Linköping Studies in Science and Technology
Dissertation No 1041
Nitrogen removal in treatment wetlands –
Factors influencing spatial and temporal variations
Sofia Bastviken
Department of Biology, IFM
Linköping University, SE-581 83 Linköping, Sweden
Linköping 2006
Nitrogen removal in treatment wetlands –
Factors influencing spatial and temporal variations
ISBN 91-85523-12-7
ISSN 0345-7524
© Sofia Bastviken
Front cover photo by Jonas Andersson, WRS i Uppsala
Printed by LiU Tryck, Linköping, 2006
ABSTRACT
Decreasing the nitrogen transport from land to surrounding seas is a
major task throughout the world to limit eutrophication of the coastal
areas. Several approaches are currently used, including the establishment
of wetlands, to decrease the transport of nitrogen. Wetlands represent
ecosystems where the nitrogen removal from water can be efficient given
that they are appropriately designed. The aim of this thesis was to
investigate and quantify the effect of critical factors that regulate the
nitrogen removal in wetlands, and to develop better guidelines for
wetland design. Studies were performed at different scales, from
microcosms to full scale wetlands, and methods included modelling, mass
balance calculations and process studies.
A first-order rate model was used to simulate the nitrogen
transformations in two large wetlands treating wastewater containing both
ammonium and nitrate nitrogen. It was found that the dynamics of the
main nitrogen transformation processes could not be satisfactorily
described using this approach. Large wetlands containing vegetation are
complex ecosystems, and the process rates vary in both time and space.
The great diversity of microenvironments favours different nitrogen
processes, and large differences in potential nitrification and
denitrification rates were found between different surface structures
within a wetland. The results from microcosms measurements showed
that the highest potential for nitrification was on surfaces in the water
column, while the denitrification capacity was highest in the sediment.
For the sediment denitrification capacity, the plant community
composition was shown to be of major importance primarily by supplying
litter serving as a carbon and energy source, and/or attachment surfaces,
for denitrifying bacteria. Denitrification rates may be affected more than
three fold by different types of litter and detritus in the sediments. Intact
sediment cores from stands of the emergent plants Glyceria maxima and
Typha latifolia had higher denitrification potential than sediment cores
from stands of the submersed plant Potamogeton pectinatus. However,
the quality of the organic material for the denitrifying bacteria was
highest in G. maxima and P. pectinatus stands. All sediment cores from
the wetland were limited by carbon, and the lower denitrification capacity
of the submersed plant, P. pectinatus, was likely due to lower amounts of
organic matter. However, in another wetland, intact cores from stands of
the submersed plant Elodea canadensis had a higher denitrification
capacity than the cores from stands of T. latifolia and Phragmites
australis. This was possibly due to a larger biomass, and better quality, of
the organic matter from that submersed specie, or to epiphytic biofilms on
the living plants. Those microcosms studies showed that both the quality
of the organic matter as a substrate for the microbial communities, and
the amount of organic material produced were important for the
denitrification capacity.
In pilot-scale wetlands, the composition of the plant community was
also a more important factor for high nitrate removal than the differences
in hydraulic loads (equivalent of 1 or 3 d retention time), despite the cold
climate. The greatest removal was found in wetlands with emergent
vegetation dominated by P. australis and G. maxima, rather than in
wetlands with submersed vegetation. In brief, the results presented in this
thesis emphasize the importance of dense emergent vegetation for high
annual nitrate removal in treatment wetlands.
LIST OF PAPERS
The following papers are included in the thesis. They are referred to in
the text by their Roman numerals.
I
Kallner S. and Wittgren H. B. 2001. Modelling nitrogen
transformations in free water surface wastewater treatment wetlands
in Sweden. Wat. Sci. Tech. 44 (11-12): 237-244.
II
Kallner Bastviken S., Eriksson P. G., Martins I., Neto J. M.,
Leonardson L. and Tonderski K. 2003. Potential nitrification and
denitrification on different surfaces in a constructed treatment
wetland. J. Environ. Qual. 32:2414-2420 and J. Environ. Qual.
33:411 (2004) (errata).
III Kallner Bastviken S., Eriksson P. G., Premrov A. and Tonderski K.
S. 2005. Potential denitrification in wetland sediments with different
plant species detritus. Ecol. Eng. 25:183-190.
IV Kallner Bastviken S., Eriksson P. G., Ekstöm A. and Tonderski K. S.
Seasonal denitrification potential in wetland sediments with organic
matter from different plant species. Submitted to Water, Air and Soil
Pollution
V
Kallner Bastviken S., Weisner S. E. B., Thiere G., Svensson J. M.,
Ehde P. M. and Tonderski K. S. Effects of vegetation and hydraulic
load on seasonal nitrate removal in treatment wetlands. Manuscript
Papers I, II and III are reprinted with permission from the publisher.
CONTENTS
Introduction...............................................................................................1
The nitrogen transformation processes in wetlands ................................2
The role of vegetation for nitrification and denitrification .....................6
The role of nutrient and hydraulic load for nitrification and
denitrification ..........................................................................................7
Constructed wetlands in the landscape....................................................9
Rationale of this thesis............................................................................11
Methods....................................................................................................12
Microcosm process studies....................................................................12
Model approach for describing nitrogen removal.................................14
Spatial variability of the nitrogen removal processes .........................17
The effect of organic matter on denitrification....................................19
Critical hydraulic load for nitrate removal..........................................23
Comparing wetland performance .........................................................24
Conclusions and future studies..............................................................27
Conclusions ...........................................................................................27
Future studies.........................................................................................28
Acknowledgement ...................................................................................28
References................................................................................................29
Tack ..........................................................................................................35
INTRODUCTION
Eutrophication is a major problem in many coastal areas and lakes of the
world including the coastal areas surrounding Sweden. Nitrogen and
phosphorus limit the primary production of different parts of the seas and
it is likely that an increase of nitrogen and phosphorus transports from the
drainage areas has enhanced the eutrophication process (Boesch et al.
2006; Granéli et al. 1990; Håkansson 2002). The direct effect of nitrogen
and phosphorus loads on the seas is hard to measure as the primary
production of the seas varies with weather conditions and water
circulations. However, to improve the trophic status of the coastal areas
surrounding Sweden in a long-term perspective, measures to limit the
load of nitrogen and phosphorus are critical.
At the Swedish west coast, Kattegat and Skagerrak, a decrease of
particularly the nitrogen transport is thought to be critical to prevent the
eutrophication process (Boesch et al. 2006; Håkansson 2002; Nielsen et
al. 2002). Håkansson (2002) estimated that the load of nitrogen from
point sources and rivers to the Swedish west coast, Kattegat and
Skagerrak, has increased by 40 % from the 1970s. Agriculture has been
estimated to be the largest single source of nitrogen to the sea (SEPA
1997). During the last two decades, measures have been taken to reduce
the nitrogen transport originating from agricultural activities. Firstly, best
management practises are increasingly used to reduce the input of
nitrogen to the agricultural fields and the nitrogen leakage from the fields.
Secondly, different water treatment approaches, such as wetlands and
ponds, are used to decrease the transport of nitrogen from land to the sea
(Fleischer et al. 1994; Svensson et al. 2004).
Drainage of large areas from the 19th century has caused decreased
residence times of the water from land to sea. The sometimes very short
residence times in catchment areas limit the contact between the nutrients
in the water and the environments that cycle these nutrients, which has
led to increased transport of nutrients to the sea. These large drainage
projects led to less lakes and wetlands in the catchment areas. Lakes and
ponds are able to store water volumes, and consequently the high floods
and the low flows become less pronounced downstreams. In this way,
lakes and wetlands increase the water residence times in catchment areas
during high floods and provide environments for nutrient cycling and
removal from the water phase. The year-round presence of water and high
biomass production make wetlands among the most biologically active
ecosystems on earth. Due to the high biological activity, wetlands
transform many compounds at a high rate and have increasingly been
1
used for wastewater treatment (Kadlec and Knight 1996). In the recent
decade, research has focused on how wetlands can be used to remove
nitrogen, both in agricultural drainage water (Braskerud 2002; Phipps and
Crumpton 1994; Raisin and Mitchell 1995; Spieles and Mitsch 2000) and
municipal wastewater (Andersson et al. 2005; Kadlec and Knight 1996).
