Biodiversity and ecosystem functioning in created agricultural wetlands

Biodiversity and ecosystem functioning in created agricultural wetlands
Biodiversity and ecosystem functioning
in created agricultural wetlands
A doctoral thesis at a university in Sweden is produced either as a monograph or as a collection of papers. In the
latter case, the introductory part constitutes the formal thesis, which summarizes the accompanying papers. These
have either already been published or are manuscripts at various stages.
Cover illustration- Thomas Prade and Geraldine Thiere
Chapter photos and illustrations- Geraldine Thiere
Printed by E-huset tryck, Lund
ISBN: 978-91-7105-295-7
Biodiversity and ecosystem functioning
in created agricultural wetlands
Geraldine Thiere
som för avläggande av filosofie doktorsexamen vid Naturvetenskapliga fakulteten vid Lunds Universitet kommer att offentligen försvaras i Blå Hallen, Ekologihuset, Sölvegatan 37, Lund, fredagen den
24 april, klockan 10.00.
Fakultetens opponent: Professor Jos T.A. Verhoeven, Section of Landscape Ecology, Institute of
Environmental Biology, Department of Biology, Utrecht University, The Netherlands
Avhandlingen kommer att försvaras på engelska.
Lund 2009
Department of Ecology, Limnology
Ecology Building
SE-223 62 Lund
Geraldine Thiere
Document name
Date of issue
April 24, 2009
Sponsoring organization
Wetland Research Centre,
School of Business and Engineering,
Halmstad University, Box 823
SE-30118 Halmstad, Sweden
Title and subtitle
Biodiversity and ecosystem functioning in created agricultural wetlands
Abstract . Wetland creation at large, regional scales is implemented as a measure to abate the biodiversity loss in
agricultural landscapes and the eutrophication of watersheds and coastal areas by non-point source nutrient pollution
(mainly nitrogen). The consequences of creating many new wetlands for biodiversity conservation and nutrient retention (ecosystem functioning) in agricultural landscapes are still relatively unknown, both on local (per wetland) and
regional (per landscape) scales. In Sweden, wetland creation has progressed already since the 1990s, and by now larger
numbers of created wetlands are present, mainly in the intensively farmed landscapes of southwestern Sweden. This
thesis aimed to investigate the following aspects in these systems: (i) their large-scale effects on biodiversity, (ii) their
functional diversity of bacterial denitrifiers, (iii) the abiotic and biotic influences on wetland ecosystem functioning,
(iv) the potential for biodiversity-function links, and (v) the potential for functional links and joint functioning.
(i) Created wetlands hosted diverse assemblages of macroinvertebrates and plants. They maintained a similar composition and diversity as natural ponds in agricultural landscapes. The environmental conditions per wetland did hardly
affect macroinvertebrate and plant assemblages, and the prerequisites for nutrient retention did neither. In landscapes
were wetland creation efforts had increased the total density of small water bodies by more than 30%, macroinvertebrate diversity of created wetlands was facilitated on both local and regional scales. (ii) Diverse communities of
denitrifying bacteria with the capacity for conducting different denitrification steps (functional types) were present in
all investigated wetlands. The richness of denitrifying bacteria communities was affected by nitrate concentration and
hydraulic loading rate, which may potentially be relevant for the nitrogen retention function of created wetlands. The
diversity across different functional types of bacterial denitrifiers increased with nitrate concentration. (iii) Both abiotic
and biotic factors influenced ecosystem functions of created wetlands. Variation in nitrogen retention was associated
to nitrate load, but even to vegetation parameters. In wetlands with constant nitrate load, planted emergent vegetation
facilitated nitrogen retention compared to other vegetation types. In wetlands with variable loads, nitrogen retention
was facilitated if nitrate load was high and many different vegetation types were present; nitrogen load could explain
the majority of the variation in nitrogen retention compared to vegetation parameters. Phosporus retention of created
wetlands was best explained by vegetation parameters. Litter decomposition was inhibited at high nitrate to phosphorus
ratios. Methane production increased with age and decreased with plant cover. (iv) Biodiversity may facilitate wetland
ecosystem functions, particularly in dynamic wetland ecosystems. Nitrogen retention increased with vegetation type
diversity, phosphorus retention capacity with plant richness, and litter decomposition with macroinvertebrate diversity.
