Biogeochemical interactions between iron and sulphatein freshwater wetlands and their implications forinterspecific competition between aquatic macrophytes
MARLIES E. W. VAN DER WELLE,* ALFONS J .P . SMOLDERS,* , † HUUB J .M. OP DEN CAMP, ‡
JAN G.M. ROELOFS* AND LEON P.M. LAMERS*
*Department of Aquatic Ecology & Environmental Biology, Institute for Water and Wetland Research, Radboud University
Nijmegen, Nijmegen, The Netherlands†Research Centre B-Ware, Institute for Water and Wetland Research, Radboud University Nijmegen, Nijmegen, The Netherlands‡Department of Microbiology, Institute for Water and Wetland Research, Radboud University Nijmegen, Nijmegen, The
Netherlands
SUMMARY
1. Wetlands are threatened by desiccation, eutrophication and changing water quality,
generally leading to greatly altered biogeochemical processes. Sulphate pollution can lead
to severe eutrophication and sulphide toxicity, but may also interact with the availability of
iron and other metals.
2. In the present study, we examined the biogeochemical interactions between sulphate
and iron availability, and their effects on aquatic macrophytes, in a field experiment with
enclosures. The natural iron supply by groundwater was mimicked by adding iron to the
sediment, and the effect of increased sulphate concentrations in the surface water was also
studied. The enclosure experiment was performed in a mesotrophic, anaerobic ditch in a
peat meadow reserve in the Netherlands. In all enclosures, three Stratiotes aloides plants
were introduced to serve as indicator species.
3. Addition of sulphate led to the mobilisation of phosphate, whereas addition of iron or
both iron and sulphate did not affect P mobilisation. Growth of S. aloides was decreased by
both iron addition and sulphate addition (sulphide toxicity). Addition of iron under
sulphidic conditions, however, led to mutual detoxification of both toxicants (iron and
sulphide) and did not decrease S. aloides growth. The uptake of metals was highest in the
treatment involving sulphate addition, probably as a result of increased mineralisation of
the peat soil.
4. Growth of Elodea nuttallii, which grew naturally in the enclosures, was stimulated by
iron or iron plus sulphate addition. It did not, however, grow in the enclosures with
sulphate addition, as a result of sulphide toxicity or sulphide-induced iron deficiency.
Under iron-rich conditions, E. nuttallii appeared to be a better competitor than S. aloides
and depressed the growth of the latter species.
5. We propose that the growth of S. aloides is directly regulated by interactions between
sulphide and iron and indirectly by the effects of both compounds on the competitive
strength of E. nuttallii. In general, we conclude that biogeochemical interactions between
sulphate and iron can have a strong influence on plant species composition in freshwater
wetlands, because of direct effects or changes in the competitive strength of plant species
related to differential sensitivity to either iron or sulphide.
Correspondence: Marlies E.W. van der Welle, Adviesgroep Water & Ecologie, Rotterdam, Haskoning Nederland B.V., a company of
Royal Haskoning, P.O. Box 8520, 3009 AM Rotterdam.
E-mail: [email protected]
Freshwater Biology (2007) 52, 434–447 doi:10.1111/j.1365-2427.2006.01683.x
434 � 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd
Keywords: Elodea nuttallii, iron, Stratiotes aloides, sulphide, sulphur
Introduction
Wetlands are being threatened by desiccation and
eutrophication on a global scale (Mitsch & Gosselink,
2000). In addition, the chemical composition of
groundwater and surface water has changed, often
severely, as a result of anthropogenic pollution. It
has been shown that changes in the macro-ionic
composition of groundwater and surface water can
greatly influence the biogeochemistry of wetlands
(Ingram, 1967, 1983; Wheeler & Proctor, 2000).
Peatlands (both fens and bogs) are particularly
sensitive to these changes, as they have large
amounts of nutrients stored in the peat, which may
be mobilised by altered biogeochemical processes
(Lamers, 2001).
Increased sulphate input into freshwater wetlands
causes serious problems. Under waterlogged condi-
tions, sulphate will be reduced to sulphide, which can
be highly toxic to plants (Koch & Mendelssohn, 1989;
Koch, Mendelssohn & McKee, 1990; Armstrong,
Armstrong & Van der Putten, 1996; Smolders &
Roelofs, 1996; Van der Welle et al., 2006). Moreover,
sulphide binds to iron, which may cause iron defici-
ency in aquatic plants (Smolders, Nijboer & Roelofs,
1995; Van der Welle et al., 2006). In addition, sulphate
can lead to increased mobilisation of phosphate,
which may lead to serious eutrophication (Patrick &
Khalid, 1974; Bostrom, Jansson & Forsberg, 1982;
Roelofs, 1991; Lamers, Tomassen & Roelofs, 1998;
Lamers et al., 2001).
Hydrological changes can markedly alter the water
chemistry of wetlands (Schot & Van der Wal, 1992;
Beltman et al., 1996; Runhaar, Van Gool & Groen,
1996; Krebs, Corbonnois & Muller, 1999; Lucassen
et al., 2004, 2005; Smolders et al., 2006). Groundwater
is often a source of iron (Eser & Rosen, 1999; Lucassen,
Smolders & Roelofs, 2000a) or base cations like
calcium and magnesium (PiPujol & Buurman, 1997;
Lamers et al., 1999; Wheeler & Proctor, 2000). In the
type of wetlands we studied, groundwater is an
important source of iron. When the input of iron by
groundwater is blocked, as a result of regional or local
desiccation, this may lead to the accumulation of
phytotoxic sulphide and to eutrophication, as a result
of the processes described above.
