Effects of gradual chemical additions on phosphorus
mobilization and macrophyte growth in
peat lake Terra Nova
Clara Chrzanowski
830701-157-130
M.Sc. Thesis
Wageningen University
Department Environmental Sciences
Aquatic Ecology and Water Management Group
Report no. ?????
Supervisors: J.J.M. Geurts (Radboud University Nijmegen)
E.H. van Nes (Wageningen University)
4
Abstract
High internal phosphorus loading annually causes eutrophication problems in peat lakes
resulting in poor biodiversity and algal blooms. To decrease phosphorus mobilization rates
from the sediment to the water layer and to improve water quality, peat lake Terra Nova was
gradually treated with FeCl₃ (33 g Fe/m2 in 1½ years). In an isolated untreated part of the
lake a mesocosm experiment was set-up to compare the effect of gradual addition of
aluminum chloride (AlCl₃), poly aluminum chloride (PAC), iron chloride (FeCl₃), Phoslock and
Phoslock+ (PL+) on phosphorus mobilization on a smaller scale.
During 56 days two undisturbed sediment cores from each of the 24 mesocosms were used
for phosphorus mobilization measurements in a dark climate room at 15 ºC. Four undisturbed
sediment cores from two locations in Lake Terra Nova and from two isolated experimental
peat ponds, one being treated with FeCl₃, are also included in the research. After that, the
sediment layer was covered with sand and the submersed macrophyte Elodea nuttallii was
planted to look for treatment effects on the vegetation. Half of the sediments were loaded
with 1600 µmol P/ L to substantially decrease the ratios of Fe:PO4, Al:PO4 and La:PO4.
Phosphorus mobilization rates were lowest in both Phoslock treatments (-5.3 ± 2.3 (PL) and 4.4 ± 2.5 (PL+) mmol m-2 yr-1) and in the PAC treatment (-5.0 ± 2.9 mmol m-2 yr-1). Highest
phosphorus mobilization rates were found in the control treatments (8.7 ± 3.6 and 5.6 ± 2.0
mmol m-2 yr-1). FeCl3 addition resulted in 4.6 ± 3.5 mmol m-2 yr-1.
Phosphate mobilization rates ranged between -1.4 ± 0.4 mmol m-2 yr -1 (PAC) and 0.3 ± 1.0
mmol m-2 yr -1 (FeCl). They were significantly higher in the control treatment compared to the
aluminum and Phoslock treatments. Lowest phosphorus and phosphate concentrations in
sediment pore water were found in treatments containing aluminum and Phoslock.
Gradual addition of iron chloride to Lake Terra Nova also resulted in phosphate fixation. One
year after starting gradual iron addition in Lake Terra Nova, phosphorus mobilization rates
were -0.94 ± 4.6 and 2.1 ± 2.0 mmol m-2 yr-1 compared to 15.8 ± 7.1 and 3.5 ± 1.7 mmol m-2
yr-1 before the addtion started.
Gradual addition of aluminum, Phoslock and iron chloride can contribute to the improvement
of the water quality in peat lakes through fixation of phosphorus. In general, algae growth is
inhibited, which improves light climate and stimulates macrophyte growth. However, addition
of FeCl3 can result in formation of iron-phosphorus colloids (< 0.15 µm) which could be
indirectly available to plants and algae.
Neither any of the treatments nor the extra sediment P loading resulted in significant effects
on plant growth and nutrient availability for the competitive species E. nuttallii. Therefore, it
can be assumed that the aquatic vegetation in peat lakes with a P-rich sediment will be
dominated by fast-growing macrophyte species after gradual chemical additions due to high
total phosphorus loads in the sediment and the bioavailability of chemically bound
phosphorus for fast-growing macrophytes.
Table of Contents
1. Introduction ............................................................................................................. 4
1.1.
Theoretical background........................................................................................................... 4
1.2.
Aim of the research ............................................................................................................... 12
2. Materials and methods ......................................................................................... 15
2.1.
Study area.............................................................................................................................. 15
2.2.
Field work .............................................................................................................................. 15
2.3.
Experimental setup and sampling ......................................................................................... 18
2.4.
Chemical analyses ................................................................................................................. 22
2.5.
Statistical analysis .................................................................................................................. 24
3. Results.................................................................................................................. 25
3.1.
Experiment 1 ......................................................................................................................... 25
3.1.1.
Mesocosm experiment .................................................................................................. 25
3.1.2.
Iron supplementation in peat ponds and Lake Terra Nova ........................................... 34
3.2.
Experiment 2 ......................................................................................................................... 40
3.2.1.
Elodea nuttallii response .............................................................................................. 40
3.2.2.
Sediment nutrient composition .................................................................................... 44
3.2.3.
Leakage through sand layer .......................................................................................... 46
4. Discussion ............................................................................................................ 47
5. Acknowledgements............................................................................................... 53
6. References ........................................................................................................... 54
7. Appendix............................................................................................................... 60
1. Introduction
1.1.
Theoretical background
Peat lakes consist mainly of dead organic matter. Waterlogged conditions create an
anaerobic and humid environment preventing bacteria and fungi to decompose dead plant
material entirely. Humans have been using peat for centuries as fuel. Anthropogenic
activities in fens, however, have created a unique semi-natural landscape with a huge
biodiversity. Typical for the distinctive pattern of Dutch peatlands are peat ponds (Dutch
“petgaten”) and baulks (“legakkers”) (Lamers et al. 2002, Lamers et al. 2010b).
Nowadays, fens and peat lakes deal with serious water quality problems caused by
anthropogenic alterations such as hydrological changes, agricultural run-off carrying nutrient
rich water and habitat fragmentation. This leads to a decrease of rare and threatened
species resulting in an undesirable loss of biodiversity (Lamers 2001, Lamers et al. 2002,
Van Der Welle et al. 2007). The bottleneck of successful lake restoration is the insufficient
reduction of external phosphorus loads and release of phosphorus from the sediment
(Bakker et al. 2012).
Changes of the groundwater levels and increased use of water for industry and agriculture
resulted in water shortage problems in Dutch peatlands causing decomposition and erosion
of the peat. However, supply of external water to keep the water level can cause external
and internal eutrophication resulting in blue-green algae blooms (Lamers 2001, Lamers et al.
2002, Smolders et al. 2006, Geurts 2010a, Lamers et al. 2010b, Poelen et al. 2011).
The main focus of this project lies on the mobilization of phosphate and phosphorus from the
sediment to the water layer. Sediment-bound P is released to the pore water which diffuses
to the surface water (Burley et al. 2001). This so called internal eutrophication (Smolders et
al. 2006) is linked to different biogeochemical processes (Poelen et al. 2011):
-
Increased decay rate of organic matter caused by alkalinity, sulfate and other water
quality parameters
Decrease in phosphate binding due to iron shortage, sulfate production
Internal nutrient cycle supported by external eutrophication
4
Faith of phosphate in the sediment-water-interface (SWI)
Orthophosphate PO43- is the most significant form of inorganic phosphorus. It is one of the
ions of orthophosphoric acid OP(OH)3. Depending on the pH levels, the various anions vary
(Reynolds and Davies 2001). Below an overview of three-step ionization of orthophosphoric
acid is given:
H3PO4(s) + H2O(l) ⇌ H3O+(aq) + H2PO4–(aq)
(aq)+
H2O(l) ⇌ H3O+(aq) + HPO42–(aq)
dihydrogen phosphate
H2PO4–
hydrogen phosphate
HPO42–(aq)+ H2O(l) ⇌ H3O+(aq) + PO43–(aq)
orthophosphate
These ions form an insoluble complex with the metal ions iron (Fe), aluminum (Afsar and
Groves 2008) and Calcium (Ca) such as apatites Ca5(PO4)3(OH, F, Cl), ferric phosphate
FePO4 , and aluminium phosphate AlPO4 . This so-called immobilization of orthophosphate
by redox-sensitive metals like iron is very important at the sediment-water layer in aquatic
systems (Reynolds and Davies 2001).
Sediments act as a sink for phosphorus, but in case of change in conditions sediments may
also serve as source preventing the improvement of the water quality (Bartram et al. 1999).
An excepted view among researchers is that lake sediments function as a net sink for
phosphorus. A long-standing paradigm among limnologists is the idea of oxygen controlling
the phosphorus release from sediments (Hupfer and Lewandowski 2008). However,
phosphorus is bound to Fe (III) under oxidized conditions whereas under anoxic conditions
Fe (III) will be reduced to Fe (II). Eventually both phosphorus and Fe dissolve again. Shallow
lakes are usually oxic due to a mixed water column and at the oxic sediment layer
phosphorus is bound to Fe(III) (Sondergaard et al. 2003).
Figure 1 explains the process of phosphorus immobilization by coupling with iron (Fe) at the
sediment-water interface. Both graphs are based on the ideas of Einsele, Mortimer, and Ohle
according to the authors. Figure 1A resembles the P sorption capacity of the aerobic
sediment surface in a well mixed system (we assume a shallow lake). Figure 1B visualizes
the situation in an anoxic environment such as a eutrophic shallow lake. Iron is reduced to
Fe(II) and together with phosphorus released into the water column.
Figure 1 Phosphorus fixation in the sediment under aerobic (A) and anaerobic (B) conditions (Hupfer
and Lewandowski 2008)
According to Hupfer and Lewandowski there are other alternative release mechanisms which
are often more important than redox processes binding P to Fe. They suggest dissolution of
5
phosphorus which is bound to calcium and decomposition of organic P under both aerobic
and anaerobic conditions. Furthermore concentrations of aluminum hydroxides [Al(OH) 3]
need to be considered since they promote P redox-insensitive sorption capacity within the
sediment under low redox conditions. Among other things aluminum treatment as a lake
restoration measure can increase P retention in sediments. Alternative phosphorus release
mechanisms from the sediment are often as important and should not be ignored.
On the one hand resuspension caused by wind or benthivorous fish increases turbidity of the
water and can be a cause for increased immobilization of phosphorus as well. Temperature
on the other hand stimulates mechanisms releasing phosphorus from the sediment.
Examples are the mineralization of organic material, release of inorganic phosphate as well
as increased sedimentation of organic material. Thickness of the oxidized sediment layer is
influenced by microbial activity and decreased with rising temperatures. This plays an
important role concerning redox-sensitive phosphorus release. Its thickness influences the
overall concentration of phosphorus in the water column. A moderate pH supports the
phosphorus binding capacity since hydroxyl ions compete with phosphorus ions for the iron
ions. Especially during the summer high photosynthetic activity leads to an increase in pH
values and higher release rates of P from the sediment. Another aspect is the Fe/P ratio
since there is a strong positive relationship between concentrations of iron and phosphorus
in surface sediment. Different studies suggested ratios higher than 10 to 15 (by weight) to
regulate phosphorus release. Similar to resuspension, bioturbation by benthic invertebrates
enhances P release. However, inhibited release due to oxygen supply by benthic
invertebrates also needs to be considered. Chemical diffusion in the interstitial water of the
sediment resembles another important upward transport of phosphorus between sediment
and water. Furthermore sediment bacteria play a significant role in the uptake, storage, and
release of phosphorus. Last but not least macrophytes contribute to the process of
phosphorus retention by oxidizing or ,if dense macrophyte beds occur, even deoxidizing the
sediment layer (Søndergaard et al. 2003).
These examples of release mechanisms shall emphasize the complexity of phosphorus
sediment release in the field of lake restoration measures. However, a thorough analysis of
the lake ecosystem and its catchment area is necessary to choose the appropriate
measures.
6
Biochemical processes in peat lakes
Figure 2 gives a good overview to understand the complexity of interacting biochemical
processes affecting the mobilization of phosphorus in peat lakes.
Figure 2 Interaction of biochemical
processes in surface water and
sediment in peat lakes. Net fluxes
of chemical variables are
symbolized by arrows (Geurts et al.
2008)
Bicarbonate (HCO3-) increases the alkalinity, also called bicarbonate alkalinity, which leads to
decay of peat because bicarbonate is able to neutralize decay-inhibiting acids. This process
is also known as mineralization. Not shown in the figure is the ability of bicarbonate to
compete with phosphate for anion adsorption sites resulting in increased release of
phosphate to the water phase. In general it can be said that bicarbonate alkalinity increases
the availability of nutrients such as phosphate, nitrate and ammonium in peat lakes
(Smolders et al 2006). Roelofs introduced the term “internal alkalinity” which is caused by
reduction of oxidants such as nitrate, sulfate and iron(hydr)oxides resulting in the formation
of bicarbonate. The increasing alkalinity further enhances decomposition (Smolders et al.
2006, Poelen et al. 2011). Oxidants are either already present or enter via inlets or surface
runoff the lake depending on the characteristics of the surrounding area as well as the land
use.
A relationship is found between alkalinity and concentrations of ammonium and phosphate in
sediment pore water (Fig. 3). It can be concluded that the decomposition of organic matter
results in an increase of inorganic carbon which can be interpreted as bicarbonate alkalinity.
Increasing bicarbonate concentrations cause release of nutrients ammonium and phosphate
(Poelen et al. 2011).
7
Figure 3 Correlation between bicarbonate (alkalinity) and ammonium (left graph) as well as phosphate
(right graph) in sediment pore water. Taken from a publication by van der Heide et al. 2010, the
graphs resemble data of a large amount of surface waters (Poelen et al. 2011).
Furthermore, the reduction of sulfate (SO42-) leads to the formation of sulfides. Phosphate is
mobilized due to the binding of sulfides with iron. Iron sulfides (FeSx) are formed and
phosphate will be released to the water layer resulting in algal blooms and turbidity (Lamers
et al. 2010b). In case of continuous sulfate reduction, rooted macrophytes are stressed by
iron deficiency and toxic concentrations of free sulfide in the sediment pore water (Smolders
et al. 2006).
Oxygen, nitrate, iron(III) ions (Fe3+) and sulfate are so called electron acceptors.
Accumulation of peat is favored by acid or very poorly buffered conditions and a lack of
alternative electron acceptors other than oxygen. The decay of organic matter is caused by
microbial processes which affect the redox state. Redox potential (Eh) is a measure for
electron activity. Oxygen has the highest (most positive) redox potential and from a
thermodynamical point of view it is the most favorable electron acceptor (Smolders et al.
2006). The higher the redox potential the stronger is the affinity of oxidants to accept
electrons. Microbes derive their energy from electron transfers. The stronger electrons are
bound the lower is the production of energy used for the growth of microbes. If oxygen is
depleted nitrate, manganese, iron, sulfate and carbon dioxide will be reduced one by one
(figure 4). Respectively nitrogen gas (N2), nitrogen oxide (N2O) or ammonium (NH4),
manganese (Mn2+), reduced iron (Fe2+), hydrogen sulfide (H2S) and methane (CH4) are
formed (Poelen et al. 2011). In fens and peat lakes organic matter is a source of nutrients
being mobilized if peat decays. Eventually this can cause eutrophication.
8
Figure 4 Electron acceptors used by
microbes to derive energy for the
breakdown of organic matter. Oxygen
provides most energy and is used first.
However, the decay of organic matter
decreases with electron acceptors shown
in the graph from right to left (oxygen to
carbon dioxide) since less energy is
derived. Graph according to Mitsch and
Gosselink, 1993 (Wienk et al. 2000).
Submerged macrophytes are able to mobilize phosphorus from sediments. At the end of the
growing season decomposition of macrophytes could lead to higher nutrient concentrations
in the surface water. This represents an important P recycle mechanism (Barko and Smart
1980). However, macrophytes also take up nutrients via leaves (Angelstein and Schubert
2008). Macrophyte dominated lakes are preferred (Scheffer et al. 1993) and macrophyte
development has become an important part of lake restoration (Bakker et al. 2012).