Wetlands have been constructed with different design and the two main
groups are free water surface wetlands, where the water is passing over
the sediment surface, and sub surface wetlands, where the water is
transported within the soil medium of a planted bed. Free water surface
wetlands are the most commonly constructed wetlands in Sweden and are
also in focus in this thesis.
Free water surface wetlands have been constructed using different
designs and with different environmental conditions, i.e. with different
water quality, water flow and load, surface area, depth and containing
dense emergent or submersed vegetation or no macrophytes at all.
Physical, chemical and biological factors influence the nitrogen removal
both spatially and seasonally and the nitrogen removal efficiency varies
among wetlands and throughout the year. Today, we can not satisfactorily
explain the observed differences in nitrogen removal between wetlands
and more knowledge is needed to develop better guidelines to improve
nitrogen removal in constructed wetlands. The aim of this thesis was to
investigate and quantify critical factors that regulate the nitrogen removal
in wetlands, and to develop better guidelines for wetland design. Studies
were performed at different scales, and methods included modelling,
mass balance calculations and process studies.
The nitrogen transformation processes in wetlands
In the biosphere, nitrogen is continuously transformed between organic,
soluble inorganic and gaseous nitrogen forms (Fig. 1). The nitrogen cycle
is very complex, and it is hard to control even the most basic
transformations within a wetland. How much nitrogen that is removed,
i.e. transformed or removed from the water phase, in a wetland and which
nitrogen processes or nitrogen fluxes that are the most important depend
on water chemistry and other wetland conditions, such as climate,
vegetation, water depth and water flow. Ammonification is the microbial
mineralisation of organic nitrogen to ammonium. This process can be
caused by heterotrophic bacteria and fungi. Ammonium is then
transformed to nitrate by the bacterial process nitrification and nitrate or
nitrite is finally reduced to gaseous end products, nitrous gas and
dinitrogen gas, through the bacterial process denitrification. Nitrate can
2
also be transformed to ammonium during low redox conditions (Prescott
et al. 1990; Vymazal 2001). Nitrogen fixation is a bacterial process,
which transfers dinitrogen gas to ammonium. However, nitrogen fixation
is generally not significant in nitrogen rich waters such as constructed
wetlands for water treatment (Kadlec and Knight 1996).
Nitrogen assimilation is the transformation of inorganic nitrogen to
organic nitrogen in cells and tissue. Plant assimilation only accounts for a
small percent of the total nitrogen removal when the nitrogen load is high
(Tanner et al. 1995), which is the case in most free water surface
treatment wetlands (Kadlec and Knight 1996). Further, plant uptake does
not play an important role in the annual nitrogen removal in the long run
as the nitrogen assimilated by vegetation usually is released during
decomposition of litter (Johnston 1991). However, plant uptake can
contribute to the seasonal dynamic of nitrogen removal and account for a
significant part of the wetland nitrogen removal during the growth period
with rapid nitrogen uptake. In temperate climates and in wetlands
receiving riverine runoff, this often coincides with periods of low flow
and nitrogen load, thus making the plant uptake more important
temporary.
Ammonia volatilisation is a physicochemical process where ammonia in
the ammonium-ammonia equilibrium is transported to the gas phase.
Ammonia volatilisation can be important in wetlands with high
temperature and pH, although volatilisation losses of ammonia usually
are small if pH is below 8 (Reddy and Patrick 1984). Other nitrogen
fluxes that can occur in a wetland are sedimentation and resuspension of
the various nitrogen forms. Burial of organic nitrogen in the sediment will
make the nitrogen less available to living plants and organisms, while
release of nitrogen from biomass during decomposition will make the
nutrients available again. Ammonium can also readily adsorb to sediment
particles and litter, ammonium adsorption, as ammonium is a positively
charged ion. The adsorbed ammonium concentration is in equilibrium
with the ammonium concentration in the water, and can be released if a
change in the water chemistry or hydrology occurs.
3
Nitrogen fixation
N2
NH3
Air
Ammonia volatilisation
Denitrification
(Anoxic)
Ammonification
Nitrification
ON
Assimilation
AN
NN
Assimilation
Sedimentation
Diffusion and adsorption
Nitrification
(Oxic)
Ammonification
ON
Assimilation
Diffusion
AN
Water
Sediment
NN
Assimilation
Figure 1. A simplified picture of the nitrogen processes and the flows of
different nitrogen forms in a wetland (modified from Wittgren H. B.,
unpublished). ON is organic nitrogen, AN ammonium nitrogen and NN
nitrate nitrogen.
Nitrification followed by denitrification is usually the most important
processes for the permanent nitrogen removal in wetlands receiving
ammonium rich waters (Johnston 1991). In wetlands receiving nitrate as
the dominant nitrogen form, such as in agricultural drainage water, the
most important process is usually denitrification (Johnston 1991; Trepel
and Palmeri 2002; Vymazal 2001). In this thesis, both nitrification and
denitrification have been studied.
Nitrification and denitrification is influenced by factors at different
spatial scales. In Figure 2, factors affecting denitrification have been
described at different spatial scales; the process scale, the wetland scale
and the landscape scale. At the process scale, the process rates performed
by the bacteria are directly regulated by factors like temperature and
redox conditions. These process scale factors can be affected by factors
on the wetland scale, such as water flow, nutrient load and different plant
communities. Finally, the wetland scale factors are affected by landscape
4
factors such as climate and land use, e.g. higher nitrogen load the more
agricultural areas there is in the upstream catchment area.
Figure 2. Factors affecting the microbiological process of denitrification
on different spatial scales (Trepel 2002).
On the process scale (Fig. 2), the rates of nitrification and denitrification
are affected by several factors. The nitrifying bacteria require oxygen, an
inorganic carbon source, as well as ammonium, and are favoured by high
soil temperature (optimum 30-40 oC) (Reddy and Patrick 1984) and high
pH (7.5-8.0) (Prosser 1989). In the denitrification process, commonly
organic material serves as an electron donor while nitrate is used as an
electron acceptor. Denitrification rates will be favoured by anoxic
conditions, high nitrate availability and bioavailable organic carbon. The
bacteria are also favoured by high temperature (optimum 60-75 oC) and
pH 6-8.5 (Knowles 1982; Reddy and Patrick 1984). Most denitrifying
bacteria are heterotrophs, but some are autotrophs and uses H2, CO2 and
reduced sulphur compounds (Knowles 1982).
5
The role of vegetation for nitrification and denitrification
On the wetland scale, plants can play an important role for the total
removal of nitrogen in wetlands. Wetlands containing plants have been
shown to remove larger quantities of nitrate than unplanted wetlands
(Bachand and Horne 2000; Lin et al. 2002; Tanner et al. 1995; Zhu and
Sikora 1995). The vegetation can affect the nitrifying and denitrifying
bacteria in the wetland in many ways. In wetlands, the organic carbon is
supplied mainly by the vegetation. This organic matter is used directly as
a carbon and energy source for heterotrophic bacteria such as the
denitrifying bacteria. The effect of vegetation on denitrification is
dependent on the quality and the amount of the plant litter (Hume et al.
2002). The decomposition rate, or the availability of the organic carbon
as an energy source for the heterotrophic bacteria, is linked to the initial
composition of nitrogen and lignin in the plant (Hume et al. 2002).
Submersed and free-floating plants generally have a lower initial lignin to
N ratio than emergent plant species (Godshalk and Wetzel 1978; Hume et
al. 2002), and consequently, the organic matter of a submersed plant is
usually more available for decomposition, and hence for denitrifying
bacteria, than the organic matter of an emergent plant. However, the
primary production is higher for the emergent plants (Westlake et al.
