Biogeochemical interactions between iron and sulphate in ... · in freshwater wetlands and their...

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.


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



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).



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.



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.



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

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nN

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ts

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ntr

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295

±17

AB

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05±

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B0.

18±

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B0.

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B16

3±

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1±

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Fe

348

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AB

3±

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AB

11

55

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3A

374

±56

AB

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44±

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5±

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A0.

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00.

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9±

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±0.

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B0.

03±

0.00

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9±

2927

3±

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Init

ial

343

±3

41±

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953

±8

674

±63

ND

24±

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67±

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1±

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ND

ND

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416

6±

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ots

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ol

87±

4A25

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B13

26±

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465

±15

0.14

±0.

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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

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AB

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72±

249

357

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0.04

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±0.

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2

Fe&

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3±

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395

±50

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1.3

±0.

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18±

0.04

0.48

±0.

070.

09±

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AB

462

±38

175

3±

383

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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).



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.



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



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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(Manuscript accepted 24 October 2006)



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