The disappearance of Stratiotes aloides L., a species
that used to be very common in the Netherlands, is
thought to be related to the increased sulphur load in
Dutch wetlands and the concomitant eutrophication
and ammonium toxicity (Smolders, Roelofs & Den
Hartog, 1996a; Smolders et al., 2003). However, it
might also be related to decreased input of iron, or
perhaps a combination of decreased iron availability
and increased sulphate load. In the present study, the
role of iron and the interactions between iron and
sulphate were studied in a field enclosure experiment,
using S. aloides as an indicator species. In addition, we
studied the effects of sulphate and iron on metal
uptake by aquatic plants, as sulphur biogeochemistry
is also known to interact with other metals than iron
(e.g. Di Toro et al., 1992; Ankley et al., 1993; Besser,
Ingersoll & Giesy, 1996; Chapman et al., 1998; Huerta-
Diaz, Tessier & Carignan, 1998; Morse & Luther, 1999;
Wang & Chapman, 1999). Sulphur and iron biogeo-
chemistry can have a strong impact on plants, in
particular because both iron and sulphide can be toxic
and are interacting with each other (Van der Welle
et al., 2006, in press).
Methods
Site description
The experiment was carried out in a small nature
reserve close to Vinkeveen, the Netherlands (52�12¢N,
4�56¢E). The reserve consists of strips of peat meadow,
separated by ditches resulting from peat digging in
the past. Some of the fields are still in agricultural use
and are sparsely grazed by cattle or sheep. The nature
reserve was established to protect fen meadow birds
like the black-tailed godwit (Limosa limosa L.) and the
black tern (Chlidonias niger L.) and is managed by the
Dutch State Forestry Service (Staatsbosbeheer). The
water level is strictly regulated and does not vary by
more than 10 cm throughout the year. The aquatic
vegetation of the ditches is dominated by Nuphar lutea
(L.) Sm. Other commonly occurring species are Rumex
hydrolapathum Huds. and Iris pseudacorus L. The soil is
composed of a thick peat layer, which was probably
formed in an alder carr vegetation, as many remnants
of trees have been found in the peat. In the ditches, the
Biogeochemical interactions between iron and sulphate 435
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
peat layer is covered by a muddy layer of highly
decomposed peat.
During the 18th and 19th centuries, people tried to
improve the use of the peat meadows for agricultural
purposes by a process called ‘toemaken’. This basic-
ally meant that a mixture of manure, urban waste,
dredging sludge and sometimes sand was applied to
the land, resulting in a specific, anthropogenic layer
on the original peat soil, which is called a ‘toemaak-
dek’ (literally: covering layer) (Lexmond et al., 1987;
Bosveld et al., 2000). As a result of this process, the
area is diffusely polluted with metals originating from
urban waste.
Experimental set-up
Twelve polycarbonate cylinders of 1 m diameter
and 1.5 m height were inserted approximately
50 cm into the sediment of one of the ditches to
serve as enclosures (Fig. 1). A ceramic cup (Eijkelk-
amp Agrisearch Equipment, Giesbeek, the Nether-
lands) with attached tube was inserted into the
upper 10 cm of the sediment in each enclosure to
allow porewater sampling. Three extra ceramic cups
were inserted in the sediment outside the enclo-
sures, as an additional control measurement and to
test for possible effects of the enclosures themselves
(outside treatment).
The enclosures were randomly divided into four
groups. One group (Fe treatment) was treated once a
year with 50 g Fe m)2, which was carefully injected
into the upper 10 cm of the sediment using a
nebuliser. The second group (Fe&SO4 treatment)
received the same iron treatment in its sediment, but
additionally received 1.5 mmol L)1 Na2SO4 in the
surface water layer. The third group (SO4 treatment)
was treated with Na2SO4 only, and the fourth group
served as a control (control treatment). Na2SO4 was
added regularly to keep the sulphate concentration at
a constant level and prevent depletion. Due to logistic
problems, the SO4 treatment started 6 months later.
However, we were still able to monitor the effects
during two growing seasons.
After 4 weeks of acclimatisation, three S. aloides
(water soldier) plants were introduced in each enclo-
sure. The plants were collected from a nearby ditch
and their fresh weight, diameter, number of leaves
and buds and leaf width were measured to obtain
initial values before they were introduced. The plants
were collected and introduced during the winter
period, to minimise the effect of disturbance and to
allow acclimatisation before the growing season.
Three plants were kept apart to determine initial
nutrient concentrations.
Every month, porewater and water samples were
collected and the number of S. aloides plants was
counted. The cover of other plants in the enclosures
was regularly estimated. We also measured redox
potential, sediment samples, water depth and the
thickness of the muddy top layer of the sediment. The
samples were further processed as described in the
‘Chemical analysis’ section.
After two growing seasons, all plants were harves-
ted and separated by species. Fresh weight, diameter,
number of leaves and buds and leaf width were
determined for the S. aloides plants. In addition, the
plants were separated into young, intermediate and
old leaves, inflorescences, roots and buds. A sub-
sample of fresh roots was kept apart for root plaque
extractions (see ‘Chemical analysis’ section). All parts
were dried for 48 h at 70 �C and weighed. All other
plant species were divided into aboveground and
belowground parts and dried.
Chemical analysis
Immediately after the porewater had been collected, a
10.5 mL portion was fixed with 10.5 mL sulphide anti
oxidant buffer (Van Gemerden, 1984) to measure
sulphide concentrations with an ion-specific electrode
(Orion type 9416 SC; ATI Orion, Boston, MA, U.S.A.).Fig. 1 Overview of the cylinders in the field.
436 M.E.W. van der Welle et al.
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
A 10mL subsample was used to measure pH and
alkalinity. Alkalinity was estimated from the amount
of HCl needed to titrate the sample to a pH of 4.2.
The remaining sample was divided into two parts.
One part was frozen with 0.5% HNO3 until analysis
for Ca, Mg, Fe, Al, P, S, SO2�4 , Cr, Ni, Cd, Pb, Cu and
Zn with Inductively Coupled Plasma Mass Spectros-
copy (ICP-MS X-series; Thermo, Waltham, MA,
U.S.A.). The other part was frozen with 0.125 g L)1
citric acid and analysed for o-PO4 (Henriksen, 1965),
NO�3 (Kamphake, Hannah & Cohen, 1967) and NHþ4(Grasshoff & Johansen, 1977) using an Auto Analyzer
(AA 3; Bran + Luebbe, Norderstedt, Germany), and
for K using flame photometry (FLM3 Flame photom-
eter; Radiometer, Copenhagen, Denmark).