1998) and hence, the organic matter lasts longer when emergent plants
are decomposed than for the submersed macrophytes. In addition, the
seasonal pattern of litter accumulation may significantly affect the
availability of organic matter, and the redox conditions in the sediments
and water. Studies have shown that dry dead leaves remain on the stems
till late autumn and only then fall off in a littoral stand of P. australis
(Dykyjová and Kvet 1978). The dry dead plant parts become available to
the bacteria first when it reaches the water. In a study of G. maxima, the
amount of loose litter, not attached to dead shoots was highest in
November to March, and then rapidly decreased to a minimum in May
(Westlake 1966). However, questions remain about what plant litter that
is most favourable for denitrification in wetlands.
Organic material can also have an indirect effect on the nitrifying and
denitrifying bacteria. The supply of litter on top of the sediment may limit
the oxygen diffusion to lower sediment layers, which causes anaerobic
sediments that is favourable for the denitrification process, but not for the
nitrifying bacteria. Also, the supply of organic carbon raises the total
heterotrophic activity, which consumes oxygen and thus indirectly
favours denitrification by lowering oxygen concentrations in the sediment
(Nielsen et al. 1990a).
6
Plant tissue provides a large amount of surface areas for microbial
growth. Attached bacteria are more active and more abundant than freeliving bacteria in aquatic systems (Hamilton 1987). This attachment of
microorganisms to submersed solid surfaces, such as sediment and
aquatic macrophytes, leads to the formation of biofilms. Different
surfaces can affect the presence of nitrifiers and denitrifiers by being
more or less suitable as attachment surfaces. Surfaces in the water phase,
to which bacteria can attach and where oxygen and ammonium
concentrations are high would favour nitrification. However, it is not yet
known if some surface structures in wetlands are more important than
others for the different nitrogen transformations. Eriksson and Andersson
(1999) showed that litter of different emergent plants was of varying
quality as habitats for nitrifying bacteria, with higher ammonium and
nitrite oxidation in litter of Scirpus sylvaticus and Carex rostrata than in
Equisetum fluviatile and Typha latifolia. Commonly, the sediment has
been considered to be the site where most of the denitrification occurs in
wetlands e.g. Seitzinger (1988), but later studies have shown that
denitrification in periphytic communities on submerged plants can make
a significant contribution to overall denitrification in a wetland (Eriksson
and Weisner 1997; Toet 2003).
Living plants can also increase the oxygen concentrations in the water
and thus create conditions that do not favour denitrification (Nielsen et al.
1990b; Toet et al. 2005), but are favourable for nitrification. Submersed
plants can release oxygen to the water through their photosynthesis, while
emergent plants can transport oxygen in aerenchymatous stem tissue to
the rhizomes and roots, that may eventually diffuse to the sediment and
water (Reddy et al. 1989; Vretare 2001). Hence, living and dead
vegetation can influence the presence and activity of nitrifying and
denitrifying bacteria in several different ways on the process scale.
However, what type of vegetation within a wetland that is most
favourable and provides the highest potential for nitrification and
denitrification is not fully understood.
The role of nutrient and hydraulic load for nitrification and
denitrification
The water flow regime of constructed treatment wetlands is mainly of
two types. Either the wetland receives a relatively constant water flow,
e.g. wetlands receiving municipal wastewater, or a water flow that is
determined by the runoff in the catchment area, e.g. wetlands receiving
drainage or storm water. At the wetland scale, the nitrogen removal in
7
treatment wetlands is usually assumed to be regulated by the supply of
nitrogen, i.e. the dominant nitrogen fraction, the concentration, and the
hydraulic load, and by the water temperature. A lower nitrogen removal
has been observed in wetlands receiving large concentrations of
ammonium nitrogen, i.e. municipal wastewater, than in those receiving
nitrogen in the form of nitrate (Andersson et al. 2005). The reason is that
in the former both nitrification and denitrification has to occur, and the
conditions in the wetland have to be suitable for both processes. In
wetlands receiving runoff water from agriculturally dominated catchment
areas, the dominating nitrogen fraction is nitrate (Ekologgruppen 2001)
and the nitrogen removal is mainly regulated by the rate of denitrification.
The hydrology of a wetland that is situated in a stream is strongly
dependent on the runoff in the catchment area. The hydraulic load, the
water volume entering the wetland divided by the wetland surface area,
and the water velocity can have extreme variations and occasionally be
very high. A high hydraulic load increases the amount of ammonium and
nitrate passing the wetland and thus the availability of these nutrients for
the bacteria. The nitrate removal rate of wetlands receiving runoff water
from agricultural catchment areas is generally thought to be regulated by
the nitrate load. The nitrate load is commonly correlated with the
hydraulic load, and mass balance studies (inlet and outlet measurements)
have shown that higher hydraulic loads result in higher area-specific
nitrate removal (kg per hectare and year) (Bachand and Horne 1999;
Kadlec 2005). However, Arheimer and Wittgren (2002) used data sets
from 8 different wetlands to model nitrate removal, and found that areaspecific nitrate removal was almost negligible at very high hydraulic
loads. Other studies have shown that at occasional high hydraulic loads
there can be larger outflows of nitrate than inflows (Raisin and Mitchell
1995; Spieles and Mitsch 2000). However, at present it is not known at
what flow velocities such negative effects will occur and override the
positive effects of an increased supply of nitrate.
The negative effects of high hydraulic loads on the nitrogen removal
could be caused by several factors. High water flows and water velocities
increases the turbulence and exchange of water within the wetland, which
favours the supply of ammonium and nitrate to the surfaces of the
wetland through turbulent diffusion. On the other hand, this also increases
the oxygen concentrations of the water, which would favour nitrification,
but limit the denitrification. In flowing water, oxygen may penetrate
through the entire biofilm community and inhibit the development of
anoxic microzones (Eriksson 2000). Christensen et al. (1989), showed
that denitrification in a stream was restricted to a depth of 1-5 mm in the
8
biofilm of the sediment. In the upper level in the sediment, the
denitrification was restricted by high oxygen concentrations, and at the
lower levels it was restricted by nitrate availability with low nitrate
concentrations in the deeper sediment. Very high water flows might also
resuspend the organic material in the upper sediment, and thus decrease
the amount of energy and carbon available for the denitrifying bacteria.
High water velocities may also restrict the biofilm development and
maturation due to physical disturbances (Claret and Fontvieille 1997). For
example, Eriksson and Weisner (1996) showed that the epiphytic
abundance on submersed plants and the denitrifying capacity was higher
at sites that were less exposed to wave turbulence or water currents, than
at more exposed sites.
The hydraulics, the internal flow patterns, can be of importance for how
efficiently the wetland area is used for nitrogen transformations (Persson
1999). Different flow patterns can cause shortcuts in the system and result
in much shorter nitrogen residence times, higher water velocities and less
efficient wetland area. This affects the contact between the nutrients in
the water and the bacteria and consequently the nitrate removal. The
hydraulics can depend on the morphology (Persson 1999) and vegetation
(Braskerud 2001) of the wetland.
Negative high flow effects on nitrogen removal have been shown to have
a large impact on the annual nitrogen removal of wetland basins (Spieles
and Mitsch 2000). Hence, to achieve high nitrogen removal on an annual
basis, high flow effects that limit the nitrogen removal in a wetland
should be avoided. The magnitude of these high flow effects are probably
dependent on the conditions in the wetlands, e.g. vegetation, depth,
hydraulics etc., and it is of great importance to identify the critical
hydraulic loads at which such nitrogen release occur in different kinds of
wetlands.
Constructed wetlands in the landscape
At the landscape scale, wetlands are affected by climate, nitrate
concentrations and water flow. Under boreal temperate conditions, the
hydraulic load and thus the nitrate transport is usually high during the
colder period of the year, when the biological activity is low, and low
during the warmer period when the biological activity is high (Fig. 3).
9
Q
Temp.
Conc.