Peat lake Terra Nova
Terra Nova is a peat lake in the western part of the Netherlands and belongs to a fen area
with other shallow lakes called “Loosdrechtse Plassen”. The lake itself has an area of 85ha
with an average depth of 0.5 to 2 meters. Since 1950 the water quality is decreasing and the
lake became highly eutrophic (Gulati and van Donk 2002). From 1987 to 2004 blue-green
algae dominated the lake during the entire year. Phosphorus loads were calculated to be
0.22g P m-² yr-1 whereas almost 50% were released from the sediment (Ter Heerdt 2009).
In 1941, more than half of the lake bottom was covered by submerged macrophytes, mostly
Characeae. Until the 1980s dense stands of Characeae, Elodea sp. Najas marina and
Potamogeton spp. were observed (Van de Haterd and Ter Heerdt 2007).
Natural condition with upwelling iron rich groundwater disappeared due to groundwater level
changes. In the 1950s, external alkaline and sulphate rich water was used to keep the water
level. This resulted in accumulation of sulphates in the sediment. Sulphides are produced
and interact with iron(hydr)oxides by forming iron-sulphides complexes like FeS and FeS2.
This stimulates the mobilization of phosphorus to water layer and the decay of peat resulting
in high nutrient concentrations in the water layer (Roelofs 1991, Smolders and Roelofs
1993a, Brouwer and Smolders 2004, Smolders et al. 2006).
Since 1988 algae blooms dominated by filamentous cyanobacteria are observed (ter Heerdt
and Hootsmans 2007). Macrophytes play an important part in lake restoration. Low
macrophyte density is mainly caused by high turbidity due to resuspension by benthivorous
fish in combination with bird grazing (Van de Haterd and Ter Heerdt 2007). Recent
restoration measures included the reduction of external nutrient loads (Liere and Gulati 1992)
and biomanipulation by reducing benthivorous and planktivorous fish stocks (ter Heerdt and
9
Hootsmans 2007). After a successful pilot experiment in the closed-off peat ponds, a wholelake biomanipulation experiment took place. However, these measures were not effective to
restore the lake on the long term (Ter Heerdt 2009).
After a pilot experiment in two closed-off peat ponds in 2009 the entire lake was gradually
treated with iron chloride (FeCl₃). The treatment is going to last 1.5 years with approximately
100 grams iron chloride having been applied per square meter. This is done in a sustainable
way by using a windmill on a floating pond fixed in the north-east of the lake. Furthermore, in
one of the closed-off peat pond 24 cylinders were implemented to investigate the effect of
different chemical treatments on the phosphorus mobilization. The chemicals were also
added gradually to the cylinders.
The following treatments are investigated:
- Aluminum chloride (AlCl₃)
- Poly aluminum chloride (8.2% PAC - solution)
- Iron chloride (40% FeCl₃ solution)
- Phoslock
- Phoslock+ (combination Phoslock and PAC)
Aluminum salts
The addition of aluminum salts is widely used as restoration measure. They are colloidal,
amorphous flocs with high affinity to coagulate and adsorb to phosphorus. When settling at
the water bottom they remove particles such as algae and phosphorus from the water
column congregating around them (Cooke et al. 1993, Rydin and Welch 1998). Dose
determination is lake specific due to alkalinity and sediment-mobile P (Kennedy and Cooke
1982). If aluminum chloride is added to the water acids are produced during hydrolysis
reaction. A low or moderate alkalinity can cause a significant decrease in pH if alum is
applied (Cooke et al. 2005).Treated waters have to have adequate buffer capacity to
neutralize acids. Otherwise pH drops immediately. Lime such as NaOH are added to prevent
drop in pH (Van Oosterhout and Lurling 2011).They are redox-insensitive and form flocs
even under anaerobic conditions. The pH determines the actual product available in the
water column: At pH < 4 hydrated and soluble Al3+ ions and at pH 4-6 various soluble
intermediate forms are present, at pH 6-8 insoluble aluminum hydroxides [Al(OH)3] and at pH
> 8 aluminum ions dominate. A low or moderate alkalinity (which can also be described as
low buffer capacity) can cause a significant decrease in pH in surface and sediment pore
water. This would result in formation of toxic aluminum hydroxides [Al(OH)3 ]. Best results are
achieved between pH 6 and 8 (Cooke et al. 1993). Furthermore manufactures produce
‘prehydrolysed’ aluminum salts, such as polyaluminum chlorides, by adding a base to a
concentrated aluminum salt solution (Jiang and Graham 1998).
Iron chloride
Iron chloride treatment is redox-sensitive. Oxidized iron (Fe3+) is able to bind phosphate,
binding is optimal between pH 5 and 7. Under anaerobic conditions iron is reduced (Fe2+)
and phosphate is released (Cooke et al. 1993). Sulfides originated from sulfate reduction in
anaerobic sediment compete with phosphate for iron adsorption sites. High iron
concentrations are favored to compensate consumption by phosphate and sulfides (Geurts
2010a). Smolder et al (2001) investigated the ability of different iron compounds to bind
phosphorus. Iron(II) and iron(III) successfully decreased phosphate concentrations. FeSO4,
10
however, was not preferred because sulfate addition increased alkalinity which stimulates
decomposition of peat.
Phoslock
Phoslock is a lanthanum modified bentonite clay developed by the Land and Water Division
of Australia’s CSIRO (Commonwealth Scientific and Industrial Research Organisation). The
aim of this product is to reduce the amount of filterable reactive phosphorus (FRP) in rivers,
lakes, and other water bodies being present in the water column and in the sediment pore
water. Phoslock is available in granular form dispersing evenly in fine particles and spreading
evenly in the water. Settling of the material takes place in a few hours causing a very turbid
water body in the first 2 to 3 hours after application (Haghseresht 2006). Lanthanum (La3+ ) is
a rare earth element and might be toxic to some aquatic organisms. However, under nonsaline conditions La3+ is strongly bound to clay particles and release to the water column is
quite unlikely (NICNAS 2001).
Phosphate molecules are adsorbed to lanthanum.
La 3+ + PO43– = LaPO4
The very stable mineral formed is called rhabdophane (LaPO4 * n H2O). The formed complex
has a very low solubility constant of KS < 10-23, therefore the bound phosphate is no longer
bioavailable. Furthermore, Phoslock is insensitive to changes in pH, redox potential and
oxygen concentrations (Afsar and Groves 2008).
Phoslock is a brand of the company Phoslock Water Solutions Ltd., Sydney, which
distributes it almost all over the world. In Germany, Austria, and Switzerland it is sold under
the name Bentophos by Institute Dr. Nowak, Ottersberg.
Many trials with Phoslock have been performed by CSIRO as well as Institute Dr. Nowak. In
Australia Phoslock has been applied to two western Australian waterways in the summer of
2001/2002. In the Vasse River it has clearly been shown that phytoplankton growth is
reduced by limitation of SRP (soluble reactive phosphorus). In comparison with the untreated
area less blue-green algae have established. However, species of blue-green algae differed
between treated (non-nitrogen fixing) and untreated areas (nitrogen fixing). On the other
hand, the Canning River did not show significant differences between treated and untreated
areas possibly caused by nitrogen being the limiting factor instead of phosphorus (Robb et
al. 2003).
The hypertrophic Hartbeespoort Dam in South Africa was location of another trial taking
place in 2006 being dominated by Microcystis aeruginosa for most of the year. A successful
reduction of phosphorus levels was recognized without influence on pH or nitrate
concentrations. Phosphorus concentrations also remained low after increasing water
circulation in the winter (Ross 2006).
In Germany Bentophos was applied to Lake Großer Bärensee in 2007 by Institute Dr.
Nowak. Ortho-phosphate levels stayed below detection level, no algae scums were formed,
and phosphorus release from the sediment could be prevented by a 1 mm thick layer of
Bentophos on the sediment. Similar results were found for the trial in Lake Silbersee
(Anonymous 2008a, b). In the Netherlands a successful trial of Phoslock in combination with
the flocculant PAC39 was performed by Wageningen University in Lake Rauwbraken in 2008
11
which was dominated by blue-green algae Aphnanizomenon and Planktothrix rubescens.
The combination is called Flock & Lock and has the purpose to first precipitate present
phosphorus (dissolved and particulate) and bind it to the modified lanthanum clay Phoslock
(Lurling and Van Oosterhout 2009).
Additionally, toxicity tests with Phoslock were conducted as part of the trials to study the
toxicity of lanthanum to aquatic organism and humans. Studies with Lanthanum proved that
due to its very low bioavailability it is not toxic to humans (Persy et al. 2006). Other acute and
chronic toxicity tests with the cladoceran Ceriodaphnia dubia and the juvenile eastern
rainbow fish Melanotaenia duboulayi were performed. They result in a minimal risk of acute
or chronic toxicity to freshwater organisms (Stauber and Binet 2000). Furthermore,
lanthanum is also used in medicine: it is contained in the medical product Fosrenol™ for
patients with urinary failure. No accumulation of lanthanum could be detected in animals and
humans due to its excretion (Afsar and Groves 2008).
1.2.
Aim of the research
The main aim of this research is to investigate the efficiency of phosphorus binding agents
added to a peat lake Terra Nova. Little is known about the addition of lanthanum containing
clays and aluminum to peat lakes and their effect on the decomposition of peat. Furthermore,
restoration measures aim to increase the biodiversity. As submerged macrophytes play an
important role in shallow lakes it is interesting to investigate the effect of different treatments
on the growth of a macrophyte.
Research questions
1) What is the effect of chemical treatments including iron chloride (FeCl3), aluminum
chloride (AlCl3), poly aluminum chloride (PAC), Phoslock and Phoslock+ on the
phosphorus mobilization in a shallow peat lake?
2) What is the relative effect of chemical treatments on the phosphorus availability for
submerse macrophyte Elodea nuttallii?
Hypotheses
1)
Phosphorus is stored in the sediment, in algae and bacteria. Sediment-bound P is
redox-sensitive if phosphorus is released to the water column under reducing conditions
(Akhurst et al. 2004). By applying phosphor-binding agents the release of phosphorus is
prevented. Phosphorus limitation will prevent algae blooms in the water bodies. These
agents have different characters and differ in efficiency at different waters and sediments. It
can be assumed that all treatments are able to bind phosphorus. Treatments with aluminum
and Phoslock have the best binding capacity. Phoslock contains lanthanum ions which bind
irreversibly phosphate by formation of the mineral rhabdophane (Haghseresht 2006).
Flocculants such as PAC (polyaluminum-chloride) and aluminum chloride contain aluminum
ions forming aluminum hydroxide in the presence of water. Aluminum hydroxide flocculates
in the water, binds phosphate and removes particles from the water column resulting in
decreasing turbidity (Cooke et al. 1993). Environmental conditions such as redox potential,
pH, sulfate and oxygen concentrations influence the efficiency of treatments. Flocculants are
more sensible to variations in pH and have an optimal wide between 6 and 8. Otherwise
12
aluminum hydroxide will convert in other forms toxic to aquatic organisms (Cooke et al.
1993). In soft waters with low buffer capacity addition of flocculants reduces pH which
increases the formation of toxic forms such as Al3+. Phoslock is not influenced by alkalinity
nor pH (Afsar and Groves 2008). Iron chloride, however, is sensitive to anaerobic conditions.
In aquatic sediments phosphorus exists in different forms, so called P fractions (Golterman
1996). The interaction of different P fractions and the influence of alkalinity as well as
sulfates are shown in figure 5. Phosphorus bound to organic matter. The organic P fraction
includes organic and other refractory P and will be mobilized if decomposition takes place.
The mobilized phosphorus will be loosely adsorbed to surfaces unless there is iron,
aluminum or calcium available and can diffuses easily from the pore water to the surface
water. Inorganic P is bound to metal oxides such as aluminum and iron. The calcium-bound
P fraction includes Ca-and Mg-bound P in the form of carbonates and apatites (Wauer et al.
2005).
Sulfate concentrations can influence the efficiency of iron chloride treatment resulting in an
increase of loosely bound phosphorus which can easily be released to the surface water.
Aluminum bound phosphorus is insensitive to sulfate concentrations. However, pH can
decrease if alkalinity is low resulting in a toxic environment due to formation of aluminum
hydroxides for flora and fauna. Acidic conditions and low buffer capacity promote peat
accumulation (Smolders et al 2006).
Phoslock and liming both result in general in long-term mobilization of nitrate and ammonium
caused by an increase in both pH and alkalinity after treatments. Eventually peat
decomposition and mobilization of nutrients increase (Geurts et al. 2011).
Figure 5 Effect of alkaline and
sulfate rich water on the different
phosphorus fractions in organic
sediment. The P fractions are
extracted according to Goltermanmethod (Golterman 1996). The
extractions used for each fraction
are given between brackets. The
oxalate extraction is needed to
differentiate between Fe- and Albound phosphorus but this step
has
no
meaning
for
the
understanding of the phosphorus
fractions. The symbol “+” stands
for ‘increase’ whereas “-“ is
interpreted
with
‘decrease’
(diagram taken from Poelen et al.
2011)
If decomposition of organic matter continues more nutrients are released and therefore more
phosphorus-binding agents are needed. This will favor even higher decomposition rates. In
case of Phoslock an increase in alkalinity (or bicarbonate) could result in mobilization of
nutrients on the long term.
2)
It is assumed that treatments with aluminum and Phoslock have a negative effect on
the amount of bioavailable phosphorus for Elodea nuttallii resulting in a lower biomass and a
higher root:shoot ratio.
13
Elodea nuttallii (Planch.) St. John (Nuttall’s waterweed) is a fast growing submersed
macrophyte being able to outcompete other species. They dominate in waters with sediment
pore water concentrations between 1 and 100 µmol P L-1 (Lamers et al. 2010b). Nuttall’s
waterweed can exist in eutrophic as well as clear oligotrophic waters putting it in an exclusive
situation compared to other submerged macrophytes.
For the experiment a fast growing plant is needed functioning as phytometer. The presence
of E. nuttallii, its availability at the university and its character make it an ideal choice for this
research project. In a highly eutrophic environment E. nuttallii takes up nutrients via the
leaves. However, in nutrient-poor water, which will be the case during this experiment, E.
nuttallii will meet its nutrient requirement by uptake via the roots. It can be assumed that E.
nuttallii does not leach phosphorus via the shoots to the water phase (Angelstein and
Schubert 2008). It can be assumed that plants will be more vulnerable in a toxic environment
under nutrient poor conditions. For this reason half of the columns will receive an extra
100µmol NaH2PO4 /L.
Addition of phosphorus-binding agents will have an effect on the bioavailable phosphorus
resulting in different shoot and root length as well as biomass of E. nuttallii.
Phoslock treatment leads to the formation of rhabdophane. Phosphate is irreversible bound
and not bioavailable anymore (Haghseresht 2006). Roots of submerged macrophytes
release oxygen to anaerobic sediment which results in enhanced phosphorus immobilization
in iron-rich sediments. Aluminum is redox-insensitive, therefore no effect of aeration by roots
is assumed.
It is known that low pH conditions occurring during and after aluminum application can
generally cause toxic effects resulting in a negative effect on the biomass (Cooke et al.
1993). However, focus of this study lies on the effect of phosphorus binding agents on the P
availability for E. nuttallii. So far, there were no studies found which investigate the P uptake
by submerged macrophytes in aluminum treated water bodies. According to Barko et al.
adaptation of aquatic macrophytes to sediment fertility by adjusting the root:shoot ratio can
be expected (Barko et al. 1991). This relationship is symbolized in figure 6.
Figure 6 Relationship between macrophyte root:shoot ratio and
sediment fertility according to
Barko et al. (Barko et al. 1991)
Lanthanum and aluminum have strong ionic binding characteristics which will lower the P
availability for macrophytes. Therefore the sediment fertility according to the idea of Barko et
al. will decrease and a higher root:shoot ratio can be expected. Furthermore, it is expected
that plants growing on sediment with extra P loading will result in lower root/shoot ratios.