The dried plant samples were ground and 200 mg
of dried plant material was weighed exactly and
dissolved with 4 mL nitric acid (65%) and 0.9 mL
35% hydrogen peroxide, using an ETHOS D micro-
wave labstation (Milestone, Sorisole, Italy). The
destructed sample was diluted and analysed with
ICP-MS as described above.
A sub-sample of fresh roots was weighed and
extracted with an anaerobic bicarbonate-dithionite
solution as described by Christensen & Sand-Jensen
(1998) to remove root plaque and determine the
concentrations of metals in the root plaque. The
supernatant was diluted and analysed by ICP-MS as
described above.
Sediment samples were dried for 48 h at 105 �C
to determine the moisture content, and then heated
to 550 �C for 4 h to estimate the organic matter
content. Dried soil samples were digested with
4 mL nitric acid (65%) and 0.9 mL 35% hydrogen
peroxide and analysed as described for the plant
samples.
In addition, fresh sediment was extracted with bi-
distilled water, CaCl2, ammonium citrate and sodium
nitrate to determine bioavailable metal fractions in the
soil (see Table 1, Chojnacka et al., 2005).
Data analysis
Possible differences between treatments were ana-
lysed with ANOVAANOVA using Tukey’s post hoc test. Data
were transformed whenever necessary to obtain equal
variances between treatments. Nutrient concentra-
tions in porewater samples were analysed with a
repeated-measures ANOVAANOVA to check for differences
between treatments in time. Regression and correla-
tion analysis were used to test for relationships
between metal concentrations in the extracted frac-
tions and plant tissue concentrations. All statistical
analyses were performed with SPSS 13.0 (SPSS,
Chicago, IL, U.S.A.).
Results
Porewater composition
The experimental design, using field enclosures,
proved to be an elegant method to study the effects
of changed surface water and porewater characteris-
tics under controlled conditions. Apart from ammo-
nium concentrations, there were no large differences
between porewater concentrations outside the enclo-
sures and the control treatment. The differences in
ammonium concentrations are probably a result of
fertilisation of the surrounding agricultural land.
Sulphate addition resulted in a greatly increased
sulphide (HS) concentration in the porewater, as a
result of sulphate reduction (Fig. 2a; Table 2). The
porewater in the SO4 treatment had a significantly
higher sulphide concentration than all the other
treatments, although sulphide concentrations were
very high in all treatments, except for the Fe treatment
(Fig. 2a; Table 2). Iron addition, in contrast, resulted
in significantly lower sulphide concentrations com-
pared with all other treatments, as a result of iron-
sulphide precipitation. The Fe and the Fe&SO4
treatments had higher free iron concentrations than
Table 1 Summary of the extractions
Extraction
Amount of
soil (g)
Amount of
solution (mL) Concentration
Extraction
time (h)
MilliQ 35 100 1 h
Calcium chloride 10 100 0.01 mol L)1 CaCl2 6 h
Ammonium citrate 10 100 2 g L)1 ammonium citrate 6 h
Sodium nitrate 10 100 0.1 mol L)1 NaNO3 6 h
Biogeochemical interactions between iron and sulphate 437
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
the control (five and three times higher, respectively).
The sulphate-only treatment resulted in increased
concentrations of phosphate (Fig. 2b) and ammonium
(Table 2) in sediment porewater, but not until the
second year. During the first year, all treatments
showed the same seasonal pattern. Iron addition also
led to increased ammonium concentrations compared
with the control treatment. Concentrations of both
nutrients, however, were not increased in the surface
water layer. The different treatments had no effect on
nitrate concentrations, which remained low during
the entire experiment. Iron and sulphide concentra-
tions in the porewater showed a significant negative
correlation [Spearman’s correlation coefficient
(SCC) ¼ )0.789, P ¼ 0.004]. No significant correla-
tions were found between the other compounds.
The concentration of phosphate in the enclosures
was inversely correlated with the iron : phosphate
ratio in the porewater (exponential correlation, R2 ¼0.41, P < 0.000). Increased iron : phosphate ratios, like
those in the Fe treatment and to a lesser extent in the
Fe&SO4 treatment, resulted in decreased mobilisation
of phosphate in the sediment, compared with the SO4
treatment. In Fig. 3, all data points of the SO4
treatments have very low ratios and very high P
mobilisation, in contrast to those of the Fe treatment,
which have higher ratios and lower P mobilisation.
The Fe&SO4 and control treatments yielded inter-
mediate values.
Plant biomass
The highest total biomass of S. aloides per cylinder was
measured in the control treatment (Fig. 4). Addition of
iron or sulphate resulted in a greatly decreased total
biomass. Addition of both compounds, however, did
not significantly decrease the total biomass compared
with the control treatments (Fig. 4). The lower total
biomass was a result of either fewer plants (Fe
treatment) or smaller plants (SO4 treatment) (Table 3).
Total biomass of S. aloides appeared to be negatively
correlated with ammonium and sulphide concentra-
tions in the porewater. However, these correlations
were not significant, unless the enclosures with high
biomass of Elodea nuttallii (Planch.) St John were
excluded from the analysis. This species was very
abundant in certain enclosures. After omission of the
0
200
400
600
800
1000
1200
30-voN
40-naJ
40-raM
40-yaM
40 -lu J
40-peS
40-voN
5 0-n aJ
5 0-r aM
50-yaM
50 -lu J
50-peS
Su
lph
ide
(µm
ol L
-1)
Ph
osp
hat
e (µ
mo
l L-1
)
0
10
20
30
40
50
60
70
80
90
100
30-voN
40-naJ
40-raM
40-yaM
4 0-lu J
4 0- peS
40- voN
50- naJ
50 -r aM
50 -y aM
5 0-lu J
50 -peS
(a)
(b)
Fig. 2 Sulphide (a) and phosphate (b) concentrations in the
porewater during the experiment for all treatments (aver-
age ± standard error of the mean). Because of logistical prob-
lems, the SO4 treatment started later. Open circles ¼ control,
open squares ¼ Fe, open triangles ¼ Fe&SO4, crosses ¼ SO4,
diamonds ¼ outside.