Water flow (m3 d-1)
50000
20
40000
30000
15
20000
10
10000
5
0
jul-99
feb-99
apr-99
nov-98
jun-98
sep-98
jan-98
apr-98
nov-97
jun-97
aug-97
jan-97
mar-97
okt-96
aug-96
maj-96
dec-95
mar-96
jul-95
okt-95
maj-95
dec-94
mar-95
0
sep-94
-10000
Conc. (mg L-1), Temp. (oC)
25
60000
Figure 3. Seasonal variations of water flow (Q), nitrogen concentration
and temperature in the outlet of the wetland basin Råbytorp, southern
Sweden, in 1994 to 1999.
To reduce the nitrogen transport to the coastal areas, an important task is
thus to achieve high nitrogen removal during the colder period. The
activity of the nitrogen transforming processes is slower in cold
temperatures, and thus the hydraulic load must be lower during winter to
achieve the same percentage nitrogen removal (%) as in summer.
Nevertheless, studies have shown that on an annual basis the area-specific
nitrate removal, expressed as kg ha-1 y-1, is higher at higher loads of
nitrate (in kg ha-1 y-1), also in colder climate (Fleischer et al. 1994).
However, the relative nitrogen removal, expressed as percentage of the
inflow of nitrogen, can be very low at high nitrogen loads. To remove
large percentages of the nitrate load from the agricultural runoff water,
the residence times of the runoff from land to sea need to be increased.
To achieve longer residence times and efficient removal, wetlands should
be constructed with a sufficiently large area to remove the amount or
percentage of nitrate that is of interest. The relation between hydraulic
load and wetland area is usually decided through estimating the size of
the catchment area, and predicting the hydraulic load from that catchment
size. Of great importance to achieve an efficient result is the position of
the wetland in the landscape. This includes land use, geology and the
distance from the recipient. The nitrogen processes are more efficient at
higher nitrogen concentrations in the water. Large agricultural fields, e.g.
sandy soils planted with potatoes, can loose high concentrations of
nitrogen to the water (Hoffmann 1999) and could thus be a good choice
to include in the catchment for a planned efficient wetland. Arheimer
(1998) showed that nitrogen that was discharged to the surface waters of
10
a catchment close to the recipient of interest, e.g. the coastal areas, would
have much greater impact than the nitrogen that was discharged far
upstream from the recipient. The reason is that nitrogen that is transported
a long distance by rivers and streams will have a long residence time
before it reaches the recipient and thus there is a long time for the
nitrogen transformation processes to act. On the opposite, nitrogen in the
runoff discharged close to the recipient has a very short residence time
and a larger proportion will affect the recipient. This leads to the
conclusion that wetlands that are established close to the recipient are
more efficient than those established far upstream.
RATIONALE OF THIS THESIS
The aim of the study in Paper I was to investigate if a fairly simple firstorder area-based kinetic model could describe the fate of nitrogen
satisfactorily. Another intention was to discuss the fate of nitrogen in two
wetlands, receiving different forms of nitrogen, and analyse what
environmental factors that could be most critical for the nitrogen removal
processes in the respective wetland. The model in Paper I was sitespecific and insufficient to describe the fate of the different nitrogen
forms in the wetlands. This indicated that the spatial heterogeneity within
the wetlands, i.e. differences within and between the wetland basins, were
important. Using the first-order area-based equation, the nitrogen
transformations are assumed to occur only at the sediment surface in a
wetland. However, substantial nitrification and denitrification have also
been documented in biofilms of surfaces in the water column, such as
submersed plants (Eriksson and Weisner 1997; Eriksson and Weisner
1999). Hence, in Paper II, the spatial distribution of the nitrogen
processes within a wetland was investigated. This was studied using a
microcosm approach.
The aim of Paper II was to investigate the capacity of nitrification and
denitrification on different surfaces in a wetland and also to discuss the
importance of these surfaces for the overall nitrogen removal in the
wetland. The results showed great spatial variability with highest
nitrification on twigs laying in the water and highest denitrification in the
sediments. This shows that the first-order area based model was a good
choice for describing denitrification in the wetlands as the sediment had
the greatest potential for denitrification and very low denitrification
potential was found at surfaces in the water column. However, different
characteristics of the sediments might cause large variations in
denitrification capacity within and between wetlands. The high
11
denitrification in the sediment compared to the other surfaces indicated
that organic matter availability may regulate the distribution of the
denitrifying bacteria when nitrate concentrations are high.
This emphasised the importance of organic matter in the sediment. Plants
supply the denitrifying bacteria with organic material when decomposing
in the sediments and constructed wetlands contain different types and
species of vegetation. Consequently, the aim of Paper III and IV was to
investigate if the denitrification capacity differed between wetland
sediments formed in stands of different plant species and if these
denitrification rates varied with season. The study showed a large
variation in denitrification capacity between the sediments from different
plant species and between seasons. However, these studies were
performed in microcosms under laboratory conditions and other more
important factors could override the effect of vegetation under field
conditions. Hence, in Paper V we wanted to quantify the effect of
vegetation relative to another important factor for nitrate removal, i.e.
hydraulic load, at the field scale.
METHODS
Microcosm process studies
To study the nitrogen transformation processes, laboratory microcosm
studies where environmental factors can be controlled are useful. Using
this approach, selected samples are collected from the wetland and the
processes are measured during laboratory incubations. These methods can
be carried out during in situ conditions to measure the actual process rates
in field, or under optimal conditions, to measure potential process rates.
Nitrification rates are commonly measured as an increase of nitrite and
nitrate per hour. The isotope-dilution technique (Koike and Hattori 1978;
Rysgaard et al. 1993) is one common method to measure the production
of nitrate. 15N-nitrate is added to the aquatic phase of the samples.
Nitrifying bacteria dilute the 15N-nitrate pool by oxidizing 14Nammonium to 14N-nitrate. The dilution of 15N-nitrate is considered
proportional to nitrification. The 14N- and 15N-nitrate content can be
measured according to Davidsson et al. (1997) after converting nitrate to
nitrogen gas using a denitrifying culture (Pseudomonas nautica). 28N2,
29
N2 and 30N2 are measured using a masspectrometer (Hewlett-Packard
12
MS engine quadropole mass spectrometer). In the study of Paper II,
potential nitrification was measured using this isotope-dilution technique.
There are several methods to measure denitrification activity. Among the
most common ones are the acetylene inhibition technique, the 15N tracer
technique and the N2 flux technique (Leonardson 1992; Seitzinger et al.
1993). The acetylene inhibition technique is a method that is used to
indirectly measure the denitrification rate. Acetylene blocks the last step
in the denitrification process, the reduction of nitrous oxide to dinitrogen
gas. If denitrification occurs in a sample, addition of acetylene to the
sample causes an accumulation of nitrous oxide. This nitrous oxide can
easily be measured using a gas chromatograph. However, the acetylene
also inhibits the nitrification. If the denitrification process is mainly
supplied with nitrate from the nitrification process in the sample, this can
result in severe underestimations and the 15N tracer technique or the N2
flux technique is preferable.
The 15N tracer technique is a method where both denitrification and
nitrate produced through nitrification can be measured (Nishio et al.
1983). The isotope 15NO3 is added to the denitrifying sample and the
dilution of the isotope is measured as described in the isotope-dilution
technique for nitrification. This is measured with masspectrometer by
converting all nitrate to dinitrogen gas and measure the weight of the gas.
This method needs high concentrations of nitrate and is thus not a good
method when measuring actual activity of samples in which
denitrification is limited by nitrate concentrations. The 15N tracer
technique is also a costly method. The N2 flux technique is a direct
measurement of the end product in the denitrification process, the
dinitrogen gas (Seitzinger et al. 1993). The background concentration of
dinitrogen gas is lowered from 79 to 1% using helium and molecular
oxygen and then it is possible to measure an increase in dinitrogen gas
with a gas chromatograph. This technique needs gas tight incubations and
is sensitive to contamination, as there is high concentration of dinitrogen
gas in the surroundings. In summary, when the nitrate concentrations are
high in the water and the denitrification process is not mainly supplied
with nitrate from the nitrification process, the acetylene inhibition method
is an easy and cheap method to use.