(v) Created wetlands have the capacity of sustaining several parallel ecosystem services. Some wetland functions were
coupled; nitrogen retention increased with fast litter decomposition. On the other hand, methane emission and nitrogen retention were independent of each other, as were nitrogen and phosphorus retention.
In conclusion, created wetlands have the potential to at least partly abate the lost biodiversity and multifunctionality
caused by the past extensive destruction of natural wetlands in agricultural landscapes.
Key words:
constructed ponds; eutrophication abatement; biodiversity conservation; functional diversity; macroinvertebrates; plants; bacterial denitrification; watershed scale; nitrogen removal; phosphorus retention
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I, the undersigned, being the copyright owner of the abstract of the above-mentioned dissertation, hereby grant
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March 23, 2009
Table of Contents
Papers included in this thesis and my contributions.............................................................6
Shortsighted decisions & wetland loss.............................................................8
The concept of wetland creation......................................................................8
Thesis scope.........................................................................................................9
TERMS & CONCEPTS..................................................................................................... 10
Wetland & watershed scales........................................................................................ 10
Biodiversity.................................................................................................................... 11
Ecosystem functioning................................................................................................ 13
Environmental influences .......................................................................................... 15
Potential interactions................................................................................................... 16
Summary: Created agricultural wetlands................................................................... 17
Wetland creation - a potential solution?........................................................ 17
Multiple purposes............................................................................................. 18
Studied ecosystem functions & services....................................................... 18
Thesis relevance................................................................................................ 19
RESULTS & DISCUSSION.............................................................................................. 19
Biodiversity results....................................................................................................... 19
Functional diversity results......................................................................................... 23
Ecosystem functioning results................................................................................... 24
Biodiversity - function links........................................................................................ 25
Function - function links............................................................................................. 28
SYNTHESIS......................................................................................................................... 30
What explains diversity and ecosystem function of created wetlands?... 30
IMPLICATIONS & APPLICATIONS............................................................................ 31
REFERENCES.................................................................................................................... 32
Thank you! Tack! Danke!........................................................................................................ 38
Swedish Summary.................................................................................................................... 43
German Summary.................................................................................................................... 47
PAPER I................................................................................................................................. 53
PAPER II............................................................................................................................... 67
PAPER III............................................................................................................................. 85
PAPER IV........................................................................................................................... 105
PAPER V............................................................................................................................. 121
Doctoral theses of the Limnology department in Lund................................................. 143
Biodiversity and ecosystem functioning in created agricultural wetlands
This thesis is based on the following papers:
Thiere G., Milenkovski S., Lindgren P.-E., Sahlén G., Berglund O. &
Weisner S.E.B. (2009) Wetland creation in agricultural landscapes: biodiversity benefits on local and regional scales. Biological Conservation.
II Milenkovski S., Thiere G., Weisner S.E.B., Berglund O. & Lindgren P.-E.
Variation of eubacterial and denitrifying bacterial biofilm communities
among constructed wetlands. Submitted to Environmental Microbiology.
III Thiere G., Stadmark J. & Weisner S.E.B. Nitrogen retention versus methane
emission: Environmental benefits and risks of large-scale wetland creation.
Submitted to Ecological Engineering.
IV Weisner S.E.B. & Thiere G. Effects of vegetation states on biodiversity and
nitrogen retention in surface-flow wetlands. Submitted to Freshwater Biology.
V Thiere G. & Weisner S.E.B. Influence of biotic and abiotic parameters on
ecosystem functioning of created wetlands. Manuscript.
Paper I is reprinted with permission from the publisher.
Biodiversity and ecosystem functioning in created agricultural wetlands
My contribution to the papers
I planned the wetland surveys together with
Stefan Weisner (my supervisor). I conducted
the macroinvertebrate sampling and identification with assistance of Göran Sahlén
(Halmstad University). I and Susann Milenkovski (PhD student, Lund University) carried out the vegetation survey (in 36 wetlands). I compiled all data and carried out the
analyses and interpretation of the data set.
The final scope of the manuscript took shape
during regular discussion meetings with the
working group “Wetland Ecology & Biotechnology (WEB)”, in which all coauthors
are members. I wrote the manuscript, with
contributions from the other authors.