Table 2 Average concentrations, ±
standard error of the mean, in porewater
(lmol L)1) during the experiment
Treatment HS Fe o-PO4 NH4 NO3
Control 118 ± 18B 5 ± 1A 15 ± 2A 157 ± 21A 5.7 ± 2.0A
Fe 11 ± 2A 36 ± 6C 14 ± 2A 327 ± 36B 7.4 ± 1.0A
Fe&SO4 102 ± 21B 18 ± 3B 16 ± 2A 231 ± 23AB 4.8 ± 0.4A
SO4 569 ± 40C 3 ± 1A 39 ± 4B 517 ± 45C 6.2 ± 0.5A
Outside 128 ± 13B 10 ± 2AB 21 ± 2A 323 ± 24B 3.9 ± 0.1A
Superscript letters indicate significant differences between treatments for each species
(A N O V AA N O V A, P < 0.05).
438 M.E.W. van der Welle et al.
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
enclosures with high biomass of E. nuttallii, sulphide
concentrations, but not ammonium concentrations,
were significantly correlated with total S. aloides
biomass (Fig. 5; SCC ¼ )0.71, P ¼ 0.025). Another
remarkable effect of sulphate addition was that the
plants did not completely emerge in the SO4 treat-
ment in the second year (Fig. 6). Although they grew
roots, the plants remained submerged and did not
grow to more than approximately 10–20 cm above the
sediment (which was 40–50 cm below the water level)
during the entire second growing season.
Elodea nuttallii, which was the second most common
species in the experiment and spontaneously
appeared in the enclosures, showed a completely
different response to the treatments (Fig. 4). This
species had its greatest biomass in the Fe treatment
and was not found at all in the SO4 treatment.
Remarkably, E. nuttallii only formed roots when iron
was added (Fe and Fe&SO4 treatments, data not
shown). There was, however, a significant negative
correlation between the biomasses of S. aloides and E.
nuttallii (SCC ¼ )0.788, P ¼ 0.002).
The S. aloides plants in the Fe&SO4 treatment had
relatively less biomass per plant in the roots and more
in the buds, compared with the control treatment and
the initial measurements. In the SO4 treatments, the
plants had invested relatively more biomass in the
roots and were never found to be flowering, in
contrast to all other treatments (data not shown).
Although root length was not significantly affected,
both the SO4 treatment and the Fe&SO4 treatment
yielded a lower shoot : root ratio than the control
(Table 3). The shoot : root ratio of the plants was
negatively correlated with the average sulphide con-
centrations in the porewater [Pearson’s correlation
coefficient (PCC) ¼ )0.643, P ¼ 0.044].
Morphology of S. aloides
Plants in the SO4 treatment had significantly fewer
buds and narrower leaves than the initial and control
plants. In the Fe&SO4 treatment, however, the
numbers of buds and the leaf width were significantly
increased compared with the initial values (Table 3).
In addition, we noticed that the leaves in the SO4
treatment were very thin and almost translucent. The
morphology of plants in the Fe treatment did not
differ from that in the control treatment, or from that
in the initial measurements.
Many morphological traits showed strong negative
correlations with the actual sulphide concentrations.
Fresh biomass (PCC ¼ )0.718, P ¼ 0.015), number of
buds (PCC ¼ )0.771, P ¼ 0.008), number of inflores-
cences (PCC ¼ )0.783, P ¼ 0.006) and leaf width
(PCC ¼ )0.861, P ¼ 0.001) were negatively correlated
with sulphide concentrations. Despite the negative
correlation between iron and sulphide concentrations
in the porewater, no significant correlations between
morphology and iron were found. Another potential
phytotoxin, ammonium, was found to negatively
affect the number of leaves (PCC ¼ )0.624, P ¼0.036) and the number of inflorescences (PCC ¼)0.583, P ¼ 0.05).
0
10
20
30
40
50
60
Fe : P ratio
Ph
osp
hat
e (µ
mo
l L-1
)
Fig. 3 Phosphate concentration in the porewater related to the
iron : phosphate ratio (mol : mol). The ovals indicate the iron
treatments (open squares) and the sulphate treatment (crosses).
Open circles ¼ control, open squares ¼ Fe, open triangles ¼Fe&SO4, crosses ¼ SO4, diamonds ¼ outside.
FeControl
0
50
100
150
200
250
300
Fe and HS- (µmol L-1)
Bio
mas
s (g
)
Fig. 4 Biomass of Stratiotes aloides (open bars) and Elodea nuttallii
(black bars) in the second year, per enclosure, at different con-
centrations of iron (Fe) and sulphide (HS) in the porewater
(lmol L)1). Values are averages ± standard error of the mean.
The different treatments are indicated above the bars.