In Paper II, III and IV, the acetylene inhibition technique was used as the
nitrate concentrations were very high and potential denitrification only,
i.e. no coupled nitrification and denitrification, was of interest. In the
incubations of intact cores, the acetylene and nitrate might diffuse more
or less rapidly into the sediment depending on the sediment
13
characteristics, the water movements of the overlying water and the time
of incubation. This might both underestimate and overestimate the
potential denitrification that would occur in field. In the experiments of
Paper III and IV, the potential denitrification might be overestimated by
faster stirring of the overlying water, causing a more rapid diffusion than
in situ, while this might have been underestimated in Paper II. In all
incubations, the incubation time might influence the diffusion of
acetylene and nitrate. The longer time, the lower down into the sediment
acetylene and nitrate would diffuse. However, even though this
compromise comparison between the rates measured in the different
experiments, the relationship between the samples within each
experiment is not likely to be affected.
Model approach for describing nitrogen removal
The nitrogen removal in treatment wetlands is usually described using
first-order equations based on input and output data of the wetland, i.e.
mass balance models. Higher nitrate concentrations in the inflow result in
a steeper concentration gradient in the sediment which would enhance the
diffusion rate of nitrate. Increased rate of nitrate diffusion into the
sediments where the denitrification occur increase the nitrogen removal
when nitrate is the limiting factor. This has been shown in flooded soils
where diffusion of nitrate into the sediments was limiting denitrification,
and the nitrate removal rate was best described as being linearly related to
the nitrate concentration (first-order) (Reddy and Patrick 1984). Hence,
the first-order equation (Eq. 1) is most commonly used in many wetland
models to describe overall total nitrogen removal when diffusion is not
considered explicitly in the model (Kadlec and Knight 1996; Knight et al.
2000).
J = k ⋅C
(Equation 1)
where J is the rate of the nitrogen transformation process (g m-2 y-1), k is
the first-order area-based reaction rate coefficient (m y-1) and C is the
concentration of nitrogen (g m-3).
First-order design models are either area dependent (Eq. 1), and thus
estimate nitrogen removal per unit wetland area, or volume dependent,
and thus estimate nitrogen removal per unit wetland volume (Kadlec
2000; Kadlec and Knight 1996; Knight et al. 2000; Spieles and Mitsch
2000). Models used in a cold climate region usually also include
temperature as a factor e.g. Arheimer and Wittgren (1994). Temperature
14
effects on nitrogen removal are most commonly described using a
modified Arrhenius temperature relation (Paper I), where the temperature
has stronger effect on the nitrogen process at lower temperatures. In the
models, the water column is usually assumed to be totally mixed every
time step or to follow a plug flow, where the water volume of each time
step is treated separated from the others. These assumptions are however
simplifications of the water flows in a wetland and tracer studies have
shown that the water flows in wetlands usually follow patterns that can be
described as a combination of these two assumptions. After evaluating
tracer studies in wetlands, the tank-in-series model was developed to
better mimic the hydraulics (Kadlec and Knight 1996). In the tank-inseries model, the wetland basin is modelled as consisting of several cells,
tanks, that are assumed to be totally mixed every time step. These cells
are connected and the nitrogen discharged from the first cell is entering
the second. Tracer studies can be used to analyse the hydraulic efficiency,
i.e. how well the water is distributed over the wetland (Persson 1999).
Using tracer studies, the number of cells (N) that would be suitable for a
certain wetland can be estimated and used in the first-order equation
(Paper V).
The first-order reaction rate coefficients are then used to compare the
efficiency of different wetlands. However, the use of this approach has
been questioned (Kadlec 2000). Published area-based nitrogen removal
rates vary largely between wetlands (0.5 < k < 70 m yr-1) (Kadlec and
Knight 1996) making the prediction of nitrogen removal in specific
wetlands quite uncertain. The large variation has several causes. All
nitrogen transformation processes and affecting factors are lumped
together and seasonal changes, like temperature, are not considered in all
calculations. Studies conducted in wetlands receiving mainly, or high
concentrations of, nitrate might be more comparable as one nitrogen
process, denitrification, is usually dominating for the nitrate removal. But
also the nitrate removal rates in such wetlands have resulted in a very
large variation (10 > k > 360 m yr-1) (Arheimer and Wittgren 1994;
Burgoon 2001; Kadlec and Knight 1996). These rates are, however, also
influenced by several different factors that could be limiting nitrate
removal, such as organic carbon availability and hydraulic efficiency,
which are not included in the calculations.
Researchers have tried to describe the seasonal variation of nitrogen
removal using the lumped model approach. Spieles and Mitsch (2000)
were able to explain 80-96% of the average annual nitrate removal in
three wetlands using the four parameters, hydraulic loading, nitrate
loading, wetland volume and temperature. The model was on the other
15
hand poor in describing the seasonal variation (R2=0.40), probably due to
flooding events. Arheimer and Wittgren (1994) used the same model
approach as in Paper I and succeeded to describe the seasonal variation
very accurately (R2=0.92). However, the wetlands had occasionally very
short residence time and a very low relative nitrate removal, which made
the incoming concentrations closely related to the outlet concentrations
and thus the description of the nitrogen processes was not as critical.
Other studies have presented more complex models for wetlands
receiving both ammonium and nitrate, containing more detailed
hydraulics and several nitrogen transformation processes (Dørge 1994;
Jørgensen 1994; Martin and Reddy 1997; Werner and Kadlec 2000). A
serious drawback of these more complex approaches is that a lot of data
of the processes are needed to properly calibrate and validate the models.
There are not many studies that have used models of intermediate
complexity, between lumped mass balance nitrogen removal models and
complex process focused models. To improve the treatment wetland
design modelling, models containing more internal processes than in the
lumped total nitrogen removal models, and less than in the complex
models, are needed. In Paper I, one such model was developed from data
commonly available from monitoring programs. The model used was
intermediate in both complexity and realism between lumped mass
balance nitrogen removal models and complex process focused models.
First-order area-based equations were set up to describe three major
processes in two different wetlands; ammonification, nitrification and
denitrification. The water in the wetland of Oxelösund was modelled as
six basins with completely mixed conditions within each basin, while the
wetland of Hässleholm was considered as only one totally mixed basin.
An advantage of this model approach was that it recognised the reactivity
of individual nitrogen transformations, and for the wetland of Oxelösund,
also that the water regime was considered.
The study of Paper V was performed in the experimental wetlands of
Halmstad. These are pilot-scale wetland basins, and a great advantage is
that replicates have been used, providing unique opportunities to
statistically test the effects of different factors. The area-based first-order
kinetic coefficients for nitrate removal were calculated using the tank-inseries model approach. The first-order rate coefficients were used to
compare the performance of the basins with different treatments. All
basins were small and similar in shape and the hydraulics was assumed to
be of minor importance for the differences between the basins. Further,
different temperature dependencies were applied and the effect of these
on the first-order rate coefficients was analysed.
16
SPATIAL VARIABILITY OF THE NITROGEN REMOVAL
PROCESSES
Large constructed wetlands with large nitrogen removal (kg ha-1 y-1) are
usually very heterogeneous with basins, or parts of basins, containing
different vegetation and having different water flow patterns. Nitrogen
transformations in such wetlands are complex in both time and space, and
different factors are limiting the nitrogen transformation processes in
different parts of the wetland and during different times of the day or
year. The model used in Paper I was site-specific for the two different
wetlands. This model approach was an improvement of a lumped total
nitrogen model, by separating three different nitrogen processes;
ammonification, nitrification and denitrification, but it did not have the
flexibility in predicting outflow concentrations over a wide variety of
conditions. The calibrated model described the seasonal variation of both
ammonium and nitrate removal well in Hässleholm, but poorly in
Oxelösund. Also, the calibrated reaction rate coefficients for ammonium
removal and nitrate removal were very different (Paper I).