II I contributed to the study planning and
sampling, the interpretation of data, and the
development of the final scope of the manuscript as a member of the WEB group. I and
S.M. carried out the vegetation and microbial biofilm surveys. S.M. analyzed all samples
in the lab and compiled the microbial data.
I compiled the other wetland data and carried out the analyses of the combined data
set. S.M. wrote the manuscript in close cooperation with me and contributions from the
other authors. This manuscript was included
in the thesis of S.M.
III I and S.W. contributed the data on N retention. Johanna Stadmark (Lund University)
had the idea to this study and collected the
methane samples. Together we developed the
manuscript into its current form. I compiled
and analyzed the data and wrote the paper
in close cooperation with J.S. and S.W. This
manuscript was included in the thesis of J.S.
IV I planned and coordinated the biotic field
sampling during the four study years. I conducted parts of the macroinvertebrate, and
all of the plant identification. Chemical samples were taken and analyzed by Per Magnus
Ehde (Halmstad University). S.W. designed
the experimental setup of the experiment. I
compiled the dataset and S.W. did the statistical analyses and wrote the paper in cooperation with me.
I and S.W. planned the study and sampling
design. I coordinated and conducted most
of the field work with assistance from undergraduate students. Chemical analyses were
performed by Per Magnus Ehde (Halmstad
University). I compiled and analyzed the
data set. I wrote the manuscript with contributions from S.W.
Biodiversity and ecosystem functioning in created agricultural wetlands
Shortsighted decisions & wetland loss. The
topic of my thesis (created wetlands, see definition Box 1) would be irrelevant if not for a
common mistake in the history of mankind:
shortsighted decisions. Such decisions do not
consider the large-scale or long-term consequences of an action, but solely focus on the
local-scale/short-term benefits.
An example of this is the past (and ongoing) destruction and reduction of natural
wetland areas in many parts of the world.
The primary aim, i.e. often the extension of
areas suitable for agricultural production or
other human needs, seemed to easily outweigh any potential benefits associated with
the original wetland areas. As a consequence,
Box 1. Wetland definitions.
Mitsch & Gosselink (2000) define
physicochemical environment and prevalent
biota: “Wetlands are distinguished by the
presence of water, either at the surface or
within the root zone; they often have unique
soil conditions (poorly aerated and/or watersaturated soil) that differ from adjacent
uplands; wetlands support vegetation
adapted to the wet conditions (hydrophytes)
and, conversely, are characterized by an
absence of flooding-intolerant vegetation.”
The Ramsar convention (
includes a wide variety of aquatic
ecosystems in the term wetland: “Wetlands
are areas of marsh, fen, peatland or water,
whether natural or artificial, permanent or
temporary, with water that is static or
flowing, fresh, brackish or salt, including
areas of marine water not exceeding six
metres depth.”
The wetlands described in this thesis are
man-made, pond-like systems with an inlet
and outlet (hence, connected to a watershed),
which are located in landscapes with
intensive agriculture and receive high nitrate
regions with intensive agriculture have lost
up to 90% of the historic wetlands (Finlayson & Spiers 1999; Mitsch & Gosselink 2000;
Biggs et al. 2005) mainly during the last 200
years (Hoffmann et al. 2000). The excessive
wetland destruction was realized at great financial expenses, and led to the installation
of extensive drainage measures (e.g. pipe
systems, river channeling, and groundwater
level manipulation), effectively transforming
wet ecosystems into productive agricultural,
forested or urban land.
The concept of wetland creation. Today,
wetland functions are better understood
(Mitsch & Gosselink 2000), and wetlands
have finally established a reputation as one of
the world’s most productive ecosystem types
providing services of invaluable ecological
(e.g. biodiversity) and high economic value
(e.g. nutrient retention, flood control, food
production) (Costanza et al. 1997; Zedler
2005; Costanza et al. 2008). Consequently,
wetland protection is now common in the
industrialized world and wetlands of international importance are protected by the Ramsar convention (
Apart from wetland protection, strategies
to actively abate the loss of wetland ecosystem services involve wetland restoration and
creation. Wetland restoration aims at restoring damaged sites which to some extent still
serve as wetlands or have done so until rather
recently. The concept of wetland creation
(Fig 1) on the other hand, can include the
establishment of wet areas from scratch, i.e.
on land which has been under another type
of usage for long periods of time.