14
2. Materials and methods
2.1.
Study area
The study is performed with sediment from Lake Terra Nova (52º13’N, 5 º02’E), also called
Lake Loenderveen West (Van de Haterd and Ter Heerdt 2007). The lake is part of the
‘Loosdrechtse Plassen’ located between Amsterdam and Utrecht in the Western part of the
Netherlands (Fig. 1). The Loosdrecht lake system is managed by water company ‘Waternet’.
Amsterdam
Utrecht
Figure 1 Location of peat lake Terra Nova between Amsterdam and Utrecht in the Loosdrechtse
Plassen (source: Google Maps).
2.2.
Field work
In summer 2010 a total of 24 mesocoms was installed in the untreated peat pond to
determine the gradual addition of five different chemical treatments on phosphorus release
from peaty sediments (Fig. 2).
15
plas 2
plas 1
**
**
Figure 2 Aerial image of Lake Terra Nova (source: Google Maps). The yellow pins show the location
of the two experimental closed-off peat ponds (Dutch: petgat). The 2 locations in lake Terra Nova are
marked by stars and called “plas1”and “plas2”.The picture on the lower right presents a close-up of the
two experimental ponds. The white circle points out the location of the mesocosm experiment, white
asterisks represent core sampling sites in both peat ponds.
In a pilot experiment performed in summer 2009 in two closed-off peat ponds of similar size,
the effect of gradual addition of iron on the phosphorus immobilization was investigated. In
total 85 g Fe m-2 were added gradually to the treated peat pond. The untreated peat pond
functioned as control (Saris 2011). After successful pilot the entire lake was treated with iron
chloride. In Lake Terra Nova approximately 33 g Fe m-2 were added to the surface water
within 1.5 years (table 1). Supplementation was performed with a floating windmill located in
the North-East of Lake Terra Nova (Fig. 2).
Table 1 Overview sampling locations and treatments
Location
Closed-off peat ponds
(pilot)
Lake Terra Nova (wholelake experiment)
Mesocosm experiment
Amounts (mol m-2)
Period
August – September
2009
May 2010 –
November 2011
August 2010 –
November 2011
1.5 mol m
-2
~ 85 g Fe m
0.6 mol m
-2
~ 33 g Fe m
1.79 mol Fe/Al/
-2
La m
16
Amounts (g m-2)
-2
-2
~ 100 gFe/ 54 g Al/
-2
250g La m
Until sampling approximately 20 g Fe m-2 were applied to Lake Terra Nova. The mesocosms
were treated gradually with different binding agents between May 2010 and November
2011(Table 2). In each enclosure (0.785 m-2) 1.41 mol Fe/Al/La will have been added after
10 applications. This results in a total amount of 1.79 mol Fe/Al/La m-2 which corresponds
with suggested dose of 100 g Fe m-2 used by Boers et al (1994). Time of sediment core
sampling occurred after the sixth addition which correspondents with 0.843 mol Fe/Al/La per
enclosure.
Table 2 Amount of binding agents added monthly per cylinder
Chemical
100%
Unit
10% agent/ mesocosm/ month
April 2011
FeCl3
0,42
L
0.042
0.250
AlCl3
PAC
Phoslock
Phoslock+
0,19
0,46
3,90
1,31
kg
L
kg
kg
0.019
0.046
0.390
0.131
0.112
0.278
2.342
0.785
The experimental setup of the mesocosms in the untreated peat pond is shown in figure 3.
On Monday, April 4th 2011, and Thursday, April 7th 2011 undisturbed sediment cores were
sampled with a piston sampler (Ø 6cm) and sediment cores were immediately transferred to
glass cylinders (Ø 6cm, height 50cm). Two sediment cores were sampled in each of the 24
mesocosms. A maximum amount of two sediment cores per mesocosm was set to avoid too
much disturbance of the sediment. Within the untreated peat pond two sediment cores were
taken at each location from 25 to 28.
28
28
Cilinderexperiment Terra Nova
24 con
23 PL
22 FeCl
21 PAC
2626
20 PL+
19 AlCl
18 PL
7 con
17 PL+
Figure 3
The setup for the mesocosm
experiment.
PAC=Poly aluminum chloride,
AlCl=Aluminum chloride, FeCl=Iron
chloride, PL=Phoslock,
Pl+=Phoslock+,
Con=Control.
8 AlCl
16 PAC
9 PL
10 PAC
11 PL+
27
27
12 AlCl
13 PL
14 con
6 FeCl
15 FeCl
5 AlCl
4 PL+
3 FeCl
2 con
1 PAC
25
25
Furthermore, in both experimental peat ponds four replicates were taken along the side to
allow comparison with a mobilization study done in 2010 (Voerman 2010) (Fig 2, lower right
picture “asterisks”). Sediment cores were also taken from 2 locations (Fig. 2 “plas1”, “plas2”)
in the lake Terra Nova each with 4 replicates, to investigate the effect of gradual addition of
iron chloride to Lake Terra Nova on phosphorus mobilization rates. In total this results in 68
sediment columns.
17
Also on April 7th 2011, one sediment sample was taken with a piston sampler in each
mesocosm and at 3 locations in the untreated peat pond (locations 25-28). They were stored
in air tight bags at 4°C until further digestion and P-fractionation analysis.
The sediment cores were placed in a dark climate room with a constant temperature of 15°C
at the Radboud University Nijmegen.
2.3.
Experimental setup and sampling
A week after sampling the sediment cores, the original lake water was carefully replaced by a
standard solution containing 1.2 mmol/L CaCl₂, 2.0 mmol/L NaHCO₃ and 0.25mmol/L
MgCl₂*6H₂O (height 20cm, V ~ 0.57L).
Surface and sediment pore water were collected anaerobically using 40 mL vacuum serum
bottles according to scheme given in table 3. Bottles are connected to Rhizon soil moisture
samplers (Eijkelkamp Agrisearch Equipment, Giesbeek, Netherlands) placed 1 cm below and
5 cm above the sediment layer and fixed by a wooden stick (Fig. 4 & 5).
After each sampling the water layer was carefully filled up with the standard solution.
Transpiration losses were compensated with demineralized water.
Table 3 Sampling scheme experiment 1 and type of analyses for surface water and sediment pore water
Surface water
pH
Alkalinity
CO₂, HCO₃
ICP, Auto Analyzer
Sediment pore water
pH
Alkalinity
CO₂, HCO₃
ICP, Auto Analyzer
Sulfides
week 0
x
x
x
x
X
x
week 1
x
x
x
x
x
x
x
x
week 2
x
x
week 4
x
x
x
x
x
x
x
18
week 6
x
x
x
x
week 8
x
x
x
X
x
x
x
x
x
x
x
x
x
Figure 4 Setup of a sediment tube in the
climate room and locations of the Rhizons
(white) to sample sediment pore water
and surface water.
20 cm
5 cm
sediment layer
1 cm
6 cm
Figure 5 Experimental setup of sediment tubes
19
At the end of the first experiment the decomposition rate was measured. In each cylinder a
syringe with a known diameter and volume (V = 60ml) was put on top of the water layer and
fixed to avoid any movement. Gas samples were taken at T0, after 1 hour (T1), 2 hours (T2)
and 4 hours (T4) by using syringes with a volume of 1 ml. Inorganic carbon and methane of
the gas samples were measured immediately with Infrared Gas Analyzer (IRGA, ABB
Advance Optima) (Fig. 6).
Figure 6 Measuring of decomposition rate
In the second experiment the effect of different chemical treatments on the growth of Elodea
nuttallii was tested. The remaining water layer from the first experiment was removed. Half of
the cylinders (even numbers) were loaded with 100 µmol NaH2PO4 L-1 by using a syringe
with a volume of 60 ml and an injection needle of 10 cm length attached to a plastic tube to
be able to load the entire sediment column. Sediment columns from the cylinder mesocosms
at Terra Nova (24 with each two replicates) and from both experimental peat ponds (each
four replicates) were used for inserting the plants.
E. nuttallii plants were collected from a storage basin in the Botanical garden of Radboud
University. The plants originated from “De Bruuk” in Groesbeek and were cultivated in the
botanical gardens. Plants were washed with tap water to remove algae as far as possible.
Apical shoot tips with a length of 13 cm were planted without roots and ramifications into the
sediment columns (total FW per column 0.31 ± 0.014 g). The remaining 12 columns taken at
2 locations in Lake Terra Nova (“plas 1”, “plas 2”) and the untreated peat pond (location 2528) were used to investigate the leaching of phosphorus via the sand layer. This resulted in
56 sediment columns, which each contained one individual of E. nuttallii, and 12 sediment
columns without plants.
A layer of sand with a thickness of 3 cm (140 ± 1.6 g) was added to all 68 sediment columns
without covering E.nuttallii. The layer of sand with low lime and P content is expected to
cover the sediment layer and to avoid mobilization of nutrients via the sediment. The sand
originated from the “Hatertse Vennen” and has been studied before by research institute BWARE as a measure to cover lake sediment (Van Diggelen et al. 2010). An extra rhizon was
added approximately 3 cm above the sediment layer to sample surface water (Fig. 7 & 8).
The water layer of approximately 20 cm consisted of the same standard solution as the first
experiment and contained 1.2 mmol/L CaCl₂, 2.0 mmol/L NaHCO₃ and 0.25mmol/L
MgCl₂*6H₂O.
20
Figure 7 Setup of a sediment tube in
the climate room and locations of the
rhizons (white) to sample sediment
pore water and surface water. The
plant is located in the middle of the
cylinder. A layer of approximately 3
cm sand covers the sediment.
20 cm
3 cm sand
6 cm
.
Figure 8 Beginning of experiment 2 showing
sediment columns with sand layer and Elodea
nuttallii
21
The sediment column was covered with dark foil to avoid algal growth. Columns were kept in
the same climate room at constant temperature of 15°C and irradiated in a 10/14h dark/light
cycle with a light intensity 5 cm below water surface of 230 ± 21 µmol m –2 s–1 during the
experiment. Sampling was done as described in the first experiment following the scheme
given in table 4. After sampling, the water layer was filled up with standard solution and
transpiration losses were compensated with demineralized water.
Table 4 Sampling scheme experiment 2 and type of analyses for surface water and sediment pore water
Surface water
pH
Alkalinity
CO₂, HCO₃
ICP, Auto Analyzer
Sediment pore water
pH
Alkalinity
CO₂, HCO₃
ICP, Auto Analyzer
week 0
x
x
x
x
X
x
week 1
x
x
x
x
x
x
x
x
week 2
x
x
week 3
x
x
x
x
week 4
x
x
x
x
x
x
x
x
At the end of the second experiment the following biological variables were measured:
Lengths of shoot (including ramifications), number of ramifications, fresh weight of aboveand belowground biomass. Root lengths and diameters were determined by analyzing the
images with WinRHIZO (Reg 2005c, Regent Instruments Inc., Quebec, Canada). All roots
and shoots were dried at 70°C for at least 48 h and dry weight was determined. Sediments
were stored in air tight bags at 4°C until further analysis.
2.4.
Chemical analyses
After sampling, 10 ml sample were used to measure alkalinity and pH with a TIM800 titration
manager (Radiometer, Copenhagen, Denmark). Alkalinity was determined by titration to pH
4.2 with 0.01 M HCl using an ABU901 Autoburette.
Colorimetrical analysis of PO4, NO3 (including NO2), NH4 and Cl of water samples, which had
been stored in polyethylene bottles (30 mL) at -20ºC, was done with the Autoanalyzer 3
(Bran + Luebbe, Norderstedt, Germany) according to Geurts et al (2008). For determination
of Al, Ca, Fe, La, Mg, Mn, S, Zn and P, a subsample of 10 ml surface water and pore water
were stored in polyethylene sample tubes at 4ºC. 100µL nitric acid (65% HNO3) were added
to prevent metal precipitation and to conserve the sample. Furthermore, addition of nitric acid
improves nutrient analysis which was carried out with an ICP Spectrometer (IRIS Intrepid II,
Thermo Electron Corporation, Franklin, MA).
On the last sample day, in addition to the other analyses, 10.5 mL of pore water sample was
immediately fixed after sampling with 10.5 mL of sulphide antioxidant buffer (SAOB). On the
same day, sulphide concentrations were measured using an ion-specific Ag electrode
(9416BN Orion Research, Beverly, CA, USA) and an Ag/AgCl double junction reference
electrode (900200 Orion Research, Beverly, CA, USA).
Sediment samples were dried for at least 24 h at 70°C to determine the moisture content.
The organic matter content was determined after heating dry sediment samples at 550°C for
4 h. Homogenized portions of 200 mg dry sediment were digested with 4 mL HNO 3 (65%)
22
and 1 mL H2O2 (30%), using an Ethos D microwave (Milestone srl, Sorisole, Italy). Digestates
were diluted and concentrations of Al, Ca, Fe, K, La, Mg, Mn, Na, P, S, Si and Zn were
determined by ICP as described above.
Chlorophyll-a was measured at 649, 665 and 750nm with a spectrophotometer (Shimadzu
UV-1205, Kyoto, Japan).The vitality of the plants was investigated with a chlorophyll
fluorometer (Junior PAM, Walz GmbH, Effeltrich, Germany). Plant material was dried for at
least 24h at 70 °C and weighed to determine aboveground and belowground dry weight.
Homogenized portions of 200 mg dry plant material was digested as described above.
Digestates were diluted and nutrient concentrations of plant material were determined by ICP
as described above.
Sediments were analyzed according to the Golterman P-fractionation method to determine
the different phosphorus fractions (Golterman 1996). Homogenized portions of 5g fresh
sediment were sequentially extracted with NH4Cl (1M), Ca-EDTA (0.05M) and Na-EDTA
(0.1M). Samples were centrifuged (Sorvall 10,000 rpm) after every step and supernatants
were stored separately at 4°C. The remaining pellet was digested and diluted. Supernatants
and pellet were analyzed by ICP as described above.
23
2.5.
Statistical analysis
Potential phosphorus and phosphate mobilization rates were calculated in mmol m-2 y-1 using
linear regression of surface water concentrations between t=0 and t=57 days. By
investigating the data set of each individual column the best fit of phosphorus and phosphate
measurements to the regression line was determined and regression coefficient calculated.
In some cases the presence of high peaks especially for phosphate concentrations
complicated the determination of regression coefficients and it was decided to use the first
part of the peak for analysis. The same method was applied for nitrate and ammonium
mobilization rates. Both rates were added per column to calculated total nitrogen mobilization
rate. Duplicate samples (pseudoreplica) were averaged for each cylinder. For every time
step ratios of Fe:TP and Fe:PO4 were calculated. Averaged total phosphorus and phosphate
concentrations and iron ratios of sediment pore water were used for statistical analysis.
Mobilization rates were statistically analyzed in µmol L-1d-1 and nutrient concentrations in
µmol L-1. Data were log10(x+1) transformed if homogeneity of variances was violated and
otherwise Welch’s F was used. Differences in treatments were investigated using one-way
ANOVA followed by LSD or Bonferroni post-hoc test. If transformation was not acceptable
nonparametric Kruskal-Wallis test was performed followed by multiple Mann-Whitney U tests.
An independent t-test was used to test differences between control and untreated peat pond
which functions as outside control to investigate cylinder effects. Experimental data for
phosphate and total phosphorus mobilization were tested by one-sample t-test and test value
0 to investigate possible real mobilization.
Time effects of treatments on biochemical variables were tested with GLM repeatedmeasures ANOVA followed by 2-sided Dunnett t post-hoc. All data were log10(x+1)
transformed for a better fit to normal distribution. However, Levene’s test was still significant
in most cases. If Mauchly’s test was significant and therefore assumption of sphericity of data
violated, the Greenhouse-Geisser correction was used to produce a valid F-ratio. Repeated
measures were done with time steps T0, T1, T2, T4, T6 and T8 for surface water, and for
pore water time steps T0, T1, T4, T6 and T8 were used.
