and Zn are
sorbed by apatite with different mechanisms. Pb is primarily removed with this mechanism: the dissolution
of the apatite that provides PO 4
3-
and the following precipitation of hydroxyl fluoropyromorphite on Apatite
surface. Minor otavite (CdCO 3 ) precipitation was observed in the interaction of the apatite with aqueous Cd.
However other sorption mechanisms, such as surface complexation, ion exchange, and the formation of
amorphous solids, are primarily responsible for the removal of Zn and Cd [5].
There are four general factors that aid in the effective sequestration of metals by apatite II, depending upon
the type of metal and its concentration and the groundwater chemistry: 1) heterogeneous nucleation; 2)
buffer acidity to pH 6.5 to 7; 3) surface chemi-adsorption; 4) biological stimulation; [21]
2.2.2.1.1 Heterogeneous nucleation
A small, but sufficient, amount of phosphate is provided to solution in order to overtake the solubility limits
of the metal-apatites.
The (K sp ) of the resulting metal-apatites is very low for the metals Pb, U, Cd, Zn, Cu and Al and in particular
for Pb that is equal to K sp =10 -167 [21].
Pb-pyromorphite can precipitate only through heterogeneous nucleation, namely, a seed crystal with the
apatite crystal structure is necessary for the precipitation , unless Pb concentrations exceed about 10 ppm
( not frequent in environmental conditions) [18][19].
The Apatite II amorphous structure is relatively reactive and supplies a slight excess of phosphate ions to
solution. The random nanocrystal inclusions of crystalline apatite, provide the seeds for nucleating the
precipitates.
Pb in the system precipitates only as microscope Pb-pyromorphite minerals which over the time will grow
and join together eventually forming larger mineral bunches [22]. The presence of apatite II- supplied
phosphate in this process ,that can take many years, keeps the concentration of Pb in solution extremely low,

limits[16]. The degree of protonation, namely the number of hydrogen ions attached to the PO 4
3-
, in the
intermediate reactions, depends upon the solution pH. The example above is for the range of acid soils or
acid mine drainage, pH

phases by precipitation that it has resulted being a mixture of sphalerite and recrystallized apatite. Within the
first few feet of the PRB Pb and Cd vanish completely from solution, Zn concentration instead diminishes
continuously along the barrier, finally dropping to about 0.1 mg/L or below once exited from it.
In order to understand which mechanisms are operating and how much time the barrier can last, analysis of
material captured from barrier and geochemical modelling using MINTEQ and PHREEQ were performed.
MINTEQ indicates that pyromorphite and sphalerite (ZnS) are the steadiest phases within the barrier for Pb
and Zn. These results are based to the dates displayed in the table 3. The other metals such as Mn,Cu, Al and
U are also being reduced from ppm/ subppm level to the pbb level or below detection. This indicates that
biological reduction, pH buffering, precipitation and surface adsorption all act in the PRB. Cd can undergo
adsorption onto the Apatite II or precipitate as a sulphide similar to Zn.
Table 3- Changes in some groundwater constituents entering/exiting the Apatite II PRB in August 2002 [14]
2.2.2.3 Benefits of use of Apatite II [6]
The use of Apatite II is a new remediation technology that has several advantages.
First at all Apatite II can sequester most heavy metals and radionucleotides for geologically long time
periods. Many metals and radionuclide contaminants precipitate as mineral phases in which the metals are
not bioavailable. Apatite II is one of the best non-specific surfaces sorbent for species that do not precipitate
as a separate phase and is the most cost-effective reactive media for most metals and radionuclides, e.g.,
immobilizes up to 20% of its weight in Pb, U, Pu and perhaps other metals.
Apatite II can be mixed into contaminated soil or waste, emplaced in PRB or used as a disposal liner and can
be combined with most other reactive media e.g., grout, bentonite, zero valent iron, and other remediation
technologies, e.g., bioremediation, etc.
xiii

3. CONCLUSIONS AND REMAINING RESEARCH QUESTIONS
The feasibility of PRB’s installations depends on the chemical, geological and environmental settings of the
PRB’s capture zone. The reactivity of the barriers can be impaired by the remediation processes themselves
(e.g. oxidation), by the reaction products like for example precipitates and by exhaustion of sorption
capacity[11]. The types of materials used in barriers are those changing redox potential (e.g. zero valent iron
– degradation of chlorinated hydrocarbons), those causing precipitation, materials with high sorption
capacity, those releasing nutrients/oxygen to enhance biological degradation [12];
Zero Valent Iron is the reactive medium in Permeable Reactive Barrier system most used because of its vast
capacity to remediate a large number of contaminants through its corrosion and donation of electrons for the
reduction of the pollutants in a harmless state. Hence, are available numerous terms of comparison from
which can be extrapolated the guide lines for the planning and the emplacement of new installations[15].
ZVI is very effective towards the chlorinated aliphatic hydrocarbons acting as reductant and degrading them
through a dehalogenation. For the contaminants in the form of inorganic cations and anions, their reduction
and their consequent precipitation are the processes involved when is utilized the ZVI as reactant[23].
To date the mechanisms of remediation of ZVI towards several compounds are fairly noted and they have
been studied in depth. Not completely clear and requiring further studies are the long term effects of some
precipitates like maghemite (Fe 2 O 3 ), ferric hydroxide ( Fe(OH) 3 ), ferrous hydroxide ( Fe(OH) 2 , a mixed
valence (Fe 2+ /Fe 3+ )compound, said “green rust”, on the maintenance of the PRB’s functionality.
Apatite II is a particular configuration of apatite that appears very suitable for the remediation of
groundwater contaminated by metals, including uranium and plutonium [14].
In the case study considered in this paper, apatite II has demonstrated high efficiency in the abatement of the
heavy metals Pb, Cd and Zn. A not negligible benefit in the utilization of apatite II as medium in PRB is the
fact that it has not environmental impacts and there is no phosphate loading to the environment due to its low
solubility, K sp

4. AKNOWLEDGEMENT
Particular thanks to my tutor who has supported me with patience in developing this term paper, giving me precious
hints and councils. Further thanks to my reviewers.
5. REFERENCES
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cluin.org.
[2]- Long-term Performance of Permeable Barriers used for the Remediation of contaminated Groundwater - PEREBAR EVK1-CT-
1999-00035- 5 th Framework Programme Research and Technological Development Project – Literature Review: Reactive
Materials and Attenuation Processes for Permeable Reactive Barriers by National Technical University of Athenes, Department of
Mining And Metallurgical Engineering, laboratory of Mettallurgy, GR-157 80 Zografos, Athens, Greece- August 2000.
www.perebar.bam.de/PereOpen/pdfFiles/Review_Reactive_Materials.pdf
[3] Patricia D.Mackenzie, David P.Horney and Timothy M. Sivavec – Mineral precipitation and porosity losses in granular iron
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Water, Air and Soil Pollution. 98: 57-78.
[6] Conca, J. L., N. Lu, G. Parker, B. Moore, A. Adams, J. Wright and P. Heller. 2000. “PIMS–Remediation of Metal Contaminated
Waters and Soils.” In G.B. Wickramanayake, A.R.Gavaskar, J.T. Gibbs and J.L. Means (Eds.), Remediation of Chlorinated and
RecalcitrantCompounds, vol. 7. p. 319-326. Battelle Memorial Inst., Columbus, OH.
[7] Cornell,R.M. and Schwertmann, U., 1996. The iron oxides. VCH Publishers: New York
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University, Prague, Czech Republic – http://www.cgts.cz/5e_journal_documents/jirasko.pdf
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reactive barrier. Journal of Hazardous Materials 68, 78-101.
[10] Farrell, J., Kason, M., Melitas, N., Li, T., 2000. Investigation of the long-term performance of zero-valent iron for reductive
dechlorination of trichloroethylene. Environmental Science Technology vol.34, n°3, 514-521.
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permeable reactive barriers used for the remediation of contaminated groundwater - Radioactive Waste Management and
Environmental Remediation – ASME 2001.
[12] Franz – George Simon, Tamás Meggyes – Removal of organic and inorganic pollutants from groundwater using permeable
reactive barriers – published in Land Contamination & Reclamation, Vol. 8, Issue 2, 103-116 (2000).
[13] Gianluca Ambrosini – Reactive Materials for Subsurface Remediation through Permeable Reactive Barriers. Diss. ETH
No.15784 (2004).
[14] James L. Conca, Judith Wright – An Apatite II permeable reactive barrier to remediate groundwater containing Zn, Pb and Cd
– Applied Geochemistry 21 (2006) 2188-2200
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laurea specialistica in ingegneria per l’ambiente e il territorio, corso di bonifica dei terreni contaminati A.A. 2006-2007,
professore: Raga Roberto – Facoltà di ingegneria, università degli studi di Padova (on – line)- www.image.unipd.it.
[16] Wright, J., Rice, K. R., B. Murphy, and J. Conca (2004) “PIMS Using Apatite II: How It Works To
Remediate Soil and Water,” in Sustainable Range Management-2004. Proceedings of the Conference on
Sustainable Range Management, January 5-8, 2004, New Orleans, www.battelle.org/bookstore, ISBN 1-
57477-144-2, B4-05.
[17] Koeppenkastrop, D. and E. J. De Carlo. 1990. “Sorption of rare earth elements from seawater onto synthetic mineral phases.”
Chem. Geol. 95: 251-263.
xv

[18] Lower, S. K., P. A. Maurice, S. J. Traina and E. H. Carlson. 1998. “Aqueous lead sorption by hydroxylapatite: Applications of
atomic force microscopy to dissolution, nucleation and growth studies.” American Mineralogist. 83: 147-158.
[19] Lower, S. K., P. A. Maurice and S. J. Traina. 1998b. Simultaneous dissolution of hydroxylapatite and precipitation of
hydroxypyromorphite: Direct evidence of homogeneous nucleation. Geochim Cosmochim. Acta. 62: 1773-1780.
[20] Ma, Q. Y., T. J. Logan and S. J. Traina. 1995. “Lead immobilization from aqueous solutions and contaminated soils using
phosphate rocks.” Environ. Sci. Technol. 29: 1118-1126.
[21] Manecki, M., P. A. Maurice and S. J. Traina 2000. “Kinetics of aqueous Pb reaction with apatites.” Soil Science. 165(12): 920-
933.
[22] Morse, J. W. and W. H. Casey. 1988. “Ostwald processes and mineral paragenesis in sediments.” American Journal of Science.
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[23] Robert M.Powell, Robert W. Puls, David W.Blowes, John L.Vogan, Robert W.Gillham, Patricia D.Powell, Dale Schultz,
Timothy Sivavec, Rich Landis - Permeable reactive barrier technologies for contaminant remediation- Environmental Protection
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Washington DC 20460, EPA/600/R-98/125, September 1998.
[24] Ruby, M. V., A. Davis and A. Nicholson. 1994. “In situ formation of lead phosphates in soils as a method to immobilize lead.”
Environ. Sci. Technol. 28: 646-654.
[25] Tim White, Cristiano Ferris, Jean Kim and Srinivasan Madhavi – Apatite – An Adaptive Framework Structure – School of
Materials Engineering, Nanyang Technological University, Singapore 639798 - 2004.
[26] Wright, J. V., L. M. Peurrung, T. E. Moody, J. L. Conca, X. Chen, P. P. Didzerekis and E. Wyse.1995. In Situ Immobilization of
Heavy metals in Apatite Mineral Formulations, Technical Report to the Strategic Environmental Research and Development
Program, Department of Defense, Pacific Northwest Laboratory, Richland, WA, 154 p.
[27] James Conca, Judith Wright, 2004 - PIMS-Apatite II treatment of Metal-Contaminated Water. Minutes of the Permeable
Reactive Barriers Action Team. www.rtdf.org/PUBLIC/PERMBARR /minutes/102704/pdf/conca.pdf
[28] Werner Stumm, James J. Morgan – Aquatic Chemistry, Chemical Equilibria and Rates in natural waters – Third edition. John
Wiley and Sons Inc. New York USA.
[29] Li, L. and C. Benson, Univ. of Wisconsin-Madison– Reactive transport in the saturated zone: case histories for permeable
reactive barriers - Water Encyclopedia, Vol 2: Groundwater. John Wiley and Sons, New York. ISBN: 0-471-73683-X, 23 pp,
2005
[30] Nichole Ott - Permeable Reactive Barriers for inorganics - 2000 National Network of Environmental Management Studies
Fellow for U.S. Environmental Protection Agency - Office of Solid Waste and Emergency Response - Technology Innovation
Office - Washington, DC - http://www.cluin.org
[31] U.S. Environmental Protection Agency - 40 CFR Part 247 - Comprehensive Guideline for Procurement of Products Containing
Recovered Materials; Proposed Rule - Federal Register / Vol 63. No. 165/ Wednesday. August 26 1998/ Proposed rules.
http://www.epa.gov/fedrgstr/EPA-WASTE/1998/August/Day-26/f22793.pdf
[32] http://www.epa.gov/OCEPAterms/aterms.html
xvi

Response of Coccolithophorids and Diatoms to Climate
Change
Term Paper in Biogechemistry and Pollutant Dynamics
By Fabio Marconi
Tutors: Dr. Jacqueline Flückiger and Prof. Nicolas Gruber
Handed in: June 22. 2007
Abstract
Among other effects, climate changes induced by increased atmospheric CO 2 levels affect the
species composition of marine phytoplankton. Most of the primary production of the phytoplankton
resulted from the activity of two functional groups: coccolithophorids and diatoms. Public concern
has risen about how the expected changes in environmental factors will affect the marine
productivity. Therefore, impact of carbonate chemistry, nutrient and light availability on the
ecophysiology of these two groups, have been widely assessed. This paper will focus on how
ecophysiological differences between these two groups may lead to different responses in front of
the climate change. A review of the literature shown that coccolithophorids may benefit from the
environmental changes (principally the reduction of mixed layer depth) in particular in higher
latitudes and become more competitive at the expense of diatoms, which was dominant at these
latitudes. However, in tropics and mid-latitudes the marine phytoplankton composition will more or
less remain the same and coccolithophorids continue to prevail on diatoms. Because of the
important role of coccolithophorids in the marine carbon cycle, shift in ratio of this calcifying
phytoplankton group versus non-calcareous primary producers seems to affect the partitioning of
CO 2 between ocean and atmosphere having consequently important feedback on future climate
change.
Table of contents
1. INTRODUCTION ........................................................................................................................................................................- 1 -
2. GLOBAL CHANGES DUE TO RISING ATMOSPHERIC PCO 2 ON OCEAN WITH PARTICULAR ATTENTION ON
TROPICS AND SUBPOLAR REGIONS.......................................................................................................................................- 2 -
2.1 DIRECT EFFECT: CHANGES IN CARBONATE CHEMISTRY............................................................................................................... - 2 -
2.2 INDIRECT EFFECT: INCREASED STRATIFICATION ......................................................................................................................... - 3 -
2.2.1 Effects of climate warming in tropics and mid-latitudes compared to higher latitudes ..................................................- 4 -
3. ECOPHYSIOLOGY OF DIFFERENT PHYTOPLANKTON SPECIES: COMPARISON BETWEEN
COCCOLITHOPHORIDS AND DIATOMS.................................................................................................................................- 4 -
3.1 GROWTH PROCESSES................................................................................................................................................................. - 5 -
3.1.1 Light availability and growth..........................................................................................................................................- 5 -
3.1.2 Nutrient supply and growth.............................................................................................................................................- 6 -
3.1.3 Dissolved inorganic carbon (DIC) and growth...............................................................................................................- 8 -
3.1.4 Temperature and growth...............................................................................................................................................- 10 -
3.2 LOSS PROCESSES..................................................................................................................................................................... - 10 -
4. RESPONSE OF COCCOLITHOPHORIDS AND DIATOMS TO PREDICTED GLOBAL CHANGES CONSIDERING
THEIR SPECIFIC ECOPHYSIOLOGY .....................................................................................................................................- 11 -
4.1 TROPICS AND MID-LATITUDES ................................................................................................................................................. - 12 -
4.2 HIGHER LATITUDES................................................................................................................................................................. - 13 -
4.3 ROLE OF GROWTH AND LOSS PROCESSES IN DETERMINING THE RESPONSE OF COCCOLITHOPHORIDS AND DIATOMS TO THE GLOBAL
CLIMATE CHANGES ....................................................................................................................................................................... - 14 -
4.4 EFFECTS ON BIOLOGICAL CARBON PUMP.................................................................................................................................. - 15 -
5. CONCLUSION...........................................................................................................................................................................- 16 -
6. LITERATURE............................................................................................................................................................................- 16 -
7. ANNEX........................................................................................................................................................................................- 18 -

1. Introduction
Since the industrial revolution rising atmospheric pCO 2 induced global changes in the environment.
In the past 20 million years the CO 2 level in the atmosphere seems to have never exceeded 300 parts
per million (ppm), fluctuating between 180 ppm during glacial periods and 280 ppm during
interglacial periods (Rost and Riebesell, 2004). Today the pCO 2 level in the atmosphere reaches
about 370 ppm and this concentration is expected to double until the end of this century. Due to
these large-scale perturbations induced by human activities environmental condition change at an
unprecedented rate. To predict impacts of these global changes on marine ecosystems (chapter 4),
and more precisely on the phytoplankton composition and succession, a first requirement is to
predict changes in abiotic factors in the ocean; like carbonate chemistry, nutrient supply and light
availability, due to rising atmospheric CO 2 level (chapter 2). Secondly, to asses the response of
phytoplankton to environmental changes, the physiological characteristics should be examined
(chapter 3). The marine pelagic system is very complex and represents one of the largest
ecosystems on our planet. Because the whole ecosystem consists of more than 5000 phytoplankton
species it will be impossible to discuss all details and relevant key species have to be identified.
Within all these species only few taxonomic groups are responsible for most of the primary
production of the system (Rost and Riebesell, 2004). One of these so-called phytoplankton
functional groups is represented by siliciers. To this group belong diatoms, one of the most
successful eukaryotic algae in the contemporary ocean, which is responsible for around 40 % of the
net primary production and 50 % of the organic carbon exported to deep sea (Falkowski et al.,
2004). Diatoms dominate microphytoplankton (20-200 µm) and have a unique absolute requirement
for orthosilicic acid used to form strong opal shells called frustules. The second important group
considered in this term paper are calcifying phytoplanktons, which are dominated by
coccolithophorids. These are smaller than diatoms and belong to nanophytoplankton (5-10 µm).
They are characterized by an outer sphere of calcite plates known as coccoliths. Emiliania huxleyi is
the most abundant coccolithophorids and also the most productive lime-secreting organism
(Sarmiento and Gruber, 2006). Coccolithophorids, performing the calcification, can affect
significantly the oceanic carbon cycle. Know their future abundance relative to non-calcifying will
give important information on CO 2 partitioning between atmosphere and ocean and finally on
carbon dioxide level in atmosphere.
Figure 1 Phytoplankton functional
groups considered in this term
paper: silicifiers (diatom, left) and
calcifiers (coccolithophorids, in
this case Emiliania huxleyi, right).
Pictures have not the same
magnitude: the diameter of
diatom is about 60 µm, while E.
huxleyi is 10 times smaller
(pictures are from:
http://www.salem.k12.va.us/staff/j
wright/vocabulary/diatom2.jpg;
http://www.soc.soton.ac.uk/soes/st
aff/tt/eh/pics/cocco9.jpg)
- 1 -

2. Global changes due to rising atmospheric pCO 2 on ocean
with particular attention on tropics and subpolar regions
The effects of rising atmospheric pCO 2 on the surface ocean can be distinguished in direct and
indirect effect. The former occur since the surface layer of the ocean and the atmosphere constantly
exchange CO 2 and consequently change in the atmospheric CO 2 level will affect the surface ocean
carbonate system. The latter are a consequence of the global warming caused by increase
atmospheric CO 2 level, a greenhouse gas; and consist on changes in nutrients and light availability
due to stratification of the upper ocean.
2.1 Direct effect: changes in carbonate chemistry
Carbon dioxide obeys Henry’s law, which means that an increase in the atmospheric pCO 2 will be
reflected in the concentration of CO 2 in the surface oceans. Dissolved CO 2 combines with water
forming carbonic acid (H 2 CO 3 ). This weak acid releases protons (H + ) into solution by dissociating
in bicarbonate (HCO 3 - ) and to a lesser extent in carbonate ions (CO 3 2- ) according to the logarithm of
acid dissociation constants (pK a -value). Also a small fraction of the carbonic acid as well as
dissolved carbon dioxide remains in solution. The result of dissolving carbon dioxide in seawater is
a decrease in pH. The pH of pristine water measures about 8-8.3 (Doney, 2006a). However, the
uptake of anthropogenic CO 2 during the last 200 years (oceans seem to have absorbed half of the
CO 2 produced by human activities) has caused a decrease of pH by 0.1 unit, which corresponds to a
30 % increase in concentration of H + (Raven, 2005). If the global emission of anthropogenic CO 2
continue to rise with the current trends the concentration of protons in the surface ocean will triple
until 2100 and the pH will drop by 0.5 units (Raven, 2005). Changing pH also the speciation of
carbonate species will be shifted: a decrease of pH in the range of pristine seawater will increase the
fraction of dissolved CO 2 , decrease the fraction of carbonate ions (CO 3 2- ), while the large pool of
bicarbonate (HCO 3 - ) remains more or less constant (Figure 2). More quantitative, considering the
expected increase in atmospheric CO 2 until the end of this century, the concentration of CO 3
2-
will
have dropped to the half while the dissolved CO 2 will triple relative to pre-industrial values (Figure
2).
Figure 2 Fraction of the three inorganic species of CO 2 dissolved in seawater versus pH (left, Raven, 2005);
course of seawater pH, dissolved CO 2 and CO 3 2- concentrations assuming a “business as usual” anthropogenic
CO 2 emission scenario until the year 2100 (dashed lines represent the changes in carbonate chemistry if CO 2
emission decreased as prescribed in the Kyoto Protocol) (right, Rost and Riebesell, 2004)
- 2 -