Wetland creation is nowadays implemented at large spatial scales (Mitsch et al. 2001;
Paludan et al. 2002; Zedler 2003; Hoffman
& Baattrup-Pedersen 2007), and is financed
by international (European Union) agri-environment schemes or national (e.g. Sweden,
Denmark, USA) political entities. In Scandinavia and particularly Sweden, large-scale
wetland creation was implemented early on,
starting in the 1990s (Lindahl 1998). The
national environmental objectives (www.
Biodiversity and ecosystem functioning in created agricultural wetlands
Fig 1. Created agricultural wetlands. (a) Illustration of the creation progress: After excavation, water is
collected via an inlet (here: drainage pipe) slowly establishing a permanent water table. The wetland outlet
discharges into a small agricultural stream. (b) A wetland three years after establishment. state that 12,000 ha wetland
area are to be created until 2010 (SJV 2000).
Hence, by now large numbers of created
wetlands are already established in Swedish agricultural landscapes and watersheds.
The majority of these wetlands have been
created in the intensively farmed landscapes
stretching along the southwestern coast of
Sweden (Scania/Halland). In these areas,
historic landscape changes and habitat losses
were drastic (up to 95% wetland losses; Krug
1993; Ihse 1995; Hoffmann et al. 2000) and
today’s nutrient export to the coast is nationally highest (; Kyllmar 2006),
as is the number of threatened species (Artdatabanken 2009).
Thesis scope. The background outlined
above forms the base for research questions
in the area of ‘applied science’, aimed to provide knowledge relevant for decision making,
policy and management strategies for created
agricultural wetlands. Applied research questions of this thesis deal with if and how created wetlands can compensate for the lost
multifunctionality and biodiversity of natural wetlands.
Apart from that, the many small created
water bodies appearing in the agricultural
landscape may also serve as interesting model
systems to answer questions of general ecological relevance, i.e. serving the purposes of
‘fundamental science’. Wetland creation is an
example of a large-scale manipulation of the
agricultural management, providing an opportunity for landscape-scale experiments
which are otherwise practically impossible to
carry out (Herzog 2005). Such fundamental
research questions deal with biodiversity and
ecosystem functioning, their controls and interactions.
In my research on created agricultural
wetlands, I tried to regard both applied and
fundamental aspects. Further, I tried to avoid
shortsighted recommendations, and I aimed
to integrate several aspects and consequences
of wetland creation instead. More specifically, I studied the following aspects in created
• their large-scale effects on biodiversity (Papers I & II; unpublished data),
• their functional diversity of bacterial denitrifiers (Paper II),
• the abiotic and biotic influences on wetland
ecosystem functioning (Papers III, IV & V),
• the potential for biodiversity-function links
(Papers IV & V), and
• the potential for functional links and joint
functioning (Papers III & V) in created wetlands.
Biodiversity and ecosystem functioning in created agricultural wetlands
Wetland & watershed scales
Wetland destruction is associated with dramatic consequences for single wetlands, but
also for entire watersheds. For the purpose
of this thesis I will introduce two main environmental problems arising from the past
diminishing of wetland areas, particularly in
agricultural watersheds: (i) habitat destruction causing loss of species (biodiversity loss),
and (ii) nutrient export from agricultural to
aquatic systems (retention loss) causing eutrophication.
Biodiversity loss. The accumulated local loss
of wetland habitat may affect the regional
species pool, i.e. the metacommunities (Leibold et al. 2004) that are sustained by all the
regional freshwater habitats together. Although local species diversity and composition of small isolated wetlands/ponds typically vary in space, over time, and with season,
this habitat type is hypothesized to be particularly important for the local and regional
diversity (Fig 2) of certain organism groups,
including aquatic macroinvertebrates and
plants (Scheffer et al. 2006). For maintaining regional diversity of these groups, small
lentic water bodies are crucial as their heterogeneity and spatial turnover is high, i.e. local
species assemblages are highly distinct from
each other (Oertli et al. 2005; Robson &
Clay 2005; Scheffer et al. 2006; Céréginho et
al. 2008). Further, the regional species pools
hosted by different aquatic environments
(ponds, lakes, rivers, streams, and ditches) in
the agricultural landscape, differ in quantity
and quality: small ponds exceed regional diversity of all other water body types (Davies
2005), they harbor over 70% of all aquatic
plant and macroinvertebrate species in agricultural landscapes (Williams et al. 2004),
and also most rare species (Biggs et al. 2007;
Davies et al. 2008). The local destruction of
wetlands may thus have large impacts on re10
gional scale, particularly as rare species loose
habitat refuges and may disappear from the
region. In southernmost Sweden, where extensive wetland loss occurred, diversity loss is
particularly threatening. In Scania alone, 380
wetland- and freshwater species are red-listed
(; this corresponds
Fig 2. The spatial components of biodiversity. (A)
Local or a diversity. Species number and composition per wetland is shown (four examples). (B) Spatial turnover or b diversity. The wetlands in a defined
region can be compared pair-wise (arrows) with each
other to determine the overall degree of differentation among wetland assemblages. (c) Regional or g
diversity. If the species present in all the wetlands of
a region (the four examples in (a) are pooled, the cumulative richness can be determined.