The reaction rate coefficients for nitrate removal in Oxelösund were very
high compared with other studies, but similar rates have been shown in
wetlands with similar water quality, i.e. high ammonium concentrations
and low nitrate concentrations (Gerke et al. 2001). This could be a result
of a very intense denitrification process, but more likely it reflected that
the first-order equation was not suitable for such low nitrate
concentrations since diffusion was not included in the model. The choice
of equation has been discussed in other studies (Liehr et al. 2000; Reddy
and Patrick 1984), also pointing out the dependence on nitrate
concentration. At low nitrate concentrations (<2 mg NO3-N L-1),
denitrification can be regulated by nitrate concentration (Kadlec and
Knight 1996) and diffusion, and nitrogen removal might be better
explained using a second-order equation. In a wetland receiving very high
nitrate concentrations (>400 mg NO3-N L-1), the first-order model clearly
overestimated the nitrogen removal (Liehr et al. 2000), and the authors
suggested the use of a zero-order model under such conditions. At such
high nitrate concentrations, denitrification may be nitrate saturated and
hence regulated by neither concentration nor diffusion.
The wetland of Oxelösund could be considered more complex than that
of Hässleholm both in water quality and vegetation, as the wetland of
Oxelösund in large areas was covered by dense emergent vegetation. The
Oxelösund wetland was also receiving mainly ammonium nitrogen and
thus both nitrification and denitrification has to occur to remove nitrogen.
17
In contrast, the Magle wetland in Hässleholm received almost equal
amounts of ammonium and nitrate, had mainly open water surface areas
and, in parts of the wetland, dense submersed vegetation. The varying
environment in these wetlands probably favour different nitrogen
processes, and it is likely that the spatial heterogeneity within a wetland
ecosystem is an important component when trying to describe the
nitrogen transformation processes in a model. In less complex wetlands,
where only one factor is limiting the wetland system within the whole
wetland and throughout the year, e.g. nitrate concentrations or
temperature, simple models might better describe the nitrogen removal.
In Paper II, the spatial differences within a wetland of different nitrogen
processes were investigated. Large differences in potential nitrification
and denitrification rates were found between different structures in a
wetland, i.e. submersed macrophytes, filamentous algae, twigs and the
sediment. This indicated a substantial spatial variability for the nitrogen
transformations in wetlands. The results showed that the potential for
nitrification was highest on different surfaces in the water column, while
the denitrification capacity was highest in the sediment. Hence, the use of
an area-based rate coefficient as in Paper I, was less suitable for
describing the transformation of ammonium, while for denitrification the
area-based assumption was acceptable, although sediments might be very
diverse and denitrification may occur in thin or thick layers of the
sediment (Nielsen et al. 1990a).
The nitrification potential was shown to be highest on twigs, followed by
sediment and plants (Fig. 4). In the sediment, nitrifying bacteria might be
affected by the presence of other heterotrophic bacteria through
competition for oxygen which would explain the lower potential (Zhang
and Bishop 1996). Surprisingly, the rates (kg N ha-1 d-1) on twigs were
100 times higher than those noted for Myrriophyllum spicatum. The
photosynthetic activity of plants producing oxygen could be assumed to
favour the nitrifying bacteria. Interestingly, other investigators have
reported values representing nitrification on litter of emergent
macrophytes that were also much higher than the corresponding values
for the biofilms on the living submersed plants in the present study
(comparison based on activity per dry weight of substrate) (Eriksson and
Andersson 1999). It has been suggested that plants might restrict biofilm
development and nitrification in epiphytic microbial communities by
release of inhibitory chemical compounds (Lodhi 1982; Rice and
Pancholy 1972). Such impeding effects might explain the varying
nitrification capacity displayed on twigs, litter and living submersed
plants. For denitrification, the sediment was the most important site,
18
followed by twigs and plants (Fig. 5). The submersed plants showed very
low denitrifying capacity expressed per shoot area compared to the other
surfaces. However, expressed per wetland area and assuming the highest
abundance of submersed vegetation, the denitrification capacity of the
submersed plants with epiphytic biofilms was in the same range as the
sediments, which has also been shown by Eriksson and Weisner (1997)
and Toet (2003). The greater denitrification capacity in the sediment was
probably due to a high amount of available organic matter.
Figure 4. Potential nitrification (black bars), potential denitrification
(striped bars), and nitrate removal (white bars) on the surfaces of
Eurasian watermilfoil (Myrriophyllum spicatum), twigs, sediment, and
macroalgae (biofilm area-1 in A; dry wt.-1 in B). The samples were
collected in Basins B and D (n = 5) in Magle wetland, Hässleholm. Error
bars show SD. (Reprinted from Paper II).
THE EFFECT OF ORGANIC MATTER ON DENITRIFICATION
Organic carbon is supplied to the sediment by plants in a wetland and
several studies have shown that the supply of the organic matter often
limits the denitrification in treatment wetlands that receive high
concentrations of nitrate (Burgoon 2001; Ingersoll and Baker 1998;
19
Kozub and Liehr 1999; Zhu and Sikora 1995). Sediments have different
characteristics depending on their organic content. Wetland sediments
with low organic content have been shown to have much lower nitrogen
removal and denitrification capacity compared to those with high organic
content (Craft 1997). The vegetation is thus important as a source of the
detritus that fuels the denitrification and by creating a favourable
environment in other ways, e.g. lowering the oxygen concentrations
through increased heterotrophic activity in the sediment, or supplying
surfaces for microbial growth.
In the study of Paper III and IV, the effect of organic matter from
different plant species on denitrification was investigated. The results of
Paper IV showed that the organic matter in stands of Glyceria maxima
and Potamogeton pectinatus was more available, i.e. easily degradable, to
the denitrifying bacteria than that of the stands of Typha latifolia and
Phragmites australis. The litter from a submersed plant is usually more
easily degradable than the litter from an emergent plant (Godshalk and
Wetzel 1978). However, when measuring the denitrification capacity in
intact sediment cores, more closely resembling wetland conditions, the
organic sediment of the emergent plant G. maxima showed the highest
denitrification capacity, followed by T. latifolia (Fig. 5). The
denitrification in the sediment cores was limited by organic carbon and
the lower denitrification capacity in the stands of the submersed plant P.
pectinatus was likely due to less amount of organic matter and not the
availability of the organic matter.
The potential denitrification rates in the intact cores of Paper IV were
highest in autumn (November), while when measuring potential
denitrification in slurries, the highest denitrification rates were found in
May and August. This indicates that the organic matter was more
available, i.e. easily degradable, in May and August, and that the higher
denitrification rates in November for the intact cores were due to larger
quantities of organic material. The higher quality of the organic matter of
P. pectinatus was, in this study, not enough to compensate for the much
lower organic matter production. However, in a study of another wetland
(Paper III), intact cores of the submersed plant Elodea canadensis had a
higher denitrification capacity than the cores from the emergent plants T.
latifolia and P. australis in November. This was probably explained by
larger amounts of available organic matter from that submersed plant.
Further, some cores of E. canadensis included fresh plants with epiphytic
biofilms that could have contributed to high denitrification. The impact of
submersed plants and epiphytic biofilms was addressed by Weisner et al.
(1994) showing that microcosms containing the fresh submersed plant P.
20
Incubation 3 ( x 10 g N2O-N m-2, h-1)
Incubation 1 and 2 (g N2O-N m-2, h-1)
pectinatus with epiphytic biofilm, had higher denitrification rates than
litter of G. maxima. Altogether, these studies suggest that both the
availability and the amount of organic material are critical for the
denitrification capacity in a wetland, and emphasize the importance of
dense vegetation, and thus high amounts of available organic material, for
high annual wetland nitrate removal.
0.4
0.35
incubation 1
incubation 2
incubation 3
0.3
0.25
0.2
0.15
0.1
0.05
0
c
c
c
c
c
c
c
c
c
c
c
c
May Aug Nov Nov Apr Apr May Aug Nov Nov Apr Apr May Aug Nov Nov Apr Apr
G. maxima
P. pectinatus
T. latifolia
Figure 5. Potential denitrification rates (g N m−2 h−1) in intact cores
containing organic material of different plant species from the sampled
Källby wetland (Experiment I). C denotes acetate addition in incubations
2 and 3. In incubation 1, no external carbon source was added. Error bars
indicate SD. (Reprinted from Paper IV).