2.2 Indirect effect: increased stratification
The density is the main parameter causing stratification of ocean and is a function of temperature,
salinity and pressure (ρ = f(T, S, p)). In order to evaluate how the density of seawater profile will
change as a result of increasing atmospheric level of CO 2 one has to consider the modellingpredicted
variations in sea temperature and salinity. Knowing the density profile of the ocean,
prediction on the stratification will be possible.
Global warming increases the sea surface temperature (SST) at all latitudes (Figure 3a). The sea
surface salinity (SSS, Figure 3b) response to global warming reflects the overall enhancement of the
hydrologic cycle, due to the increased capacity of warmer air (Manabe, 1991; reviewed in
Sarmiento et al., 2004). The high evaporation in subtropical regions causes an increase of the SSS,
while at higher latitudes the greater rainfall (Sarmiento et al., 2004) and enhanced freshwater input
from melting ice cover (Figure 3c) lower surface layer salinity. Assuming a rule of thumb, which
establishes an increase in seawater density by one unit when temperature increases by 5 °C or when
salinity increase by one unit (Sarmiento and Gruber, 2006), it can be assess that the changes in
density (Figure 3) seem to be affected mainly by temperature.
Figure 3 Response of the physical ocean-atmosphere system to climate change (warming simulation minus
control) calculated with six different coupled climate models (called AOGCMs: Atmosphere-Ocean General
Circulation Models). The figure shows averaged over the period 2040 to 2060 by different latitudes. a) sea
surface tempature (SST) in °C, b) sea surface salinity (SSS) in practical salinity units (PSU), c) ice cover as the
ocean area of maximum wintertime sea ice extent in 10 12 m 2 per degree, d) sea surface potential density (SSD), e)
wintertime maximum mixed layer depth in m. For a extensive description of the parameters considered by each
coupled climate model consult Sarmiento et al. (2004).
Figure 4 Effects of increased
atmospheric pCO 2 on the upper layer
of the oceans. Increased stratification
reduced mixing layer depth, which
causes a reduction of nutrients input
from deep ocean and an increased
light availability (Rost and Riebesell,
2004).
- 3 -

The increased vertical stability of the ocean profile reduces the mixed layer depth. Assuming that
most of the nutrients (particularly macronutrients) are provided by cold, nutrient-rich, deep ocean, a
shoaling of the upper mixed layer will reduce the nutrients supply (Sarmiento et al., 2004; Rost and
Riebesell, 2004; Figure 4). Also the light availability will increased in response to changes in mixed
layer thickness and in cloudiness (Sarmiento et al., 2004; Rost and Riebesell, 2004; Figure 4).
2.2.1 Effects of climate warming in tropics and mid-latitudes compared to
higher latitudes
The tropics and the mid-latitude regions are
already characterized by water columns, which
are thermal-stratified and consequently are
nutrient-limited because of reduced vertical
mixed layer depth. Climate warming will
further increase the stratification inhibiting
vertical mixing and therefore reducing the
nutrient supply coming from deeper nutrient-
rich layers (Figure 5 a). At higher latitudes the
mixing layer depth is often deeper than the
euphotic zone, so that phytoplankton is carried
to layers, where the light-limited conditions
occur. The predicted global warming will
increase the stratification due to rising sea
surface temperature but also due to increased
Figure 5 Effects of increased stratification due to precipitation and melting of ice, which reduced
climate change in the tropics and mid-latitudes (a) the sea surface salinity. The shoaling mixed
compared with them occurred at higher latitudes (b)
layer depth at high latitudes may increase
(Doney, 2006b).
productivity of phytoplankton, because of
reduced light limitation condition (Figure 5 b).
At lower latitudes, on the other hand, phytoplankton productivity will decline substantially, as a
result of increased nutrient limitation. (Doney, 2006b)
3. Ecophysiology of different phytoplankton species:
comparison between coccolithophorids and diatoms
In order to asses the response of two plankton functional groups – calcifiers and silicifiers – with
regard to the predicted global environmental changes, it is crucial to know the specific
physiological characteristics of each plankton group. Local environmental factors select specific
organisms affecting the competitive advantage of determinate groups or species of phytoplankton.
With detailed information about the physiological and ecological characteristics of the individual
phytoplankton species and predictions of the future environmental changes it is possible to
speculate about changes in the distribution and succession of phytoplankton species.
The ecological success of a phytoplankton species will be mainly determined by its ability to
optimise the balance between growth and loss processes. Availability and optimal use of essential
resources like light and nutrient (carbon, nitrogen, phosphorous and plus micronutrient such as iron
and zinc) are the major factors controlling growth processes while loss processes can be caused by
cell sinking, cell mortality due to grazing, viral and parasite infection as well as autolysis (Rost and
- 4 -

Riebesell, 2004; Sarmiento and Gruber, 2006). This chapter is principally subdivided into these two
processes.
Before starting the description of the physiological issues of coccolithophorids (calcifiers) and
diatoms (silicifiers) it is important to emphasize that most of the data on phytoplankton presented in
this term paper are based on laboratory experiments, which are normally conducted on individual
species easy to culture. One has to consider that data of singular species are not necessarily
characteristic for a whole phytoplankton functional group (Le Quéré et al., 2005). Another
important point is that experiments may not be really representative for the behaviour of a species in
a complex community in the ocean consisting of many species (Sarmiento and Gruber, 2006). More
in detail, Sarthou et al. (2005) reviewed many experiments on different diatoms species so that
average data of this group of silicifiers are usually available. On the other hand for
coccolithophorids lots of experiments are done on Emiliania huxleyi, but for other
coccolithophorids the physiological characteristics are poorly described. As a result many
coccolithophorids issues described in this paper are based principally on E. huxleyi, which is,
however, the most important bloom-forming coccolithophorids in present geological period
(Iglesias-Rodriguez et al., 2002).
3.1 Growth processes
3.1.1 Light availability and growth
Vertical mixing of the upper layer of ocean modifies the light intensity. Phytoplankton species have
consequently experienced high variability in light conditions. Differences in photosynthetic
characteristics between phytoplankton species cause specific affinity to light or ability to deal with
light irradiance variability. In order to evaluate the photosynthesis physiology and, more specific,
the efficiency of light utilisation, photosynthetic rate versus irradiance curves (PI-curves) are widely
used. PI-curves represent the response of photosynthesis at various light intensities. Typically these
curves can be subdivided in three regions: light-limited, light-saturated and photoinhibition (Figure
6). For light limitation, the photosynthetic rate increases asymptotically by increasing the light
intensity from zero at a threshold light level, by which the photosynthesis is light-saturated and it
reaches the maximal rate. At higher light irradiance than light-saturated condition photoinhibition
occurs and the photosynthesis rate decrease by increasing light intensity (Figure 6). The slope of PIcurves
in the light-limited condition representing the fraction between the maximal photosynthesis
rate (V max ) and the light-saturation parameter (I k ) is called the affinity for light (Figure 6): a higher
slope will mean a higher affinity (Le Quéré et al., 2005).
Figure 6 Typical response
curve of phytoplankton by
increasing irradiance. The
initial slope corresponds to
the light affinity. At V max
the photosynthesis is
saturated by light. By high
light
irradiance
photoinhibition occurs
causing a decrease of the
photosynthetic rate
(Sarmiento and Gruber,
2006)
- 5 -

Comparing PI-curves of various phytoplankton groups clearly show that they have specific response
of photosynthesis to varying light irradiance. Green algae (flagellates), which include also
coccolithophorids, have the highest affinity to light and become light-saturated at lower light
irradiance (Figure 7) than diatoms and dinoflagellates. Diatoms are less efficient than green algae,
but more than dinoflagellates (Figure 7).
Figure 7 Photosynthesis
response at different light
intensities (P-I curves) for
different phytoplankton
group (note that
coccolithophores are
included in the group of
flagellates) (Sarmiento
and Gruber, 2006)
Another important characteristic of E. huxleyi is that this species of coccolithophorid does not show
any sign of photoinhibition even at irradiances of 1700-2500 μmol -1 m
-2 s -1 , which are equal or
higher than light intensity by full sunshine (Paasche, 2002). The ability to survive in the dark is also
a parameter which changes in different phytoplankton groups. Laboratory experiments have shown
that silicifiers can survive for weeks without light supply, while calcifiers begin to die already after
one day (unpublished data from R. Geider; reviewed in Le Quéré, 2005). This difference should be
considered especially during winter, when the mixing depth is deeper than the euphotic zone.
3.1.2 Nutrient supply and growth
The response of the phytoplankton species to nutrient supply is another factor determining their
competitive ability. Analogous to the determination of light affinity using light saturation parameter
and maximal rate of photosynthesis, by the nutrient supply the affinity to a specific nutrient is
quantified using the half-saturation constants (Ks). These are derived from Michaelis-Menten (Eq.
1) and Monod (Eq. 2) saturation functions, which determine specific uptake rate (Vmax) and the
specific growth rate (μ), respectively.
[ Nut]
[ Nut] Ks
V
V Nut
= max Eq. 1
+
μ =
μ max[ Nut]
[ Nut] + Kμ
Eq. 2
- 6 -

Figure 8 Rate of nutrient uptake
(example of nitrate) by increasing
nutrient concentration. K N is the
half-saturation constant of nitrate
and is determined by the
concentration of nitrate at which the
uptake reaches the half of maximal
rate V max (Sarmiento and Gruber,
2006).
Half-saturation constants are specific for a given nutrient and represent the concentration Ks of the
nutrient that limits V Nut to the half of Vmax (Eq. 1) and the concentration Kμ of the nutrient that
limits μ to the half of μmax (Eq. 2) (Sarthou et al., 2005; Figure 8). Species with a high halfsaturation
for a determined nutrient need a higher concentration of a specific nutrient to grow.
Coccolithophorids show an exceptional inorganic phosphorous (P) assimilation capability (Riegman
et al., 2000;
Table 1). One reason of this high affinity to P could be that coccolithophorids have many receptors
for P (Le Quéré et al., 2005) and the availability of two types of alkaline phosphatase bound to the
cell surface, which enables the coccolithophorids to use dissolved organic P (DOP; organic
phosphate esters) at nM concentration level (Paasche, 2002; Rost and Riebesell, 2004). On the other
hand diatoms have a higher half-saturation constant for P (
Table 1), which means a higher requirement of P in comparison to coccolithophorids to grow.
Table 1 Half-saturation constants of essential macro- (P and N) and micro- (Fe) nutrients for coccolithophorids
and diatoms. Also represented are taxum-specific nutrients like orthosilicic acid for diatoms and dissolved
organic phosphorous (DOP) for coccolithophorids ( ± Buitenhuis et al. (2002) ‡ Riegman et al. (2000): value found
for E. huxleyi; ; † Le Quéré et al. (2005); Iglesias-Rodríguez et al. (2002) ).
Phytoplankton
group
Ks (P)
[nM] ±
Ks (N)
[μM]
Kμ (Fe)
[pM] ±
Ks (Si)
[μM] ±
Other nutritional
source
Coccolithophorids 4 0.2 ‡ 20 - 1.9 DOP †
Diatoms 75
< 0.2 ‡
> 0.2 120 2
Considering N affinity some contradictions between scientists can be found. Riegman et al. (2000) describe E.
huxleyi as not to a good competitor for nitrogen in contrast to phosphorous. Their experiments have shown a
rather low maximal uptake rate of nitrate for E. huxleyi in comparison to most other algae species (also diatoms)
and also the affinity for nitrate uptake was not extremely high (Riegman et al., 2000). Therefore the growth at N-
limited condition of E. huxleyi is less exceptional than under phosphorous limitation. On the other hand Iglesias-
Rodríguez et al. (2002) suggest that areas with sea surface presenting decreasing nitrate concentrations are
selective for coccolithophorid blooms. Additionally they consider coccolithophorids as having an exceptionally
high affinity for dissolved inorganic nitrogen, with half-saturation constants being approximately half that of
diatoms of comparable size (
Table 1).
- 7 -

Iron and z inc are essential micronutrients for phytoplankton taxa. As well for these elements coccolithophorid E.
huxleyi shows lower requirements and therefore a higher affinity in comparison to diatoms (Rost and Riebesell,
2004;
Table 1).
The cell size of phytoplankton seems to play an importa nt role in the uptake and affinity for
nutrients. Small planktons have a higher surface to volume ratio (S/V) than larger cells, which
implies a better exchange of nutrients across the cell surface. The limitation in diffusion of nutrients
uptake is consequenty lower for small plankton species than for larger ones (Sarthou et al., 2005).
This theory could therefore explain the difference between diatoms and coccolithophorids in
nutrient uptake and affinity: the cell size of coccolithophorids is in the range of 5-10 μm, while that
of diatoms is between 20 and 200 μm (Le Quéré et al., 2005). Experiments have shown a significant
relationship between size and growth within a plankton functional group (Sarthou et al., 2005),
however considerations across different plankton functional groups do not necessarily be correct
(Le Quéré et al., 2005). Actually different plankton species have developed specific adaptive
strategies to overcome deficits in the nutrient uptake. For example diatoms can reduce their size
(increasing their S/V ratio) in order to perform a better nutrient uptake (Sarthou et al., 2005).
Another explanation of the difference in the ecophysiology between diatoms and coccolithophorids
is based on the cellular design (Iglesias-Rodríguez et al., 2002). In diatoms a relative large fraction
(approximately 35 %) of the cell volume is dedicated to storage vacuoles (Falkowski et al., 2004),
which allow excess “luxury” uptake of macronutrients and their storage in the internal reservoirs.
This mechanism allows to perform several cell divisions without the need of external input of
nutrients (Iglesias-Rodríguez et al., 2002). In contrast the accumulation of macronutrients is lower
for the smaller coccolithophorids than for diatoms (Rost and Riebesell, 2002; Iglesias-Rodríguez et
al., 2002). Consequently growth rate in diatoms may be better buffered due to internal nutrients
storage under changing nutrients supply (for example under unstable environment with pulsed
nutrient supply), while by coccolithophorids the growth rate is directly coupled with the nutrient
concentration in water. Because of the high nutrient uptake capacities, diatoms prevail under high
nutrient conditions (mainly in spring and early summer) (Iglesias-Rodríguez et al., 2002; Sarthou et
al., 2005). But under nutrient limiting conditions coccolithophorids out compete diatoms, as a result
of their higher affinity to nutrients (Falkowski et al. 2004; Iglesias-Rodríguez et al., 2002; Paasche,
2002; Riegman et al., 2000).
3.1.3 Dissolved inorganic carbon (DIC) and growth
In order to evaluate the potential response of coccolithophorids and diatoms to global changes and
the related change in dissolved inorganic carbon supply it is also fundamental to consider the
mechanism of carbon acquisition of both plankton groups and emphasize the differences between
them. In comparison with other phytoplankton nutrients in seawater dissolved inorganic carbon is
always in excess and for this reason in the last decades the role of inorganic carbon acquisition has
been partially considered in phytoplankton ecology and evolution (Rost et al., 2003). However, the
CO 2 fixation by the carboxylating enzyme RuBisCO (Ribulose-1,5-bisphosphate
carboxylase/oxygenase) is restricted due to his low substrate affinity to CO 2 (half- saturation
constant K M of 20-70 μmol L -1 ) (Rost and Riebesell, 2004). Because of the low seawater CO 2
concentration ranging from 5 and 25 μmol L -1 , photosynthesis of marine phytoplankton may be
limited by CO 2 (Rost et al., 2003). To overcome the limitation caused by the low affinity of
carboxylating enzyme to CO 2 , most microalgae have developed mechanisms which increased the
concentration of CO 2 at the site of carboxylation. These CO 2 concentrating mechanisms (CCMs)
allow the active uptake of CO2 and/or HCO - 3 into the algal cell and/or in the chloroplast (Riebesell,
2004). Another enzyme called carbonic anhydrase (CA) is coupled to CCMs and accelerates the
rate of conversion between HCO - 3 and CO 2 (Rost et al., 2003). As a result of the CCMs and CA
- 8 -

most marine phytoplanktons reach a high affinity for inorganic carbon and consequently
photosynthetic carbon saturation under ambient CO 2 levels (Raven and Johnston, 1991; reviewed in
Rost and Riebesell, 2004).
Recent experiments considering various groups of phytoplankton indicate differences in CO 2
senstivity. Rost et al. (2003) compares three marine bloom-forming microalgae: the diatom
Skeletonema costatum, the coccolithophorids Emiliania huxleyi and the flagellate Phaeocystis
globosa. The photosynthetic carbon fixation of diatom and flagellate considered in this experiment
is close to CO 2 -saturation considering typical marine CO 2 concentration at present days of 5-25
μmol L -1 (Figure 9). While coccolithophorid E. huxleyi has comparatively lower affinity and seems
to be carbon limited at present CO 2 level (Figure 9).
The low affinity of E. huxleyi for
inorganic carbon has led to the
hypothesis that this species of
coccolithophorids may not rely on
an active carbon uptake, but only on
diffusive CO 2 supply for
photosynthesis (Raven and
Johnston, 1991; reviewed in Rost
and Riebesell, 2004). Though recent
studies have confirmed the presence
of CCMs in E. huxleyi, but its
efficiency appear to be lower than
in other phytoplankton groups (Rost
et al., 2003).
Although intensive research on
coccolithophorids the function of
calcification (and consequently of
coccoliths) is not well understood
(Paasche, 2002). It has been
Figure 9 Relative photosynthesis response at different CO 2 hypothesized that coccoliths protect
concentration of the diatom Skeletonema costatum, the the cell from strong light or they are
coccolithophorids Emiliania huxleyi and the flagellate Phaeocystis
used to redirect light into the cell
globosa. K 1/2 (CO 2 ) represents the half-saturation constant for
CO 2 in μmol L -1 (Riebesell, 2004).
(useful in light-limited condition)
(Young, 1994; reviewed in Paasche,
2002). But the removal of coccoliths did not confirm these hypothesises: no change of light
inhibition by high irradiance and of light-saturation parameter (I k ; see Figure 6) has been found
(Nanninga and Tyrrell, 1996). Experiments with 14 C-labeled DIC indicate that E. huxleyi makes the
coccolith from bicarbonate (HCO - -
3 ) rather than from carbonate or CO 2 and that both HCO 3 and
CO 2 in the medium are used for photosynthesis (Paasche, 2002). These experiments seem to prove
the close coupling between photosynthesis and calcification and to support a “trashcan function” of
calcification, in which calcite (CaCO 3 ) precipitation serves as mechanism to facilitate the use of
HCO - 3 in photosynthesis (Riebesell, 2004). According to the following reactions (Eq. 3 and Eq. 4)
2+
−
Ca + 2 HCO3
→ CaCO3
+ CO2
+ H
2O
2+
−
+
Ca + HCO3
→ CaCO3
+ H
Eq. 3
Eq. 4
the process of calcification is beneficial to photosynthesis since it releases CO 2 or protons (H + ),
-
which could be used directly in photosynthesis or in the conversion of HCO 3 to CO 2 , respectively
(Rost and Riebesell, 2004; Paasche, 2002). However, non-calcifying cells of E. huxleyi seem also to
- 9 -

e able to direct uptake of HCO 3
-
and this could mean that the use of bicarbonate is not tied to
calcification (Rost and Riebesell, 2004). Experiments conducted by Riebesell et al. (2000) indicate
that the rate of photosynthesis decreases with decreasing CO 2 concentration although a concomitant
increase in the calcification rate. In conclusion the beneficial of calcification for photosynthesis is
not yet clear, but what can be said is if the primary role calcification is the CO 2 supply for
photosynthesis, this mechanism seems to be rather inefficient in comparison with CCMs of noncalcifying
phytoplankton like diatoms (Rost et al. 2003).
3.1.4 Temperature and growth
Under ideal nutrient supply and light availability the maximal growth of phytoplankton is given by
the temperature. However the growth rates found in laboratory experiment for different taxa seem
not to change so much, and therefore a generic growth rate of 0.6 d -1 at 0 °C is widely used for
phytoplankton (Banse, 1992; reviewed in Le Quéré, 2005). For example also in The Dynamic
Green Ocean Model for all six plankton functional groups (picophytoplankton, nanophytoplankton,
coccolithophorids, diatoms, phaeocystis and N 2 -fixers) is based on a maximum specific growth rate
of 0.6 d -1 at 0 °C is used (Buitenhuis et al., 2002). For this reason temperature seems to be a factor,
which does not differ significantly in coccolithophorids and diatoms.
3.2 Loss processes
Grazing by zooplankton is - like nutrient availability and light irradiance - another factor which can
control or limit the growth of phytoplankton functional group. The plates of calcium carbonate
called coccoliths covering the cell of coccolithophorids do not seem to provide a protection against
various types of microzooplankton (Harris, 1994; reviewed in Rost and Riebesell, 2004). Whereas
the frustules by diatoms seem to hold a more important role as mechanical protection against
grazers (Hamm et al., 2003; reviewed in Riebesell and Rost, 2004). Also the smaller size of
coccolithophorids makes them more susceptible to microzooplankton grazing in comparison with
diatoms.
Cell lysis is another important process influencing the plankton dynamic. Two types of cell lysis are
recognized: the viral lysis and death due to nutrient stress. Against the viral lysis diatoms appear to
be quite well protected. Autolysis rates determined through experiments are relatively low and
therefore can significantly reduce net growth rate particularly under nutrient limitation when growth
rates are low (Sarthou et al., 2005). While for coccolithophorids limited research has been
performed on the cell lysis.
Sinking rate of plankton is an additional loss processes. In low nutrient environment movement
through the water could be essential for the survival. For phytoplankton species without motility
apparatus a mechanism to compensate this lack is to regulate the sinking rate by changing their
density or forming aggregate (Paasche, 2002; Sarthou et al., 2005). An overproduction of calcite
coccoliths by coccolithophorids or a thickening of siliceous frustules increase the cell density in
relation to water density and consequently increase the sinking rate. Because sinking rate is also a
function of the size, the aggregation of cells has the purpose of increase the sinking rate too. For the
reason that both mentioned phytoplankton species allow to similar properties, the sinking rate as
loss process will to a less degree considered.
- 10 -