In experiment 2 statistical analysis has only been performed with data from cylinder
treatments (df=5), unless explicitly mentioned. Differences in biomass as well as nutrient
concentration in biomass and sediment were tested with two-way ANOVA with treatment
(consisting of levels AlCl, PAC, PL+, PL, FeCl and control) and PO 4 loading
(loaded/unloaded) as fixed factors. If homogeneity of variances was not given (significant
Levene’s test), a non-parametric Scheirer-Ray-Hare test was performed instead.
After splitting data in unloaded and loaded sediments, differences between treatments were
tested with one-way ANOVA. If variances were unequal, nonparametric Kruskal-Wallis test
followed by multiple Mann-Whitney U tests was performed.
For clarity of presentation, results and figures present non-transformed data expressed as
mean ± standard error of mean (SEM). For all tests a significance level of α=0.05 was
assumed. SPSS (ver. 17.0, 2008, SPSS, Chicago, IL, USA) outputs are given in the
appendix.
24
3. Results
3.1.
Experiment 1
3.1.1. Mesocosm experiment
Mobilization rates
Treatments containing aluminum and/or Phoslock had significant lower phosphate
mobilization rates compared to the control treatment, which had the highest mobilization rate
(7.7 ± 3.5 mmol m-2 yr -1). Lowest mobilization rates were found in PAC and Phoslock
treatment with respectively -1.4 ± 0.4 mmol m-2 yr -1 and -1.1 ± 0.5 mmol m-2 yr -1. Both AlCl
and Phoslock+ showed the same phosphate mobilization rate (-0.6 mmol m-2 yr -1).
Treatment with FeCl (0.3 ± 1.0 mmol m-2 yr -1 ) was not significant different from control or
other treatments (Fig. 1).
Phosphorus mobilization rates of PAC, Phoslock and Phoslock+ were negative and
significantly lower than in the FeCl treatment (4.6 ± 3.5 mmol m-2 yr -1) and control (8.7 ± 3.6
mmol m-2 yr -1). Treatments with aluminum and/or Phoslock obviously resulted in negative
mobilization rates (so called “fixation”) of phosphorus and phosphate whereas FeCl
treatment, control and untreated peat pond showed positive mobilization rates. The untreated
peat pond functioned as outside control to investigate possible cylinder effects. Mobilization
rates of phosphorus (2.6 ± 0.6 mmol m-2 yr -1) and phosphate (1.2 ± 1.5 mmol m-2 yr -1) in
sediments of the untreated peat pond did not differ significantly from the control cylinder (Fig.
1).
To test if there was a significant mobilization of phosphate or phosphorus from the sediment
a one-sample t-test with test value 0 was performed. However, concerning phosphate
mobilization only PAC (df=3) was significant. In this case it can be assumed that during the
experiment in columns from PAC cylinders a significant fixation of phosphate to the sediment
has taken place. The one-sample t-test with phosphorus data was significant only for the
untreated peat pond (df=3). It can be assumed that during the experiment a significant
phosphorus mobilization from the sediment only occurred in sediment columns from the
untreated peat pond.
25
C
b
BC
ab
a AB
a
A
a
A
a
A
Figure 1 Effect of treatments on total phosphorus and phosphate mobilization rates (mean
mobilization ± SEM) in cylinders in comparison with the untreated peat pond PG-Fe (N=4). Differences
between cylinders were tested. Different lowercase letters indicate significant differences in phosphate
rates (Kruskal-Wallis, p < 0.05, multiple Mann-Whitney U tests). Different capital letters indicate
significant differences in total phosphorus rates (one-way ANOVA, p<0.01, followed by LSD post-hoc
α = 0.05). There was no significant difference between control and untreated peat pond (Independent
t-test).
Variances of phosphate mobilization rates differed remarkably from each other between
treatments in comparison to variances of total phosphorus rates. In general total phosphorus
mobilizations rates had homogeneous variances as shown in figure 2 on the basis of mean
mobilization rates and their standard deviations. Also Levene’s test was not significant for
total phosphorus mobilization rates (FTP(6,21)=1.75, p=0.16).
For phosphate mobilization rates Levene’s test was highly significant (FPO4(6,21)=5.86,
p=0.001) and data transformation did not result in homogeneity of variances:
FlogPO4(6,21)=5.96, p=0.001). Especially untreated peat pond PG-Fe, control and FeCl
treatment showed much bigger variances compared to other treatments. As mentioned
earlier, determination of regression coefficients for phosphate mobilization rates was
complicated due to high phosphate peaks in columns from control, untreated peat pond and
in columns with iron supplementation.
26
Multiple independent t-tests were performed to investigate if variances of phosphate and total
phosphorus mobilization rates differed significantly between treatments (table 1). In general
variances of total phosphorus mobilization rates did not differ significantly between
treatments. However, variances of phosphate mobilization rates of untreated peat pond and
control differed significantly from other treatments. Columns with iron supplementation did
not show significantly larger variances in comparison with other treatments. This will further
be reviewed in the discussion.
treat1
PL+
AlCl
PAC
PL
PL
con
con
con
con
PL+
con
AlCl
PAC
con
PL
FeCl
PL
PL
PAC
PL+
PAC
Figure 2 Error bars with mean phosphorus and phosphate
mobilization rates ± standard deviations of cylinders in comparison
with the untreated peat pond PG-Fe (N=4).
treat2
PG-Fe
PG-Fe
PG-Fe
PG-Fe
PL+
PL+
FeCl
PAC
PL
FeCl
FeCl
FeCl
FeCl
PG-Fe
AlCl
PG-Fe
PAC
FeCl
PL+
AlCl
AlCl
p (PO4)
0,002
0,005
0,006
0,009
0,015
0,031
0,038
0,039
0,049
0,072
0,088
0,142
0,157
0,174
0,191
0,200
0,256
0,309
0,337
0,358
0,922
p (TP)
0,124
0,411
0,077
0,159
0,845
0,386
0,054
0,574
0,295
0,497
0,878
0,089
0,700
0,032
0,281
0,056
0,607
0,392
0,748
0,217
0,133
Table 1 Bold values indicate
significant differences in
variances (p < 0.05, Levene’s
test as tested by multiple
independent t-tests).
Sediment pore water concentrations
Pore water phosphate concentrations were highest in control (12.8 ± 5.7 µmol L-1) and FeCl
treatment (2.0 ± 0.8 µmol L-1). Treatments with aluminum and/ or Phoslock resulted in much
lower average phosphate concentrations (0.2 - 0.4 µmol L-1). In general. phosphate pore
water concentrations of treated cylinders were significant lower than control cylinders (Fig 3).
Phosphate concentration of untreated peat pond (8.9 ± 1.0 µmol L-1) did not differ
significantly from control.
Pore water total phosphorus concentrations of control (23.5 ± 8.2 µmol L-1) and FeCl treatment
(22.2 ± 5.7 µmol L-1) were significantly higher compared to other treatments (1.3 – 4.4 µmol L-1).
Total phosphorus concentrations of control did not differ significantly from untreated peat
pond (28.9 ± 3.3 µmol L-1). Lowest pore water concentrations where found in Phoslock+
0.2 PO4 ± 0.03 µmol L-1) and Phoslock (1.3 TP± 0.06 µmol L-1).
27
B
B
b
A
a
A
a
A
a
a
A
a
Figure 3 Effect of treatments on total phosphorus and phosphate sediment pore water concentrations
(mean ± SEM) in cylinders in comparison with the untreated peat pond PG-Fe (N=4). Differences
between cylinders were tested on transformed data with Bonferroni post-hoc (α=0.05). Different
lowercase letters indicate significant differences in phosphate concentrations (Welch’s ANOVA. p<
0.05). Different capital letters indicate significant differences in total phosphorus (one-way ANOVA.
p<0.01). There was no significant difference between control and untreated peat pond (Independent ttest).
Nitrogen mobilization
Repeated measures analysis did not show significant differences in nitrate and ammonium
concentrations in pore water (table 2). Surface water nitrate concentrations were significantly
lower in the Phoslock treatment (2.3 ± 1.1 µmol L-1) in comparison with the control treatment
(16.4 ± 3.7 µmol L-1). The iron treatment resulted in the highest nitrate concentrations in
surface water (19.2 ± 4.3 µmol L-1).
Nitrogen (NO3 + NH4) mobilization from the peat sediments to the water layer varied between
103.3 ± 18.2 mmol m-2 yr -1 (Phoslock treatment) and 413.7 ± 163.4 mmol m-2 yr -1 (control).
However, there were no significant differences between treatments (one-way ANOVA. p >
0.05). Low mobilization rates were also measured in the PAC treatment (118.2 ± 31.2 mmol
m-2 yr -1). Iron and aluminum chloride addition resulted in 259.8 ± 30.5 and 296.2 ± 108.3
mmol m-2 yr -1, respectively. Furthermore, nitrogen mobilization in the untreated peat pond
(156.7 ± 26.6 mmol m-2 yr -1) did not differ significantly from the control (Independent t-test, p
> 0.05).
28
Changes in chemical variables
In general, almost all chemical variables in surface water and sediment pore water changed
significantly during the experiment (Table 2, “time”).
All treatments resulted in significant lower phosphate concentrations in sediment pore water
(0.2 – 2.0 µmol L-1) and in surface water (0.14 – 0.3 µmol L-1) in comparison to control
(respectively 12.8 ± 5.7 and 2.6 ±1.5 µmol L-1, treatment effect, Table 2). In all treatments
with aluminum and/ or Phoslock total phosphorus concentrations were significant lower in
surface water (0.9 – 1.2 µmol L-1) and in pore water (1.3 – 4.4 µmol L-1) compared to control
(respectively 4.4 ± 1.6 and 23.5 ± 8.2 µmol L-1). Iron supplementation did not result in
significant lower total phosphorus concentration in surface water (2.0 ± 0.5 µmol L-1) and
pore water (Fig. 3). Lowest average concentrations in surface water were found respectively
in Phoslock (0.9 TP ± 0.2 µmol L-1) and AlCl treatment (0.2 PO4 ± 0.05 µmol L-1).
Table 1 Statistical analysis of time effects (time). treatment effects (treat) and interaction effects
(time*treat) for several chemical variables. Bold p-values show significant results of GLM repeatedmeasures ANOVA followed by Dunnett t post-hoc test with α=0.05 on log10(x+1) data. Chemical
variables are marked with asterisk (*) if Levene’s test was non-significant after transformation.
Otherwise Levene’s test was significant.
Chemical variable
time
time*treat
treat
Dunnett t with control
Surface water
-1
TP (µmol L )
-1
PO4 (µmol L )
-1
Fe:TP (mol mol )
-1
Fe:PO4 (mol mol )
-1
alkalinity (meq L )
pH*
-1
Fe (µmol L )
-1
Al (µmol L )
-1
Ca (µmol L )
-1
Cl (µmol L )
-1
Mn (µmol L )
-1
NO3 (µmol L )
-1
NH4 (µmol L )
-1
S (µmol L )
0.000
0.017
0.000
0.000
0.000
0.000
0.000
0.003
0.000
0.002
0.000
0.000
0.000
0.000
0.008
0.080
0.012
0.000
0.104
0.815
0.466
0.370
0.698
0.024
0.901
0.071
0.063
0.466
0.004
0.006
0.036
0.003
0.001
0.379
0.525
0.495
0.004
0.000
0.263
0.001
0.191
0.525
AlCl, PAC, PL, PL+
AlCl, FeCl, PAC, PL, PL+
PL
AlCl, FeCl, PL, PL+
AlCl, PAC
Sediment pore water
-1
TP (µmol L )
-1
PO4 (µmol L )
-1
Fe:TP (mol mol )
-1
Fe:PO4 (mol mol )
-1
alkalinity (meq L )
pH
-1
Fe (µmol L )
-1
Al (µmol L )
-1
Ca (µmol L )*
-1
Cl (µmol L )
-1
Mn (µmol L )*
-1
NO3 (µmol L )
-1
NH4 (µmol L )
-1
S (µmol L )
0.001
0.000
0.056
0.000
0.000
0.000
0.002
0.000
0.000
0.000
0.000
0.004
0.000
0.008
0.068
0.007
0.000
0.007
0.171
0.026
0.281
0.717
0.717
0.131
0.503
0.192
0.411
0.057
0.000
0.000
0.000
0.000
0.000
0.001
0.002
0.609
0.021
0.001
0.032
0.330
0.450
0.095
AlCl, PAC, PL, PL+
AlCl, FeCl, PAC, PL, PL+
AlCl, FeCl, PAC, PL, PL+
PL, PL+
AlCl, PAC
AlCl, PAC
FeCl
29
AlCl, FeCl, PAC, PL, PL+
AlCl, FeCl, PAC, PL+
PL
FeCl
AlCl, FeCl, PAC
PL
Pore water alkalinity was significantly lower in treatments with AlCl and PAC (both 1.8 meq L -1)
compared to the control treatment (2.3 meq L-1). Phoslock treatment resulted in the highest
average pore water alkalinity (2.8 meq L-1). Similarly, surface water alkalinity was significantly
lower in treatments with AlCl and PAC (both 1.7 meq L-1) in comparison with control (1.9 meq
L-1). Highest surface water alkalinity was also measured in Phoslock treatments (2.1 meq L-1).
Pore water pH was significantly lower in AlCl and PAC treatments (respectively 6.9 and 7.0)
compared to control (pH 7.2).
Iron concentrations in the pore water of the FeCl treatment (129.2 µmol L-1) were 5 times
higher than average iron concentrations in the control treatment (23.3 µmol L-1). However,
pore water Fe/PO4 ratios were 20 to 40 times higher in all treatments (143.0 – 238.8 mol mol1
) compared to the control (6.3 mol mol-1, Fig 4.). Furthermore, pore water Fe/TP ratios in all
treatments (6.0 – 31.2 mol mol-1) were significantly higher than in the control (1.5 mol mol-1,
treatment effect, Table 2).
a
a
a
a
a
A
A
A
A
AB b B
Figure 4 Average Fe/PO4 and Fe/TP ratios in sediment pore water (±SEM) in cylinders in comparison
with the untreated peat pond PG-Fe (N=4). Differences between cylinders are tested. Differences
between cylinders were tested with one-way ANOVA followed by Dunnett t post-hoc (α=0.05).
Different lowercase letters indicate significant differences in phosphate rates (log-transformed,
p<0.01). Capital letters belong to total phosphorus concentrations (p<0.01). There was no significant
difference between the control cylinder and the untreated peat pond (Independent t-test).
30
The five treatments were characterized by a different chemical composition.
Aluminum and lanthanum were the main component of the Phoslock+ treatment (Phoslock
Water Solutions Ltd). Digestion analysis of 200 mg Phoslock resulted in 5% La. 1% Ca. 0.8%
Al. 0.5% Fe. 0.3% Na. 0.3% Mg and 0.1% K (Geurts et al. 2011). The polyaluminum chloride
solution (Kemira PAX-14 & PAX-18) was mainly characterized by chloride and aluminum.
The iron chloride (40% solution) was characterized by iron and chloride ions whereas the
aluminum chloride (Merck-Schuchardt) was dominated by aluminum and chloride ions.
This resulted in significant differences in sediment pore water concentrations of calcium,
chloride, iron and lanthanum. Differences in aluminum concentrations were not significant. In
general, addition of these chemicals did not result in significant differences in surface water
concentrations, with the exception of chloride concentrations.
Pore water calcium concentrations were significantly higher in the FeCl treatment (1.4 ± 0.06
mmol L-1) compared to the control (1.2 ± 0.08 mmol L-1 ), but did not differ significantly from
the other treatments ranging between 1.2 ± 0.02 mmol L-1 (PAC) and 1.4 ± 0.02 mmol L-1
(Phoslock). Pore water calcium concentrations in the control also differed significantly from
the untreated peat pond (1.7 ± 0.08 mmol L-1). Surface water calcium concentrations ranged
between 1.0 ± 0.03 mmol L-1 (control) and 1.2 ± 0.03 mmol L-1 (FeCl) (Fig. 5a).