Biogeochemical interactions between iron and sulphate 439
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
Tissue concentrations and extractions
In all treatments, S concentrations in aboveground
parts of S. aloides were slightly increased compared
with the initial measurements, except in the Fe
treatment, where concentrations were only half of
those in the other treatments. The pattern for iron
was exactly the opposite: tissue Fe concentrations
decreased in all treatments, except in the Fe treat-
ment, compared with the initial values. The highest
concentrations of metals in aboveground parts were
found in the SO4 treatment for all other metals
(except cadmium) when comparing the four treat-
ments (not including the initial measurements). The
lowest metal concentrations in aboveground parts
were generally found in the Fe&SO4 treatment
(Table 4). Both the roots and the aboveground parts
in the SO4 treatment had remarkably high concen-
trations of aluminium. In the Fe treatment we
measured increased iron and manganese concentra-
tions in the roots when comparing the four treat-
ments. However, this appeared to be to a large
extent related to the formation of root plaque in the
iron treatment (Table 5). In the Fe treatment, but not
Table 3 Morphological characteristics of the Stratiotes aloides plants before (initial) and after the experiment for the different treat-
ments
Treatment Total no. of plants Initial (n ¼ 36) Control (n ¼ 12) Fe (n ¼ 6) Fe&SO4 (n ¼ 10) SO4 (n ¼ 8)
Fresh weight of shoots* (g) 300 ± 29B 243 ± 42B 201 ± 77AB 335 ± 27B 45 ± 13A
Fresh weight of roots (g) 0 ± 0A 2.8 ± 1.6B 5.5 ± 2.6B 6.3 ± 2.1B 2.8 ± 1.1AB
Shoot : root ratio ND 98 ± 10 73 ± 25 59 ± 7 30 ± 4†
Diameter (cm) 50 ± 1B 39 ± 3A 40 ± 4AB 40 ± 2AB 31 ± 4A
No. of leaves 56 ± 1 50 ± 3 57 ± 4 63 ± 3 49 ± 8
No. of buds 2.8 ± 0.3AB 4.8 ± 0.6BC 4.8 ± 1.2BC 7.0 ± 0.7C 1.3 ± 0.6A
No. of inflorescences ND 1.1 ± 0.3AB 1.2 ± 0.6AB 1.6 ± 0.3B 0 ± 0A
Leaf width (mm) 14.4 ± 0.4B 15.4 ± 0.6BC 15.3 ± 1.2BC 17.7 ± 1.0C 8.9 ± 0.9A
Root length (cm) ND‡ 48 ± 8 48 ± 19 28 ± 8 26 ± 7
Root hair length (cm) ND‡ 2.4 ± 0.6 4.1 ± 2.0 1.3 ± 0.5 1.0 ± 0.5
Rooting depth (cm) ND‡ 19 ± 4.4 18 ± 8.2 5 ± 1.9 11 ± 3.1
Values are averages per plant ± standard error of the mean. Superscript letters indicate significant differences between treatments and
the initial values (A N O V AA N O V A, P < 0.05).
*There was a significant correlation between shoot fresh weight and shoot dry weight (linear regression; R2 ¼ 0.921; P < 0.000; dry
weight ¼ 0.0831 · fresh weight).†The shoot : root ratio of plants in the SO4 treatment was significantly lower than in the control treatment (t-test, P ¼ 0.012).‡No initial measurement of roots was performed, since at that time of year the plants do not have roots.
R2 = 0.71
0
100
200
300
400
500
600
sulphide (µmol L-1)
Bio
mas
s (g
)
Fig. 5 Correlation between the total biomass of Stratiotes aloides
per enclosure and the sulphide concentration. The enclosures
with a high biomass of Elodea nuttallii are indicated by triangles.
The regression line is only for the enclosures without dominance
of E. nuttallii.
0
1
2
3
4
5
6
7
8
9
10
40-guA
40-peS
40-tcO
40-voN
40-ceD
50-n aJ
50-beF
50-raM
50-rpA
50- yaM
50- nu J
50- luJ
50- guA
gnita
olf.o
Nse
diola.
S
Fig. 6 Number of floating Stratiotes aloides plants during the
second year of the experiment (no data available for the first
year). Open circles ¼ control, open squares ¼ Fe, open trian-
gles ¼ Fe&SO4, crosses ¼ SO4.
440 M.E.W. van der Welle et al.
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
in the Fe&SO4 treatment, high concentrations of iron
were measured in the root plaque extract. When
iron concentrations in the roots are corrected for the
amount of iron in the root plaque, however, no
treatment effect seems to remain on iron uptake in
the roots. Phosphorus concentrations in the S. aloides
plants were higher than the initial concentrations in
all treatments.
Iron concentrations in E. nuttallii were also in-
creased by iron addition (ANOVAANOVA, P ¼ 0.002). Iron
concentrations in aboveground parts were approxi-
mately 30 times higher in the Fe treatment than in the
control (Table 6). In addition, very high iron concen-
trations were found in the roots after iron plaque
removal (data not shown). In contrast to S. aloides,
phosphorus concentrations were lower in the Fe
and Fe&SO4 treatments compared with the control
(Table 6). Moreover, sulphur concentrations in
E. nuttallii were not decreased in the Fe treatment,
unlike those in S. aloides. No differences in sulphur
concentrations were found between the treatments
involving E. nuttallii. The various metal extractions
did not show a clear relation with the concentrations
in the plants (data not shown).
Discussion
Eutrophication
Sulphate addition led to increased mobilisation of
phosphate from the sediment to the porewater,
although this effect was not measurable in the surface
water. This is a well-known process, which has been
described for several different systems (Patrick &
Khalid, 1974; Bostrom et al. 1982; Roelofs, 1991;
Lamers et al., 1998). Iron addition, mimicking natural
iron influx by groundwater discharge, on the other
hand, is supposed to lead to immobilisation of
phosphate (Sperber, 1958; Bostrom et al., 1982; Boers,
1991; Smolders et al., 1995, 2001). This was confirmed
by our results from the Fe&SO4 treatment. When both
iron and sulphate were added, there was no increased
P mobilisation in the porewater compared with the
control treatment. However, we found less mobilisa-
tion of phosphate in the Fe treatment. This might be
related to the fact that all added iron was probably in
the reduced (Fe2+) form, and reduced iron has a lower
affinity for phosphate than oxidised iron (Fe3+)
(Lamers et al., 1998). In that case, the positive effectTab
le4
Tis
sue
con
cen
trat
ion
sin
abo
veg
rou
nd
and
bel
ow
gro
un
dp
arts
of
Str
atio
tes
aloi
des
(lm
ol
g)
1,
aver
age
±st
and
ard
erro
ro
fth
em
ean
).In
itia
lco
nce
ntr
atio
ns
wer
em
easu
red
inp
lan
tsth
atw
ere
kep
tap
art
atth
eb
egin
nin
go
fth
eex
per
imen
t.
Mg
Al
KC
aC
rF
eM
nN
iZ
nC
uC
d(·
10–3)
Pb
PS
Sh
oo
ts
Co
ntr
ol
385
±20
AB
4±
0.5
B10
90±
61A
295
±17
AB
0.03
±0.