Table 2 Summarizing and comparative table of physiological characteristics of coccolithophorids (C) and
diatoms (D). For a more detailed description of each trait see the relative chapter about ecophysiology. (
Divergence between Riegman et al. (2000) and Iglesias-Rodríguez et al. (2002)).
ocesses
Growth pr
Loss
processes
Coccolithoporids (C)
vs Diatoms (D)
C
C < D
Affinity (based on halfsaturation
P
C > D
constant) N C > < D
Nutrient
supply
Fe, Zn
C > D
Cellular design: storage Nutriment
vacuoles
uptake capacity
C < D
Diffusion limitation S/V ratio C < D
Light saturation parameter (a lower value
Light means a higher “affinity” for light)
C < D
Photoinhibition
C < D
Temperature Growth rate dependence to temperature C = D
Sinking rate
C ≈ D
Grazing by zooplankton (depends on mechanical protection
given by coccoliths (C) and frustules (D))
C > D
4. Response of coccolithophorids and diatoms to predicted
global changes considering their specific ecophysiology
Changes in seawater carbonate chemistry, affected directly by increasing atmospheric pCO 2 , have
consequences on the availability of carbon for photosynthesis and calcification for
coccolithophorids. At the present CO 2 levels, the rate of photosynthetic carbon fixation by E.
huxleyi seems to be below saturation, in contrast to diatom species (e.g. Skeletonema costatum)
which are more efficient in carbon uptake (Figure 9). Therefore coccolithophorids may benefit from
the predicted increase in the dissolved CO 2 concentration in seawater fulfilling their higher carbon
requirements. This implies that an increase in CO 2 availability may improve the resource utilisation
of E. huxleyi and other coccolithophorids with similar physiological characteristics, causing an
increase in their contribution to overall primary production. On the other hand, the continued
acidification of the upper ocean due to increasing CO 2 leads to deteriorated conditions for biogenic
calcification. Riebesell et al. (2000) observed a reduced calcite production at increased CO 2 levels
in two bloom-forming coccolithophorids, E. huxleyi and Gephyrocapsa oceanica. Because of higher
dissolved CO 2 concentration in surface seawater, these species of coccolithophorids have no to
compensate relatively inefficient CO 2 concentrating mechanisms by performing calcification (Eq.
3), which release carbon dioxide for photosynthesis. This theory could explain the observed
coccoliths under-calcification and malformation in both natural environments and under laboratory
conditions. Although the role of CaCO 3 formation in coccolithophorids is not clear, reduced
calcification seems to affect both resource utilization and cellular protection with possible adverse
effects on the competitive fitness of coccolithophorids.
Although under natural conditions CO 2 limitation seems to be less important compared to other
limiting resources for E. huxleyi (Riebesell, 2004), which should be more CO 2 -sensitive than
- 11 -

diatoms because of higher requirements. Consequently, additional possible limiting factors have to
be considered to asses the potential changing fitness of coccolithophorids and diatoms due to
climate change. As explained before, raising atmospheric pCO 2 has not only direct effects on
carbonate chemistry of surface seawater, but also indirect effects that are correlated with changing
climate due to increased carbon dioxide levels. A first effect is the enhanced stratification of surface
ocean equivalent to an increase in vertical stability of the water profile, thus decreasing nutrient
supply. In the previous chapter half-saturation constants for major macronutrients (P and N) as well
as micronutrients like iron are illustrated. Considering P or Fe half-saturation the difference in the
affin ity of coccolithophorids and diatoms are significant. For
nitrogen some contradictions between
scientists can be found. Riegman et al. (2000) referred to the coccolithophorid E. huxleyi as a poor
competitor with regard to nitrogen. Whereas Iglesias-Rodríguez et al. ( 2002) indicate areas with sea
surface presenting decreasing nitrate concentrations as selective for coccolithophorid blooms and
consider coccolithophorids as having an exce ptionally high affinity for dissolved inorganic
nitro gen, with half-saturatio n constants being approximately half that of diatoms of comparable
size.
To simplify the assessment of the response of the considered phytoplankton groups to changes in
nutrients supply, no different iation between different nutrients will be made. A value for the affinity
to nutrients of coccolithophorids and diatoms as a whole should be taken into account to determine
their response to global changes. Assuming that P is the limiting factor in seawater of many regions,
the half-saturation constant for P is taken as an approximation for the whole nutrient supply. This
w ill mean that an enhanced stratification occurring on a global scale and the resulting reduced
nutrient supply mainly increases the competitive ability of coccolithophorids, which can take up
nutrients at very low concentrations. This assumption is also supported if micronutrients like Fe or
Zn are considered. For this class of nutrients, Coccolithophorids also display lower half-saturation
constants, which indicates better affinity. Finally, considering the more recent paper of Iglesias-
Rodríguez et al. (2002), evidence suggests that coccolithophorids are also more competitive at low
nutrients supply in regard to N.
Enhanced stratification due to global climate change and subsequently shoaling of the mixed layer
depth seems to occur globally in oceans. However, the strength and the impact of this process
change by considering different latitudes, as a result of already different conditions characterizing
these areas today. For example mid-latitudes or tropics are nutrient-limited because of a highly
thermal-stratified surface seawater profile. Whereas at higher latitudes the mixed layer depth is
often deeper than the eutrophic zone and phytoplankton species are transported down to layers
where the light does not penetrate, hence phytoplankton is mainly light-limited at higher latitudes at
present. A shoaling mixed layer depth and the consequently reduced nutrient supply at these two
latitudes will therefore have a different impact on the present conditions. These changed
environmental conditions could affect the composition, distribution and succession of
phytoplankton species.
4.1 Tropics and mid-latitudes
The predicted enhanced stratification in the tropics and mid-latitudes decreases the already low
nutrient supply and consequently, surface seawater will be even more nutrient-limited. The
planktonic composition of these areas should therefore not change so much in the next decades.
Low mixed layer depths in these regions allow the plankton species to colonize the shallow
superficial layer, which is characterized by high solar irradiance and low nutrient supply. Both
parameters are selective for coccolithophorids, species showing a higher affinity to nutrients and no
or little photoinhibition. The mixed layer depth and thus the nutrient supply vary little at these
latitudes, because of relatively constant temperatures during the year. Even in winter, when the
mixed layer depth reaches its maximum, low nutrient supply and high light intensities are not
favourable conditions for diatoms. The fraction of the sediments dominated by CaCO 3 at these
- 12 -

latitudes could confirm the planktonic composition prevailed by coccolithophorids (see annex,
Figure A 1).
Because of increased nutrient limitation, marine productivity in the tropics and mid-latitudes seems
to decrease substantially. Bopp et al. (2001) investigated the response of the productivity of
phytoplankton to global warming using coupled atmosphere-ocean general circulation models and
ocean biogeochemical schemes. At doubled atmospheric pCO 2 (700 ppm) Bopp et al. indicate a
decline in productivity in the tropics reaching -15 to -20 %.
4.2 Higher latitudes
At higher latitudes phytoplankton are not nutrient limited but light limited. This is due to the water
columns being less stabilized by thermal stratification and, consequently this low density gradient in
surface seawater causes the mixed layer depth to be large. This factor does not allow an efficient
light use, since phytoplanktons are carried to deeper depths with prohibitive light condition for
photosynthesis. The maximal mixed layer depth occurs in winter when the surface water profile is
minimally stabilized. Warming of the surface water during the year increases his stratification and
so decreases the mixed layer depth. The nutrient supply course during the year follows mainly the
mixed layer depth. Assuming that nutrients come from the deep ocean, a higher mixed layer depth
in winter leads to higher nutrient supply and higher concentrations in the surface layers. Analogous,
shoaling mixed layer depth during spring and summer reduces the nutrient supply. These eutrophic
conditions particularly at the beginning of the spring are favourable and selective for diatoms. This
plankton species does not have a high nutrient affinity, but the evolution of nutrient storage
vacuoles allows the cells to retain high concentrations of nutrients. Comparison experiments with
diatoms also indicate a good ability to survive in the dark for several weeks (unpublished data from
R. Geider; reviewed in Le Quéré, 2005). These two features allow competitive advantage of
diatoms compared to coccolithophorids, which are better adapted to oligotrophic condition and
show a poor ability to survive in the dark (die after one day without light; unpublished data from R.
Geider; reviewed in Le Quéré, 2005).
The predicted increasing stratification in higher latitudes due to global warming will mean a better
photosynthetic efficiency during spring and summer. Because of a shallow mixed layer depth
phytoplanktons experience lower light limited conditions and this implies better as well as longer
light utilization in the euphotic zone. The shallowing of the mixed layer occurs earlier in the year
and the deepening occurs later, resulting in an expansion of the growing season length. At present
and under global warming conditions diatoms bloom earlier in the growing season than
coccolithophorids, because of the high survival ability in the dark, which allows them to maintain
relatively high biomass over the winter months. During spring and summer the mixed layer depth is
expected to shallow more than at present. As a result of high nutrient sequestration by diatoms at
the beginning of the season and of reduced mixed layer depths, which also limit the nutrient supply;
surface seawater becomes progressively more nutrient limited. Diatoms seem to partially adapt
themselves to these conditions reducing their size, in order to increase the surface to volume ratio
and consequently meliorate their nutrient uptake in oligotrophic conditions. However the lower
mixed layer depth in late spring or early summer and the subsequently low nutrient availability
could be favourable conditions for coccolithophorids. Due to their higher affinity to nutrients these
calcifiers may out compete other phytoplankton species. In this way, under climate warming, the
productivity at higher latitudes could increase in comparison to today. This poleward shift of
phytoplankton production is also observed in the simulation performed by Bopp et al. (2001), who
estimates an increase in productivity in subpolar regions (south of 50°S and north of 60°N) by more
than 10 %.
At higher latitudes the fraction of opal, which can be used as an indicator for diatoms, dominates
the sediment content. Although, due to melioration of abiotic factors in regard to coccolithophorids
physiological characteristics and the migration of this phytoplankton group to higher latitudes, the
opal to calcite ratio in the sediment may decrease during the next decades (see annex, Figure A 2).
- 13 -

4.3 Role of growth and loss processes in determining the response of
coccolithophorids and diatoms to the global climate changes
In the determination of the competitive ability and selective advantages of coccolithophorids and
diatoms considering changing abiotic factors, not all growth and loss processes seem to have the
same relevance. Generally one can say that the larger the difference for a physiological trait
between coccolithophorids and diatoms, the more relevant is the considered characteristics to assess
their response to global changes. Considering the affinity to nutrients of these two groups a
significant difference in the values of half-saturation constants principally for P and Fe can be
observed (
Table 1). On the other hand coccolithophorids and diatoms do not show large difference in their
“affinity” to light. For example, in The Dynamic Green Ocean Model the “affinity” for light of
coccolithophorids in relationship with that of diatoms slightly differs and the ratio is 3:2. In
comparison, the ratio for affinity of P and Fe is 19:1 and 6:1, respectively. So the different affinities
for nutrients exhibited by these two phytoplankton groups seem to be more relevant. However,
specific characteristics like survival ability in the dark or photoinhibition as well as availability of
storage vacuoles in the cell, which increase the nutrient uptake capacities without changing their
affinity to nutrients, should not be neglected, since they complete/influence significantly the affinity
value for nutrients and light.
In order to assess a shift in composition and succession of coccolithophorids and diatoms due to
global changes, composition of the nutrient supply is not taken into account in this term paper and
for nutrients as a whole the half-saturation of P will be considered in the assessment. However not
all oceans are P-limited. In addition, considering only P, an extreme scenario will be found, due to
extremely high P affinity by coccolithophorids compared to diatoms. The predicted changes in
composition and succession in this term paper will not change significantly considering also iron.
But, if diatoms have a better affinity to N as reported by Riegman (2000) and for these oceans
which are N-limited, an enhanced stratification and consequent reduction of nutrient supply (also N)
will change considerably the future predictions. Another factor which should be taken into account
to obtain a more realistic scenario, is the supply of orthosilicic acid. Diatoms are unique among
photoautotrophic taxa, which have an absolute requirement for orthosilicic acid to form shells called
frustules (Falkowski et al., 2004). Silica is introduced into oceans by continental weathering but
also from deep ocean. How an enhanced stratification affects the silica availability is still not well
known.
The loss processes are also determinant in assessing the ecological success of a phytoplankton
species, which is principally determined by the ability of a species to optimise the balance between
growth and loss processes. For many of these processes little information is known and further
research will be useful for better assessment of competitive ability. In the considerations above
about the response of coccolithophorids and diatoms at different latitudes to global changes loss
processes are not taken in account.
Mortality rate is a function of temperature. So the predicted climate warming may increase
mortality rate of both coccolithophorids and diatoms. Another loss process is sinking rate. Diatoms
are generally larger than coccolithophorids and in addition they form agglomerates. Both reasons
lead to the conclusion that diatoms may have a higher sinking rate. However coccolithophorids, and
likely diatoms, can raise their density by an overproduction of coccoliths and increase the sinking
rate. For this process it is difficult to weigh its specific importance in both phytoplankton groups.
Against autolysis and viral lysis diatoms seem to be quite well protected, but for coccolithophorids
limited research has been performed on this topic. Among all loss processes grazer control by
higher trophic level may mostly influence the composition and seasonal succession of
phytoplankton groups. While frustules in diatoms seem to serve as an effective mechanical
protection against various predators, coccoliths of coccolithophorids may not provide the same
degree of protection. Zooplankton has the potential to propagate at similar rates as its prey and
therefore reduces or impedes phytoplankton mass development. Also in The Dynamic Green Ocean
- 14 -

Model zooplankton grazing is a point which is being worked on, because of the importance of this
process in structuring the marine food web. Also other scientists like Rost and Riebesell set the
understanding of the complex structure and regulation of marine pelagic ecosystem as a major
challenge for marine sciences in the next decades.
4.4 Effects on biological carbon pump
In the previous chapter effects of changing climatic conditions on two important phytoplankton
groups like coccolithophorids and diatoms have been described. In turn, their activity can have a
direct impact on the climate. They can mitigate or amplify climate change by driving many of the
global elemental cycles. Since CO 2 represents the most important greenhouse gas leading to global
warming, knowing the role of phytoplankton for the oceanic carbon cycle is important. The effects
of coccolithophorids on this cycle differ greatly from other phytoplankton species and seem to have
a relevant feedback on climate. That is the reason why coccolithophorids are principally dealt in this
chapter.
The fixation of inorganic carbon through
photosynthesis in the upper ocean and the
subsequent transport of organic matter
(“ideally” CH 2 O) to depth is termed the
organic carbon pump (Figure 10). This
process is performed by all primary
producers and causes a net draw down of
CO 2 from the atmosphere into the ocean.
Additionally coccolithophorids produce
calcite (CaCO 3 ) coccoliths through the
reaction of calcification. The production and
export of CaCO 3 to the deep sea binds
dissolved inorganic carbon (DIC) in
particulates and therefore reduces total
dissolved carbon. This process also lowers
alkalinity and changes the equilibrium Figure 10 Biological carbon pumps: organic carbon
pump and carbonate counter pump. The strength of
between different species of DIC, causing a
these two processes determines the rain ratio, an
net release of CO 2 to the atmosphere (Rost important parameter for ocean atmosphere CO 2
and Riebesell, 2004; Figure 10). Because of exchange (Rost and Riebesell, 2004).
the counteracting effect on the CO 2 flux, this
process is called the carbonate counter pump (or also simply carbonate pump). The rain ratio gives
the relative strength of these two biological carbon pumps and corresponds to the ratio of particulate
inorganic carbon (CaCO 3 ) to organic carbon (C org) exported to the deep sea (Figure 10). This ratio
assumes a non-negligible significance since it determines the biologically mediated partitioning of
CO 2 between ocean and atmosphere.
Coccolithophorids, and generally calcifying phytoplankton species, compared to non-calcifying
primary producers are a smaller sink for CO 2 or could even act as a net source (Zondervan et al.,
2001). Consequently, in determining the rain ratio it seems to be crucial to assess the proportion of
calcareous to non-calcareous primary production. So that the marine phytoplankton composition
and succession impact on carbon cycle and so on global climate. If the predicted reduced nutrient
supply and the increase availability of dissolved CO 2 in surface seawater due to rising atmospheric
pCO 2 are advantageous to coccolithophorids, causing an increase in their fraction among primary
producers, calcification and then relative global export of calcite to the deep sea would increase.
Increasing rain ratio, by increased calcification and more efficient carbon counter pump, would
result in a positive feedback to rising atmospheric CO 2 . On the other hand, experiments conducted
by Riebesell et al. (2000) have shown reduced calcification with increasing pCO 2 . This observed
- 15 -

process may neutralise or even reverse the positive feedback described before into a negative one,
leading to an increase in the CO 2 storage capacity of the surface ocean. A quantification of these
two counteracting processes, increased abundance of coccolithophorids and decreased calcification,
will help to predict the ocean atmosphere CO 2 exchange in the future and consequently also the
atmospheric CO 2 level, which is finally a determinant for global climate changes. Further research
is required on these processes and Riebesell (2004) confirms this need by claiming: “despite the
potential importance of global change-induced biogeochemical feedback, our understanding of
these processes is still in its infancy”.
5. Conclusion
In the determination of the response of coccolithophorids and diatoms to future climate change the
nutrient supply may have the most important relevance among all considered parameters. Predicted
enhanced stratification of the upper ocean due to global warming and consequently a shoaling
mixed layer depth will reduce nutrient supply. The impact and the strength of this process will vary
considering different latitudes. In the tropics and mid-latitudes the present conditions may change
little and therefore the planktonic composition at these latitudes will remain dominated by
coccolithophorids. However, at higher latitudes this process seems to lead to favourable conditions
for coccolithophorids particularly by low nutrient availability in late spring or early summer. The
increase in the fraction of calcifying phytoplankton like coccolithophorids among primary
producers raises the rain ratio causing a net release of CO 2 to the atmosphere (positive feedback to
rising atmospheric CO 2 ). However the observed reduced calcification with increasing pCO 2 could
reverse the positive feedback into a negative one, which would mean an increase of CO 2 storage
capacity of the surface ocean. A better quantification of these two counteracting processes could
help to predict CO 2 partitioning between ocean and atmosphere in the future.
6. Literature
BOPP L. AND OTHERS (2001): Potential impact
of climate change on marine export production.
Global Biogoechemical Cycle, 15, 81-99.
BUITENHUIS E. T., C. LE QUÉRÉ, O. AUMONT (2002): The Dynamic Green Ocean Model: 6
plankton functional groups in an ocean global circulation model. Poster of workshop on “Global
ocean productivity and the fluxes of carbon and nutrients: combining observation and models”,
24.-27. June, Ispra (Italy). (http://ijgofs.whoi.edu/GSWG/Ispra_modelling/Buitenhuis.pdf)
[13.06.2007]
DONEY S. C. (2006a): The dangers of ocean acidification. Scientific American, March 2006, 58-
65.
DONEY S. C. (2006b): Plankton in a warmer world. Nature, 4447, 695-696.
FALKOWSKI P. G., O. SCHOFIELD, M. E. KATZ, B. VAN DE SCHOOTBRUGGE, A. H.
KNOLL (2004): Why is the land green and the ocean red? In: THIERSTEIN H. R., J. R.
YOUNG (2004): Coccolithophores from molecular processes to global impact. Springer Verlag,
427-453.
- 16 -