Biodiversity and ecosystem functioning in created agricultural wetlands
to more than half of Sweden’s endangered
species of these habitat types, and greatly exceeds the average numbers (170 wetland and
freshwater species) being red-listed in Swedish regions with less intensive agriculture and
lower farmland proportions (
Retention capacity loss. Similarly as above,
the repeated loss of local wetland buffering capacity may affect the watershed-scale
nutrient retention capacity. Wetlands act
as transforming landscape units (Mitsch &
Gosselink 2000) and are effective nutrient
traps (Saunders & Kalff 2001; Zedler 2003).
In agricultural areas, nutrient levels (nitrogen
N and phosphorus P) and the associated eutrophication risk are often raised throughout
entire watersheds (e.g. in Europe, Kroeze &
Seitzinger 1998), and the excess nutrients are
exported to the estuaries and coastal shelfs.
Monitored watersheds in southern Sweden
(7 to 15 km2 in size) today sustain nutrient
exports of 2,400 to 4,400 kg N km-2 yr-1, and
30 to 60 kg P km-2 yr-1 (Kyllmar 2006), originating largely from anthropogenic agricultural use (Fig 3). The high regional nutrient
efflux (Scania and Halland) constitutes up to
one third of the total Swedish N and P emis-
sions to the Baltic Sea (125,000 Mg N and
3,200 Mg P per year in 2007;
se), ensuring Sweden a top per-capita emission rank among the countries in the Baltic
watershed (Helcom 2004).
The extent of nutrient export is proportional to the extent of wetland destruction in
the watersheds (Mitsch et al. 2001). Apart
from the increased eutrophication risk within agricultural watersheds, the N export in
particular also states a threat for the marine
recipients. Mitsch et al. (2001) link the N
export (21,000,000 Mg N yr-1) of the Mississippi river basin to the area with anoxic sea
bottoms in the Bay of Mexico (2,000,000
ha). Similarly, anoxic bottoms along the
Swedish and Danish coasts are increasing,
and algal blooms are common in the Baltic
Sea (Helcom 2004).
Thus, wetland loss had consequences for
the biodiversity and functional integrity in
natural wetlands, entire agricultural watersheds, and even the recipient coastal habitats.
Consequently, as wetland creation is now
implemented at large spatial scales, there is a
need to assess the potential effects on diversity and function on local to regional scales
(Wagner et al. 2008).
Fig 3. Anthropogenic
N leakage from watersheds in Sweden.
Shading denotes extent of N leakage; the
darkest shade in the
south indicates >500
kg km-2 of annual N
export to the coast
Realizing the increasing threat arising from
habitat and species loss, the European Commission has set the target to halt biodiversity
loss in Europe until 2010, in order to avoid
profound consequences for the natural world
and human well-being (
multimedia/vnr-biodiversity). The term biodiversity covers many aspects of biological
variation, ranging from genes and species
over microhabitats to ecosystems (Gaston
1996). In common words it describes the living species in a defined space (often an area),
that may vary from very small (e.g. a soil
sample; a water droplet) to very large (e.g. a
continent; the whole earth). Scientific definitions of biodiversity are complex and cover
Biodiversity and ecosystem functioning in created agricultural wetlands
‘the variety of life forms, the ecological roles
they perform, and the genetic diversity they
contain’ (Wilcox 1984; Murphy 1988).