00B
3.5
±0
.4A
11
±1A
0.03
±0.
00B
0.7
±0.
10.
05±
0.00
B0.
18±
0.03
B0.
01±
0.00
B16
3±
1622
1±
34B
Fe
348
±31
AB
3±
0.4
AB
11
55
±4
3A
374
±56
AB
0.02
±0.
00A
42±
7.5B
44±
7B0.
02±
0.00
A0.
5±
0.0
0.03
±0.
00A
0.07
±0.
02A
0.01
±0.
00A
B15
0±
1511
0±
5A
Fe&
SO
44
61
±2
1B
2±
0.3
A10
11±
39A
237
±14
A0.
03±
0.00
B2
.7±
0.2
A1
2±
1A0.
03±
0.00
AB
0.5
±0.
00.
03±
0.00
A0.
17±
0.02
B0.
01±
0.00
A14
2±
1223
1±
16B
SO
431
0±
37A
12
±2
.5C
16
30
±8
3B
441
±62
B0.
05±
0.01
B3
.9±
0.5
A84
±15
B0.
06±
0.01
C0.
9±
0.1
0.07
±0.
01B
0.14
±0.
04A
B0.
03±
0.00
C14
9±
2927
3±
46B
Init
ial
343
±3
41±
2.6
953
±8
674
±63
ND
24±
1.5
67±
2N
D1.
1±
0.0
ND
ND
ND
109
±1.
416
6±
1.2
Ro
ots
Co
ntr
ol
87±
4A25
±2A
B13
26±
171
465
±15
0.14
±0.
0222
±2A
12±
2A0.
12±
0.01
1.0
±0.
20.
22±
0.03
0.51
±0.
080.
10±
0.01
B23
5±
5170
1±
116
Fe
116
±35
AB
21±
5A12
72±
249
357
±9
0.04
±0.
0115
9±
15B
27±
6B0.
07±
0.01
0.4
±0.
10.
15±
0.06
0.49
±0.
250.
10±
0.05
AB
137
±87
490
±36
2
Fe&
SO
414
3±
6B
11±
2A18
25±
102
415
±25
0.10
±0.
0212
±4A
17±
3AB
0.07
±0.
020.
7±
0.4
0.10
±0.
020.
31±
0.19
0.04
±0.
01A
439
±32
841
3±
158
SO
496
±4
A45
±15
B19
35±
216
395
±50
0.12
±0.
0417
±4A
15±
2A0.
10±
0.02
1.3
±0.
40.
18±
0.04
0.48
±0.
070.
09±
0.03
AB
462
±38
175
3±
383
Val
ues
inb
old
are
sig
nifi
can
tly
dif
fere
nt
fro
mth
ein
itia
lco
nce
ntr
atio
ns.
Su
per
scri
pt
lett
ers
ind
icat
ed
iffe
ren
ces
bet
wee
nth
ed
iffe
ren
ttr
eatm
ents
.N
oin
itia
lm
easu
rem
ents
on
roo
ts
wer
ep
erfo
rmed
,as
the
pla
nts
wer
eco
llec
ted
du
rin
gth
ew
inte
ran
dso
did
no
th
ave
any
roo
ts.
ND¼
no
td
eter
min
ed.
Biogeochemical interactions between iron and sulphate 441
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
of iron addition in the Fe&SO4 treatment was merely
the effect of precipitating iron sulphides, which
prevents the replacement of phosphate by sulphide
at phosphate-binding sites.
Previous studies have shown that there was a
clear correlation between iron concentrations in the
porewater and sulphide and phosphate concentra-
tions (Smolders & Roelofs, 1993) and that phosphate
mobilisation depended on the iron : phosphate ratio
in the porewater (Fe : P ratio, Smolders et al., 2001).
We found that phosphate mobilisation (porewater
concentration above 25 lmol L)1) occurred mainly
in the SO4 treatment (Fe : P ¼ 0.16) and, to a lesser
extent, in the outside treatment (Fe : P ¼ 0.50),
where surface water quality was less constant as
water was let in from external sources during drier
periods to maintain a constant water level. The Fe
treatment had the lowest phosphate mobilisation
(Fe : P ¼ 10.4), while the Fe&SO4 and the control
treatments resulted in intermediate values (Fe : P ¼3.34 and 1.40, respectively). The Fe treatment and
the SO4 treatment differed significantly in Fe : P
ratio. From our results, it appears that phosphate
mobilisation takes place at Fe : P ratios in the
porewater below 1. The ratios in the control treat-
ment were just above this level (1.40) and in this
treatment led to lower phosphate mobilisation
(porewater concentrations below 20 lmol L)1). A
study by Smolders et al. (2001) showed that phos-
phate mobilisation from the sediment to the water
layer occurred when Fe : P ratios in the porewater
were below 1. In that study it was also shown that
when Fe : P ratios exceed 10, hardly any phosphate
is mobilised from the sediment, as was the case in
our Fe treatment.
Interactions between iron and sulphate
Iron addition can counteract sulphide toxicity and,
conversely, sulphate addition can counteract iron
toxicity. In the Fe&SO4 treatment, sulphide concen-
trations were much lower than in the SO4 treatment,
and the plants flourished. The precipitation of iron
sulphides is a well-known process, which has been
described by several other authors (Murray, 1995;
Smolders et al., 1995, 2001; Lucassen, Smolders &
Roelofs, 2000b; Van der Welle et al., 2006, in press).
However, not much is known about the interactions
between iron and sulphate and their effects on
freshwater plants. It has become clear from the
present study that mutual detoxification of iron and
sulphide – compounds which are both potentially
toxic to aquatic plants – takes place (e.g. Wheeler,
Al-Farraj & Cook, 1985; Cook, 1990; Snowden &
Wheeler, 1993; Lucassen et al., 2000a; Kamal et al.,
2004 for iron toxicity and Tanaka, Mulleriyawa &
Yasu, 1968; Koch & Mendelssohn, 1989; Koch et al.,
1990; Armstrong et al., 1996; Smolders & Roelofs,
1996; Lamers et al., 1998; Van der Welle et al., 2006
for sulphide toxicity).