Pore water chloride concentrations in the FeCl treatment (3.0 ± 0.2 mmol L-1) and AlCl
treatment (2.9 ± 0.2 mmol L-1) were significantly higher compared to the control (2.2 ± 0.03
mmol L-1) and Phoslock treatment (2.3 ± 0.04 mmol L-1). Treatments with PAC and
Phoslock+ resulted in respectively 2.7 ± 0.06 mmol L-1 and 2.5 ± 0.02 mmol L-1 (Fig 5b).
Surface water chloride concentrations differed significantly from each other. Treatments with
AlCl (2.9 ± 0.07 mmol L-1 ) and FeCl (2.9 ± 0.07 mmol L-1 ) resulted in significantly higher
chloride concentrations compared to the Phoslock treatment (2.6 ± 0.05 mmol L-1) and the
control (2.5 ± 0.02 mmol L-1). The PAC treatment (2.8 ± 0.02 mmol L-1) also differed
significantly from the control but not from the Phoslock treatment. The Phoslock+ treatment
(2.7 ± 0.01 mmol L-1) did not differ significantly from the other treatments. Surface water
chloride concentrations in the control were significantly lower compared to the untreated peat
pond (2.6 ± 0.04 mmol L-1).
Pore water lanthanum concentrations (measured only at T8) were significantly higher in the
Phoslock treatment (0.046 ± 0.0058 µmol L-1) compared to the control (0.0052 ± 0.0029 µmol
L-1). However, higher La concentrations were also found in the Phoslock+ treatment (0.023 ±
0.0036 µmol L-1) compared to the treatments with PAC (0.0099 ± 0.0064 µmol L-1), FeCl
(0.0062 ± 0.003 µmol L-1) and AlCl (0.0055 ± 0.0025 µmol L-1) (Fig. 5c). Surface water
lanthanum concentrations ranged from 0.0038 ± 0.001 µmol L-1 (PAC and control) to 0.0054
± 0.0017 µmol L-1 (Phoslock) and 0.0059 ± 0.001 µmol L-1 (Phoslock+).
Pore water aluminum concentrations did not differ significantly from each other. Treatments
with aluminum chloride, PAC and Phoslock showed the highest aluminum concentrations
with respectively 1.78 ± 0.72 µmol L-1, 2.93 ± 2.33 µmol L-1 and 1.36 ± 0.42 µmol L-1. Lowest
Al concentrations were found in the Phoslock+ treatment (0.54 ± 0.15 µmol L-1). The control
and FeCl treatment resulted in 0.99 ± 0.23 µmol L-1 and 0.68 ± 0.06 µmol L-1, respectively
(Fig 5d).
Surface water aluminum concentrations also did not differ significantly from each other.
Highest aluminum concentrations were found in the control (0.75 ± 0.50 µmol L-1), FeCl
treatment (0.72 ± 0.37 µmol L-1) and Phoslock+ treatment (0.69 ± 0.56 µmol L-1). Lowest
31
surface water aluminum concentrations were measured in the Phoslock treatment (0.34 ±
0.18 µmol L-1) and PAC treatment (0.13 ± 0.03 µmol L-1 ).
Pore water iron concentrations were significantly higher in the FeCl treatment (129.2 ± 47.0
µmol L-1). Lowest iron concentrations were measured in the control (23.3 ± 4.2 µmol L-1)
followed by PAC (26.5 ± 3.7 µmol L-1), Phoslock+ (30.8 ± 2.3 µmol L-1), AlCl (32.1 ± 2.7 µmol
L-1) and Phoslock treatment (38.1 ± 3.1 µmol L-1) (Fig. 5e). Surface water iron concentrations
were higher in the FeCl treatment (6.4 ± 4.8 µmol L-1) compared to the control (1.0 ± 0.5
µmol L-1). Lowest surface water iron concentration were measured in the PAC treatment with
0.9 ± 4.2 µmol L-1.
32
a
*
a
ab
ab
ab
a
*
a
A
ab
b
ab
b
A
AB
ABC
ab
*
B
b
*
C
b
c
d
b
a
a
a
a
a
e
b
a
a
a
a
a
33
Figure 5a-e Concentrations of calcium,
chloride, lanthanum (mean T8 ± SEM),
aluminum and iron (mean ± SEM) in sediment
pore water and surface water in the cylinders
in comparison with the untreated peat pond
(PG-Fe; N=4).
Differences between cylinders were tested
with one-way ANOVA. Different letters
indicate significant differences in pore water
concentrations (p<0.05, Bonferroni post-hoc
test). Pore water iron concentrations were
tested with Kruskal-Wallis followed by multiple
Mann-Whitney U tests. Except for chloride,
surface water concentrations did not differ
significantly from each other. There was no
significant difference between the control
cylinder and the untreated peat pond except
for calcium and
chloride pore water
concentration
(indicated
with
asterisk;
Independent t-test).
3.1.2. Iron supplementation in peat ponds and Lake Terra Nova
Neither phosphate nor total phosphorus mobilization rates differed significantly from each
other. The untreated peat pond showed the highest phosphate mobilization rates (4.1 ± 1.9
mmol m-2 yr-1). The treated peat pond (1.6 ± 0.5 mmol m-2 yr-1) showed higher phosphate
mobilization rates in comparison with the two locations in lake Terra Nova (-2.5 ± 2.5 mmol
m-2 yr-1 for “plas 1” and -0.4 ± 0.5 mmol m-2 yr-1 for “plas 2”; Fig. 6)
Highest total phosphorus mobilization rates were measured in the untreated peat pond (5.6 ±
2.0 mmol m-2 yr-1) and “plas 2” (2.2 ± 2.0 mmol m-2 yr-1). Lowest phosphorus mobilization
rates were found in “plas 1” (-0.9 ± 4.6 mmol m-2 yr-1).
Figure 6 Total phosphorus and phosphate mobilization rates (mean ± SEM) of the untreated peat
pond (PG-Fe, N=8) and the treated peat pond (PG+Fe, N=4) in comparison with two locations in lake
Terra Nova (N=4). No significant differences were found (ANOVA p>0.05).
Pore water total phosphorus and phosphate concentrations differed significantly between
locations (Fig. 7). Highest phosphate concentrations were found in the untreated peat pond
(11.8 ± 1.8 mmol m-2 yr-1) followed by “plas 1” (5.0 ± 0.7 mmol m-2 yr-1) and the treated peat
pond (2.1 ± 0.9 mmol m-2 yr-1). Lowest phosphate concentrations were measured in “plas 2”
(1.4 ± 0.4 mmol m-2 yr-1).
Highest pore water phosphorus concentrations were also found in the untreated peat pond
(38.7 ± 5.7 mmol m-2 yr-1) being the two times higher than in the two locations in lake Terra
Nova (“plas 1” with 22.4 ± 2.2 mmol m-2 yr-1 and ”plas 2” with 16.0 ± 2.0 mmol m-2 yr-1). In
comparison with lake Terra Nova, the treated peat pond also showed higher pore water
phosphorus concentrations (35.5 ± 2.2 mmol m-2 yr-1).
34
A
A
AB
B
a
b
bc
c
Figure 7 Total phosphorus and phosphate sediment pore water concentrations (mean ± SEM) of the
untreated peat pond (PG-Fe, N=8) and the treated peat pond (PG+Fe, N=4) in comparison with two
locations in lake Terra Nova (N=4). Differences between locations were tested on log-transformed
data with a Bonferroni post-hoc test (α=0.05). Different lowercase letters indicate significant
differences in phosphate concentrations (One-way ANOVA, p < 0.01). Different capital letters indicate
significant differences in total phosphorus concentrations (Welch’s ANOVA, p < 0.01).
In general it can be concluded that mobilization rates as well as pore water concentrations
were higher in the untreated peat pond compared to the treated peat pond and Lake Terra
Nova.
Comparison of recent mobilization rates with data from 2010
Gradual FeCl3 addition in the treated peat pond in summer 2009 resulted in 2010 in lower
PO4 mobilization rates (0.11 ± 0.15 mmol m-2 yr-1) compared to untreated peat pond (5.9 ±
3.1 mmol m-2 yr-1). Comparison of PO4 mobilization rates resulted in an increase in both peat
ponds from 2010 to 2011. Differences in PO4 mobilization rates between 2010 and 2011
were significant for treated peat pond. Lake Terra Nova resulted in lower PO4 mobilization
rates in both locations after application of nearly 20 g Fe m-2 until April 2011 (Fig. 8)
35
PO4 mobilization (mmol m-2 yr-1)
25
April 2010
20
April 2011
15
*
10
*
5
0
-5
-10
PG-Fe
PG+Fe
plas 1
plas 2
Figure 8 Phosphate mobilization rates from April 2011 in comparison with April 2010 (Voerman 2010).
Mobilization rates in treated peat pond and plas 1 changed significantly after one year (Asterisk,
Independent t-test).
Phosphorus mobilization rates showed the same pattern as phosphate mobilization rates. An
increase in phosphorus mobilization rate was found in both untreated and treated peat pond.
Differences within the treated peat pond between 2010 and 2011 were significantly different.
Gradual addition of FeCl3 in Lake Terra Nova had positive effects on phosphorus mobilization
rates. At location 1 (plas 1) phosphorus mobilization rates decreased significantly. Also
location 2 (plas 2) resulted in lower phosphorus mobilization rates after approximately 20 g
Fe m-2 had been applied (Fig. 9).
25
*
TP mobilization (mmol m-2 yr-1)
20
*
April 2010
April 2011
15
10
5
0
-5
-10
PG-Fe
PG+Fe
plas 1
plas 2
Figure 9 Phosphorus mobilization rates from April 2011 in comparison with April 2010 (Voerman
2010). Mobilization rates in treated peat pond and plas 1 changed significantly after one year
(Asterisk, Independent t-test).
36
Chemical variables
Significantly higher pore water iron concentrations were found in the treated peat pond
(299.6 ± 77.7 µmol L-1). Compared to the untreated peat pond (50.6 ± 5.5 µmol L-1), pore
water iron concentrations of “plas 2” (106.2 ± 9.5 µmol L-1) were significantly higher whereas
“plas 1” resulted in average concentrations of 61.7 ± 20.7 µmol L-1.
Surface water iron concentrations did not differ significantly, but were 20 to 50 times lower
than pore water iron concentrations. However, highest iron concentrations were also
measured in the treated peat pond ( 12.1 ± 6.5 µmol L-1) compared to the untreated peat
pond (1.07 ± 0.32 µmol L-1), “plas 1” (1.56 ± 0.1 µmol L-1) and “plas 2” (2.96 ± 1.07 µmol L-1).
Significantly higher pore water sulfate concentrations were found in “plas 2” (25.8 ± 4.2 µmol
L-1) compared to “plas 1” (12.9 ± 1.5 µmol L-1) as well as the treated and untreated peat pond
(respectively 12.1 ± 1.6 and 13.2 ± 0.7 µmol L-1). Surface water sulfate concentrations
ranged between 9.2 ± 0.7 µmol L-1 (treated peat pond) and 23.4 ± 5.8 µmol L-1 (“plas 1”).
Ammonium concentrations differed significantly in surface water and pore water and were
highest in the treated peat pond (83.0 ± 20.4 and 349.9 ± 56.7 µmol L-1, respectively).
Ammonium concentrations in surface water and pore water from the untreated peat pond
(respectively 58.5 ± 7.5 and 273.4 ± 32.6 µmol L-1) were almost twice as high as
concentrations measured in sediments from Lake Terra Nova.
Nitrogen mobilization in the treated peat pond (375.2 ± 85.8 mmol m-2 yr-1) was significantly
higher compared than in “plas 1” (52.6 ± 51.7 mmol m-2 yr-1). In comparison, also high N
mobilization rates were measured in the untreated peat pond and in “plas 2” (237.6 ± 32.6
and 145.9 ± 24.2 mmol m-2 yr-1, respectively).
Chloride concentrations did not differ significantly and ranged between 2.1 ─ 2.2 mmol L-1 in
pore water and 2.5 – 2.6 mmol L-1 in surface water. However, pore and surface water
chloride concentrations were in general lower in comparison to the cylinder experiment.
Indicators for phosphate and phosphorus mobilization
Within the mesocosm experiment, PO4 mobilizations correlated best with pore water Fe/PO4
ratios (Fig. 10, R=0.57). At low ratios (<3.5 mol mol-1, control treatment), high PO4
mobilization rates were found. Sediments with phosphate pore water concentrations above 5
µmol L-1 had increased phosphate mobilization rate (Fig. 11, R=0.94). Phosphorus
mobilization rates correlated fairly with pore water Fe/TP ratios (Fig 12, R=0.45). Both,
control and untreated peat pond, showed high mobilization rates at low Fe/TP ratios (< 1.5
mol mol-1). Also sediments treated with FeCl3 resulted in relatively low Fe/TP ratios of
averaged 6.0mol mol-1. However, phosphorus mobilization rates were much higher compared
to untreated peat pond and similar to the control treatment. AlCl3 and PAC treatment
resulted in both low Fe/TP ratios as well as low phosphorus mobilization rates.
37
20
PO4 mobilization (mmol m-2 yr-1)
AlCl
PAC
15
PL+
PL
FeCl
10
con
y = -2,161ln(x) + 10,354
R² = 0,5709
5
0
-5
0
50
100
150
200
250
300
350
Pore water Fe:PO4 (mol mol-1)
Figure 10 Correlation between phosphate mobilization rates and pore water Fe/PO4 ratio within the
mesocosm experiment.
20
PO4 mobilization (mmol m-2 yr-1)
15
AlCl
10
PAC
y = 0,6503x - 1,0115
R² = 0,9353
PL+
PL
5
FeCl
con
0
-5
0
5
10
15
20
25
pore water PO4 (µmol L-1)
Figure 11 Phosphate mobilization rates as function of phosphate pore water concentrations within
mesocosm experiment
38
30
20
AlCl
PAC
15
PL+
TP mobilization rate (mmol m-2 yr-1)
y = -4.113ln(x) + 7.3404
R² = 0.445
PL
10
FeCl
con
5
PG-Fe
0
0
5
10
15
20
25
30
35
40
45
-5
-10
-15
-20
Pore water Fe:TP (mol mol-1)
Figure 12 Correlation between phosphorus mobilization rates and pore water Fe/TP ratio within the
mesocosm experiment and untreated peat pond
Decomposition rates
Decomposition (mmol m-2 yr-1)
Neither iron supplementation nor addition of other chemicals to the mesocosms in the field
did have an effect on the decomposition rates (Fig. 13).
Figure 13 Decomposition rates (mean ± SEM) in the cylinders, two locations in Lake Terra Nova, the
treated peat pond (N=4) and the untreated peat pond (N=8). There were no significant differences
(One-way ANOVA).
39
3.2.
Experiment 2
In this paragraph Elodea nuttallii response and sediment nutrient composition in the different
cylinder treatments (AlCl, PAC, Phoslock+, Phoslock, FeCl and control) with and without
extra phosphate loading are presented. In table 5 and 6 (Appendix) mean end results of
biomass, nutrient composition of biomass and sediment for different treatments of unloaded
and loaded sediment columns can be found including statistical analysis. Unless not explicitly
mentioned, statistical analyses have only been performed with data from cylinder treatments
(df=5).
3.2.1. Elodea nuttallii response
In general there were not many significant differences in plant biomass or sediment nutrient
concentrations between treatments and between the two phosphate loadings (Table 3).
Table 3 Effects of different treatments and extra PO4 loading on E. nuttallii biomass, shoot:root ratio
and nutrient concentrations of biomass and sediment.