For diversity investigations, this broad biodiversity term needs to be specified for differing applications and quantified with suitable
measures. For this purpose, it is useful to distinguish between biodiversity components
related to structural, functional, spatial and
temporal aspects. Structural aspects account
for different amounts of individual organisms present in a population or community,
i.e. that species or genetic diversity is related
to number (richness) and relative abundance
(amount). Functional aspects regard (i) the
specializations of individual organisms for
their environment and their influence on
the ecosystem, and (ii) the organism/species interactions (e.g. competition, trophic
relations). The spatial components of biodiversity regard the variation of community
structure among continents, landscapes, ecosystem types, and individual sites of the same
ecosystem type. ‘Landscapes’ usually contain
assemblages of different ecosystem types (e.g.
forests, lakes, agricultural fields), which in
turn are represented by several individual
sites. Finally, the temporal aspects of biodiversity regard the variation of the structural,
functional and spatial components over time.
The species identity and number of organisms present in an ecosystem, and their interactions change on a daily, seasonal or annual
basis, or on longer evolutionary time-scales.
All these biodiversity aspects are of relevance for created wetlands, and in the following section I introduce several specific
research issues for the separate diversity components.
Structure & function. Diversity mechanisms depend on organism size (Finlay 2002;
Cottenie 2005; Beisner et al. 2006), and biotic group/organisation level (Beisner et al.
2006; Prosser et al. 2007). To allow comparisons, I included several distinct biotic groups
of differing organisation level in my diversity
investigations on created wetlands, namely
plants (primary producers), macroinverte12
brates (consumers and predators), and bacteria (mainly decomposers).
Within the main biotic groups, further
structural and functional aspects can be differentiated, e.g. by applying different concepts of defining the biotic units of diversity,
either based on phylogeny (species/strains)
or function (functional groups). Species are
defined by close phylogenetic relationships
of individual biotic units; for macroorganisms the ‘biological species concept’ applies
(see Coyne & Orr 2004). For microbes, phylogeny can be evaluated based on genetic
markers, universal, e.g. the eubacteria (16S
ribosomal RNA gene), but specific for bacterial strains (Dahlöf 2002). In contrast, functional groups are defined by the presence of
similar morphological or genetic features that
are relevant for a defined function. Phylogenetically distant organisms may fulfill similar
roles in an ecosystem (e.g. predators) or have
similar prerequisites/genes for carrying out
ecosystem processes (e.g. bacterial denitrifiers). Based on specific features supporting a
function, they can be grouped together in a
functional group.
In this thesis, plant- and macroinvertebrate
assemblages were differentiated according to
species or function. Further, eubacterial communities were differentiated based on genetic
variation (within the 16S rRNA gene). The
functionally important group of denitrifying bacteria was differentiated based on three
enzyme genes (required for different denitrification steps) and and their genetic variation (within the different enzyme genes).
I investigated both diversity and structural
composition of the biotic assemblages in created wetlands.
Spatial aspects. As outlined before, largescale wetland creation may affect diversity
on local to regional scales (Fig 2). Local or
alpha (α) diversity (per created wetland) can
be measured by richness (number of distinct
biotic units), or by the Shannon diversity index, integrating both the number of biotic
units and their relative abundance (Spellerberg & Fedor 2003). Beta (β) diversity or
Biodiversity and ecosystem functioning in created agricultural wetlands
spatial turnover rates among several created
wetlands can be assessed directly (degree
of differentiation) or indirectly (degree of
similarity) by pair-wise comparisons of assemblages between wetlands (Koleff et al.
2003). Gamma (γ) or regional diversity is
the diversity that is cumulatively hosted by
all wetlands, located in a defined geographic
area (landscape). Integrating all three spatial
aspects of diversity allows a basic evaluation
of the biodiversity conservation value of created wetlands, but also to investigate the factors influencing diversity, at different spatial
Temporal aspects. Species diversity is timedependent and increases with observation
period (Adler & Lauenroth 2003). To investigate temporal changes, repeated measurements in the same system are required; the
time scale of observation needs to be adapted
to the ecological question. The time frame of
this thesis covers interannual variation of created wetlands during the first few years after
Habitat heterogeneity. The environmental
conditions prevalent among individual farmland ponds/wetlands are often highly distinct;
this habitat heterogeneity is hypothesized
to be one of the mechanisms maintaining
the high regional species diversity of small
water bodies (Briers & Biggs 2005; Biggs et
al. 2005; Scheffer et al. 2006). The environmental conditions in aquatic habitats may
reflect the local assemblages of various biotic
groups (Declerck et al. 2005; Lindström et al.