Table 5 Metal concentrations in root plaque extracts (lmol g)1, average ± standard error of the mean)
Treatment Fe Mn Ni Cu Zn Pb
Control 15 ± 2A 7 ± 1 0.11 ± 0.01 2.3 ± 0.2AB 1.8 ± 0.2AB 0.03 ± 0.00
Fe 193 ± 22B 24 ± 11 0.12 ± 0.01 3.3 ± 0.6BC 2.5 ± 0.4B 0.03 ± 0.01
Fe&SO4 7 ± 2A 23 ± 3 0.11 ± 0.01 1.5 ± 0.0A 1.3 ± 0.0A 0.01 ± 0.01
SO4 15 ± 4A 16 ± 6 0.18 ± 0.06 3.7 ± 0.3C 2.8 ± 0.3B 0.03 ± 0.01
Superscript letters indicate differences between the different treatments (A N O V AA N O V A, P < 0.05).
Table 6 Tissue concentrations in aboveground parts of Elodea nuttallii (lmol g)1, average ± standard error of the mean)
Mg Al K Ca Fe P S
Control 79 ± 17 5 ± 1 423 ± 54 447 ± 123 3 ± 1A 74 ± 36B 35 ± 8
Fe 94 ± 5 29 ± 6 457 ± 26 702 ± 98 103 ± 11B 48 ± 6AB 66 ± 9
Fe&SO4 82 ± 19 9 ± 3 436 ± 42 1061 ± 327 10 ± 3A 21 ± 4A 55 ± 6
SO4 NA NA NA NA NA NA NA
Superscript letters indicate differences between the different treatments.
NA, data not available (no Elodea in the SO4 treatment).
442 M.E.W. van der Welle et al.
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
Toxicity
In the SO4 treatment, the plants were suffering from
sulphide toxicity, which led to a greatly decreased
biomass, smaller and fewer plants and very thin,
almost translucent leaves. Another remarkable effect
was that in the SO4 treatment, the plants did not
completely emerge during the growing season. This
means that the species will easily lose the competition
with algae, as a result of light deprivation. In previous
experiments by Smolders & Roelofs (1996) it was
found that the roots of S. aloides were strongly affected
by sulphide concentrations higher than 10 lmol L)1.
Surprisingly, sulphide did not affect root biomass in
the present study, despite much higher sulphide
concentrations. Increased sulphide concentrations
did, however, greatly affect aboveground biomass
and morphology. This is remarkable, as most studies
of sulphide toxicity have found a negative effect on
the roots (e.g. Koch & Mendelssohn, 1989; Armstrong
et al., 1996; Smolders & Roelofs, 1996). The decreased
growth of the plants can then be attributed to the
plants being unable to take up nutrients. In the
present study, no root decay was observed, so the
decreased growth must have been caused by other
processes. Armstrong et al. (1996) suggest that sul-
phide toxicity may lead to blockage of the gas space
system in Phragmites australis (Cav.) Steud., which can
induce the accumulation of ethene and carbon dioxide
and thus disturb physiology. This might also be an
explanation for the decreased growth and thinner
leaves in our plants. In addition, disturbed gas
transport may also explain the fact that the plants
did not emerge during the growing season. It is also
possible that the emergence of the plants was ham-
pered by the presence of algae or Lemna sp. Photo-
synthesis would then be decreased as a result of light
deprivation, resulting in a lower buoyancy. Another
possible explanation for the smaller biomass is that
the plants had to invest all their resources in gas
transport to the roots and in root biomass, to reduce
the effects of toxic sulphide in the rhizosphere, which
might explain the relatively high root biomass. As
there was no E. nuttallii present in the enclosure with
the SO4 treatment, competition cannot have caused
the small biomass of S. aloides.
Sulphide-induced iron deficiency, which we have
proposed to be a major negative effect of sulphide
(Van der Welle et al., in press), might also play a role,
but not in the present study. In the Fe&SO4 treatment,
iron concentrations in the plant did not differ from
those in the SO4 treatment or the controls (despite
much lower sulphide concentrations), which indicates
that there was only a direct effect of sulphide on plant
growth. This also appeared to be the case for
E. nuttallii, which did not occur at all in the SO4
treatment. In fact, E. nuttallii only occurred in high
densities in the treatments involving iron addition to
the sediment. It appears that under conditions of
increased iron availability, E. nuttallii is a better
competitor than S. aloides. As sulphide has a direct
negative effect on both species, this leads to the
assumption that the performance of S. aloides in
systems where iron–sulphur interactions play an
important role might be determined by interactions
between competition (with E. nuttallii) and toxicity.
Fig. 7 summarises these interactions. In this model,
sulphide has a negative effect on both plant species,
while iron negatively affects S. aloides but has a
positive or no effect on E. nuttallii. The negative effect
of iron addition on S. aloides is increased by the
positive effect on E. nuttalli, which, at increased iron
availability, is a stronger competitor for light than
S. aloides. Nutrient limitation is less probable, given
the high phosphate concentrations in the porewater.
The negative effect of sulphide on E. nuttallii might
S. aloides E. nuttallii
Fe
HS
Toxicity
Competition
Competitive advantage
Precipitation
Toxicity?Toxicity
- +--
-
-
PO4PO4
algae
+ +- -
Light deprivationLight deprivation
Increased growth?
Increased growth?
+
+
-
-
Fig. 7 Conceptual model of the interactions between dissolved
iron (Fe), sulphide (HS), phosphate (PO4), Elodea nuttallii,
Stratiotes aloides and algae. Dashed lines indicate indirect effects,
through lines indicate direct effects. Plus and minus signs
indicate positive and negative effects, respectively. A positive
effect of increased P availability will only occur if P is the
growth-limiting nutrient.