IGLESIAS-RODRÍGUEZ M. D. AND OTHERS (2002): Representing key phytoplankton functional
groups in ocean carbon cycle models: Coccolithophorids. Global Biogeochemical Cycles, 16,
1100, doi: 10.1029/2001GB001454.
LE QUÉRÉ C. AND OTHERS (2005): Ecosystem dynamics based on plankton functional types for
global ocean biogeochemistry models. Global Change Biology, 11, 2016-2040.
NANNINGA H. J., T. TYRRELL (1996): Importance of light for the formation of algal blooms by
Emiliania huxleyi. Marine Ecology Progress Series, 136, 195-203.
PAASCHE E. (2002): A review of the coccolithophorid Emiliania huxleyi (Prymnesiophyceae),
with particular reference to growth, coccolith formation, and calcification-photosynthesis
interactions. Phycologia, 40, 503-529.
RAVEN J. AND OTHERS (2005): Ocean acidification due to increasing atmospheric carbon dioxide.
The Royal Society, June 2005.
RIEBESELL U., I. ZONDERVAN, B. ROST, P. D. TORTELL, R. E. ZEEBE, F. M. M. MOREL
(2000): Reduced calcification of marine phytoplankton in response to increased atmospheric
CO 2 . Nature, 407, 364-367.
RIEBESELL U. (2004): Effects of CO 2 enrichment on marine phytoplankton. Journal of
Oceanography, 60, 719-726.
RIEGMAN R., W. STOLLTE, A. A. M. NOORDELOOS, D. SLEZAK (2000): Nutrient uptake
and alkaline phosphatase (EC 3:1:3:1) activity of Emiliania huxleyi (Prymnesiophyceae) during
growth under N and P limitation in continuous cultures. Journal of Phycology, 36, 87-96.
ROST B., U. RIEBESELL (2004): Coccolithophores and the biological pump: responses to
environmental changes. In: THIERSTEIN H. R., J. R. YOUNG (2004): Coccolithophores from
molecular processes to global impact. Springer Verlag, 99-125.
ROST B., U. RIEBESELL, S. BURKHARDT, D. SÜLTEMEYER (2003): Carbon acquisition of
bloom-forming marine phytoplankton. Limnology and Oceanography, 48, 55-67.
SARMIENTO J. L. AND OTHERS (2004): Response of ocean ecosystems to climate warming.
Global Biogeochemical Cycles, 18, GB3003.
SARMIENTO J. L., N. GRUBER (2006): Ocean biogeochemical dynamics. Princeton University
Press.
SARTHOU G., K. R. TIMMERMANS, S. BLAIN, P. TRÉGUER (2005): Growth physiology and
fate of diatoms in the ocean: a review. Journal of Sea Research, 53, 25-42.
ZONDERVAN I., R. E. ZEEBE, B. ROST, U. RIEBESELL (2001): Decreasing marine biogenic
calcification: a negative feedback on rising atmospheric pCO 2 . Global Biogeochemical Cycles,
15, 507-516.
- 17 -

7. Annex
Figure A 1 Fraction of calcite in surface sediments (% weight: mass calcite to mass total solid dry sediments)
(Sarmiento and Gruber, 2006).
Figure A 2 Fraction of opal in surface sediments (% weight: mass opal to mass total solid dry sediments)
(Sarmiento and Gruber, 2006).
- 18 -

Term paper, June 2007
Mechanisms of oxidative vs. reductive nitrobenzene
transformation by microorganisms
Term paper in Biogeochemistry and Pollutant Dynamics,
Master Studies in Environmental Sciences, ETH Zürich
By
Ivo Caravatti
Tutor: Dr. Thomas Hofstetter
Abstract
Nitroaromatic compounds such as nitrobenzene (NB), are widespread environmental
contaminants. In natural attenuation processes selected microorganisms are able to
degrade nitrobenzene via two different pathways, that is oxidation of nitrobenzene to
catechol and reduction to aniline, respectively. Since both processes occur under oxic
conditions but the latter leads to products of similar toxicity than the initial
contaminant, it is necessary to differentiate between the two reactions at contaminated
sites. Using compound-specific isotope analysis (CSIA), it should, in principle possible
to identify NB degradation pathways by measuring changes of stable isotope signature
of the remaining NB during its degradation. This hypothesis relies on the different
enzymatic mechanisms that have been reported for enzymatic nitrobenzene
degradation. However, since data on isotope effect are lacking, this term paper reviews
the elementary reaction steps of the competing enzymatic nitrobenzene transformation
pathways on a molecular scale to obtain some estimates of the potential magnitude of
carbon and nitrogen isotope effects. Oxidation of nitrobenzene by Comamonas sp.strain
JS765, should result in a moderate C isotope effect. The magnitude of the isotope effect,
however, depends on some elementary reaction steps in the dioxygenation mechanisms,
which are not fully understood to date. In addition, depending on the reaction
mechanisms, also a strong N isotope effect might be expected. On the other hand,
reduction of NB by Pseudomonas pseudoalcaligenes strain JS45 is expected to result in
negligible C and substantial N isotope fractionation. The theoretical evidence for
isotope fractionation obtained in this term paper suggest that the two enzymatic
nitrobenzene transformation pathways might be distinguished in the field on the basis
of compound-specific isotope analysis.
1. Introduction
Nitroaromatic compounds (NACs) constitute a very important class of environmental contaminants.
The stability of nitroaromatic compounds make them valuable as chemical intermediates in
industrial chemistry (8). Nitroaromatic compounds are used especially for the production of
pesticides, dyes, and polymers. On the other hand, nitroaromatic explosives, which are derived from
nitration of toluene, constitute a big problem at sites of military explosive production, ammunition
testing and storage. The widespread application NACs and their improper disposal has resulted in
the release of NACs into the environment (10). Innumerable sites are contaminated worldwide. In
1

Term paper, June 2007
addition, NACs used as agrochemicals such as pesticides are spread deliberately into the
environment (9).
Fortunately, many nitroaromatic compounds accumulated in the environment are biodegraded by
bacteria. Bioremediation can be defined as any process that uses microorganisms to return a
polluted environment to its original conditions. Adding specific substrate is possible to induce or
accelerate a determinate biodegradation process. This process is called stimulation. Since
mechanical cleanup of nitroaromatic polluted areas is nowadays an economic issue, bioremediation
is an important alternative (9).
However, compounds like nitrobenzene (NB) have more than one biodegradation pathway.
Comamonas sp.strain JS765 is able to oxidise NB, on the other hand Pseudomonas
pseudoalcaligenes strain JS45 is able to reduce NB (14, 15). Fig 1 shows the main step of the NB
degradation for both microorganisms. Although both degradation pathways lead to mineralization
of the compound, some products of the reductive pathway are transient more toxic products
(especially aniline) Many of this products are cytotoxic and/or mutagenic (20).
NO 2
A) Oxydation
B)
Reduction
OH
Catechol
OH
-
CO 2 +NO 2
NH 2
CO 2 +NH 3
Aniline
Figure 1: Main steps of the degradation of NB: A) Oxidation leads to the formation of catechol; B)
Reduction leads to the formation of transient more toxic products. Both pathways lead to the complete
degradation of NB. Modified from (14, 15).
It is important to determine which pathway occurs in the contaminated site. However, a direct
measurement of the biodegradation products is labour intensive and not always conclusive for the
distinction of the pathways. An increasingly important tool for qualitative and quantitative
assessment of abiotic and enzymatic transformations of organic contaminants in the environment is
the Compound Specific Isotope Analysis (CSIA)(6). CSIA is based on isotopic signature. Isotopic
signature is determined through the relative isotopic compositions of a given element within a
compound. Isotope signatures for a given element E is defined as follow (6):
2

Term paper, June 2007
⎛ ⎛
⎜
⎜
⎜
h ⎝
δ E = ⎜
⎜⎛
⎜
⎜
⎝⎝
h
l
h
l
E ⎞
⎟
E ⎠
E ⎞
⎟
E ⎠
sample
reference
⎞
⎟
⎟
−1⎟
⋅1000
⎟
⎟
⎠
(1)
As isotope signature values are rather small, the number is converted into 0 / 00 . h E and l E represent
the relative abundance of heavy and light isotope of element E, respectively, in the sample and in an
international reference standard (4). If isotope signatures are studied on more than one element,
information on specific reaction mechanisms can be derived.
As during the course of a reaction (e.g. a bond breaking or formation) the isotopes of a given
element E present in the substrate do not react with the same rate constant, one of the isotope can be
enriched in the substrate. This effect is called Kinetic Isotopic Effect (KIE E ). Generally, the light
isotope reacts faster than the heavy one, so that the heavy isotope is enriched in the substrate.
Normally, this enrichment is independent of the concentration of the compound; therefore an
enrichment factor for the element E (ε E ) can be calculated, ε E is commonly reported in per mil
(‰)(6):
h
⎛ 1000 + δ E ⎞ ε
E
ln
⎜
= ⋅ ln
h
1000 E
⎟
⎝ + δ
0 ⎠ 1000
f
(2)
Where f is the fraction of compound that has not reacted (usually c(t)/c(t=0)), δ h E and δ h E 0 are the
isotopic signatures of the compound for the element E at times t and zero, respectively. As
explained above changes in isotope signature (∆δE) are due to different reaction kinetics of the
isotopes. ε E -values can dependent on environmental conditions.
Considering the reaction mechanisms it is possible to calculate an apparent kinetic isotope effect
(AKIE E ). This can be directly derived from the experimentally observed ε E . AKIE E can be
calculated as follow (6):
AKIE E
=
l k obs
h k obs
=
1
1+ λ ⋅ε E
1000
If only one atom of the element E is present in the molecule and this atom is taking part to the
reaction, λ is equal 1. If several atoms of the same element are present in a contaminant, those not
taking part in the reaction will decrease the observable isotope fractionation and the corresponding
ε E -value due to isotopic dilution (7). It is important to know the reaction mechanisms so that the
element involved in the different biodegradation reactions can be defined. Bond cleavage or
creation have different effects on KIE and therefore on ε E. Enrichment factors are directly dependent
on reaction mechanisms (4). In addition, it is important to distinguish between primary isotope
effects and secondary isotope effects. Primary isotope effects are observable on atoms involved in a
bond cleavage or formation. Secondary isotope effects are observable on atoms that are not directly
taking part to the cleavage or formation reaction. Normally primary isotope effects are bigger than
secondary. Hydrogen has a big weight difference between its isotopes, so that often secondary
isotope effects for this atom are experimentally accessible (4).
(3)
3

Term paper, June 2007
Unfortunately, the exact values for the kinetic isotope effect for NB degradation reactions are not
known yet. As point out above, the products of the NB biodegradation are not always conclusive
for the distinction of the pathways; one of the advantages of using CSIA is that you measure only
the change in isotopic signature in the substrate (NB).
2. Goals
Goal of this term paper is to asses the magnitude of isotope fractionation of the two alternative NB
degradation pathways that is reductive and oxidative. This means to discuss the main elementary
reactions steps that lead to a substantial isotope effect.
In the following chapters, I review the elementary reaction steps, for NB oxidation and reduction,
where bond are broken or bonding between the elements is changed significantly. In other words,
this means that I try to identify where elementary reaction steps can cause a primary isotope effect.
Since secondary isotope effects are usually one order of magnitude smaller and are more difficult to
exploit for to the identification of isotope fractionation (4).
Transport processes shows a negligible isotope effect and therefore masks isotope effect caused by
bond cleavage if transport is limiting. In this work, we assume that transport processes do not limit
the kinetic of NB transformation. Thus, I discuss isotope effects that are linked to bond cleavage
and formation. The results of this research can than be applied for the design isotope fractionation
experiment in laboratory and field.
3. Results and Discussion
3.1. Oxidative Pathway
Comamonas sp.strain JS765 is able to oxidise NB: product of this reaction is catechol, whereas
nitrite is released. Figure 2 gives an overview of the elementary reaction steps taken in
consideration. First NB enter into the bacteria (1); second NB binds to the enzyme responsible for
the oxidation (2). After the formation of the enzyme substrate complex, two atoms of oxygen are
added (3), finally catechol is produced and nitrite is released (4). To understand how elementary
reaction can influence isotopic enrichment factors and which kind of enrichment factors can be
expected, a closer look at the processes influencing this reaction is necessary. Therefore one needs
to understand enzymatic reaction.
Figure 2: overview of the elementary reaction steps occurring until nitrite is released (breaking of the C-N
bond). E = NBDO (Nitrobenzene dioxygenase), S = NB, P = Catechol. Elementary reaction steps: 1) NB
4

Term paper, June 2007
enters in to the bacteria and it is transported to the reactive-site of NBDO (E); 2) NB bind to the active site of
NBDO, the complex “ES” is built; 3) Dioxygenase reaction: two atoms of oxygen are added 4) Catechol is
produced, Nitrite is released.
Unfortunately, less is known about the transport of NB into the bacteria as well as the transport to
the active site. The rate constant of each elementary reaction step is important for determine isotope
effects. The overall reaction rate for the dioxygenation is determined by the rate of the slowest step
(18). In part 3.1.1. the effects of elementary reaction steps 3 and 4 on isotope fractionation are
discussed in detail.
3.1.1. Enzyme description, active-site and substrate binding
In this chapter, a closer look to the elementary reaction step 2 in figure 2 will be taken.
Nitrobenzene dioxygenase (NBDO) from Comamonas sp.strain JS765 is able to add two atoms of
oxygen to NB. Two electron and two protons are consumed during this process. As result a
dihydroxyl intermediate (EP in figure 2) is produced. Through spontaneous rearrangement a
molecule of nitrite is released and a molecule of catechol (P in figure 2) is formed (5).
NBDO from Comamonas sp.strain JS765 was molecularly characterised by Lesser et al.(2001) (13).
Successful purification and characterisation of all enzyme components were reached by Parales et al
(2005) (17). Based on gene order and degree of sequence similarity NBDO belongs to the
naphthalene family of Rieske nonheme iron dioxygenases (RDOs) (13). Three main components, a
reductase, a ferrodoxin and a dioxygenase, build the multicomponent enzyme system. NBDO (the
dioxygenase component) is a hexamer composed of three α and three β subunits. The active site is
in a α subunit. Figure 3 shows how the multicomponent enzyme system works. The dioxygenase
reaction takes place only when the active-site iron is in the reduced form (5).
Figure 3: Electron transfer chain in a Rieske nonheme iron dioxygenase (RDO) enzyme system. (Modified
from (2)) Both the ferrodoxin component and the dioxygenase component contain a [2Fe2s] iron sulfur
cluster responsible for the electron transfer. The dioxygenase contain also an iron atom in the active-site
The active site, which has an oval-shaped form, is formed by seventeen amino acid residues. Most
of them are hydrophobic and provide an appropriate environment for the binding of aromatic
substrate (5). However only three residues distinguish the NBDO activity from the activity of the
others RDOs. Those are: an aspargine whose amino group is able to form a hydrogen bond with the
nitro groups of nitrobenzene, an isoleuctine and a phenylalanine which are important for the correct
positioning of the substrate and the dimension of the active pocket (10, 12). The formation of the
hydrogen bond is represented in figure 2 as “ES” after elementary reaction step 2.
5

Term paper, June 2007
NBDO has high similarity with other RDOs. For example naphthalene dioxygenase (NDO) has
80% of sequence identity with NBDO (5) the reductase and ferrodoxin components of NBDO are
identical of this of 2-Nitrotoluene Dioxygenase (2-NTDO) (13). This explain the ability of NBDO
to use as substrate compound such as naphthalene, 2-nitrotoluene, 3-nitrotoluene, 2, 4-
dinitrotoluene. Figure 4 shows the iron molecule in the active site and the hydrogen bond between
NB and the amino group of aspargine.
Figure 4: in red, the iron at the active site coordinated by two histidines, one aspargine and two molecule of
water (not shown). The amino group of the aspargine is able to form a hydrogen bond with the nitro group of
NB (Modified from (5))
3.1.2. Reaction mechanism: dioxygenation and release of Nitrite
Here elementary reaction 3 and 4 in figure 2 are discussed. Boyd et al (2005) (2) investigated the
possible mechanisms for the dioxygenation of aromatic compounds. Figure 5 shows the two
possible dioxygenation pathways for NB. It is demonstrate that both oxygen atoms come from the
same oxygen molecule, so that two consequent monooxygenation reactions are to exclude (15). In
case A oxygen atoms are added sequently, while in case B both atoms are added simultaneously. It
is not known which of the two mechanisms is used by the enzyme; however since it is suggested
that case B might be the used one (2), here isotopic fractionation consideration are based on case B.
During the reduction reaction a C-C double bond is transformed to a C-C single bond, two O-C
bonds are formed and a C-N bond is broken. For the analysis of the relevant elementary reaction
steps the time of the release of nitrite plays a central role. It is not clear yet if nitrite is released from
the dihidroxyl intermediate (ES in figure 2) during the addition of oxygen (case 1) or after the
oxygen addition (case 2). However, the fact that the release of nitrite is involved in the limiting step,
supports the thesis that nitrite is released during the addition of oxygen (15).
If we assume that the C-N bond is broken during the addition of oxygen (case 1) primary isotope
effect will be measurable for carbon and for nitrogen. On the other hand, if nitrite is released after
the addition of oxygen (case 2), primary isotope effect will be observable only for carbon.
Indicative substantial AKIE values for the cleavage of a double C-C bond is 1.01. AKIE value for
the cleavage of a C-N bond is 1.03 for both elements (carbon and nitrogen) (4).
Important to say is that hydrogen might undergo to an important secondary isotope effect.
Hydrogen is not directly involved in a bond cleavage or formation reaction. However due to the big
6

Term paper, June 2007
difference in the weight between the two hydrogen isotopes, even a small isotope effect might leads
to measurable enrichment factor.
Figure 5: possible catalytic mechanisms for NB dioxygenation; case A: sequently addition of the O atoms,
case B: simultaneously addition of the O atoms. In blue the dihidroxyl intermediate. (Modified from (2))
3.2 Reductive Pathway
Pseudomonas pseudoalcaligenes strain JS45 is able to reduce NB: a product of the reduction
reaction is nitrosobenzene. As the formation of nitrosobenzene is the only step which cause a
measurable isotope effect, only this reaction is here taken in consideration. Figure 6 shows the
elementary reaction steps occurring until the first bond is broken (N-O bond). First NB enter into
the bacteria (1); second NB binds to the enzyme responsible for the reduction (2). After the
formation of the enzyme substrate complex, two electrons are added (3), finally nitrosobenzene is
released (4) (14). In the same way as for the oxidative pathway, we assume that the rate limiting
step is the step where a bond is broken. In this case elementary reaction steps 4. A primary isotope
effect will so be observable only for two elements. Those are nitrogen and oxygen.
7

Term paper, June 2007
Figure 6: overview of the elementary reaction steps occurring until nitrosobenzene is produced (breaking of
the N-O bond). E = nitrobenzene nitroreductase, S = NB, P = Nitrosobenzene
Elementary reaction steps: 1) NB enters in to the bacteria and it is transported to the reactive-site of
nitrobenzene nitroreductase; NB bind to the active site of nitrobenzene nitroreductase, the complex “ES” is
built; 3) Reduction reaction: addition of two electrons; 4)The product nitrosobenzene is released.
The reaction can be summarized as follow:
Ar-NO 2 + 2e - +2H + ! Ar-NO + H 2 O
3.2.1 Enzyme description: reaction mechanism
Nitobenzene nitroreductase was purified and characterized by Somerville et al.(1995)(20).The
enzyme does not have metal cofactors. It was found that two FMN (Flavin mononucleotide)
molecule are bonded to the protein as cofactor (20). The reduction reaction is NADH dependent
(14). Nitrobenzene nitroreductase belongs to the oxygen insensitive (type 1) enzyme (20). This
category of enzyme uses a two electrons reduction. In contrast oxygen sensitive enzyme (type 2)
catalyses a one electron reduction of the nitro group: a nitro anion radical is built. The radical reacts
than spontaneously with oxygen building a superoxyde radical so that the nitro group is regenerated
and no net reduction of the nitro group is observable (19). A two electrons reduction reduces the
nitro group irrespective of the presence of oxygen.
To figure out the mechanisms of the NB reduction scientists studied the reduction of NB by
NAD(P)H nitroreductase by Enterobacter cloacae. Unfortunately the mechanism of two electrons
reduction of nitroaromatics is not enough understood to provide any information about isotope
fractionation (16). The nitro group is a facile electron acceptor. Many enzymes are able to catalyse
the reduction of aromatic nitro group. For example cytochrome c reductase, cytochrome P-450
reductase, glutathione reductase, hepatic NAD(P)H reductase and many others are able to catalyse
the reduction. One should note that the physiological role of this enzyme is not the reduction of
aromatic nitro group (20). Interestingly cytochrome P-450 reductase has analogously to
nitrobenzene nitroreductase a FMN and a FAD (flavin adenin dinucleotide) molecules as cofactors.
In fact both enzyme are flovoenzyme (11). NB is a substrate for cytochrome P-450 reductase (3).
The proposed mechanisms for the reduction of NB for both enzyme is a ping pong mechanisms (1,
3, 11). Figure 7 represents a possible mechanism. For cytochrome P-450 reductase is suggested that
electrons transfer follows the pathway NADH! FAD ! FMN ! electrons acceptor. Probably the
binding-site for NADH is a cysteine residue (3).
8