Treatment
P loading
df = 5
df = 1
Treatment x P loading
df = 5
Total biomass
mg
F
1.309
p
0.282
F
0.040
p
0.842
F
0.294
P
0.913
Root biomass
Shoot biomass
Shoot:root ratio
mg
mg
-1
gg
1.165
1.253
1.179
0.345
0.305
0.339
0.245
0.066
0.821
0.624
0.798
0.371
0.361
0.337
0.602
0.872
0.887
0.698
Root length*
Shoot length
mm
cm
10.839
1.307
0.170
0.138
0.713
0.436
µmol/ g DW
2.876
0.680
0.713
37.561 <0.001
2.914
0.993
P biomass
0.055
0.283
0.028
0.733
0.604
Al biomass
µmol/ g DW
1.170
0.025
0.876
0.256
0.934
La biomass*
Fe biomass*
P sediment
Al sediment
La sediment*
Fe sediment
µmol/ g DW
µmol/ g DW
µmol/ g DW
µmol/ g DW
µmol/ g DW
µmol/ g DW
32.545
4.793
1.790
1.607
34.292
2.184
0.343
<0.001
0.442
0.140
0.183
<0.001
0.078
0.072
0.123
6.346
2.811
0.680
3.303
0.789
0.726
0.016
0.102
0.409
0.077
1.766
7.293
0.081
0.041
0.915
0.058
0.880
0.200
0.995
0.999
0.969
0.998
Data were analyzed with two-way ANOVA (F) or non-parametric Scheirer-Ray-Hare (*) test with levels of
treatment (AlCl, PAC, PL+, PL, FeCl, Control; N=8) and PO4 loading (loaded/ unloaded) as fixed factors. Bold
values indicate p ≤ 0.05.
Over the 4 week period that the experiment was running, E. nuttallii experienced a six fold
increase of biomass from 17.10 ± 0.76 mg dry weight to 102.46 ± 6.11 mg dry weight
(unloaded sediments) and 96.61 ± 5.72 mg DW (P loaded sediments) (including data from
both peat ponds).
Shoot and root dry weight did not differ significantly between treatments and P loadings nor
did shoot/root ratio (Table 3). Lowest shoot biomass was measured in the PAC treatment for
both P loaded and unloaded sediment with respectively 70.8 ± 19.6 mg DW and 88.5 ± 17.4
mg DW. Highest shoot biomass was measured on unloaded sediments of the FeCl treatment
(107.5 ± 7.6 mg DW) and the untreated peat pond (137.0 ± 71.0 mg DW) (Fig. 14).
Highest root dry weights were found in the unloaded sediment of the Phoslock treatment (8.8
± 1.8 mg DW) and in the untreated peat pond (8.5 ± 2.6 mg DW). In general, root biomass
was lower in P loaded sediments ranging from 5.0 ± 1.2 mg DW (AlCl) to 7.8 ± 1.9 mg DW
40
(Phoslock). Root biomass in P loaded sediments of the untreated peat pond was almost
twice as high compared to other treatments with P loaded sediments (13.0 ± 1.0 mg DW).
Root/shoot ratio was in general higher for E.nuttallii growing on sediments with higher fertility
(loaded sediments) except for treatments with Phoslock and Phoslock+ (Fig 15).
Figure 14 Shoot and root dry weight (mean ± SEM) of E. nuttallii in the different treatments with and
without extra sediment phosphate loading after 4 weeks. There were no significant differences
between treatments (cylinders) and between the two phosphate loadings (without PG-Fe and PG+Fe,
two-way ANOVA).
Root/shoot ratio (mg DW mgDW -1)
0.09
unloaded
0.08
loaded
0.07
0.06
0.05
0.04
0.03
0.02
0.01
0
AlCl
PAC
PL+
PL
FeCl
con
PG-Fe
PG+Fe
Figure 15 Mean root/shoot ratio of E. nuttallii in the different treatments with and without extra
sediment phosphate loading after 4 weeks. There were no significant differences between treatments
(cylinders) and between the two phosphate loadings (without PG-Fe and PG+Fe, two-way ANOVA).
41
Nutrient composition of Elodea nuttallii
Extra loading with phosphate in the sediment resulted in significant differences between P
loaded and unloaded treatments for phosphorus concentrations in both biomass (df=1,
p<0.001) (Table 3). Biomass phosphorus concentrations differed within unloaded treatments
(One-way ANOVA, p=0.094) ranging between 107.51 ± 20.35 µmol g DW-1 (Phoslock) and
205.6 ± 12.4 µmol g DW-1 (FeCl). In comparison, P loaded treatments generally resulted in
higher biomass phosphorus concentrations ranging from 211.8 ± 28.0 µmol g DW -1
(Phoslock+) to 307.0 ± 13.9 µmol g DW -1 (FeCl). Treatments with PAC, PL and PL+ showed
the lowest biomass phosphorus concentrations for both phosphate loadings, whereas the
highest biomass phosphorus concentrations were found in the FeCl treatment (Fig. 16C,
Appendix table 3 & 4).
Aluminum concentrations in E. nuttallii biomass did not differ significantly between treatments
or between phosphate loadings (Table 3, Fig 16A).
Iron concentrations in biomassa did not differ significantly between the two phosphate
loadings and within P loaded sediments. However, in unloaded sediments, the iron
concentration in biomass of the FeCl treatment was significantly higher (84.8 ± 50.6 µmol g
DW -1) compared to Phoslock (9.8 ± 1.8 µmol g DW -1) (Kruskal-Wallis, p<0.05, Fig. 16D).
Other treatments ranged from 12.2 ± 3.1 µmol g DW -1 (PAC) to 39.9 ± µmol g DW -1 (AlCl).
E. nuttallii grown on sediments of the untreated peat pond also resulted in generally higher
iron concentrations (76.1 ± 15.7 µmol g DW -1 in unloaded sediments and 74.9 ± 21.2 µmol g
DW -1 in P loaded sediments).
42
A
B
b
C
B
a
ac A ac
AB
c A ac A
C
D
a
a
ab
ab
ab
b
Figure 16 Concentrations (mean ± SEM) of aluminum (A), lanthanum (B), phosphorus (C) and iron (D) in biomass of
E. nuttallii in the different treatments with and without additional phosphate loading. Significant differences in lanthanum
concentrations among cylinder treatments are indicated by different letters for both groups: unloaded (small) and P loaded
(capital) (without PG-Fe and PG+Fe, Kruskal-Wallis, each p<0.001, followed by multiple Mann-Whitney U tests). Different
letters indicate significant differences in iron concentrations of E. nuttallii for unloaded sediments (without PG-Fe and
PG+Fe, Kruskal-Wallis, p<0.05, followed by multiples Mann-Whitney U tests). Aluminum and phosphorus concentration in E.
nuttallii did not differ significantly between treatments and between loadings (Two-way ANOVA).
Lanthanum concentrations in biomass differed significantly between treatments (Scheirer–
Ray–Hare test, p<0.001) but not between the two phosphate loadings (Table 3). Statistical
analysis of treatment effects in either unloaded or P loaded sediments both resulted in
significant results (Kruskal-Wallis test, Table 3 & 4). E. nuttallii growing on sediments treated
with Phoslock resulted in significantly higher lanthanum biomass concentrations with 0.704 ±
0.155 µmol g DW -1 (unloaded sediment) and 0.752 ± 0.064 µmol g DW -1 (P loaded
sediment).
Lanthanum biomass concentrations in both unloaded and P loaded sediments of the
Phoslock+ treatment were significantly different from the Phoslock treatment (respectively
0.098 ± 0.024 and 0.201± 0.080 µmol g DW -1) and the FeCl treatment (respectively 0.015 ±
0.003 and 0.019 ± 0.013 µmol g DW -1). Lanthanum biomass concentrations in the PAC
treatment in unloaded and P loaded sediments (respectively 0.030 ± 0.011 and 0.046 ±
43
0.025 µmol g DW -1) were only significantly different from the Phoslock treatment but not from
the Phoslock+ treatment (Fig. 16B).
3.2.2. Sediment nutrient composition
Extra loading with phosphate in the sediment resulted in significant differences between P
loaded and unloaded treatments for phosphorus concentrations in sediment (df=1, p<0.05,
table 3).
Sediment phosphorus concentrations were not affected by the different treatments. However,
within the cylinder experiment, additional phosphate loading resulted in significantly different
sediment phosphorus concentrations (Two-way ANOVA, df=5, p<0.05, Table 3). In general,
sediment phosphorus concentrations in P loaded columns were significantly higher (13.7 ±
0.6 µmol g DW -1; N=34) compared to unloaded columns (10.6 ± 0.6 µmol g DW -1; N=34)
(including peat ponds and lake Terra Nova, Table 4, Fig. 17C).
In general, sediment phosphorus concentrations were lowest in both unloaded and loaded
sediment columns without plants with respectively 8.0 ± 0.9 µmol g DW -1 and 12.3 ± 1.0 µmol
g DW -1 (Table 4).
Table 4 Effects of additional loading with phosphate on sediment phosphorus concentrations (mean
±SEM), (Independent t-test). Cylinder exp. = sediment columns taken in mesocosms with treatments
AlCl, PAC, PL+, PL, FeCl and control. TN stands for Terra Nova, PG for peat ponds. Bold values
indicate significant results (p<0.05).
(1) Cylinder exp. + plants (N=48)
(2) + Plants (all treatments, N=56)
(3) No plants (all treatments (TN, PG).
N=12)
(4) TN, PG + plants (N=8)
(5) all columns (N=68)
Unloaded
-1
(µmol gDW )
11.24 ± 0.65
11.08 ± 0.59
Loaded
-1
(µmol gDW )
13.79 ± 0.76
14.02 ± 0.69
df
t
P
46
54
-2.537
-3.226
0.015
0.002
8.04 ± 0.86
12.30 ± 0.97
10
-3.291
0.008
10.13 ± 1.48
10.55 ± 0.55
15.44 ± 1.66
13.72 ± 0.60
6
66
-2.390
-3.911
0.054
<0.001
Sediment lanthanum concentrations differed significantly between treatments (Scheirer-RayHare test, p<0.001, Table 3). Additional loading with phosphorus did not have an effect on
sediment lanthanum concentrations. Lanthanum concentrations were significantly higher in
the treatments with Phoslock+ (2.1 µmol g DW-1, average loaded and unloaded) and
Phoslock (8.8 µmol g DW-1, average loaded and unloaded) (Kruskal-Wallis test, Table 3 & 4).
In general, lanthanum concentrations were a hundredfold higher in sediments treated with
Phoslock and Phoslock+ than in sediments treated with AlCl, PAC, FeCl and the control
(ranging between 0.02 and 0.03 µmol La g DW -1). Sediment columns treated with Phoslock
had 3 to 6 times higher lanthanum concentrations in comparison with Phoslock+ (Fig. 17B).
Sediment aluminum concentrations did not differ significantly between treatments or between
phosphorus loadings (Fig. 17A). Highest aluminum concentrations were found in the
Phoslock+ treatment (average 147.5 µmol g DW-1) followed by the AlCl treatment (average
134.2 µmol g DW -1). Nevertheless, the untreated peat pond also had similar aluminum
concentrations (average 145.5 µmol g DW -1). Lowest aluminum concentrations were found in
the control and the PAC treatment (Appendix Table 3 & 4).
Sediment iron concentrations differed between treatments but also between phosphorus
loadings (Two-way ANOVA, Table 3). Iron concentrations were generally higher in unloaded
columns (100.58 ± 6.00 µmol g DW -1) compared to P loaded columns (86.76 ± 4.83 µmol g
44
DW -1). This did not apply to location ”plas 2” (unloaded: 85.72 ± 8.93 µmol g DW -1; loaded:
97.44 ± 8.30 µmol g DW -1) and both peat ponds (Appendix Table 3 & 4, Fig 17D).
Highest iron concentrations were found in the Phoslock+ treatment in both P loaded and
unloaded sediment columns (respectively 104.78 ± 11.40 and 118.33 ± 7.99 µmol g DW -1)
followed by the FeCl treatment (respectively 95.00 ± 11.81 and 113.03 ± 23.41 µmol g DW -1).
A
B
b
B
b
aA a A
C
B
aA aA
D
Figure 17 Sediment concentrations (mean ± SEM) of aluminum (A), lanthanum (B), phosphorus (C)
and iron (D) in the different treatments with and without additional phosphate loading. Significant
differences in lanthanum concentrations between cylinder treatments are indicated by different letters
for both groups: unloaded (small) and loaded (capital) (without PG-Fe and PG+Fe, Kruskal-Wallis,
each p<0.001, followed by multiple Mann-Whitney U tests). Sediment aluminum, iron and phosphorus
concentrations did not differ significantly between treatments and between loadings (Two-way
ANOVA).
45
3.2.3. Leakage through sand layer
In general, phosphorus mobilization rates in experiment 2 range from -1.1 ± 1.1 mmol m-2yr-1
(“plas 1”, loaded) to 1.3 ± 0.4 mmol m-2 yr-1 (“plas 2”, unloaded). Phosphorus mobilization rates
were almost two times lower compared to the rates in experiment 1. Highest phosphorus
mobilization rates were measured in the untreated peat pond (2.6 ± 0.6 mmol m-2 yr-1) and
“plas 2” (2.2 ± 2.0 mmol m-2 yr-1) (Fig. 12).
The same is true for the phosphate mobilization rates. In experiment 2 they ranged from -1.4 ±
1.0 mmol m-2 yr-1 (“plas 1”, loaded) to 0.1 ± 0.1 mmol m-2 yr-1 (“plas 1”, unloaded). Experiment
1 resulted in much higher phosphate mobilization rates in the untreated peat pond (4.1 ± 1.9
mmol m-2 yr-1). The two locations in lake Terra Nova resulted in -2.5 ± 2.5 mmol m-2 yr-1 for
“plas 1” and -0.4 ± 0.5 mmol m-2 yr-1 for “plas 2”.
No algae growth was observed in these control columns without Elodea nuttallii.
mobilization rate (mmol m-2 yr-1)
6
4
Experiment 1
Experiment 2 unloaded
Experiment 2 loaded
2
0
-2
-4
-6
-8
plas1
plas2
PG-Fe
plas1
PO4
plas2
PG-Fe
TP
Figure 18 Phosphate and phosphorus mobilization rates (±SEM) of sediment columns from two
locations in Lake Terra Nova and from the untreated peat pond. Mobilization rates from experiment 1
(N=4) and experiment 2 were compared (loaded and unloaded, N=2).
46
4. Discussion
Mesocosm experiment: Mobilization rates and pore water concentrations
The gradual addition of treatments containing aluminum, Phoslock or iron chloride resulted in
lower pore water phosphate concentrations compared to untreated locations because these
added substances were binding PO4, as expected. Measurements show that the PO4
mobilization increases with higher pore water phosphate concentrations. Pore water PO 4
diffuses to the surface water due to insufficient Fe or other binding agents in the sediment
(Geurts et al. 2010). Other studies showed an increase in surface water PO4 concentrations at
locations with pore water PO4 concentrations greater than 5-10 µmol L-1 (Geurts et al. 2008). It
can be assumed that addition of chemical binding agents in general result in lower pore water
PO4 concentrations and therefore lower PO4 mobilization rates.
Besides pore water PO4 concentrations, also Fe/ PO4 ratios are used as prognostic tools to
predict mobilization rates. Below pore water Fe/ PO4 ratios 1 – 3.5 mol mol-1 the risk of
phosphate mobilization to the surface water increases (Jensen et al. 1992, Smolders et al.
2001, Geurts et al. 2008, Geurts 2010b, Geurts et al. 2010). Treatments increased pore water
Fe/PO4 ratios above target levels of 3.0 mol mol-1 which prevents mobilization of phosphate to
the water layer.
Gradual addition of AlCl3, PAC, Phoslock, and Phoslock+ to the water layer also decreased
internal phosphorus and phosphate release from the peaty sediment.