The plants in a wetland make the site very complex and diverse
biologically, chemically and physically. In Paper III and IV the effect of
the organic matter on denitrification was investigated in a laboratory
environment, and the denitrification rates were measured during optimum
environmental conditions. In the field, several environmental factors vary
and affect the denitrifying bacteria. The quantitative importance of
vegetation might thus be different depending on the type of vegetation
and the season, and in addition, other more important factors could
override the effect of vegetation under in situ wetland conditions.
In the study of Paper V, nitrate removal was investigated in pilot-scale
wetlands with different vegetation and water flows. The mass balances
(inflow and outflow measurements) showed vegetation to be an important
factor for high nitrate removal despite high hydraulic load and cold
21
climate, and the highest removal was found in dense stands of emergent
vegetation dominated by P. australis and G. maxima (Fig. 6). Also other
studies have shown that plants are important under in situ nitrate removal
in wetlands (Bachand and Horne 2000; Lin et al. 2002). The wetland
systems in Paper V achieved higher nitrate removal as the wetlands
matured and the density of vegetation increased. This effect has been
suggested by Craft (1997), who observed higher nitrate removal in
wetlands where organic material had accumulated for about 10 years than
for 1-5 years. The organic matter was probably important for the
denitrifying bacteria as a source of organic carbon, which was also shown
in the microcosm studies of Paper IV. Further, thicker organic matter
layer at the sediment would probably lower the oxygen concentrations,
through high heterotrophic activity (Nielsen et al. 1990a). In the study of
Paper V, the first-order area based rate coefficients were highest during
autumn, which was similar to the results presented in Paper IV, where the
denitrification potentials in intact sediment cores were highest in
November, probably due to more organic matter in the sediments.
Emergent
Submersed
Free development
d-1)
0.5
Nitrate-N removal (g m
0.6
-2
0.7
0.4
0.3
0.2
0.1
0
-0.1
1
3
2003
4
1
3
2004
4
1
3
2005
4
low
high
low
high
low
2
high
low
high
low
high
low
high
low
2
high
low
high
low
high
low
high
low
2
high
low
high
high
low
-0.2
1
2006
Figure 6. Nitrate nitrogen removal (g m-2 d-1) in wetlands containing
emergent, submersed and freely developed vegetation with different
hydraulic loads (high: 0.13 m d-1 and low: 0.39 m d-1) and during
different seasons (Period 1: December to February; period 2: March to
May; period 3: June to August; period 4: September to November). Error
bars show standard errors.
22
CRITICAL HYDRAULIC LOAD FOR NITRATE REMOVAL
Mass balance studies (inlet and outlet measurements) have shown that
higher hydraulic loads, and usually also higher nitrate loads, result in
higher area-specific nitrate removal (kg per hectare and year) (Bachand
and Horne 1999; Kadlec 2005). Surprisingly, the mass balances in Paper
V showed no significant difference in either nitrate removal (kg ha-1 y-1)
or first-order rate coefficients between the basins of different hydraulic
loads. Very high hydraulic load and short residence times of the water has
been shown to be crucial for the nitrate removal, causing negative effects
for nitrate removal (Raisin and Mitchell 1995; Spieles and Mitsch 2000).
Arheimer and Wittgren (2002) suggested that the critical mean residence
time for positive annual nitrogen removal would be more than two days,
based on data sets from Swedish wetlands receiving hydraulic loads of
0.26-6.8 m d-1. The two different hydraulic loads in Paper V were thus set
to 0.13 and 0.39 m d-1, which resulted in residence times of 3 and 1 days,
respectively. The results showed no negative effects of those hydraulic
loads. However, the wetlands in Paper V received a constant hydraulic
load, while the wetlands in the study of Arheimer and Wittgren (2002)
received very varying hydraulic loads during the year. The 2 days critical
residence time for significant nitrate removal suggested by Arheimer and
Wittgren (2002), was an annual mean value, and the residence times
could occasionally have been much shorter in all the studied wetlands,
causing such negative effects that the resulting annual nitrate removal
was insignificant in some of them.
Other studies, summarized in Table 1, also suggest that there is an upper
limit for how large the hydraulic loads can be to achieve positive nitrate
removal. Critical hydraulic loads for positive nitrate removal in these
wetlands was lower than 1.5 m d-1 (Braskerud 2002; Koskiaho et al.
2003; Svedin 2006). One mass balance (inflow and outflow
measurements) study of a wetland receiving pre-treated municipal
wastewater was performed in September and November (Svedin 2006),
and the results of that study showed no significant nitrate removal in
November at hydraulic loads more than 1.4-2.0 m d-1, which was
equivalent to 6 to 7 hours residence times (Table 1). Braskerud (2002)
also showed no significant nitrate removal in wetlands with hydraulic
loads larger than 1.7-1.8 m d-1. However, the critical residence time might
vary with the wetland conditions. The hydraulic load, residence times and
water velocities is affected by the hydraulic efficiency in the wetland, i.e.
the wetland area that can be used for nitrogen transformations. It is thus
important to achieve high hydraulic efficiency to attain considerable
nitrogen removal in wetlands. Estimated critical residence times might
23
also be related to the precision in the measurements of nitrate
concentrations and water flows. As the difference between inlet and
outlet concentration is very small when the residence time is short,
precise measurements are very difficult to perform.
Table 1. Characteristics of Nordic constructed wetland systems receiving
high hydraulic loads.
Site
Finland
Norway
Sweden
Age
Area
Depth
Hydraulic
load
Conc. in
NO3
Conc. in
Tot-N
(y)
(ha)
(m)
(m d-1)
(mg l-1)
(mg l-1)
Nitrate
removal
Vegetation
Scirpus sylvaticus, T.
latifolia (Biomass: 44
g m-2)
T. latifolia, Juncus
filiformis (Biomass:
91 g m-2)
1
2
8
0.6
0.5-1
0.41
7.9
9.8
Positive
removal
10
0.48
0.5-1
2.9
6.8
8.4
No
removal
7
0.030.09
0.20.8
0.7-1.6
2.2
3.2-3.5
Positive
removal
Scirpus lacustris,
Acorus calamus
3-4
0.030.09
0.20.8
1.7-1.8
0.75-2.8
1.6-5.1
No
removal
G. fluitans,
Sparganium erectum
9
0.3
0.20.3
1.3-1.5
19
20
Positive
removal
P. australis
9
0.3
0.20.3
1.4-2.0
9.4
20
No
removal
P. australis
1) Koskiaho et al. (2003)
2) Braskerud (2002)
3) Svedin (2006)
COMPARING WETLAND PERFORMANCE
In Table 2, some design characteristics are summarized in pond-like free
water surface wetlands in Sweden with similar temperate boreal climatic
conditions. All these wetlands receive water with high concentrations of
nitrate and thus denitrification is likely to be the limiting process for
nitrate removal.
24
Ref
.
3
Table 2. Characteristics and nitrate removal rate coefficients of Swedish
constructed wetland systems. Summarised below are wetland age, area
and vegetation, mean water depth, hydraulic load and inlet nitrate and
total nitrogen concentrations, first-order area-based rate coefficients for
nitrate removal and temperature coefficients.
Site
Eskilstuna
Age
Area
Depth
Hydraulic
load
Conc. in
NO3
Conc. in
Tot-N
(y)
(ha)
(m)
(m d-1)
(mg l-1)
(mg l-1)
7
28
1
0.16
15
20
1.08
30
(MayOct)
0.1-3.0
0.72.1
0.26-6.8
4.6-17
4.6-17
1
Southern
Sweden
Temp
coeff.
Vegetation
Ref
Emergent
(20%); G.
maxima and
Typha sp.