Term paper, June 2007
During the reductive degradation pathway of NB, Nitrosobenzene reacts fast to the next metabolite
hydroxylaminobenzene (11, 14). Due to this fast reaction it is hypothesized that nitrosobenzene
does not leave the enzyme and through the consumption of another NADH, nitrosobenzene is
reduced to hydroxylaminobenzene (11).
1)
E + NADH E---NADH EH -
E + S 1
ES 1
E 1
2)
EH - +ArNO 2
EH - ---ArNO 2 E+ArNO
E 1 + S 2
E 1 S 2 E + P
Figure 7: Possible mechanism for NB reduction: Ping Pong mechanism. (Modified from (1)). 1): The first
substrate (NADH) binds to the enzyme and reduces it, conformational change occurs. 2): The reduced
enzyme is now able to bind the second substrate (NB). After the reduction of NB and production of
nitrosobenzene, the enzyme returns to the beginning state.
Summarizing during the reduction of NB an N-O bond is broken. The N atom and the O atom will
undergo to a primary isotope effect. Since we assume that elementary reaction step 4 in fig. 6 is the
rate limiting step, the primary isotope effect will be measurable. Indicative substantial AKIE values
for the cleavage of a N-O bond are about 1.03 for nitrogen and 1.04 for oxygen (4).
3.3. Expected bulk enrichment factor for NB
Using equation 3 it is possible to calculate the enrichment factors (ε E ) for atom E. Due to the
different enzymatic reaction mechanisms, different dilution factors are expected.
3.3.1. NB oxidation
Parameters:
C-C double bond transformation to C-C single bond:
• Dilution factor: 3 (λ-value: 6/2)
• AKIE C =1.01(4)
•
ε
C
≅ −10 0 00
C-N cleavage:
• Dilution factor for element C: 6 (λ-value:6)
• AKIE C =1.03 (4)
•
ε
C
≅ −30 0 00
• Dilution factor for element N: 1 (λ-value:1)
• AKIE N =1.03 (4)
• ε
N
≅ −30 0 00
9

Term paper, June 2007
Enrichment factor for carbon (ε C ):
Case 1: nitrite is released during the addition of oxygen
0 1 ⎛ 2
10 0 ⎞
ε 30
00
00
⎟ ≅ −8 0
C
≅ − ⋅ + ⎜−
⋅
6 ⎝ 6 ⎠ 00
Case 2: nitrite is released after the addition of oxygen
2
ε 0 3
00
0
C
≅ −10 ⋅ ≅ −
6 00
Enrichment factor for nitrogen (ε N ):
Case 1: nitrite is released during the addition of oxygen
1
ε 0 30
00
0
N
≅ −30 ⋅ ≅ −
1 00
Case 2: nitrite is released after the addition of oxygen
ε = 0
N
3.3.2. NB reduction
Parameters:
N-O bond cleavage:
• Dilution factor for element N: 1 (λ-value:1)
• AKIE N =1.03 (4)
• ε
N
≅ −30 0 00
• Dilution factor for element O: 2 (λ-value:2)
• AKIE O =1.04 (4)
• ε ≅ −40 0
O
00
Enrichment factor for nitrogen (ε N ):
1
ε 0 30
00
0
N
≅ −30 ⋅ ≅ −
1 00
Enrichment factor for oxygen (ε O ):
1
ε 0 20
00
0
O
≅ −40 ⋅ ≅ −
2 00
4. Conclusion
The results of this review show that a determination of the fate of NB in contaminated areas with
CSIA should be possible. Different elements are involved in the oxidation and reduction reaction.
ε C is expected to be between -8 0 / 00 and -3 0 / 00 for the oxidative pathway and 0 0 / 00 for the reductive
pathway. ε N is expected to be about -30 0 / 00 for the reductive pathway. It is not possible to predict if
nitrogen will be enriched in the oxidative pathway. If it is the case, the ε N value will be also of -
30 0 / 00 . Oxygen is enriched in reductive pathway; unfortunately nowadays it is not possible to
measure its isotopic signature with certainty. Therefore carbon is the best candidate for the
determination of the pathway.
10

Term paper, June 2007
5. References
1. Anusevicius, Z., A. Soffers, N. Cenas, J. Sarlauskas, M. Martinez-Julvez, and I.
Rietjens. 1999. Quantitative structure activity relationships for the electron transfer
reactions of Anabaena PCC7119 ferredoxin-NADP(+) oxidoreductase with nitrobenzene
and nitrobenzimidazolone derivatives: mechanistic implications. Febs Letters 450:44-48.
2. Boyd, D. R., and T. D. H. Bugg. 2006. Arene cis-dihydrodiol formation: from biology to
application. Organic & Biomolecular Chemistry 4:181-192.
3. Cenas, N., Z. Anusevicius, D. Bironaite, G. I. Bachmanova, A. I. Archakov, and K.
Ollinger. 1994. The Electron-Transfer Reactions of Nadph-Cytochrome P450 Reductase
with Nonphysiological Oxidants. Archives of Biochemistry and Biophysics 315:400-406.
4. Elsner, M., L. Zwank, D. Hunkeler, and R. P. Schwarzenbach. 2005. A new concept
linking observable stable isotope fractionation to transformation pathways of organic
pollutants. Environmental Science & Technology 39:6896-6916.
5. Friemann, R., M. M. Ivkovic-Jensen, D. J. Lessner, C. L. Yu, D. T. Gibson, R. E.
Parales, H. Eklund, and S. Ramaswamy. 2005. Structural insight into the dioxygenation
of nitroarene compounds: the crystal structure of nitrobenzene dioxygenase. Journal of
Molecular Biology 348:1139-1151.
6. Hartenbach, A., T. B. Hofstetter, M. Berg, J. Bolotin, and R. P. Schwarzenbach. 2006.
Using nitrogen isotope fractionation to assess abiotic reduction of nitroaromatic compounds.
Environmental Science & Technology 40:7710-7716.
7. Hofstetter T., H. A., Tobler N. and Bolotin J. 2007. Labguide on the course Compound-
Specific Isotope Analysis (CSIA). ETH Zürich.
8. Hughes J.B., K. H. J., Nishino S.F., Spain J. C., and Zhongqi H. . 2000. Biodegradation
of nitroaromatic compounds and explosives. Chapter 2, Strategies for Aerobic Degradation
of Nitroaromatic Compounds by Bacteria: Process Discovery to Field Application.
9. Hughes J.B., K. H. J., Spain J. C. 2000. Biodegradation of nitroaromatic compounds and
explosives. Chapter 1, Introduction.
10. Ju, K. S., and R. E. Parales. 2006. Control of substrate specificity by active-site residues in
nitrobenzene dioxygenase. Applied and Environmental Microbiology 72:1817-1824.
11. Koder, R. L., and A. F. Miller. 1998. Steady-state kinetic mechanism, stereospecificity,
substrate and inhibitor specificity of Enterobacter cloacae nitroreductase. Biochimica Et
Biophysica Acta-Protein Structure and Molecular Enzymology 1387:395-405.
12. Lee, K. S., J. V. Parales, R. Friemann, and R. E. Parales. 2005. Active site residues
controlling substrate specificity in 2-nitrotoluene dioxygenase from Acidovorax sp strain
JS42. Journal of Industrial Microbiology & Biotechnology 32:465-473.
13. Lessner, D. J., G. R. Johnson, R. E. Parales, J. C. Spain, and D. T. Gibson. 2002.
Molecular characterization and substrate specificity of nitrobenzene dioxygenase from
Comamonas sp strain JS765. Applied and Environmental Microbiology 68:634-641.
14. Nishino, S. F., and J. C. Spain. 1993. Degradation of Nitrobenzene by a Pseudomonas-
Pseudoalcaligenes. Applied and Environmental Microbiology 59:2520-2525.
15. Nishino, S. F., and J. C. Spain. 1995. Oxidative Pathway for the Biodegradation of
Nitrobenzene by Comamonas Sp Strain Js765. Applied and Environmental Microbiology
61:2308-2313.
16. Nivinskas, H., S. Staskeviciene, J. Sarlauskas, R. L. Koder, A. F. Miller, and N. Cenas.
2002. Two-electron reduction of quinones by Enterobacter cloacae NAD(P)H :
nitroreductase: quantitative structure-activity relationships. Archives of Biochemistry and
Biophysics 403:249-258.
17. Parales, R. E., R. Huang, C. L. Yu, J. V. Parales, F. K. N. Lee, D. J. Lessner, M. M.
Ivkovic-Jensen, W. Liu, R. Friemann, S. Ramaswamy, and D. T. Gibson. 2005.
11

Term paper, June 2007
Purification, characterization, and crystallization of the components of the nitrobenzene and
2-nitrotoluene dioxygease enzyme systems. Applied and Environmental Microbiology
71:3806-3814.
18. Rene P. Schwarzenbach, P. M. G., Dieter M. Imboden. 2003. Environmental Organic
Chemistry, Chapter 13: Chemical Transformation I: Hydrolysis and Reactions Involving
Other Nucleophilic Species.
19. Schenzle, A., H. Lenke, J. C. Spain, and H. J. Knackmuss. 1999. Chemoselective nitro
group reduction and reductive dechlorination initiate degradation of 2-chloro-5-nitrophenol
by Ralstonia eutropha JMP134. Applied and Environmental Microbiology 65:2317-2323.
20. Somerville, C. C., S. F. Nishino, and J. C. Spain. 1995. Purification and Characterization
of Nitrobenzene Nitroreductase from Pseudomonas Pseudoalcaligenes Js45. Journal of
Bacteriology 177:3837-3842.
12

Biogenic manganese oxides:
Formation mechanisms,
mineralogy and environmental
relevance
Term paper: Biogeochemistry and pollutant dynamics
Thomas Ulrich – 02 923 233
tulrich@student.ethz.ch
Tutor: Dr. Ruben Kretzschmar
Submitted: 12.11.2007
Abstract
Biogenic manganese oxides can be a tool for soil remediation. Recent
papers are reviewed to gather information on their formation mechanisms,
mineralogy and environmental relevance. Characterisations of the manganese
oxidation products of the bacteria groups Bacillus sp., Pseudomonas putida
and Leptothrix discophora are described and compared. The biochemical pathways
of bacterial manganese oxidation are discussed. It seems, that bacteria
are able to produce a soluble Mn(III) complex, for use in redox processes and
iron nutrition. The structures of the biogenic manganese oxides are in the
nanoscale, with sizes around 4 to 100 nm and specific surface areas around 98
to 224 m 2 /g. The biogenic oxides are phyllomanganates with mostly hexagonal
and sometimes triclinic or pseudo-orthogonal structure. The potential of
biogenic manganese oxides in heavy metal scavenging is considerable. Metal
ions of Cd, Co, Hg, Ni, Pb, Zn and U can be adsorbed inexchangeable by biogenic
manganese oxides, and thus become immobile. The adsorption range
of biogenic manganese oxides lies, in the case of Pb, between 200 and 520
mmol P b /mol Mn . The efficiency of adsorption are 2 to 5 times higher than
that of synthetic ones. These differences originate from the poor crystallinity,
small size, high surface area and vacant sites in the crystal structure of biogenic
manganese products.

Contents
1 Introduction 3
2 Manganese oxides in soils 3
2.1 Redox cycle of manganese in soils . . . . . . . . . . . . . . . . . . . . 6
2.2 Relationship of manganese oxides and microorganisms in soils . . . . 7
3 Mechanisms of biogenic manganese oxidation 8
4 Structures of biogenic manganese oxides 11
5 Role of biogenic manganese oxides in soil remediation 16
6 Conclusion 19
List of Figures
1 Strains of manganese oxidising bacteria . . . . . . . . . . . . . . . . . 4
2 Structure of manganese oxides . . . . . . . . . . . . . . . . . . . . . . 5
3 Manganese cycle in the environment . . . . . . . . . . . . . . . . . . 6
4 Manganese oxidation endproduct . . . . . . . . . . . . . . . . . . . . 10
5 Crystal structure of biogenic and synthetic manganese oxides . . . . . 14
6 Complexes of metal cations and manganese oxides . . . . . . . . . . . 16
7 Efficiency of sorption on biogenic and synthetic oxides . . . . . . . . . 18
8 Extractability of biogenic manganese oxides . . . . . . . . . . . . . . 18
List of Tables
1 Types of manganese oxides . . . . . . . . . . . . . . . . . . . . . . . . 5
2 Manganese oxides by different bacteria . . . . . . . . . . . . . . . . . 14
3 Comparison between biogenic and synthetic manganese oxides . . . . 15
4 Comparison of specific surface area . . . . . . . . . . . . . . . . . . . 15
5 Pb adsorption capacity of manganese oxides . . . . . . . . . . . . . . 17

3
1 Introduction
Biogenic manganese (in the following also called Mn) oxides can be good tool for
soil remediation and heavy metal scavenging. The absorption efficiency of biogenic
manganese oxides surpasses those of synthetic oxides and commercial manganese
products (pyrolusite). Improvements in synchrotron based techniques like XANES 1
and EXAFS 2 lead to more detailed information on the small scale properties of
manganese oxides. In this review the specific structures of biogenic manganese
oxides and also the intracellular processes are discussed, to give an understanding
why they have better sorption properties.
Manganese oxides can adsorb arsenic (As), cadmium (Cd), cobalt (Co), mercury
(Hg), nickel (Ni), plutonium (Pu), uranium (U) and Zinc (Zn) cations. In addition
to this, they are also capable of oxidising a lot of organic substances like pesticides.
The production right at the contaminated site may be a further advantage. A major
disadvantage is the oxidation of chromium (Cr) by manganese oxides, a reaction
leading to a toxic product.
Oxidation of manganese is known to be catalysed by a great number of bacteria
and also fungi of different phylogenetic lineages (figure 1). Studies on biogenic
manganese oxides have focused on four model organisms: Bacillus sp. strain SG1
is gram positive and spore-forming. This marine bacterium oxidises Mn(II) on its
spores. Leptothrix discophora strain SS-1: A sheet-forming β-proteobacterium, living
in fresh water. Pseudomonas putida strain MnB1 and GB-1: A γ-proteobacteria
forming biofilms and living in fresh water and soil. KR21-2: a ascomycete fungus
strain.
In the following, the relevance of natural manganese oxides in soils will be discussed.
After that, the intracellular properties which lead to a specific biogenic
product, will be analysed. To understand the potential of biogenic manganese oxides
in contaminated soils, the environmental relevance of different products will be
shown.
2 Manganese oxides in soils
Manganese is an important trace element in soils, which occurs at low concentrations
(1/50 of Fe concentration). By weathering of igneous and metamorphic rock, Mn(II)
is released. This Mn(II) is oxidised to Mn(III) and Mn(IV) in presence of oxygen.
Manganese can form more than 30 known oxide and hydroxide minerals. Sometimes
1 X-ray absorption near edge structure. The absorption spectra are analysed, to identify the
species and their oxidation state.
2 Extended X-ray absorption fine structure. The characteristics of atomic neighbours can be
analysed.

4 Manganese oxides in soils
Figure 1: Strains of bacteria which oxidise Mn(II). (Tebo et al., 2004).
they also appear in combined valence forms, containing both Mn(III) and Mn(IV).
Manganese oxides are characterised by open crystal structures, large surface areas
and high negative charges (Tebo et al., 2004).
Basically manganese oxides are MnO 6 octahedra. These octahedra can be corner-,
edge- or face sharing and thus forming secondary structures like phyllomanganates
and tectomanganates (figure 2). The phyllomanganates consist of layered sheets of
edge-sharing MnO 6 octahedra. Between those sheets water or small cations can be
stored in exchangeable form. Phyllomanganates exist in two forms, separated by the
width of their interlayer: with 7Å (birnessite, chalcophanite and δ-MnO 2 ) and 10Å
(lithiophorite or buserite) d-spacings. The structure with 10Å d-spacing contains an
additional layer of water. 10Å manganese oxides collapse to 7Å phyllomanganates
when they are protonated or dehydrated (Scheinost and Daniel, 2005; Tebo et al.,
2004).
Manganese oxides can also appear in chain/tunnel structures, called tectomanganates.
An example is todorokite (figure 2). Tectomanganates differ in single chain

5
(pyrolusite) to double or triple chains (like todorokite and hollandite). They consist
of edge-sharing octahedra, which are linked by corner-sharing MnO 6 octahedra,
forming tunnels or pseudo-tunnels (see Table 1).
Figure 2: Structures of manganese oxides. Todorokite, ramsdellite and hollandite
are chain/tunnel forms. Birnessite is a phyllomanganate. The barium ions in the
central tunnels of hollandite and the Na, Ca and K cations are not shown. (Tebo
et al., 2004).
.
Table 1: Types of manganese oxides and their structure. (Scheinost and Daniel,
2005).

6 Manganese oxides in soils
In semi arid soils, two forms of manganese oxide minerals are most common:
birnessite and lithiophorite, which are both phyllomanganates. δ-MnO 2 is most often
compared to biogenic manganese oxides (see also section 4). Biological conversions
from phyllomanganates to other structures, like tectomanganates, may occur (Tebo
et al., 2004; Webb et al., 2005b).
2.1 Redox cycle of manganese in soils
Manganese in soils is present in three oxidation states: Mn(II), Mn(III) and Mn(IV).
In absence of oxygen, Mn(II) appears most often in solution or adsorbed to minerals.
Whereas in presence of oxygen Mn(III) and Mn(IV) appear as solid oxides or
hydroxides (Tebo et al., 2004).
Figure 3: Manganese cycle in the environment. In the absence of oxygen Mn(II)
is more abundant. In presence of oxygen, Mn(IV) is more abundant, followed by
Mn(III). (Tebo et al., 2004).
1/2MnO 2 + 2H + + e − −→ 1/2Mn 2+ + H 2 O (1)
Mn 2+ + 1/2O 2 + H 2 O −→ MnO 2 + 2H + (2)
Depending on the local environment manganese can be oxidised and reduced abiotically
in soils. Figure 3 shows the redox cycle of the abiotic manganese oxidation.
The oxidation of Mn(II) to Mn(IV) (equation (2)) is favourable under a high pH
and the presence of oxygen. If the concentration of Mn(II) is low, the oxidation to
Mn(IV) is enforced by the dissociation of the intermediate product Mn(III) (Tebo
et al., 2004). Although the oxidation is thermodynamically possible, the kinetics of

2 Relationship of manganese oxides and microorganisms in soils 7
this reaction are very slow with a half life in the order of hundreds of years. The
rate limiting step is the oxidation of Mn(II) to Mn(III) (Webb et al., 2005a).
The reduction process (equation (1)) is favourable under a low pH and in the
absence of oxygen. Thermodynamically manganese oxides have after oxygen (O 2 )
and nitrate (NO 3 ) one of the highest oxidation potentials. Regarding the kinetics,
they even surpass oxygen. This makes manganese oxides one of the most important
oxidants in soils (Tebo et al., 2004).
2.2 Relationship of manganese oxides and microorganisms
in soils
Bacteria are capable of oxidising Mn(II) to Mn(III) and Mn(IV) (equation (2))
enzymatically. The kinetics of this enzymatically catalysed reaction are a lot faster
than the abiotic reaction, with a half life in the order of days (compared to hundreds
of years). This process will be described in section 3.
Also the reduction (equation (1)) can be catalysed enzymatically by bacteria.
Mn(IV) is used as an electron acceptor in the anaerobic respiration by a lot of
microorganisms. In the following, the focus will lie strictly on the oxidation of
Mn(II) by bacteria.
Observing biogenic manganese oxidation, one of the questions is, why bacteria
oxidise Mn(II). There are assumptions, that manganese oxidation by bacteria is an
accidental occurrence or an evolutionary remainder (Tebo et al., 2005). But there are
also some indications, that bacteria can use manganese oxidation for their benefit.
One possibility is, that bacteria use manganese oxidation to derive energy for
chemolithoautotrophic growth. Oxidising Mn(II) to Mn(III) or Mn(IV) is thermodynamically
favourable, but there is still no distinct evidence, that bacteria oxidise
Mn(II) for ATP generation. Producing storage for an electron acceptor is a more
likely option. Thus an electron acceptor could be stored, until carbon and energy
become available again (Tebo et al., 2005).
Another reason for biogenic manganese oxidation could be the use of manganese
in cellular mechanisms. Manganese can be used as an antioxidant, protecting cells
from reactive oxygen species (ROS). These are, e.g. the superoxide radical, hydroxyl
radical, hydrogen peroxide or singlet oxygen. ROS can be produced by cells as
by-products of respiration. With Mn(II) oxidation, bacteria are able to protect
themselves, even if they do not posses a superoxide dismutase, which is the option
most often used for ROS scavenging. An open question here is, whether the oxidation
of Mn(II) is used intracellular or on the innermembrane, because the bacteria seem
to precipitate manganese oxides extracellular. It was speculated, that the bacteria
use manganese oxidation to protect themselves from oxidants in their environment
(Tebo et al., 2005).