Biogeochemical interactions between iron and sulphate 443
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
have been a result of iron deficiency, although we
have no proof of this, as the species did not occur at all
in the SO4 treatment, where iron availability was
lowest. In addition, sulphide may lead to P mobilisa-
tion, which can either stimulate or decrease macro-
phyte growth, depending on the development of algal
blooms and whether or not P is the growth-limiting
nutrient.
Iron addition (Fe treatment) also led to a decreased
biomass of S. aloides. As described above, this might
have been related to increased competition by
E. nuttallii. On the other hand, iron toxicity may also
have played a role. Not much is known about the
tolerance of S. aloides to iron, but Smolders et al.
(1996a) suggest that iron limitation may be one of the
causes of the decline of S. aloides in the Netherlands. In
that case, increased iron availability might have a
positive effect on S. aloides. According to field data on
the distribution of S. aloides in the Netherlands, vital
populations of this species are found only in waters
with porewater iron concentrations between 60 and
275 lmol L)1 (Smolders et al., 1996a). Iron toxicity
therefore seems unlikely, as the iron concentrations in
our experiment never exceeded 300 lmol L)1.
Nutrient and metal uptake by S. aloides
Phosphorus concentrations did not differ between
treatments, despite the increased phosphate mobili-
sation in the SO4 treatment. This was probably caused
by the fact that P availability was very high in all
treatments and the plants probably did not need take
up more P to be able to increase their growth rate.
Tissue iron concentrations were increased in the Fe
treatment, compared with the initial concentration,
while iron concentrations were decreased in all other
treatments. This is a result of iron being immobilised
by iron-sulphide precipitation, leading to a decreased
availability of iron for the plants, which, in sulphidic
environments, may even lead to sulphide-induced
iron deficiency (Smolders et al., 1996a; Van der Welle
et al., in press). However, iron deficiency does not
appear to have occurred in our experiment. We found
no signs of chlorosis, which is a common indicator of
iron deficiency (e.g. Bienfait, 1989; Van Dijk &
Bienfait, 1993; Mengel, 1994; Smolders et al., 1996a;
Alvarez-Fernandez et al., 2004), and the concentra-
tions we measured were well above (>25 times higher
than) those found for iron-deficient S. aloides plants by
Smolders et al. (1996a) and within the range of
concentrations measured in healthy plants in the field
(Smolders et al., 1996b).
Despite the fact that other metals than iron can
also precipitate with sulphide (e.g. Di Toro et al.,
1992; Ankley et al., 1993; Besser et al., 1996; Chapman
et al., 1998; Wang & Chapman, 1999), we found the
highest tissue concentrations of Al, Cr, Ni, Mn, Zn,
Cu and Pb in the SO4 treatment. This might have
been caused by sub-optimal conditions for metal-
sulphide precipitation. According to Drever (1997),
however, sulphides of Zn, Cd, Pb and Cu are readily
formed in the presence of sulphur under reducing
conditions, which means that sulphides of these
elements should have formed in our experiment in
all treatments. The lowest metal concentrations in
plant tissue were found in the Fe and the Fe&SO4
treatments. This indicates that co-precipitation of
these metals with iron sulphides may have taken
place. It is known that metals can be incorporated
into pyrite (FeS2) or be adsorbed at the pyrite matrix
(Huerta-Diaz et al., 1998; Morse & Luther, 1999;
Muller, Axelsson & Ohlander, 2002). In that case,
co-precipitation of heavy metals with iron sulphides
would have further decreased metal availability.
Another important effect of increased sulphate
concentrations is increased decomposition (Roelofs,
1991; Brouwer et al., 1999). This may lead to addi-
tional mobilisation of organically bound heavy
metals from the peaty sediment, as could have been
the case in the SO4 treatment. The same process can
also explain the increased heavy metal concentra-
tions in the root plaque in the SO4 treatment.
Conclusions
Both sulphate and iron availability can have a strong
influence on plant growth in freshwater wetlands.
Surprisingly, we found both direct effects via toxicity
and indirect effects via the modification of interspe-
cific competition. When iron and sulphide concentra-
tions are in balance, neither of these species will have
toxic effects. However, both compounds can have
phytotoxic effects and can depress the growth of
specific plant species, thereby stimulating the growth
of other, more tolerant species.
The findings of the present study suggest that the
growth of S. aloides can be regulated by interactions
between sulphide, iron and competition with
444 M.E.W. van der Welle et al.
� 2007 The Authors, Journal compilation � 2007 Blackwell Publishing Ltd, Freshwater Biology, 52, 434–447
E. nuttallii. In the presence of iron, E. nuttallii proved
to be a better competitor than S. aloides and was able
to overgrow the latter species. High sulphide concen-
trations had a negative influence on both species, but
resulted in increased phosphate mobilisation to the
water layer. In addition, metals were mobilised to the
plants from the sediment in the sulphate treatment,
which can be potentially dangerous for the entire
trophic system.
As a result of altered hydrological conditions,
agricultural activities and increased atmospheric sul-
phur deposition, many freshwater wetlands nowadays
receive sulphate-enriched water, which can lead to
serious problems (Lamers, 2001). We showed that
increased sulphate load can lead to changes in species
composition, toxicity and mobilisation of heavy
metals. Moreover, in many wetlands, changed hydrol-
ogy can lead to decreased groundwater flow, which
may lead to decreased input of iron-rich seepage. As
we have shown, iron strongly interferes with the
processes described above. Iron depletion will there-
fore amplify the effects of increased sulphate load.
Acknowledgments
We would like to thank Martin Versteeg, Rick Kuiperij
and Karla Niggebrugge for their help with the field
work, Germa Verheggen, Roy Peters, Jelle Eygenstein,
Liesbeth Pierson, Ine Hendriks and Rien van der Gaag
for technical assistance and Bert van Dijk (Staats-
bosbeheer, the Netherlands) for allowing us to do the
fieldwork in the Ronde Venen reserve. This study was
funded by the Netherlands Organization for Scientific
Research (NWO), through its stimulation programme
on system-oriented ecotoxicological research (SSEO).
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Biogeochemical interactions between iron and sulphate 447
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