1
17
No or
submersed
vegetation
2
Rate
coeff.
kA20
(m yr-1)
Hässleholm
(Paper I)
11
20
0.5-2
0.057
14
21
1.116
29
Submersed; M.
picatum and E.
canadensis
3
Nykvarn
13
0.6
0.30.5
0.03-0.05
5.8-14
21-26
1.08
33
Emergent
(80%); P.
australis, T.
latifolia and
submersed; E.
canadensis
4
Halmstad
(Paper V)
4
0.002
0.4
0.13 and
0.39
11
11
1.088
21
(emerge
nt), 15
(submer
sed)
Emergent; P.
australis, G.
maxima and
submersed; E.
canadensis
5
NADB
1.088
1-34
Woddy
vegetation,
forested
6
NADB
1.088
26-55
Nonwoddy,
emergent
vegetation
6
1) Andersson et al. (2005)
2) Arheimer and Wittgren (2002)
3) Kallner and Wittgren (2001), Paper I
4) Tonderski (2000)
5) Paper V
6) Kadlec (2005)
The highest first-order rate coefficients were in the wetland of Nykvarn
(Tonderski 2000). This wetland contained the highest density of emergent
25
vegetation and probably had the highest accumulation of organic matter
in the sediments. Surprisingly, the wetland of Hässleholm (Kallner and
Wittgren 2001) showed similar performance as Nykvarn, though this
wetland contained submersed vegetation. Also, the nitrate removal rates
in Eskilstuna were in the same range, containing much larger areas of
open water than in Nykvarn. The wetland of Eskilstuna was only
measured during the warm season, and thus the annual removal should be
lower.
The first-order rate coefficients of the wetlands in Arheimer and Wittgren
(2002) and in Paper V were low. The wetlands in Arheimer and Wittgren
(2002) received hydraulic loads with strong seasonal variation, which
would decrease the removal rate coefficient as the main hydraulic load
occurs during the cold season. In addition, the wetlands in Arheimer and
Wittgren (2002) and Paper V contained more sparse vegetation and were
not as matured systems, and organic material might not have had the time
to accumulate in the sediments. The latter was most likely the reason for
the lower rate coefficients in the wetlands in Paper V, as there were
higher nitrate removal rate coefficients in the wetlands of dense emergent
vegetation than in the wetlands of submersed vegetation between
replicated wetland basins. Further, the wetland-specific hydraulics were
not considered in the first-order area-based rate coefficients in Table 2. In
the large complex wetlands, short-circuiting of the water, i.e. less
efficient wetland area, can cause much lower wetland performance
(Persson 2005).
All these Swedish wetlands have a nitrate removal, described with firstorder rate coefficients, that are in the lower range of the wetlands
described in Kadlec (2005). Possible reasons for this could be the climatic
differences. Wetlands that are located in warmer climate have a longer
period with high biological activity. This is critical for the primary
production of the vegetation, and the supply of organic carbon would be
larger in warmer climates than in colder. This indicates that the nitrate
removal of these wetlands might be possible to increase and that design is
crucial. To make better predictions of nitrate removal in constructed
wetlands, the accumulation of organic matter in the sediment might be
important to consider.
26
CONCLUSIONS AND FUTURE STUDIES
Conclusions
Simple models not accounting for spatial heterogeneity and vegetation
performs poorly to describe the seasonal nitrogen removal in complex
wetlands (Paper I). In such complex wetlands, different factors are
limiting the nitrogen transformation processes in different parts of the
wetland and during different times of the day or year. In less complex
wetlands, where only one or a few factors are limiting the nitrogen
removal within the whole wetland and throughout the year, e.g. nitrate
concentrations or temperature, simple models might better describe the
nitrogen removal.
The results in Paper II emphasised the importance of spatial differences
for the nitrogen processes, and suggested that the most active sites for
nitrification and denitrification were separated spatially in a wetland.
Nitrification took place largely on twigs in the water column, and this
process might thus be possible to enhance through supplying the bacteria
with more dead structures in the wetland water column. However, the use
of an area-based rate coefficient as in Paper I was not accurate for
ammonium removal, as the sediment was not the main site for
nitrification. For denitrification, the most important site in the wetland
was the sediment, and thus the area-based assumption would be
acceptable, although different characteristics of the sediments might
explain differences in rate coefficients between wetlands.
Sediments have different characteristics depending on their organic
matter content. Denitrification capacities may be affected more than three
fold by different types of litter and detritus in the sediments (Paper III and
IV), and the highest denitrification capacities were found in sediments
with organic matter from E. canadensis (Paper III) and G. maxima (Paper
IV). The vegetation was thus important by supplying organic matter that
fuels the denitrification, or by creating a favourable environment in other
ways, e.g. supplying surfaces for microbial growth.
The vegetation provided highest quality of the organic matter during
summer, but the wetland area based potential denitrification rates,
measured in intact cores, were greatest during autumn, when the organic
matter was most abundant (Paper IV). Hence, both the availability and
the amount of organic material were critical for the denitrification
capacity in a wetland.
27
The vegetation composition was a critical factor for nitrate removal in
wetlands receiving high hydraulic loads (Paper V). The nitrate removal
was higher in the basins of emergent plants than in the basins of
submersed plants and freely developed vegetation, and the removal
increased with wetland age and more dense vegetation. However, no
difference in nitrate removal in basins with 1 and 3 days residence times
could be detected. In summary, these studies emphasised the importance
of dense emergent plants for high annual wetland nitrate removal.
Future studies
For better prediction of nitrate removal in future wetlands, models are
great tools. To improve the prediction capacity of the models, evaluation
of existing wetlands is of major importance for the understanding of the
treatment performance during field and full-scale conditions. Analyzing
nitrate removal at field and full-scale conditions can allow better
understanding of the effect of hydraulics and vegetation and how the
wetlands change with age. The studies in this thesis suggest that the
characteristics of dead organic matter in the sediment and water column
could be important to include in the first-order area-based kinetic model
for nitrate removal. The organic matter changes between seasons and a
description of this might improve the general applicability of the model.
To further test the effect of different factors that could improve the
treatment performance, e.g. different plant species, pilot-scale wetlands
are very useful. Interesting studies would be to investigate the amount of
available carbon in relation to total carbon biomass of different plant
species required for optimal denitrification at different conditions, e.g.
amount of nitrate. Finally, the possibility of optimizing different wetland
services in one multipurpose wetland, e.g. nitrogen, phosphorous and
BOD removal, biodiversity and biomass production, would be interesting
to explore.
ACKNOWLEDGEMENT
I thank Karin Tonderski, Lars Leonardson and David Bastviken for
valuable comments on this thesis. This thesis was financed by MISTRA
(the Foundation for Strategic Environmental Research) through VASTRA
(Swedish Water Management Research Programme).
28
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34
TACK
Jag vill först tacka min handledare Karin Tonderski och mina
bihandledare HB Wittgren och Lars Leonardson, för att ha introducerat
mig in i våtmarkernas värld. Jag är väldigt tacksam för ert tålamod,
engagemang, alla råd och våra mycket roliga och givande diskussioner.
Tack också till Peder Eriksson för ett givande samarbete och att du
tillsammans med flera på limnologen i Lund lärde mig metoder för att
mäta kväveprocesser.
Jag är mycket glad för att ha varit en del i VASTRA. Genom våra
diskussioner i doktorandgruppen har jag fått en bredare syn på
miljöproblemen och förstått mer hur samhällsvetare kan se på saker och
ting.
En viktig och inspirerande del av mitt arbete har varit mer tillämpad.
Stort tack till Jonas Andersson, WRS i Uppsala, för ett samarbete som
gett mig inspiration och engagemang och en bättre förståelse för hur
våtmarker fungerar i verkligheten.
Mina rumskompisar under min tid som doktorand, Elisabeth Lundkvist,
Jenny Bäckman, Hristina Bojcevska och Nicklas Jansson ska ha ett stort
tack för att ni gör vårt rum till en så rolig plats. Tack även till mina
kollegor på biologiavdelningen för många trevliga fikaraster.
Till sist, men mest av allt vill jag tacka min familj, David, Emanuel,
Valdemar och Selma. David, min vetenskapliga rådgivare, men även min
bästa vän. Tack för allt ditt arbete med att läsa manuskript, analysera
problem och för att du alltid stöttar mig. Emanuel, Valdemar och Selma,
tack för att ni finns.
35