2005; Shade et al. 2008). In created wetlands,
nutrient concentrations and loads can be expected to be high, as they are prerequistites
to sustain the retention function (Kadlec &
Knight 1996; Kadlec 2005). Thus, habitat
heterogeneity of created wetlands could be
lower than that of natural wetlands, and high
local nutrient concentrations in constructed
wetlands may interfere with simultaneous
diversity aims (Hansson et al. 2005). In this
thesis, I therefore study the extent of environmental influence in general, and of retention
requirements in particular, on the diversity
and composition of local assemblages.
Ecosystem functioning
Hooper et al. (2005) define ecosystem
functioning as the sum of all processes provided by a given ecosystem. In this thesis,
the general term ecosystem functioning is
applied to cover all ecosystem functions sustained parallelly by a system (Fig 4). Natural wetlands are multifunctional ecosystems
(Costanza et al. 1997; Zedler 2005) and
there is a need to investigate if created agricultural wetlands restore lost multifunctionality. Also, small water bodies were traditionally assigned a minor role for functioning on
regional to global scales (compared to large
freshwaters); however, more recently, small
waterbodies were suggested to exceed larger
lakes both in number and cumulative area
(Downing et al. 2006), thus being highly
relevant for regional/global cycles. The ecosystem functioning of small water bodies in
particular thus requires more research attention (Downing et al. 2006).
Each ecosystem function, i.e. its quality
and absolute/relative quantity, relies on ecosystem properties and process rates; often
several properties/processes contribute to a
given ecosystem function, and properties/
processes may influence each other. The flow
chart (Fig 4) also illustrates that ecosystem
process rates and properties in turn depend
on (i) the type and quantity of abiotic ecosystem components (resource pool sizes), defined by external supplies, internal consumption, and physicochemical interactions of
abiotic resources. Further, they depend on (ii)
the abundance and activities of biotic units
(i.e. species or functional groups) and their
resource utilization rates, respectively. Interactions between biotic units (e.g. competition or predation) may directly or indirectly
alter ecosystem process rates, properties, or
functions, and consequently ecosystem functioning. The complexity and the temporal
Biodiversity and ecosystem functioning in created agricultural wetlands
Ecosystem functioning
Level V
Level IV
Level III
Function 1*
Process rate 1
Function 2
Function 3
Process rate 2
Property 1*
(Pr) Species 3
(Co) Species 2
Level II
Level I
(Pp) Species 1
Abiotic factor 1
Abiotic factor 2
Trophic link
Indirect effect
Direct effect
Fig 4. Schematic flow chart on ecosystem functioning.
Abiotic factor 1 is a resource to producer species 1, which sustains both process rate 1 and ecosystem property
1. Abundance/activity of species 1 is directly controlled by consumer species 2 and indirectly by predator
species 3. Process rate 1 controls in turn the supply of abiotic factor 1. Process rate 1 and property 1 both
contribute to ecosystem function 1; propperty 1 also contributes to function 2.
Abiotic factor 2 directly limits process rate 2, which is sustained by consumer species 2. Consumer species 2 is
directly controlled by predator species 3, which in turn is indirectly controlled by abiotic factor 1. Property 1
serves as a resource for process rate 2. Process rate 2 sustains function 3 entirely and function 2 partly.
The whole system maintains three simultaneous ecosystem functions; functions and properties marked with
* are of direct value to humans (ecosystem services). The complexity of level II (biotic units) determines the
species diversity of the ecosystem.
(and spatial) stability of biotic interactions
and activities depend on the composition
and diversity maintained in the system itself,
but also the conditions for external species
recruitment and recolonization after disturbance.
Regarding the ecosystem functioning of
created wetlands, these systems are created
to serve one or several specified functions,
or ecosystem services. The term ecosystem
service is applied for an ecosystem function/
process/property which contributes to human welfare (Boyd & Banzhaf 2007; Costanza 2008); an ecosystem service is a desired
capacity of an ecosystem. Often, ecosystem
services are assigned a direct economical val14
ue (Boyd & Banzhaf 2007), and include, for
example, the food production by crops, the
carbon dioxide assimilation by a forest, or the
the nutrient retention by wetlands.