8 Mechanisms of biogenic manganese oxidation
In addition to the mentioned gains of biogenic manganese oxidation, bacteria can
protect themselves with a coating of manganese oxides. This could protect them
from UV radiation, predation, viral attacks or heavy metal toxicity (Tebo et al.,
2005).
An other advantage of manganese oxidation is the ability of manganese oxides
to degrade humic substances oxidatively to smaller compounds. These compounds
can be used by the entire microbial community for growth. Evidence for this is the
faster rate of biomass accumulation in a monoculture actively oxidising manganese,
compared to a monoculture without manganese oxidation (Tebo et al., 2005).
Also an option could be the usage of Mn(III) as competitor for Fe(III) on
siderophores. Iron (Fe) is a critical micro nutriment for almost all bacteria. Some
organisms, including plants, use siderophores to make Fe(III) in rocks bioavailable.
To transport iron into the cell, it has to be released from the siderophore and reduced
to Fe(II). Mn(III) could as well compete Fe(III) on siderophores, as reducing
Fe(III) to Fe(II). This will also be discussed later in section 4.
3 Mechanisms of biogenic manganese oxidation
Basically, the oxidation of Mn(II) by bacteria could be reached by two ways. Bacteria
can catalyse the oxidation indirectly by influencing chemical parameters or they
can directly oxidise the manganese enzymatically. This review will focus on the
enzymatic oxidation.
It is most probable, that enzymes of the multi copper oxidase (MCO) family are
involved in the oxidation of manganese. The MCO family includes a great number
of proteins, which are copper dependent. They are found in prokarya as well as in
eukarya. Where they are for example responsible in the oxidation of Fe. To prove
that the oxidation step of manganese in Bacillus sp. is copper dependent, Francis
and Tebo (Francis and Tebo, 2002) used the copper chelator o-phenanthroline at
a concentration of 50 µM. After addition of o-phenanthroline the activities of all
proteins were inhibited, showing that manganese oxidation is copper dependent.
Tebo et al. (Tebo et al., 2004) found the MCO gene mnxG in Bacillus sp.. Very
similar genes were also found in Pseudomonas putida and Leptothrix discophora and
many other bacteria. This suggests that they all use a similar pathway for oxidising
manganese, utilising enzymes from the multicopper oxidase enzyme family. However,
the exact proteins involved in this process have not yet been purified (Tebo et al.,
2004). Differentiation has to be made, between the direct catalysis of Mn(II) by
MCO enzymes and their mere involvement in the pathway of Mn(II) oxidation.
Proteins of the MCO family could be expressed as precursors or side effects of
manganese oxidation without direct involvement in the oxidation step itself (Tebo
et al., 2004).

9
Three pathways of oxidising Mn(II) would be possible:
Mn(II) enzymatic
−−−−−−→ Mn(IV ) (3)
Mn(II) enzymatic
−−−−−−→ Mn(III) chemical
−−−−−→ Mn(IV ) (4)
Mn(II) enzymatic
−−−−−−→ Mn(III) enzymatic
−−−−−−→ Mn(IV ) (5)
Enzymatic oxidation of Mn(II) to Mn(IV) in a two electron transfer step (equation
(3)). Enzymatic oxidation of Mn(II) to Mn(III) in a single electron transfer
step, followed by a chemical dissociation of Mn(III) into Mn(II) and Mn(IV) (equation
(4)). Enzymatic oxidation of Mn(II) to Mn(III) in a one electron transfer step,
followed by a second one electron transfer step to Mn(IV) (equation (5)).
In antagonism to equation (3), Toner et al. (Toner et al., 2005) showed the
presence of an Mn(III) intermediate in Pseudomonas putida. Webb et al. (Webb
et al., 2005a) showed the presence of a Mn(III) intermediate in spores of Bacillus
sp. strain SG-1. So the oxidation of manganese with a two electron transfer is
improbable.
Comparing equation (4) and (5) most authors support the pathway of two enzymatically
catalysed one electron transfers (equation (5)) (Francis and Tebo, 2002;
Tebo et al., 2004, 2005; Toner et al., 2005; Webb et al., 2005a). This pathway would
also be analogous to the pathway, with which bacteria and eukarya are oxidising
Fe(II). Evidence has also been found, that the second step of the Mn(II) oxidation
works independently from the first step (Tebo et al., 2004).
These findings lead to the question regarding speciation of the Mn(III) intermediate.
Tebo et al. (Tebo et al., 2004, 2005) and Toner et al. (Toner et al.,
2005) differ in their arguments. Toner et al. (Toner et al., 2005), who used XANES
and STXM 3 , argue there is a membrane bound intermediate of Mn(III).Tebo et al.
(Tebo et al., 2004, 2005) used a chemical trapping step with pyrophosphate, which
complexes Mn(III) in solution. They argue that the Mn(III) phase is in solution as
an intracellular Mn- enzyme- complex.
Mn(II) oxidation seems to take place inside the cell and by a membrane bound
process. The one-electron transfer process from Mn(II) to Mn(III) seems to take
place inside the cell plasma. The oxidation of Mn(III) to Mn(IV) appears to be a
membrane bound process and the final product precipitates around the cell (Toner
et al., 2005). This would include proteins and enzymes for the transport of Mn(III)
from cell plasma to the membrane. This proteins have not yet been found. The
transport would also account for the fact, Tebo et al. (Tebo et al., 2004) found
3 Scanning transmission X-ray microscopy. Used to study the spatial resolved characterisation
of manganese oxides.

10 Mechanisms of biogenic manganese oxidation
Mn(III) in solution. Mn(III) could be complexed by pyrophosphate, before it reaches
the membrane by enzyme / pyrophosphate competition in this study. Perhaps bacteria
even have some mechanisms to store Mn(III)-enzyme complexes in or at the
membrane, to be released when needed. Considering the time, when the product is
precipitated, several authors showed that first Mn(II) in solution is diminished and
then Mn(III) and Mn(IV) concentration rises (Tebo et al., 2004; Toner et al., 2005).
In the authors opinion, Mn(III) complexes produced by bacteria, could be very
important for the environment. Mn(III) is a strong oxidant, which may be used to
free Fe(III) from siderophores, using Mn(III) as a competitor for Fe(III) and thus
making Fe(III) bioavailable (see section 2.2). It could also be used by bacteria in
redox reactions. Thus Mn(II) oxidation may be a mechanism, developed by bacteria
to produce Mn(III) in solution. The proteins responsible for this processes have not
yet been found.
Figure 4: Electron microscope pictures of the manganese oxidation endproduct of
P.putida. A and B scanning electron microscope images of the aggregates (10 - 30
µm). C and D show biogenic manganese oxide particles using transmission electron
microscopy. (Toner et al., 2005).
The final product of Mn(II) oxidation outside the cell seems to be almost pure
precipitated Mn(IV), at least at low Mn(II) initial concentrations. An electron microscope
picture of the final product can be seen in figure 4. After some time the
concentration of Mn(III) in the precipitated product rises (Toner et al., 2005). It is
likely, that only Mn(IV) precipitates are produced and later reduced in the presence
of Mn(II) to Mn(III). Evidence for this could be, that at a higher initial Mn(II)

11
concentration, Mn(III) is produced sooner and in higher concentrations (enzyme
bound Mn(III) in solution should not be confused with Mn(III) in the precipitate.).
The precipitation product after 48h consists of Mn(III) and Mn(IV) with an average
oxidation number of 3.7 to 3.9 (Villalobos et al., 2003; Webb et al., 2005b).
The average oxidation numbers are obtained with XANES analysis. Three or more
species with known oxidation state are analysed, than the oxidation state of the
sample is determined with a regression approach. The average oxidation state of
the endproduct depends on the initial Mn(II) concentration.
4 Structures of biogenic manganese oxides
In this section, the structure of biogenic manganese oxides will be discussed. To obtain
information about the structural properties, synthetic manganese oxides were
compared to biogenic manganese oxides. δ-MnO 2 and acid birnessite were formed
through reduction of permanganate in MnCl 2 respectively HCl (Villalobos et al.,
2006). Webb et al. (Webb et al., 2005b) also synthesised triclinic birnessite and
todorokite for comparison. These synthetic oxides were compared to biogenic manganese
oxides produced by bacteria in culture. The manganese oxides were analysed
with XANES, which gives a high resolution spatial characterisation of the elements
and their oxidation state. EXAFS analyse lets determine the atomic neighbours of
certain elements and the type of complexation. Synchrotron based XRD 4 gives high
resolved information about the crystal mesh of the manganese oxide. The surface
areas were measured by a N 2 specific BET 5 method in all cases.
The biogenic manganese oxides produced by the model organisms (see section
1) have general properties. All biogenic manganese oxides showed a very poor crystallinity
(Villalobos et al., 2003; Webb et al., 2005b). Villalobos et al. (Villalobos
et al., 2003) speak of “extremely defective layered compounds with turbotratic
stacking and random stacking faults”. The primary manganese oxides are in the
nanoscale, ranging from 4 to 100 nm (Kim et al., 2004; Nelson et al., 1999). The
manganese oxides form aggregates outside the cell, as shown in section 3. These
aggregates consist of Mn(II), Mn(III) and Mn(IV), where Mn(IV) seems to be the
most frequent species (Webb et al., 2005b; Villalobos et al., 2006). Dependent on
time and initial Mn(II) concentration (or the concentration of other species which
could reduce Mn(IV)), the concentration of Mn(III) may rise in the aggregates. The
most common structure of the aggregation is a hexagonal phyllomanganate (Villalobos
et al., 2003, 2006; Webb et al., 2005b). The nanoscale size of the manganese
oxides results in high specific surface areas of the aggregates in the range between
98 and 224 m 2 /g (Nelson et al., 1999; Villalobos et al., 2005b).
4 X-Ray diffraction. Gives information about the crystal properties.
5 Brunauer-, Emmet-, Teller- theory of adsorption of gas molecules on solid surfaces

12 Structures of biogenic manganese oxides
Biogenic manganese oxides have been compared to synthetic manganese oxides
by several authors (Nelson et al., 2002; Villalobos et al., 2003, 2006; Webb et al.,
2005b). Several synthetic manganese oxides were produced with the methods mentioned
above. Webb et al. (Webb et al., 2005b) found a biogenic manganese oxide
similar to δ-MnO 2 . Villalobos et al. (Villalobos et al., 2003, 2006) stated, that
the physical and chemical attributes of the biogenic manganese oxides are between
δ-MnO 2 and acid birnessite (see also Table 3). Villalobos et al. (Villalobos et al.,
2003, 2006) paid special attention to vacant sites in the crystal. In synthetic triclinic
structures the negative charge, which is responsible for the exchangeable binding of
cations like Na + , Ca 2+ or water, derives from the content of Mn(III). The poor crystalline
structure of biogenic manganese has its negative charge through vacant sites
in the crystal structure (see figure 5) (Villalobos et al., 2003, 2006). Because of this
vacant sites inside the crystal mesh, innersphere complexes can be promoted. They
are responsible for the strong binding of heavy metals (Villalobos et al., 2003, 2006).
The biggest differences between synthetic and biogenic manganese oxides are found
by comparing specific surface areas. The specific surface area of biogenic manganese
oxides is two to four times higher than that of synthetic oxides (Nelson et al., 2002;
Villalobos et al., 2003). Nelson et al. (Nelson et al., 2002), who studied Leptothrix
discophora, found a specific surface area of the biogenic product twice as high as
that of Pseudomonas putida, studied by Villalobos et al. (Villalobos et al., 2003).
See also tables 3 and 4.
Discussing the appearance of manganese oxides in nature, it is of interest, whether
the natural oxides are of biotic or abiotic origin. As shown in section 2.1 the kinetics
of the abiotic reaction (equation (2)) are very slow (Tebo et al., 2004). Comparing
biogenic manganese oxides in culture with manganese oxides found in nature, it is
probable, that most of the natural manganese oxides are of biogenic origin. Products
very similar to the biogenic manganese oxides produced in the laboratories,
have been found in the oxic / anoxic transition zone in the sea and lakes (Granina
and Callender, 2006; Webb et al., 2005b). This gives an evidence, that biogenic
manganese oxides are abundant in nature. From primary biogenic oxides secondary
manganese oxides can develop. This reaction depends on physical and chemical
conditions. Secondary structures could also include tectomanganates (Webb et al.,
2005b).
In the following the more detailed results and the differences of diverse manganese
oxide products will be discussed. In contrast to the hexagonal structure proposed
above, Jurgensen et al. (Jurgensen et al., 2004), who studied Leptothrix discophora,
found a triclinic oxide, which consists of a single octahedral layer of microcrystals
composed of Mn(IV), Mn(III) and Mn(II). This product also showed a very high
specific surface area of 224 m 2 /g (Jurgensen et al., 2004). The products of Leptothrix
discophora were in the range of 4 nm to 100 nm (Kim et al., 2004). Kim et al. (Kim
et al., 2004) suggested, there could be tectomanganates like todorokite, when they

13
studied Leptothrix discophora. These results have later been challenged by Web
et al. (Webb et al., 2005b) and Villalobos et al. (Villalobos et al., 2006), who
stated that there are no primary biogenic manganese oxides with a tectomanganate
structure. They argue, that the use of UV Raman techniques, applied by Kim et
al. (Kim et al., 2004), are not capable of showing an evidence for todorokite, and
the NEXAFS analysis alone showed no distinct evidence for todorokite. Tani et
al. (Tani et al., 2004) however supported the hypothesis of Kim et al. (Kim et al.,
2004).
Webb et al. (Webb et al., 2005b) studied the structure of the manganese oxide
of Bacillus sp.. The product they found, was a biogenic Ca-birnessite phyllomanganate.
In addition to the hexagonal structure mentioned above, they also found a
pseudo-orthogonal phase. After the first 12 h of the reaction, the hexagonal phase
was the dominant form. After some additional time, more of the pseudo-orthogonal
phase seemed to have been formed (Webb et al., 2005b). They assume for thermodynamically
reasons, that secondary biogenic manganese oxides are formed from
primary manganese oxides. In sea water samples both forms were found and supported
to be of biogenic origin (Webb et al., 2005b). The difference between the two
phases is, that pseudo-orthogonal birnessites consist of up to one third of Mn(III),
whereas hexagonal forms mainly consist of Mn(IV) (Webb et al., 2005b). The oxidation
state of manganese found in this study, was between 3.7 and 4.0. The spacing
between the layers was 10Å in both forms (Webb et al., 2005b).
Villalobos et al. (Villalobos et al., 2003, 2006) analysed the structure of the
manganese oxide of Pseudomonas putida. In contrast to Webb at al. (Webb et al.,
2005b), who studied Bacillus sp., they only found the hexagonal form of the oxide.
All products had a very high content of Mn(IV). Triclinic or pseudo-orthogonal forms
were not found. The d-spacing between layers was in this case 7.2 Å, the specific
surface area was 98 m 2 /g. The oxidation state was 3.9 (all data: (Villalobos et al.,
2003, 2006)). Villalobos et al. (Villalobos et al., 2006) showed that the crystals
contained three to six randomly stacked layers. The residual layer charge is thereby
balanced by Na + cations (0.15 per octahedra). The amount of interlayer H 2 O was
three times the amount of interlayer alkali cations. Further, the presence of Mn(II)
was found in the interlayer of the biogenic manganese oxide, propably a remainder of
the Mn(II) substrate. A question not answered by Villalobos et al. (Villalobos et al.,
2003, 2006) is, if the Mn(II) in the interlayer would reduce Mn(III) and Mn(IV).
As shown in this section, the products of bacterial oxidised Mn(II) can differ in
some parts, like the oxidation step (Bacillus sp.: 3.7, Pseudomonas putida: 3.9),
the type of interlayer cations (Bacillus sp.: Ca 2+ ,Pseudomonas putida: Na + ), the
number of layers (Leptothrix discophora: 1, Pseudomonas putida: 3 - 6), the content
of manganese species and the structure (Leptothrix discophora: Mn(II), Mn(III),
Mn(IV), Bacillus sp.: Mn(III) and Mn(IV) , Pseudomonas putida: Mn(II) and
Mn(IV)). The results are summarised in table 2.

14 Structures of biogenic manganese oxides
Figure 5: Idealised crystal structures of: a) Synthetic triclinic birnessite, b) biogenic
hexagonal birnessite. Darkened octahedra: Mn(III) rich zones, balanced by hydrated
cations (not shown). V: Vacancies. (Villalobos et al., 2005b).
Table 2: Structures, physical and chemical properties of manganese oxides by different
bacteria.

15
Table 3: Comparison between biogenic and synthetic manganese oxides. (Villalobos
et al., 2003).
Table 4: A biogenic and synthetic (fresh abiotic) manganese oxide are compared regarding
the specific surface area. Pyrolusite is commercial manganese oxide product.
(Nelson et al., 2002).

16 Role of biogenic manganese oxides in soil remediation
5 Role of biogenic manganese oxides in soil remediation
Biogenic manganese oxides play a significant role in sorption of heavy metals in
the environment. Their ability of inexchangeable sorption of heavy metals and
oxidation of organic substances make them an usefull tool in soil remediation. As
shown in section 4 biogenic manganese oxides have a greater specific surface area
than synthetic manganese oxides. In this section the influence of vacant sites in the
crystal mesh on sorption attributes will be discussed, as well as the general sorption
properties of biogenic manganese oxides.
Villalobos et al. (Villalobos et al., 2005a) ordered the types of possible sorption
mechanisms into two different kinds: a) Adsorption in the interlayer, with about
25 - 50% of the adsorbed metal cations and b) adsorption as surface complex at
the particle edges, with 50 - 75% of the adsorbed metal cations. Later Tebo et al.
(Tebo et al., 2004) suggested a third option: Incorporation into the vacant sites as
substituent for manganese. Metal cations are able to form two different complexes
with oxygen atoms of the crystal mesh. A double corner sharing bidentate complex
or a triple corner sharing tridentate complex (see figure 6)
Figure 6: Possible complexes of metal cations and manganese oxides. In this case
the metal cation is a lead ion. Metal cations can form a) a tridentate, three corner
sharing complex and b) a bidentate, two corner sharing complex. The distances
were measured with EXAFS. (Villalobos et al., 2005a).
Various studies have been conducted on the topic of heavy metal sorption by
biogenic manganese oxides. As, Cd, Co, Hg, Ni, Pb, Pu, U and Zn were analysed
concerning the sorption properties to manganese oxides. All showed very good
sorption qualities. In almost all the cases the adsorption efficiency of manganese

17
oxides exceeded those of iron oxides (reviewed in (Tebo et al., 2004)). Tebo et al
studied the oxidation of U(IV) to U(VI), with following incorporation of U(VI) in
the manganese oxide as a bidentate complex (reviewed in (Tebo et al., 2004)). U(VI)
also showed internal changes in the biogenic manganese oxide, forming todorokitelike
tunnel structures.
Nelson et al. (Nelson et al., 1999) studied the Pb sorption characteristics of the
manganese oxides of Leptothrix discophora. They compared biogenic manganese oxides
to freshly synthesised ones and commercial manganese oxides for heavy metal
sorption. The adsorption results of the abiotic oxides could approach the biogenic
ones, but the biogenic manganese oxides still had about 2 to 5 times higher adsorption
capacity (see Table 5). The adsorption of the biogenic products ranged from
150 (mmol Pb/ mol Mn) to 250 (mmol Pb/ mol Mn) with a maximum of 520 (mmol
Pb/ mol Mn). In comparison to the commercial product, they outperform them by
orders of magnitude (see also Table 5) (Nelson et al., 1999, 2002).
Table 5: Maximum adsorption capacities.
manganese products. (Nelson et al., 2002).
The pyrolusite oxides are commercial
Tani et al. (Tani et al., 2004) analysed the oxidation product of fungus strain
KR21-2 in terms of heavy metal sorption. They focused on Co(II), Ni(II) and Zn(II)
sorption. The study showed, that the metal ions adsorbed to the biogenic product
with higher efficiency than to the abiotic manganese oxides (figure 7). Co(II) and
Ni(II) bound irreversibly to the oxide, while the adsorbed Zn(II) was exchangeable
(figure 8). They used a two step extraction method with CuSO 4 on the manganese
oxides. In a solution with a initial concentration of 0.1mM of biogenic manganese
oxide, 20% of the adsorbed Co(II) and Ni(II) were released (figure 8). The same
extraction with synthetic manganese oxides, released 70% of the metal ions (Tani
et al., 2004). The adsorption efficiency thereby proceeded in the order Co(II) >
Zn(II) > Ni(II) on the biogenic products (figure 7). It is possible that the metal
ions are incoorporated into the crystal strucure of the manganese oxide (Tani et al.,
2004).