Insoluble aluminum hydroxide flocs are formed after addition of aluminum chloride or PAC to
the water. They are able to bind P irreversibly and prevent internal mobilization from the
sediment (Kennedy and Cooke 1982, Cooke et al. 1993).
Phoslock binds various forms of phosphorus and precipitates as stable mineral rhabdophane
which is characterized by very low solubility products (Haghseresht 2006). Phoslock+
combines the characteristics of both polyaluminum chloride and Phoslock resulting in
immobilization of phosphorus.
In general, it can be assumed that surface water treatments including aluminum and Phoslock
precipitate with P and accumulate in the sediment preventing internal P mobilization to the
water layer (Cooke et al. 1993, Reitzel et al. 2005, Haghseresht 2006, Afsar and Groves 2008,
Lurling and Tolman 2010, Geurts et al. 2011, Van Oosterhout and Lurling 2011).
A few studies have compared the effect of aluminum and iron addition on P immobilization
rates (Pa Ho 1976, Cooke et al. 1993, Burley et al. 2001, Hansen et al. 2003, Liu et al. 2009).
In general, P removal efficiency is in higher after aluminum addition due to the redoxsensitivity of iron. Under anaerobic conditions, iron(III)-phosphate-hydroxo-complexes in the
sediment go into solution as iron(II) and phosphate will be released to the water phase
(Lijklema 1980, Cooke et al. 1993). Furthermore, aluminum bound phosphorus is insensitive to
sulfide in contrast to iron bound phosphorus (Smolders and Roelofs 1993b, Smolders et al.
2006).
Oxygen concentrations have not been measured during the experiment. However, an orangebrownish layer was observed on the sediment layer and on the glass walls from sediment
taken at locations treated with iron. This suggests iron oxidation and therefore the presence of
Fe(III) with good binding capacity to bind P. Refilling of water could have a positive effect on
aeration of the water column resulting in increased binding capacity of iron. However, refilling
47
took place right after sampling; therefore a direct, measurable effect of aeration on the P
binding cannot be assumed.
Poly aluminum chloride (PAC) and aluminum chloride (AlCl3) have similar characteristics. PAC
is a pre-polymerized coagulant and more effective at a lower dose compared to aluminum
chloride. In general it can be said that longer polymers hydrolyze less and have a stronger
ability to bind to surfaces (Gao et al. 2002). This prevents drastic drops in pH after application.
It has to be noted that poorly buffered systems with a very low alkalinity (< 0.4 meq L -1) are not
suited for aluminum treatment (Cooke et al. 1993).
Addition of aluminum salts to soft-water lakes characterized by a low alkalinity (0.6 – 1.0 meq
L-1) and low buffer capacity could result in a drastic drop in pH, leading to the formation of toxic
Al(OH)2+ and Al3+ (Cooke et al. 1993). Furthermore, the initial pH also effects the formation of
aluminum forms. Preferably, a pH 6-8 is needed for aluminum hydroxide Al(OH)3 to dominate
(Cooke et al. 1993). This limits the amount of aluminum salts that can be used to bind
phosphorus. However, several studies recommend the use of sodium aluminate instead or
addition of a buffer such as calcium hydroxide to remain suitable pH (Cooke et al. 1993,
Lurling and Van Oosterhout 2009, Van Oosterhout and Lurling 2011). However, application of
aluminum in Lake Terra Nova is not hindered by any of these conditions but still regular pH
measurements are needed to prevent sudden toxic effects of aluminum treatment.
In Dutch surface waters, the maximum permissible concentration (MTR) of lanthanum is
10.1 μg L−1 (0.07 µmol L-1) (Sneller et al. 2000). Measurements of surface water
concentrations in the laboratory did not result in significant differences between treatments
and did not exceed 0.01 µmol L-1. However, lanthanum concentrations exceed concentrations
of 100 μg L−1 following the application of Phoslock and Phoslock+ in the field but are known to
drop during monitoring below 10 μg La L−1 (Anonymous 2008a, b). After formation of
rhabdophane, lanthanum will be unavailable biologically. Furthermore it can be assumed that
lanthanum is imbedded in the sediments (Lurling and Tolman 2010). However, concentrations
still exceed NOEC concentration of 100 μg La L−1 which were measured in a
21 d Daphnia reproduction test functioning as basis for the Dutch maximum permissible
concentration (Schneller et al. 2000).
Despite the efficiency of aluminum addition, water managers in the Netherlands prefer
dredging, aeration or iron chloride addition due to the possible toxicity of aluminum salts
(Boers et al. 1994). For aluminum an ad-hoc maximum permissible concentration of 48 µg L-1
(1.6 µmol L-1) in Dutch surface waters exists (Van de Plassche 2002). Treatments did not differ
in surface water aluminum concentrations and none of them exceeded the MTR norm.
Phoslock+, a combination of Phoslock and PAC, first flocculates and precipitates phosphate in
the water layer. Finally it immobilizes phosphate by formation of rhabdophane. Several studies
state that treatments with aluminum and Phoslock effectively reduce total phosphorus and
phosphate concentrations in the water layer and therefore prevent eutrophication (Egmose et
al. 2010). Furthermore, Phoslock addition leads to stabilization of the sediment. According to
Egemose et al (2010) this could be important for macrophyte colonization in organic
sediments.
Furthermore, economical reasons play an important role in lake management. The commercial
product Phoslock is merchandized worldwide and counted for many successful applications.
However, due to financial aspects it is more appropriate for small water bodies with less fluffy
sediments. Iron chloride is a cheap rest product in the metal industry. Production of PAC and
48
aluminum chloride is more expensive. Economic aspects and the preference for iron chloride
by Dutch water managers still hamper the use of other binding agents.
Advantage of gradual addition
Gradual addition of iron chloride and aluminum prevents drops in pH and alkalinity (Geurts
2010b, Saris 2011, Immers 2012). At this moment no similar studies are found in literature.
However, in almost all lake restoration projects where iron chloride or aluminum compounds
were added all at once, a drop in pH was observed with often harmful consequences for the
ecosystem (Cooke et al. 1993, Reitzel et al. 2003, Reitzel et al. 2005).
Depending on the buffer capacity of a lake system, only low doses of aluminum and iron
should be used to prevent acidification. However, most effective PO4 binding is achieved at an
appropriate dosage of 100 g Fe3+ m-2 to the sediment (Quaak et al. 1993, Boers et al. 1994,
Geurts 2010b). Therefore, iron should be added gradually until the final dosage is reached.
However, only very few studies have been performed with annual/ biannual addition of
aluminum compounds (Lewandowski et al. 2003). Lewandowski et al (2003) already
concluded that repeated additions of small amounts of aluminum compounds are more
effective for lake restoration than a single addition of the same total amount of aluminum
compounds.
Whole-lake iron treatment
Mobilization experiments from Lake Terra Nova showed significant lower phosphorus and
phosphate mobilization rates after gradual addition of 20 g Fe m-2 between April 2010 and
April 2011 compared to the situation before the treatment started. This agrees with the
assumption that addition of iron chloride leads to P immobilization.
In April 2010 and April 2011, phosphate mobilization rates were lower in the iron-treated peat
pond in comparison to the untreated peat pond, also as a result of iron chloride treatment (in
the summer of 2009).
However, two years after this application both phosphate and phosphorus mobilization
increased in both peat ponds compared to 2010. Applied concentrations in the treated peat
pond were 85 g Fe m-2 being added in a relatively short amount of time. Due to a drop in pH
the proposed concentration of 100 g Fe3+ m-2 (Cooke et al. 1993, Quaak et al. 1993, Boers et
al. 1994, Geurts 2010b) was not achieved (Saris, 2011). However, even less iron was applied
to Lake Terra Nova resulting in much lower phosphorus and phosphate mobilization rates
compared to mobilization rates before the treatment started. On the one hand it can be
assumed that addition of iron chloride has an effect which lasts only one year. This was also
concluded in other studies (Boers et al. 1994, Smolders et al. 1995). However, in these studies
it was explained by high sulphate concentrations and external phosphorus inlet. This cannot
be the explanation in Lake Terra Nova, because external inlet of phosphorus and sulphate is
low. The wooden walls used to separate both experimental peat ponds from Lake Terra Nova
showed first signs of erosion and weathering in 2012. Inlet of lake water into the experimental
peat ponds is therefore likely, but it is not expected that the amounts of incoming water are
high enough to have any effect.
Nevertheless, the effect of gradual addition of iron chloride to the water layer is not well
studied yet. It could be assumed that the efficiency of gradual addition is much higher than the
efficiency when adding all at once. The application of iron chloride in the treated peat pond
might still have been too fast which is supported by the drop in pH during this application
(Saris, 2011).
49
Phosphate mobilization rates of the untreated peat pond and the control treatment differed
significantly from the other treatments. The strong binding capacity of aluminum, iron and
lanthanum compounds could explain the high variation in phosphate mobilization rates and
also in pore water concentrations. Caused by anthropogenic alterations, upwelling iron-rich
groundwater disappeared and therefore iron concentrations in sediment of Lake Terra Nova
are not sufficient anymore to bind the phosphate. Without addition of binding agents,
phosphate is in general bound to the organic fraction or appears in the labile fraction. This
binding is very fragile and can be broken easily. However, phosphorus fractionation of
sediments taken in May 2011 did not reveal significant differences in extractable P between
sites and treatments. Steinman et al (2004) assumed that during sampling sufficient additional
P from deeper untreated layers veil chemical changes in phosphorus binding caused by
addition of binding agents. Furthermore, the number of samples for the P fractionation and
also for the mobilization measurements was limited due to the experimental setup.
Iron colloids
Phosphorus mobilization rates in the iron-treated peat pond were 9 times higher than
phosphate mobilization rates. A similar pattern was found in the mesocosm treated with iron
chloride, where phosphorus mobilization rates were almost 19 times higher compared to
phosphate mobilization rates.
Analysis of phosphorus, iron and manganese concentrations in surface water and pore water
samples revealed the same pattern consisting of coinciding high values for sediments from
FeCl3 treatments and controls. During experiments concerning iron supplementation in Terra
Nova such peaks have not been observed. However, another experiment with different iron
gradients in iron-rich ditches showed a similar pattern (Vliex 2012).
The mobilization of total phosphorus to the surface water correlated well with the mobilization
of iron. The release of phosphorus-iron particles to the water layer can therefore be assumed.
This phenomenon could be explained by the existence of iron colloids. A few studies
described the formation of iron colloids incorporating phosphate in eutrophic lakes (Buffle et al.
1989, Leppard et al. 1989, Mayer and Jarrell 1995, Gunnars et al. 2002). Mayer and Jarrell
(1995) stated that the diameter of iron colloids range from 0.05 to 1.0 µm. It can be assumed
that a part of these colloids pass filters of 0.15 µm which would agree with the pore size of
rhizons used to sample surface water and sediment pore water. The formation of iron colloids
could be an explanation for the higher pore water concentrations and mobilization rates of
phosphorus compared to those of phosphate. The colorimetrical measurement of phosphate
(SRP, soluble reactive P) (Haygarth and Sharpley 2000) does not detect phosphate
incorporated in iron colloids.
It is widely known that only a fraction of the total phosphorus pool is bioavailable (Boström et
al. 1988). This brings up questions about the bioavailability of iron-colloids and their
associated phosphorus fractions. Several studies stated that colloidal P in association with iron
affects turbidity, light climate and that it is indirectly bioavailable to algae (Mayer and Jarrell
1995). It is the question, however, if this might also be the case in Lake Terra Nova.
Macrophyte development
Objective of fen restoration is recovery and conservation of characteristic fen ecosystems
(Lamers et al. 2001, Lamers et al. 2002, Lamers et al. 2010a). Successful lake restoration is
directly connected with the development of aquatic macrophytes. They function as nutrient
sink and refuge for zooplankton and young fish but they also reduce resuspension of the
sediment. Submerged plant communities in fens are dependent on clear water (Scheffer et al.
1993, Lamers et al. 2002, Gulati et al. 2008, Lamers et al. 2010a). Macrophyte composition
50
seems to affect the ecosystem functions but its importance is not well known yet. Lake
restoration projects generally focus on improving water quality but very often fail in developing
vegetation with a high biodiversity. Creation of the right abiotic conditions such as clear water
state, an Fe/PO4 ratio above 10 mol mol-1, the availability of propagules in lake sediments and
low NH4 concentrations are important factors for the development of a diverse aquatic
vegetation (Bakker et al. 2012).
Development of Elodea nuttallii on sediments treated with aluminum chloride, iron chloride,
PAC, Phoslock and Phoslock+ was not negatively affected. No significant differences in shoot
and root dry weight were found. It can therefore be assumed that phosphorus bound to one of
the applied chemicals is still available for E. nuttallii.
By replacing the water with a standard solution without phosphorus and by covering the
sediment with a sand layer, E. nuttallii was forced to take up nutrients via the roots. Angerstein
and Schubert (2008) showed that in eutrophic systems phosphorus uptake via shoots
dominates if phosphorus supply in the surface water is sufficient. In oligotrophic and
mesotrophic lakes, however, they expect E. nuttallii to take up nutrients via roots to meet
phosphorus requirements. This uptake is slower compared to uptake via shoots (Angelstein
and Schubert 2008, Angelstein et al. 2009).
Initial P concentration in biomass of E. nuttallii is in general higher compared to almost all
treatments and loadings, because E. nuttallii was cultivated in the Botanical Gardens of
Radboud University in little ponds with probably high nutrient concentrations in the water layer.
This can be assumed due to the presence of duckweed. The decrease of P concentration in
biomass is remarkable in the treatments with PAC, Phoslock and Phoslock+. In comparison,
E. nuttallii growing on sediments treated with aluminum chloride, iron chloride or without any
treatment appear to have higher P concentrations in biomass.
It could therefore be assumed that treatments with PAC, Phoslock and Phoslock+ have a
negative effect on the P uptake by E. nuttallii. Lower P biomass concentrations could indicate
that uptake of phosphorus bound to lanthanum and aluminum is hampered by P binding.
However, P fractionation at the end of the second experiment did not result in remarkable
differences in distribution of P fractions. One the one hand, addition of a P-poor sand layer
disturbed the results of the P fractionation. Phosphorus binding compounds can only be found
in the first 10 cm due to lack of resuspension and mixture of sediment in the field mesocosms.
Phosphate, however, was added to the entire sediment column but pore water measurements
concluded that distribution appeared very heterogeneously throughout the sediment column.
Maximum total root length of E. nuttallii resulted in nearly 18 cm which is divided into several
individual roots. Therefore it can be assumed that roots of E. nuttallii have not been growing
deeper than 10 cm in the sediment layer. Additional P from deeper untreated layers and
heterogeneous P distribution can veil chemical changes in phosphorus binding caused by
addition of binding agents (Steinman et al. 2004).
E. nuttallii is a fast growing species dominating Charophytes and other aquatic macrophytes.
However, Angelstein et al. (2009) have shown that the ability of E. nuttallii to take up nutrients
from the sediment is limited by the sediment nutrient-pool. It can be assumed that
rhabdophane is not bioavailable anymore (Lurling and Tolman 2010, Groves 2012). However,
regardless the addition of binding agents it can be assumed that there is still sufficient
bioavailable P stored in the sediment stimulating the growth of E. nuttallii because no
significant differences in plant biomass were observed (Steinman et al. 2004, Geurts et al.
2010).
51
As expected, P concentrations are higher in plant biomass growing on sediments which
received an extra phosphate loading. However, both shoot and root dry weight did not differ
between treatments or between phosphorus loadings. Macrophyte growth in sediment
columns could also have been hampered by the limited growth space in the column, because
almost all macrophytes had reached the water surface at the end of the experiment.
Phosphorus and phosphate concentrations in the water layer were very low in all columns
(with and without macrophytes). In general we can assume that coverage of a peat sediment
with sand will limit phosphorus release from the sediment (Van Diggelen et al. 2010).