18 Role of biogenic manganese oxides in soil remediation
Figure 7: Efficiency of sorption of Co, Ni and Zn on biogenic and synthetic manganese
oxides as a function solid manganese (all closed lines, closed symbols: biogenic,
open symbols: synthetic) and specific surface area (synthetic only, dotted
lines, closed symbols). (Tani et al., 2004).
Figure 8: Extractability of manganese oxides with CuSO 4 . The grey bar gives
the extractability in %. Colums A) B) and C) show special treatments. A: no
treatments, B: 20mM HEPES buffer solution, C: 100 mM NaCl before injection of
metal ions. Initial Mn(II) concentration: 0.2mM. (Tani et al., 2004).

19
One big disadvantage of heavy metal scavenging with biogenic manganese oxides
is oxidability of Cr(III). Cr(VI) is soluble and much more toxic than the insoluble
Cr(III). Manganese oxides are known to be the only oxidants for Cr(III), they are
capable of oxidising Cr(III) to Cr(VI). Additionally, they oxidise Cr(III) approximately
up to seven times faster than synthetic manganese oxides. Manganese oxides
are reduced to Mn(II), which makes the Mn(II) bioavailable again for manganese
oxidation. Thus a small concentration of Mn(II), together with biogenic oxidation
of manganese, can form a lot of soluble Cr(VI) in a cyclic process. When using
biogenic manganese oxides for scavenging heavy metals, the toxicity of Cr(VI) in
the soil has to be considered (Murray and Tebo, 2007; Wu et al., 2005).
6 Conclusion
Biogenic manganese products can be a useful tool in soil remediation since they
can be produced right at the contaminated site and surpass commercial manganese
oxides in the order of magnitudes regarding sorption. Before application of biogenic
manganese oxides in soils the toxicity of chromium has to be analysed.
To fully understand, how the application of manganese oxidising bacteria influence
the soil environment, the following questions should be considered: The terms
for big scale engineering should be analysed. Also more studies on the reversability
of heavy metal sorption are needed. Bacteria, that oxidise manganese, can have a
great impact on the redox environment in nature. The possibility of the storage of
Mn(III) and Mn(IV) ions on the bacterial membrane should be analysed, including
the possible proteins, responsible for storage and intracellular transportation. For
further understanding of manganese oxidising bacteria, studies on manganese oxidising
microbial systems directly in nature should be done. Also the efficiency of
different bacteria strains in producing heavy metal sorbents should be discussed.

20 REFERENCES
References
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metabolically dormant spores of diverse bacillus species. APPLIED AND ENVI-
RONMENTAL MICROBIOLOGY, 68(2):874–880.
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and Crozier, E. D. (2004). The structure of the manganese oxide on the sheath
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OGIST, 89(7):1110–1118.
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OF PAPERS OF THE AMERICAN CHEMICAL SOCIETY, 227:U1213–U1213.
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Nelson, Y. M., Lion, L. W., Ghiorse, W. C., and Shuler, M. L. (1999). Production of
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NOLOGY, 36(3):421–425.
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Carbon nanotube
http://www.nccr-nano.org/ncc
Mobility behaviour of
anthropogenic
nanoparticles in the
hydrosphere
Term paper in Biogeochemistry and
Pollutant Dynamics
arg
arg
http://ina.unizar.es/
by Sarah Conradt
Submission date: 22th June 2007
Tutor: Prof. G. Furrer and Prof. B. Wehrli
C 60 Fullerene
Abstract
Nanotechnology is often referred to as a key technology of the 21 th century. However few is
known on the behaviour of nanoparticles and their potential risks to human health and the
environment, especially to the aquatic environment. This paper deals with the mobility of
nanoparticles in the aquatic phase. Mobility is crucial for determine the fate and finally the
risk emanating from these particles. This review is based on current research on lab studies,
which used silica beads in a column to simulate porous media and Erlenmeyer flasks to
investigate the influence of natural organic matter (NOM) on nanotubes. NOM seems to
stabilize suspended nanotubes to a great extent. The breakthrough behaviour of different
nanoparticles was deviating. Particularly hydrophobicity, size and the special alignment
behaviour seem to be influential parameters for deposition and aggregation of nanoparticles.

1. Introduction
Small, smaller and even smaller – till nothing is visible anymore to the naked eye – but still
smaller! A way down to the microcosm. Welcome in the world of dwarfs, welcome in the
nano world!
It was already in the year 1959 when Richard Feynman, an U.S. physicist and later Nobel
price winner, declared: “There’s plenty of room at the bottom”! With this speech he outlined a
vision of a miniaturized world, which plays at the very end of the size scale with new
promising technical applications and possibilities. This miniaturization found its way into our
daily life in many areas like chips, storage devices and integrated circuits. The scanning
tunnelling microscope (1981) was one outranking step towards this little world, because it
allowed a glance at the nanoscale and confirmed that there is not only room but plenty of it,
as Feynman mentioned. Within this little world, the nano world and its constituent, the
nanoparticle will be looked at in this paper.
‘Nano’ stands for part in a billion, thus 1nm equals to 10 -9 m. To get a better idea of the size,
here a short comparison: in only one human hair 50 000 particles of 1 nm size could be
arranged abreast (SWR2). Size is often used as the only criterion to delineate nanoparticles
from bulk materials and colloids are often used as a synonym (Lead and Wilkinson 2007:2).
In compliance with the IUPAC definition, it is sufficient that only one dimension of a particle
fulfills the size criterion (Table 1). As a consequence so called nanotubes, which consist of
long tubes (nanoscale diameter but microscale length), are classified as nanoparticles.
Hyung et al. (2007:179) described nanotubes as “pure carbon macromolecules consisting of
sheets of carbon atoms covalently bonded in hexagonal arrays that are seamlessly rolled
into a hollow, cylindrical shape with both ends rounded through pentagon rings inclusions”.
Carbon nanotubes (CNTs) are further distinguished between single-walled nanotubes
(SWNTs) and multi-walled nanotubes (MWNTs), whereas “the latter results from a coaxial
assembly of the multiple SWNTs” (Hyung et al. 2007:179) . Another important form of carbon
containing nanoparticles are the so called Buckminsterfullerenes (C 60 ). In contrast to
nanotubes, fullerenes are arranged in a spherical configuration, containing both five and six
carbon rings, with a total number of 60 C atoms (Cheng 2005). Fullerols are
polyhydroxylated C 60 nanoparticles which are modified to increase the affinity to the aqueous
phase. Often, nanoparticles are described to represent “an intermediate supramolecular
state of matter between bulk and molecular material” (Moore 2006:967).Thereby
nanoparticles are differentiated from bulk material by having not only a specific small size but
also “novel properties”. Novel properties mean that new physical laws or properties of
materials may evolve by making particles smaller and smaller (Table 1). This is a result of
the increasing surface to volume ratio, which leads to extremely high reactivity as the contact
area increases. For example aluminium in beverage cans is harmless but minimized to nano
size, aluminium is an explosive matter which is used as catalyst for missiles (Technology
Assessment).
Colloids are ubiquitary; they are present in every environment and they can be either
anthropogenic in origin or occur naturally. An example for the latter is the famous Lotuseffect,
where a special nanostructure on the surface of the plant causes so small adhesion
forces that the cohesion forces prevail, thus wetting cannot take place and the surface
remains clean (Technology Assessment).
In the following work only anthropogenic nanoparticles are considered. Released from e.g.
industrial sites, they are transported via the atmosphere, are deposited on soils and water or
enter directly the aquatic environment. Subsequent uptake of nanoparticles by endocytosis
or other mechanisms into the aquatic biota is of major concern (Moore 2006).
2

Table 1: Various definitions of nanoparticles and colloids
Author Nanoparticles Main criterion
(a) IUPAC
(provisional
recommendation)
(b) British
Standards
Institution
London
(BSI) 2005
“Microscopic particle whose size is measured in
nanometers”.
“Particle with one or more dimensions at the
nanoscale”.
Nanoscale: “having one or more dimensions of the
order of 100 nm or less”.
size
size
Frequently (small) colloids
are used synonymously to
nanoparticles which is also
done in this work.
Author colloids Main criterion
(a) Filella 2007:18
(b) IUPAC 1972
(c) Leppard
1992:233
(d) Filella 2007:17
(e) Lead and
Wilkinson
2006:160
(f) Lead and
Wilkinson 2007:2
(g) Lead and
Wilkinson 2007:2
“Any material which can be brought into a colloidal
solution is nowadays often referred to as a ‘colloid’.”
Colloidal: “a state of subdivision, implying that the
molecules or polymolecular particles dispersed in a
medium have at least in one direction a dimension
roughly between 1 nm and 1µm.”
“An appreciable fraction of the molecules of a colloid
is located at the boundary region between particle and
aquatic milieu.“
Particles are defined as colloids if their surface free
energy dominates the bulk free energy, which requires
a high surface to volume ratio.
“The colloidal fraction can be considered as a
discrete, non-aqueous phase for contaminant binding,
which does not sediment over reasonable timescales
in the absence of further aggregation”.
Brownian motion is in general sufficient to keep
aquatic colloids suspended in the water.
Colloids: “any organic or inorganic entity large enough
to have a supramolecular structure and properties that
differ markedly from those of the aqueous phase
alone, e.g. possibility of conformational changes or
development of an electrical surface field.”
(aqueous) solution
size
location
energy regime
(surface to volume
ratio)
stability
stability (Brownian
motion)
new properties
(surface to volume
ratio)
3

Nanoparticles are applied in many areas: in medicine, in common consumer products like
sunscreens or cosmetics, as catalysts or reductants. Nevertheless, not much is known about
nanoparticles in general. Most research in the past was concentrated on either aerosolic
nanoparticle 1 or on the impact and toxicology of nanoparticles on human health. To do so, it
is important to consider particle attributes like size, shape, structure, composition and
homogeneity (Burleson 2004). As Moore (2006:967) bewails “release of manufactured
nanoparticles into the aquatic environment is largely an unknown”.
That’s why this work will deal with anthropogenic nanoparticles in the hydrosphere, with main
focus not on toxicity but on the particle behaviour and its mobility.
Mobility is an important factor to consider as it shows the affinity of particles to aggregate and
settle down or to stay dispersed in solution. Thus, mobility determines the longevity of
particles, which is a relevant criterion for subsequent toxicology assessment 2 . Natural
organic matter (NOM) plays thereby a key role: “NOM is ubiquitous in natural aquatic
systems and adsorbs on most colloidal particles. Adsorbed NOM (…) can have a significant
effect on colloidal stability and consequently can affect the transport and fate of colloids in
aquatic environments” (Filella 2007:24). NOM is one parameter whose influence on
nanoparticles, namely nanotubes, will be considered in the following discussion in more
detail.
There exist many factors, which influence the behaviour of any particle in an aqueous phase.
Filella (2007: 21ff) focuses on four central properties: particle size, charge, its ability to
coagulate and porosity or the fractal dimension.
First, bigger (aggregated) particles tend to settle down more easily. Thereby, they are
removed from the aquatic system. It is assumed by Filella (2007:22) that the collision
frequency in a nanosystem is substantially higher than for larger particles, thus unexpected
aggregation of nanoparticles and subsequent settlement could diminish their dissolved
aqueous concentration.
Second, surface charge determines the attractive or repulsive interactions of particles. The
charge of a surface can be measured via the zeta potential, which is connected with the
electrophoretic mobility through the Henry equation (Figure 1).
Third, the rate of colloid aggregation is dependent on the frequency and the efficiency of
particle contact. This is associated with the electrostatic repulsion (Figure 1), which can be
described by the Derjaguin, Landau, Verwey and Overbeek theory (DLVO). Assuming that
particle stability depends upon the potential energy function (V T ), this theory takes the
electric double layer repulsion (V R ) and the attractive van der Waals forces (V A ) into account.
Particularly important for these interactions are parameters like zeta potential, particle size,
electrolyte concentration or Hamaker constant (Elimelech et al. 1995). The interaction of
these forces determines the characteristics of the energy curves like for example the
appearance of a secondary minimum.
Finally, particle or aggregation growth changes the density and composition of particles as
fluid is taken into pores. An increase in the cross-section increases the contact between
particles as well as the property of the whole aggregate.
1 “A great deal of research over the last decade has been dedicated to studying the health effects of
airborne ultrafine particles” (Burleson et al 2004:2708).
2 For instance Colvin (2004:1166) argues that it is essential to characterize the expected concentration
of engineered nanoparticles that may be present in the environmental compartment before interpreting
toxicology data.
4

Figure 1: Assumed relationship of some relevant values for aggregation and deposition of colloids.
The Henry equation is defined as: U E =
2 f(k a)
3η
U E = electrophoretic mobility; z = zeta potential; = dielectric constant; = viscosity; f(k a) = Henrys
function (often approximated by 1 (Huckel) or 1.5 (Smoluchowski).
In the following section, the used methods of the three key papers are explained. Afterwards
the main results are shortly listed and summarized as an overview in Table 2. Then, these
findings and observations are discussed in more detail and subsequent some conclusions
are drawn.
2. Reviewed literature and encountered experimental methods
2.a Literature search
This work will be based mainly on three papers, which were found via the Web of Science.
After the first search it was obvious that there are only few papers dealing with nanoparticles
in the hydrosphere. For example the search for “nanoparticle AND hydrosphere” yielded no
results. Thus, a search strategy had to be worked out. Therefore, an excel sheet was
prepared containing four columns: the entered search criterion, a column to be ticked if the
search was restricted to the title, the number of hits and finally the references of potential
interesting articles. This allowed a well-arranged search and reduces the risk to go round in
circles. It turned out that on the one hand, the restriction to “title search” was in general not
leading to results but on the other hand, the search criterion had to be more specific without
this restriction. The combination of terms, like “nanoparticle* AND water AND (aquat* OR
aqueou OR hydrosphere)” resulted in a manageable quantity of hits. After some papers were
found, the references of these papers were also included in the further search.
2.b Experimental methods
All studies were conducted in lab experiments using synthetic and natural water. Hyung et al.
(2007) used Erlenmeyer flasks in which the different MWNT solutions were analysed. The
degree of suspended nanotubes was measured by using Thermal Optical Transmittance
(TOT). This method is adjuvant as organic carbon, derived from NOM, exhibit different
stability properties than the elementary carbon stemming from nanotubes.
5

Lecoanet and Wiesner (2004b) and Brand et al. (2005) used batch systems in their
experiments. Porous media was simulated by silica glass beads, which were filled in a
cylindrical acryl column. The flow behaviour of a solute through a soil column varies,
depending on the interaction of the soil matrix and the solute. Breakthrough curves (BTCs)
provide information on this behaviour, whereas the relative concentration of the solute
measured in the effluent is a function of the pore volume, which is the ratio of effluent volume
to the total pore volume of a column (V/V p ) (Figure 2).
Figure 2: Typical breakthrough curves (Modified after Scott 2000 und Hartge 1978).
The dashed line one (1) represents only a theoretical line, requiring the same flow velocity in the whole
pore system, a conservative 3 behaviour of the dissolved solute and assuming no dispersion. The
effluent concentration remains zero until one pore volume is reached, where the relative concentration
is discharged all at once. This type of flow cannot be observed in natural environments and is referred
to as piston or plug flow (Scott 2000).
Assuming further on a conservative behaviour but including the process of dispersion, results in line
two (2). The solute can be detected in the effluent before one pore volume is discharged but does not
reach the value of one at this point. The result is a S-shaped curve, whose slope decreases with
decreasing velocity or increasing flow distance (Hartge 1978).
If sorption occurs, the curve is retarded as shown by line three (3). When all possible sorption sites are
occupied, the relative concentration reaches one.
This is not the case if the solute is transformed or degraded or is precipitated (line four (4)) (Hartge
1978).
3. Results
This review is based mainly on three papers, whose results are shortly summarized in the
following section (Table 2).
Lecoanet and Wiesner (2004b) compared the transport and deposition properties of two
oxide nanomaterials in a porous media with three aqueous suspensions of fullerene-based
nanomaterials.The porous media was simulated by a cylindrical column filled with spherical
glass beads to which two different flow rates were applied. A theoretical approach for
deposition was established. Colloidal particle are first transported to an immobile surface
where they are attached. This process can be quantitatively described by a constant
3 Conservative means that no sorption and no microbial or chemical transformation occurs (Scott
2000).
6

attachment efficiency factor 4 . Based on observations on particle behaviour in porous media
and under a couple of assumptions, an equation to estimate this attachment efficiency factor
was derived. It was assumed that higher velocity rates lead to a higher passage of
particles.
2d
α = − ln( c / co)
3(1 − ε ) η 0L
α = attachment efficiency factor; d = diameter of a collector which is assumed to be spherical; c/co =
concentration fraction of influent particles remaining; ε = porosity; η 0 = clean bed single-collector
efficiency; L = length of porous media.
Silica nanoparticle (for which retention by the media was marginal) and anatase nanoparticle
complied with the theoretical deposition predictions. This was not the case for fullerenebased
nanoparticle where no velocity dependency was observed and consequently an equal
effluent value for both velocities was reached. For these materials, an affinity transition was
detected during the second pore volume but only at higher Darcy velocity. Fullerol
nanoparticles were only slightly retained by the porous media. To compare the results and to
estimate the hydraulic characteristics of the column all experiments were repeated twice with
a tracer, namely sodium chloride. In a further experiment, the influence of the initial influent
concentration was investigated for fullerol and SWNT by adding once only about 1/3 of the
initial concentration of 10 mg/L. Thereby it was observed that the affinity transition decreased
with lower input concentration.
Brant et al. (2005) investigated the origin of fullerene stability as well as the aggregation and
deposition properties with varying ionic strength in an indifferent electrolyte solution. In pure
water, the three different n-C 60 size fractions used showed only a marginal aggregation
tendency. For the larger n-C 60 nanoparticle (size 168 nm), further investigations with varying
ionic strength were carried out. Already a small addition of salt led to the formation of supraaggregates,
qualified as m-(n-C 60 ), followed by increased settlement of these colloids.
Stronger ionic strength even reinforced this effect. Particles remaining in suspension showed
a narrow size distribution, whereby the average diameter increased with increasing ionic
strength. A different picture arose for the highest investigated strength of 1 M NaCl. The size
distribution was broader and two peaks were observed. By increasing the ionic strength the
zeta potential of fullerenes decreased. Experiments in porous media, simulated by silicate
glass beads packed in a column, point out that with increasing ionic strength, the retention of
n-C 60 increased. Similar to the findings of Lecoanet and Wiesner (2004b), an enhanced
retention affinity during the second pore volume was observed temporally, however
retrogressive for higher ionic strength.
Hyung et al. (2007) investigated the influence of NOM on the stabilization of MNTW in an
aqueous solution. Therefore MWTN was added to three different samples: One sample
contained natural river water with a high content of NOM (Suwannee River), another sample
comprised a synthetic derived NOM solution to model natural NOM (referred to as SR-NOM)
and for the last sample the surfactant sodium dodecyl sulphate (SDS) was applied. The
concentration of suspended MWNT was quantified by using Thermal Optical Transmittance
analysis. Hyung et al. observed a varying degree of MWNT stabilization among the different
samples. The dispersion for samples containing SR-NOM or natural river NOM were similar
but substantially higher compared to SDS. Further, the findings showed that higher amounts
of NOM (natural derived or modelled) resulted in increased suspension of MWNT, which
were present to a large extent in form of single tubes. Association of MWNT to NOM
increased as the relative abundance of NOM increased.
4 “The attachment efficiency is a function of numerous phenomena including van der Waals forces,
electrical double-layer interactions, steric interactions, hydration forces, and particle/surface
hydrophobicity”. (Lecoanet 2004: 5164).
7

Table 2: Overview of the three key papers.
Experimental Results
methods
Author
and
investigated
materials
Lecoanet and
Wiesner
2004b
Silica
Anatase
Fullerene
Fullerol
SWNT
(single-walled
nanotubes)
Brant et al.
2005
Fullerene
Hyung et al.
2007
MWNT
(multi-walled
nanotubes)
Appliance of two
different flow
velocities for a
porous media
simulated by a
silicate glass filled
column.
Appliance of
variable ionic
strength to a
porous media
simulated by a
silicate glass filled
column using
indifferent
electrolyte solution.
Investigation of the
influence of NOM
on nanoparticle
behaviour.
Use of three
different samples:
1) unaltered river
water with a natural
NOM content
2) SDS solution
3) modelled NOM
(SR-NOM)
• Oxide nanoparticles followed
deposition theory but not so
fullerene based materials.
• C 60 particles showed a transitional
enhancement in retention affinity
during the second pore volume
and deposition converged to a
flow velocity independent level.
• Increasing ionic strength reduced
the zeta potential and induced the
formation of supra-aggregates
(m-(n-C 60 )), which resulted in
enhanced settlement.
• For ionic strength below 1 mg/l
NaCl, suspended colloids showed
are narrow size distribution.
Above another behaviour was
observed.
• In the porous media a temporary
increased deposition affinity was
observed during the second pore
volume.
• River derived NOM and SR-NOM
are better stabilization agent as
SDS.
• Suspension of MWNT as single
tubes
• Suspension of MWNT and
association of NOM to MWNT
increased with increasing NOM.
Open questions and
problems
Discussion of several
explanations for the
deviant behaviour of
fullerene-based material
(e.g. transitional
enhancement in affinity
for the media).
Uncertainty of the origin
of colloidal charge.
NOM is highly variable.
Further research should
take these differences of
NOM composition into
account.
4. Discussion
In the following section, some breakthrough curves of diverse nanoparticles determined by
Lecoanet and Wiesner (2004b) and Brant et al. (2005) are analysed and compared to those
of Figure 2.
Silica
The breakthrough curves of silica at both Darcy velocities were nearly similar to those of the
tracer sodium chloride, meaning a marginal retention by the porous media. Already after two
pore volumes, the relative concentration reached nearly 0.9 (Figure 3). Compared to Figure
2, one should expect a steeper slope for the higher velocity, but this velocity effect was
negligible as total removal was humble.
Size is an important factor for the behaviour of any particle. Larger silica particles (135 nm)
showed less mobility in experiments by Lecoanet et al. (2004:5167 Figure 4) and the effluent
concentration reached after two pore volumes was about the half compared to the smaller
silica particles (57nm). This result is consistent with DLVO theory because “interaction
8

energies are directly proportional to particle size” (Elimelech et al. 1995:61 5 ). Thereby greater
size implies stronger secondary energy effects, which favour aggregation. Additionally it can
be also observed in Table 3 that the distance to reduce the relative concentration C/C in of the
smaller silica particle to 0.1% was about 10 times longer than for the bigger silica particle.
Thus, smaller particles have higher mobility.
Figure 3: Breakthrough curve of silica at two Darcy velocities: “Slow” corresponds to a flow velocity of
0.04cm/s, “Fast” to 0.14cm/s (Lecoanet and Wiesner 2004b).
Anatase
As expected, the slope was steeper for the higher Darcy velocity. The striking observation for
these curves is that neither for the high nor for the low velocity, C/C in reached the value one
during five pore volumes. Anatase has a low water solubility (Milnes and Twidall 1983) and
thus a great affinity to the porous media. The small value of electrophoretic mobility (Table
3), which implies less repulsion and a smaller energy barrier for attachment (Lecoanet and
Wiesner 2004b), intensifies this affinity to the column material.
Calculations were made to get an idea of the fraction of occupied porous volume V p through
anatase (Figure 4b). After five pore volumes, less than 0.0014% of the total porous volume
was engaged by the volume of anatase particles. Thus, it can be assumed that extending the
experiment to substantial higher pore volumes, could lead to the same effluent and influent
concentration (C/C in =1). This marginal increase of the C/C in ratio appears on Figure 4a as a
“plateau” because only few pore volumes were discharged.
Anatase:
Data:
V = 100 ml (selected input)
V p = 0.02 l (Lecoanet and Wiesner 2004)
C in = 10 mg/l (Lecoanet and Wiesner 2004)
= 3.8 g/cm 3 (Table 3)
V occ = ? (with anatase occupied volume)
C in V = 1 mg
V occ = m / = 2.63*10 -7 l → 0.0013%
After 5 pore volumes, about 0.0013% of
the porous volume is occupied by
anatase.
Figure 4a: Left: Breakthrough curve of anatase at two Darcy velocities: “Slow” corresponds to a flow
velocity of 0.04 cm/s, “Fast” to 0.14 cm/s (Lecoanet and Wiesner 2004b).
Figure 4b: Right: Calculation of occupied porous volume for anatase after 5 pore volumes.
5 Although Elimelech referred this statement to particle being greater than about 1µm in diameter.
9

Table 3: Some characteristics of the considered nanoparticles (Lecoanet et al. 2004a; Lecoanet and
Wiesner 2004b)
particles
size (nm)
density
(g/cm 3 )
electrophoretic
mobility
10 -8 (m 2 s -1 V -1 ) pH zpc log
distance to
reduce C/ C in
to 0.1% (m) b
fullerol 1.2 1.69 not detectable 2.3 -3.98 14
SWNT
0.7-1.1 c * 80-200
( dh = 21 nm d ) 1 -3.98 2.2 -2.89 10
silica 57 2.65 -1.95 2 -2.1 2.4
silica 135 -2.58 -0.77 0.2
n-C 60 168 1.41 -1.99 2.3 -0.52 0.1
anatase 198 3.8 -0.27 6.1 -0.47 0.1
b Conditions assumed for calculations: T) 15 °C, H) 10-20 J, Darcy velocity = 0.003 cm/s, porosity = 0.30, mean
sand grain diameter = 350 µm.
c According to the model cross-section of an individual fullerene nanotube encased in a close-packed cylindrical
surfactant micelle, the outer diameter of this nanomaterial is close to 4 nm with a specific gravity of approximately
1.0.
d Average hydrodynamic diameter.
Fullerene-based materials
Fullerenes are extremely hydrophobic, thus prone to interact with the porous media and
unlikely to be dispersed in the aqueous phase. The plateau value of C/Cin = 1 was not
reached during five pore volumes. This observation can be explained in a similar way as for
anatase above. The influence of hydrophilicity on the mobility in the aqueous phase can be
seen well in Figure 5. The comparison of the distance to reduce the relative concentration to
0.1% was substantially higher for fullerols (14 m) than for fullerenes (0.1 m), yielding a factor
of 140 of increased mobility for the more hydrophilic fullerols.
Three striking differences from Figure 2 can be observed for all three fullerene-based
materials: First, during the first and second pore volume a transitionally increased affinity to
the porous media was discovered. Second, a similar plateau value was observed for both
velocities. Third, the effluent concentration at high Darcy velocity was (nearly) all the time
lower than those at lower Darcy velocity.
Lecoanet and Wiesner (2004b) explained this latter observation by the special structural
alignment of SWNTs, comparable to a “rope”. This implies an ideal velocity, for which the
arrangement can be formed optimally. According to this, above und below this “ideal”
velocity, alignment of nanotubes should be retarded. A similar process is conceivable for
fullerene. Transmission electron microscopy suggests that C 60 growths through the formation
of crystalline structures (Chen 2006, Brant et al. 2005). This “crystalline growth” could also
imply an optimal flow velocity for proper alignment. According to this, the “Fast” velocity
(Figure 5) was more close to this assumed “ideal” velocity.
Lecoanet and Wiesner (2004b) discuss three explanations why a similar plateau for both
velocities was reached: First, they assumed that the size of aggregates changed with
velocity. This assumption contradicted with measurements and was rejected. Second, they
supposed that the attachment efficiency factor increased with velocity in order to get the
same effluent concentration. Thereby the fluid flow should transfer forces to the particle to
overcome an assumed attachment barrier. As possible forces are rather small, this seems to
be unlikely. Finally, Lecoanet and Wiesner (2004b) proposed that deposition was not mass
10

transport limited, however another step limited the attachment rate. They assumed the highly
ordered growth of nanotubes and fullerenes to be limiting. This explanation would
correspond to the above mentioned “crystalline growth” of fullerene and “rope” like alignment
of nanotubes. To reason from these findings, the structural growth of nanoparticles seems to
be essentially important.
The dip during the second pore volume could not be explained satisfactorily. As this effect
occurred not before the discharge of one pore volume, it can be suggested that the
nanoparticles itself alter the collector surface in a way that further attachment is favoured.
Consideration of the forces between the media and fullerene and fullerene to fullerene
support this assumption. The Hamaker 6 constant between the nanoparticles itself was
calculated to be 10 times higher than the Hamaker constant of fullerene-silica-water 7 . The
degree of nanoparticle coverage of the collector was calculated by Lecoanet and Wiesner
(2004b) after the discharge of one pore volume. The result was just about 1% modified
surface, which was not sufficient to explain the drop in the curves. Additionally, it does not
explain why the phenomenon was not observed at lower Darcy velocity. Perhaps this
phenomenon was also a consequence of the “ordered alignment” or specific growth
behaviour of these nanoparticles. However, this was neither further investigated nor
discussed in the paper by Lecoanet and Wiesner (2004b).
Figure 5: Breakthrough curve of fullerene (left) and fullerol (right) at two Darcy velocities: “Slow”
corresponds to a flow velocity of 0.04cm/s, “Fast” to 0.14cm/s (Lecoanet and Wiesner 2004b).
Brant et al. (2005) made similar column experiments using two ionic strengths (0.001 M and
0.1 M) 8 (Figure 6). The breakthrough curve of lower NaCl concentration was at all times
above the curve of higher ionic strength, meaning that for the former less deposition
occurred. This detection is conformable with DLVO theory. As shown in the inset of Figure 6,
greater ionic strength intensifies the appearance of a stronger secondary minimum. A strong
(secondary) minimum is favourable for aggregation. This can be explained by the balance of
power of involved forces: Higher ionic strength reduces electrical repulsion (meaning lower
energy barrier), thus fosters the convergence of particles which favours the attractive Van
der Waals forces to dominate (Elimelech et al. 1995) (Figure 1). It can also be seen in Figure
6 that the transitional affinity was reduced by applying a higher NaCl concentration. To
conclude from these observations, electrostatic repulsion could be limiting for low ionic
strength conditions, where the retention is transitionally increased (Brant et al. 2005).
6 “The Hamaker constant characterizes the resonance interactions between electronic orbitals in two
particles and the intervening medium” (http://www.erpt.org).
7 Hamaker constants: fullerene-silica-water = 2.76*10 -20 J; fullerene-fullerene = = 2.35*10 -19 J
(Lecoanet and Wiesner 2004: 4380).
8 Lecoanet and Wiesner (2004) applied in their experiments an ionic strength of 0.01 M.
11

Figure 6: Breakthrough curve of C 60 for two different NaCl concentration (0.001M and 0.1M),
applied Darcy velocity probably 0.14 cm/s 9 . The inset on the left shows the DLVO interaction energy
profiles between n- C 60 cluster and the spherical glass collector for three different ionic strength
(IS) = 01, 0.01, 0.001 M. The interaction energy is plotted versus the surface-to-surface separation
distance h in nm. (Modified after Brant et al. 2005).
Chen and Elimelech (2006) showed that divalent CaCl 2 (Ca 2+ ) electrolyte solutions were
even more effective in reducing the energy barrier between particles than monovalent
solutions (Na + ). Thus the electric repulsion or electrophoretic mobility decreased to a higher
extent by using Ca 2+ electrolyte concentration and aggregation was reinforced (Figure 1).
In further experiments, Lecoanet and Wiesner (2004b) investigated the influence of the initial
concentration at the higher Darcy velocity (0.14 cm/s) for fullerol and SWNTs. Thereby the
affinity transition to the collector surface decreased at lower concentration (3.5 mg/l instead
of 10 mg/l) (Figure 7). This observation supported the presumption that initial nanoparticle
deposition favoured subsequent attachment and was responsible for the drop in the effluent
concentration. According to this, lowering the initial concentration would reduce these special
aligned growing sites and consequently reduce the affinity transition.
Figure 7: Breakthrough curve of fullerol at high Darcy velocity, meaning 14 cm/s, for two different
initial concentrations (Lecoanet and Wiesner 2004b).
9 It was not clear which Darcy velocity was finally applied. Brant et al. (2005) specified on page 547 an
applied Darcy velocity of 0.04 cm/s whereas in the caption of the figure 7 on page 550 a Darcy velocity
of 0.14 cm/s was mentioned. These velocities correspond to “Slow” and “Fast” in the figures of
Lecoanet and Wiesner (2004b). The authors replied not yet to the made inquiry.
12

It can be concluded from these findings that aggregation and subsequent deposition could
limit the mobility and hence the risk emanating from nanoparticles. It was stated by Brant et
al. (2005:545) that through the formation of large aggregates the “risk presented by n-C 60
toxicity due to a reduced potential for exposure” is (partially) offset. This conclusion was
relativised further on by referring to natural environments because they contain compounds
with the capability to mobilize other compounds. One important class of these substances is
natural organic matter (NOM) as it plays a key role in natural environments. It is abundant,
influences surface properties and the stability of particles, affects transport and the fate of
colloids. Thus, it is important to investigate as well the effect of NOM on nanoparticles in the
aqueous phase. As nanoparticles are a heterogenic and broad class of substances, the
influence is exemplarily discussed for nanotubes by referring mostly to the paper of Hyung et
al. (2007), who studied the impact of NOM on multiple walled carbon nanotubes (MWNT).
The influence of natural organic matter
Nanotubes are expected to aggregate to a large extent as they are highly hydrophobic and
form strong Van der Waals forces due to the fact that their axial length can exceed the
micrometer range (Hyung et al. 2007).
Hyung et al. (2007:182f) observed two striking effects of MWNT suspension and association:
First, suspended MWNT increased either by increasing the NOM content or by increasing the
addition of MWNT itself, while conserving the initial NOM content.
Second, MWNT association to NOM increased either by increasing the dose of NOM or
decreasing the initial concentration of MWNT. This means that association between NOM
and MWNT was superior if the relative abundance of NOM was increased.
What are possible reasons, which explain that NOM enhanced MWNT suspension?
Filella (2007:25ff) suggests that adsorption to NOM, bearing a high content of acidic groups,
can produce negative surface charge to the colloids. Electrostatic repulsion and formation of
repulsive double-layer forces by this charge may stabilize particles in natural waters.
Additionally, some NOM have a bulky form and high molecular mass. Hence, stability of
adsorpt nanoparticles may be enhanced through steric repulsion. Hyung et al. (2007:182f)
concluded that for the formation of MWNT-NOM suspension, two aspects have to be
considered: First “the physical dispersion of MWNT added to solution” and second the
association or “dynamic equilibrium processes, similar to adsorbate-adsorbend interactions”
of NOM to MWNTs “rendering them stable in aqueous phase”. In his study, Hyung et al.
(2007) compared the adsorption efficiency of SDS, a surfactant, with modelled and natural
organic matter. The data suggests that SDS is less effective in formation of dispersed
nanotubes than both other (natural and model) forms of NOM. Accordingly, divers properties
of the surfactant and NOM must provoke this difference in stabilization capability. SDS is an
aliphatic substance whereas NOM contains aromatic groups. The – electrons
interactions of these aromatic parts and MWNT could lead to a stabilization effect (Hyung et
al. 2007). Nanotubes are stabilized intramolecularly by having delocalized -electrons over
the whole molecule. These -electrons can interact weakly with aromatic parts of NOM. The
sum of many such weak polarization interactions could lead to considerably stabilized
aggregates of NOM and nanotubes. Futher, NOM could shield the hydrophobic nanotubes
from contact with water by directing outwards to the aqueous phase their polar tails.
“However, the exact mechanism for CNT interaction with NOM will depend on both NOM
characteristics including aromaticity, charge density, and size as well as CNT characteristics
such as aspect ratio (e.g., SWNT) and functional derivatization (…)” (Hyung et al. 2007:183).
Thus, a variety of parameters determine the degree of NOM and nanoparticle interaction and
the extent of CNT suspension in a complex manner. In this context, it must be stated that
NOM does not equal NOM: depending on the environmental compartment and the origin of
NOM, the characteristics (e.g. size, molecular mass, electric charge) and abundance of NOM
vary widely.The composition of NOM for example is important as fulvic compounds tend to
stabilize colloids whereas rigid biopolymers have the opposite effect (Buffle et al.
1998:2887). To take all these into account and to characterize more precisely NOM will be
13

ather challenging and a way of doing so in a feasible manner will be the task for further
investigations.
Nevertheless, NOM seems to stabilize nanotubes in the aqueous phase. Thus, aggregation
and adjacent settlement of nanotubes in the present of NOM will be reduced and the
longevity of these nanoparticles increased. Hyung et al. (2007:179) concludes that “dispersal
of carbon-based nanomaterials in the natural aqueous environment might occur to an
unexpected extent following a mechanism that has not been previously considered in
environmental fate and transport studies”.
5. Conclusions
Nanoparticles are produced through many anthropogenic or natural processes, for example
through vulcano explosions or industrial combustion processes. Released to the
environment, nanoparticles can be transported via the atmosphere, settle down on soils or
enter the hydrosphere. In the aquatic environment, the propensity to aggregate is one
important characteristic of nanoparticles. Aggregation leads to increased particle size which
enhances the tendency of settlement and thus the degree of particle removal from the
aquatic phase. Other important parameters are shape, hydrophilicity, charge, the
composition of nanoparticles or the environmental conditions like ionic strength, the presence
of potential collector surfaces and the pH conditions. All these parameters are linked in a
complex manner and determine the form and amount of present nanoparticles. These
parameters influence also the bioavailability and toxicity of nanoparticles. For example it can
be assumed that smaller particles are taken in unintentionally to a greater extent than
particles of bigger size. For environmental risk assessment one should consider both: the
amount produced and released to the environment and the potential threat (e.g. longevity) or
toxicitiy emanating from the different class of nanoparticles.
As it was shown by experiments of Lecoanet and Wiesner (2004), the breakthrough curves
of fullerene-based nanoparticles varied compared to silica and anatase breakthrough curves.
The observation of the enhanced affinity of these molecules to the porous media could be an
indication that the structural dimension or the proper alignement of the fullerene-based
nanoparticles could be an important factor for deposition behaviour. Another relevant
property is hydrophilicity as can be deduced from the experimental findings of Lecoanet and
Wiesner (2004). Hydrophobic fullerenes were considerably more prone to deposition and
about 140 times less mobile than their more hydrophilic counterparts, the fullerols (Figure 5,
Table 3). A further important parameter is the ionic strength. It was shown by Brant et al.
(2005) that increased ionic strength reduces the surface charge and thereby reduces the
barrier for attachment. Chen and Elimelech (2006) observed that the divalent solution of
CaCl 2 (Ca 2+ ) was much more efficient in reducing this energy barrier between particles than
the monovalent solution of NaCl (Na + ) . Regarding the environment, this is important as Ca 2+
is present in some aqueous environment in such a concentration that it could exert an
influence of the aggregation behaviour of present nanoparticles 10 . Calcium rich aqueatic
environment are, according to this, less threatened by nanoparticles as aggregation and
subsequent settlement could reduce their longevity. It should be mentioned that settlement
does not eliminate all risks of nanoparticles. Change in environmental conditions could
resuspend settled nanoparticles or organisms living or feeding on sediments could be
threatened as well. Natural organic matter, which is (nearly) omnipresent in the environment,
is reported to stabilize dispersed nanotubes (Hyung et al. 2007). Hence eutrophic aquatic
environments are more endangered by potential risks emanating from nanotubes than less
active and NOM poor ones. The enhanced degradation of organic matter in an eutrophic lake
e.g. could lead to a release of the stabilized NOM-nanotubes assemblies and accelerates
settlement. In this context, it is important to consider that NOM is also highly variable
10 The calcium concentration of Lake Zürich was reported to be about [Ca 2+ ] = 1.2*10 -3 mol/l (Filella
2007:28).
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(different NOM origin, composition, functional groups etc.). To make more reliable
predictions, the influence of the different NOM constituents should be considered separatly.
Further, NOM is only one constituent in the environment and does not allow to draw
conclusions to the influence of other organic compounds such as nitrilotriacetatate (NTA) on
nanoparticles. Also other potential collector surfaces like minerals are not considered in this
work.
It should be emphasised that nanoparticles are a broad substance class with many different
characteristics and properties as it could be seen by consideration of the varying deposition
behaviour in the porous media. Thus it cannot be assumed that NOM stabilizes other
nanoparticles to the same extent or the same manner than it does for carbon nanotubes. For
this reason no conclusions can be drawn form the behaviour of one nanoparticle class to
another and environmental impacts should be addressed rather on a case-by-case basis
(Lecoanet et al. 2004a:5168). Further, for all these experiments, it should be beared in mind
that they are lab experiments. Natural behaviour may be different to the observed behaviour.
Filella (2007:25) stated that for example the DLVO theory may not be applicable as natural
systems are too heterogeneous. Particle size and distribution, pore distribution and rough
matrix surfaces are some examples which lead to discrepancies to theoretical behaviour and
restrict the application of this theory. Additionally, the spatial heterogeneity may make the
measurement of the average surface charge (expressed by the zeta potential) useless for
mobility consideration (Filella 2007:24). In nature, conditions may change very rapidly. For
example a rainfall event could abruptly alter the water composition and influence the
deposition or aggregation behaviour of nanoparticles (Chen 2006). It should be also kept in
mind that columns containing silica beads are just an attempt to represent mineral surfaces
in groundwater. Probably this approach is sufficient for most of the groundwater systems but
digressive soil composition, like high content of clays, could lead to totally different
breakthrough curves. The transfer of lab observations to the “real“ natural environment is
often restricted as the environment is exposed to complex interactions of numerous
influencing factors.
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