Furthermore, E. nuttallii is capable of releasing nutrients via the shoots and via plant
decomposition. Nevertheless, shoots also have a high absorption capacity (Angelstein and
Schubert 2008). It can be assumed that leached nutrients were immediately taken up by the
shoots. Furthermore, algae growth has been observed on the wooden sticks holding the
rhizons. They might also account for a certain but insignificant P uptake from the surface water
since no algae growth was observed in the water layer itself.
However, decomposition of macrophytes in lakes at the end of the growing season could be
an important factor that increases eutrophication (Angelstein and Schubert 2008). This
shortcut still functionates even after application of phosphorus binding agents.
It was earlier shown that growth of macrophytes in Lake Terra Nova was not limited by lack of
propagules, cohesive strength of lake sediment and presence of phytotoxins (Van de Haterd
and Ter Heerdt 2007), (Bakker et al. 2012). Nevertheless, also several other interspecific and
intraspecific competition effects play a role in the presence and absence of specific aquatic
macrophytes in Lake Terra Nova.
During the biomanipulation experiment in 2003/2004 the populations of benthivorous and
planktivorous fish were decreased. After a successful pilot experiment in the closed-off peat
ponds, it was extended to a whole-lake biomanipulation experiment. A year after the
biomanipulation experiment in Terra Nova, which resulted in a clear water state, algae bloom
reappeared.
Generally, biomanipulation experiments in peaty lakes very often fail due to external P supply,
high internal P loadings and concomitant high P mobilization rates (Meijer et al. 1999, Lamers
et al. 2010a, Bakker et al. 2012). According to Jeppesen et al. (1990) threshold values for
summer P concentrations in surface waters of lakes > 10 ha between 2.6 – 4.8 µmol L-1 (0.08
– 0.15 mg P L-1) are needed to establish a clear water state. In July 2007, total P
concentrations of 3.9 ± 0.7 µmol L-1 (0.12 ± 0.02 mg P L-1) were measured (Lamers et al.
2010a). However, after the gradual, whole-lake treatment with iron chloride, P concentrations
have been decreasing and the surface water is clear all year round (Ter Heerdt 2012).
Therefore, it can be assumed that treatments preventing internal phosphorus loading, such as
iron chloride combined with biomanipulation, could be an effective way to restore water quality
and biodiversity of Lake Terra Nova.
52
5. Acknowledgements
First of all I would like to thank my supervisor Jeroen Geurts for his knowledge, support and
supervision during my Master thesis at the department Aquatic Ecology and Environmental
Biology. Furthermore I am very grateful for his patience during my less productive periods. I
really enjoyed the regular field trips to Lake Terra Nova and I am very glad that I was able to
get to know so many different aspects of lake restoration and peat lakes.
Secondly I would like to thank my supervisor Egbert van Nes for his effort to supervise my
progress during my final Master project, his patience during my quiet periods and the trips to
Nijmegen.
Furthermore I would like to thank Leon Lamers for his untiring motivation and input of
knowledge. Sometimes I did not see the wood for the trees or wasn’t convinced about my
statistics – thanks for pushing me back on the right path when I was lost!
Special thanks also to Germa Verheggen, Jelle Eygensteyn and Roy Peters for their support
on the chemical analysis and Dries Boxman for his technical support. Also thanks to Gerard
ter Heerdt for his infinite knowledge and enthusiasm for Lake Terra Nova. Thanks to all the
students who helped during long days in the lab or climate room, especially Melchior and
Larissa. Last but not least thanks to all the students and colleagues for the great atmosphere
at the department and the pleasant times together during our working weekends, sport hours,
bbq’s, excursions and Friday-afternoon-drinks.
53
6. References
Afsar, A. and S. Groves. 2008. Alum and Phoslock: Comparison of the factors that affect their
performances. IR 015/08, Phoslock Water Solutions Limited, Frenchs Forest, Australia.
Akhurst, D., G. B. Jones, and D. M. McConchie. 2004. The application of sediment capping
agents on phosphorus speciation and mobility in a sub-tropical dunal lake. Marine and
Freshwater Research 55:715-725.
Angelstein, S. and H. Schubert. 2008. Elodea nuttallii: uptake, translocation and release of
phosphorus. Aquatic Biology 3:209-216.
Angelstein, S., C. Wolfram, K. Rahn, U. Kiwel, S. Frimel, I. Merbach, and H. Schubert. 2009.
The influence of different sediment nutrient contents on growth and competition of Elodea
nuttallii and Myriophyllum spicatum in nutrient-poor waters. Fundamental and Applied
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7. Appendix
-2
Table 1 Descriptives of total phosphorus mobilization rates in mmol m yr
-1
PG+Fe
PG-Fe
con
FeCl
PL
PLplus
PAC
AlCl
4
4
8
4
4
4
4
4
N
4
2,195
-0,942
14,921
5,580
8,711
4,582
-5,253
-4,366
-4,981
Mean
-2,766
1,996
4,576
6,106
2,024
3,646
3,511
2,280
2,521
2,915
Std. Error
1,006
-4,158
-15,506
-4,509
0,794
-2,893
-6,590
-12,508
-12,389
-14,256
Lower Bound
-5,966
8,548
13,621
34,352
10,366
20,316
15,754
2,002
3,657
4,295
Upper Bound
0,435
-0,665
-13,576
2,927
1,369
0,153
-3,536
-11,647
-11,560
-13,579
Minimum
-5,417
8,103
8,010
31,768
19,265
17,127
13,093
-0,965
-0,575
-0,624
Maximum
-0,774
95% Confidence Interval for Mean
plas1
4
-1
plas2
-2
Table 2 Descriptives of phosphate mobilization rates in mmol m yr
plas1
PG+Fe
PG-Fe
con
FeCl
PL
PLplus
PAC
AlCl
4
4
4
8
4
4
4
4
4
N
4
-0,376
-2,446
1,626
4,122
7,728
0,259
-1,141
-0,620
-1,238
Mean
-0,609
0,508
2,449
0,506
1,911
3,489
1,016
0,527
0,206
0,391
Std. Error
0,364
-1,993
-10,239
0,016
-0,397
-3,376
-2,975
-2,818
-1,275
-2,481
Lower Bound
-1,769
1,242
5,347
3,235
8,642
18,831
3,493
0,535
0,035
0,005
Upper Bound
0,550
-1,671
-9,673
0,293
-1,685
0,840
-1,198
-2,349
-1,173
-1,875
Minimum
-1,668
0,688
0,926
2,636
15,942
16,913
3,251
-0,212
-0,236
-0,119
Maximum
-0,085
95% Confidence Interval for Mean
plas2
60
mg
mg
(unit)/treatment
PL+
PL
FeCl
con
1.187
0.286
F/H
0.354
0.915
p
8.50 ± 2.50
145.50 ± 73.50
PG-Fe
6.00 ± 1.00
89.00 ± 26.00
PG+Fe
mean (±SEM) end results
PAC
5.63 ± 0.55
109.38 ± 10.97
One-way ANOVA df=5
AlCl
5.38 ± 1.25
112.88 ± 6.88
15.00 ± 7.00
83.00 ± 27.00
8.75 ± 1.80
14.96 ± 3.95
137.00 ± 71.00
113.75 ± 8.25
0.933
6.25 ± 1.11
0.601
107.00 ± 19.29
0.253
5.50 ± 0.29
0.744
94.00 ± 17.55
18.95 ± 2.73
103.75 ± 10.94
5.25 ± 1.60
24.45 ± 6.41
107.50 ± 7.58
29.50 ± 2.50
76.05 ± 15.66
14.74 ± 4.94
81.24 ± 8.93
105.00 ± 8.40
43.50 ± 4.50
17.62 ± 3.64
0.219
100.75 ± 19.41
0.478
15.88 ± 2.66
7,06*
88.50 ± 17.35
0.941
23.30 ± 5.69
38.50 ± 1.66
36.76 ± 5.42
278.95 ± 14.58
94.56 ± 17.05
40.92 ± 16.23
231.58 ± 11.98
63.78 ± 8.17
0.094
34.25 ± 3.17
0.307
30.75 ±1.49
6,08*
2.248
158.58 ± 42.58
35.11 ± 5.08
181.73 ± 26.25
33.51 ± 3.93
68.03 ± 10.55
205.60 ± 12.44
97.82 ± 12.96
65.76 ± 28.57
107.51 ± 20.35
33.00 ± 3.19
42.77 ± 11.49
156.76 ± 31.47
119.21 ± 12.28
59.22 ± 11.98
179.64 ± 19.59
36.25 ± 2.39
70.86 ± 13.86
11.78 ± 0.45
77.79 ± 30.81
8.47 ± 2.72
79.32 ± 37.41
136.43 ± 1.90
0.0250 ± 0.012
0.656
22.01 ± 6.32
0.706
0.0259 ± 0.012
0.664
0.049
0.593
<0.001
10.22 ± 1.02
16,43*
109.65 ± 6.37
10,18*
12.08 ± 2.48
87.68 ± 32.11
0.0156 ± 0.009
21.04 ± 6.41
135.52 ± 31.23
108.14 ± 2.13
0.0398 ± 0.009
0.0261 ± 0.003
9.83 ± 2.13
0.408
<0.001
84.81 ± 50.56
121.17 ± 25.13
1.073
17,68*
0.0147 ± 0.003
13.11 ± 0.75
86.61 ± 5.44
0.0254 ± 0.005
9.76 ± 1.77
151.87 ± 10.28
0.0199 ± 0.003
0.7036 ± 0.155
10.13 ± 1.50
113.03 ± 23.41
16.01 ± 3.15
121.42 ± 20.19
85.08 ± 16.32
10.4313 ± 3.945
0.0975 ± 0.024
12.10 ± 1.43
118.33 ± 7.99
1.6663 ± 0.560
12.24 ± 3.06
143.59 ± 19.62
89.38 ± 11.96
0.0188 ± 0.004
0.0298 ± 0.011
0.0251 ± 0.002
39.86 ± 13.92
111.07 ± 15.00
0.0220 ± 0.007
199.76 ± 28.44
90.20 ± 18.50
101.75 ± 11.38
107.00 ± 12.21
mean (±SEM) end results
Table 3 Mean (±SEM) end results of biomass and nutrient composition of biomass and sediment for the different treatments of unl oaded sediment columns.
unloaded
mg
total biomass
root biomass
mm
g g-1
shoot biomass
shoot root ratio
cm
root length*
µmol/ g DW
µmol/ g DW
shoot length
µmol/ g DW
P biomass
Al biomass*
µmol/ g DW
La biomass*
Fe biomass*
µmol/ g DW
µmol/ g DW
µmol/ g DW
P sediment
Al sediment
µmol/ g DW
La sediment*
Fe sediment
Data of unloaded sediment columns from the cylinder experiment were analyzed with a one-way ANOVA (F) or non-parametric Kruskal-Wallis (H) indicated with (*). Levels of treatments (AlCl, PAC, PL+, PL, FeCl, control) were used as a
fixed factor (N=4). Different letters indicate significant differences between treatments (multiple Mann-Whitney U tests). Bold values indicate p ≤ 0.05. Peat ponds (PG-Fe; PG+Fe) with N=2 were not included in statistical analysis.
61
PL
FeCl
con
0.492
1.207
0.778
0.346
13.00 ± 1.00
94.50 ± 6.50
PG-Fe
6.50 ± 2.50
87.50 ± 46.50
PG+Fe
mean (±SEM) end results
PL+
7.50 ± 2.22
114.75 ± 15.46
p
PAC
7.00 ± 0.71
106.25 ± 12.93
F
AlCl
7.75 ± 1.89
122.50 ± 10.43
One-way ANOVA df=5
(unit)/treatment
5.75 ± 1.11
11.57 ± 2.32
81.00 ± 44.00
116.00 ± 17.67
81.50 ± 7.50
6.13 ± 1.53
0.320
76.88 ± 20.42
1.268
5.00 ± 1.22
107.25 ± 13.30
97.75 ± 10.63
99.25 ± 13.29
mg
114.75 9.37
106.08 ± 56.93
110.25 ± 16.67
33.00 ± 13.00
70.75 ± 19.58
305.56 ± 39.64
92.75 ± 10.45
28.50 ± 2.50
36.90 ± 6.95
mg
319.22 ± 22.26
0.0464 ± 0.025
6.35 ± 1.07
0.312
14.16 ± 1.88
141.62 ± 20.74
0.235
0.0188 ± 0.004
0.362
1.289
0.766
0.415
1.515
<0.001
5.18*
41.5 ± 3.52
0.509
1.170
287.65 ± 30.97
17.35*
17.09 ± 3.22
35.25 ± 3.54
51.84 ±16.31
141.26 ± 49.67
307.00 ± 13.94
0.0145 ± 0.002
15.07 ± 3.26
33.75 ± 1.38
53.79 ± 14.69
79.00 ± 5.41
254.15 ± 31.77
0.0190 ± 0.013
17.02 ± 3.52
35.75 ± 1.89
60.73 ± 11.33
142.43 ± 41.19
211.75 ± 28.01
0.7518 ± 0.064
19.85 ± 1.44
29.50 ± 5.87
36.74 ± 9.36
73.29 ± 7.08
244.97 ± 44.76
0.2006 ± 0.080
11.56 ± 2.83
34.50 ± 2.22
64.97 ± 19.66
117.615 ± 18.25
297.28 ± 17.72
0.0461 ± 0.025
21.80 ± 4.76
cm
65.39 ± 17.01
73.56 ± 16.86
µmol/ g DW
0.0222 ± 0.001
mm
µmol/ g DW
93.89 ± 17.24
µmol/ g DW
0.0138 ± 0.005
13.33 ± 0.31
154.48 ± 22.60
74.92 ± 21.24
0.0384 ± 0.006
32.52 ± 2.51
0.333
17.55 ± 2.75
<0.001
0.798
1.236
0.366
16.72*
2.60*
97.74 ± 15.18
1.161
0.0211 ± 0.005
11.99 ± 2.33
114.69 ± 17.36
26.84 ± 13.91
0.0200 ± 0.005
14.70 ± 1.63
103.97 ± 15.87
105.39 ± 5.22
58.73 ± 47.86
7.2065 ± 2.454
134.25 ± 14.12
14.85 ± 3.28
140.16 ± 14.46
0.350
12.42 ± 1.76
2.4692 ± 0.719
1.196
23.78 ± 7.17
96.29 ± 18.01
77.29 ± 12.46
16.84 ± 1.46
0.0129 ± 0.003
95.00 ± 11.81
12.01 ± 2.18
124.70 ± 10.52
77.49 ± 11.73
49.45 ± 23.39
0.0195 ± 0.005
104.78 ± 11.40
25.85 ± 9.83
µmol/ g DW
72.80 ± 12.96
14.76 ± 1.37
µmol/ g DW
93.19 ± 8.76
µmol/ g DW
µmol/ g DW
µmol/ g DW
g g-1
mg
mean (±SEM) end results
Table 4 Mean (±SEM) end results of biomass and nutrient composition of biomass and sediment for the different treatments of l oaded sediment columns.
loaded
total biomass
root biomass
shoot biomass
shoot root ratio
root length*
shoot length
P biomass
Al biomass
La biomass*
Fe biomass*
P sediment
Al sediment
La sediment*
Fe sediment
Data of loaded sediment columns from the cylinder experiment were analyzed with a one-way ANOVA (F) or non-parametric Kruskal-Wallis (H) indicated with (*). Levels of treatments (AlCl, PAC, PL+, PL, FeCl, control) were used as a
fixed factor (N=4). Different letters indicate significant differences between treatments (multiple Mann-Whitney U tests). Bold values indicate p ≤ 0.05. Peat ponds (PG-Fe; PG+Fe) with N=2 were not included in statistical analysis.
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