Man is trying to enhance desired or to decrease undesired ecosystem functions, particularly in managed ecosystems. By adjusting environmental conditions or ecosystem
processes in man-made ecosystems, ecosystem functions can be directed in a desired
way. Thereby, an ecosystem service might
be established or optimized. The creation
or management of ecosystems in order to
achieve certain ecosystem services (or environmental goals) requires the assessment of
potentially adverse consequences or unde-
Biodiversity and ecosystem functioning in created agricultural wetlands
sired functions, for example by cost-benefit
analysis. Contrary to beneficial services, undesired ecosystem functions may sustain an
environmental, ecological or human health
Environmental influences
General ecological theories predict environmental factors to influence (i) biodiversity and (ii) ecosystem functioning. The
following section introduces hypotheses on
influencing factors, but outlines also where
their direct application to small discrete
waterbodies (such as created wetlands) is
limited. In these cases, environmental factors
that are likely to influence wetland diversity
and function in particular, are specified.
Biodiversity influences. Species diversity of
a given ecosystem is constrained by species
colonization and extinction rates, as well as
the rate of speciation; these rates in turn can
be affected by factors operating on local (intrinsic), regional, or global (extrinsic) scales
(Sarvala et al. 2005). Intrinsic factors like the
local environmental conditions determine
for example the stability/disturbance and
the resource pool sizes of an ecosystem. Theories (corrobated by empirical observations)
hypothesize positive relations of biodiversity to area (island geography; MacArthur &
Wilson 1967) and time (Adler & Lauenroth
2003), a positive or hump-shaped relation
of biodiversity to productivity (Mittelbach
et al. 2001), and a facilitation of diversity at
intermediate disturbance frequency (Connell 1978). Extrinsic controls include global properties (e.g. temperature range) and
regional processes (e.g. climatic events, dispersal barriers) and the links between a local
system to its landscape context (e.g. size and
structure of the regional species pool; connectivity of habitats).
Regarding the biodiversity of pond-like
systems in particular, habitat size is most important for large, dispersal-limited organisms
(e.g. fish, Sarvala et al. 2005). For many other
organism groups (plants, plant-eating birds,
macroinvertebrates, amphibians, zooplankton), habitat size is either not of importance
(Oertli et al. 2002), or small habitat sizes (and
absence of fish) facilitate high local diversity
(secondary habitat size effects, Scheffer et al.
2006). The biotic assemblages of small pond/
wetland systems are subjected to strong environmental fluctuations (extreme events),
thus are exposed to a higher risk of local extinction (Scheffer et al. 2006). Hence, local
populations need internal and external strategies to compensate for losses, for e.g. by high
reproduction rates or resting stages (bacteria,
phyto- and zooplankton) or by good dispersal ability (flying macroinvertebrates, plant
seeds). Larger organisms (e.g. fish) may be
excluded from long-term establishment in
ponds, if local disturbance frequency is too
high to compensate extinction rate by (dispersal-limited) external recruitment (Sarvala
et al. 2005).
Further, the external recruitment potential
is determined by the landscape context, i.e.
the situation of a local habitat in relation to
(i) the quantity and quality of other habitats
present within the dispersal range of its biota,
and (ii) the size and structure of the regional
species pool available for colonization (Leibold et al. 2004). Created wetlands are likely
to be strongly dependent on the regional
factors, particularly just after establishment.
However, in landscapes with low densities
of isolated habitats, wetland creation (on
large scale) may also influence the species exchange rates among natural aquatic habitats,
interconnecting habitats as ‘stepping stones’
for species dispersal.
In addition to the local environmental
conditions reflecting the species assemblages
(Declerck et al. 2005) in small pond-like habitats, mechanisms other than environmental
factors may also explain assemblage variation. For example, stochastic events tend to
have great influence on biotic assemblages,
particularly in small water bodies (Scheffer
et al. 2006). Composition of both plants and
Biodiversity and ecosystem functioning in created agricultural wetlands
macroinvertebrates also depends on the order
in which species initially enter a community
(priority or preemption effects, De Meester
et al. 2002; Forbes & Cole 2002; Chadwell
& Engelhardt 2008). Productivity (affecting
local diversity) is even influencing diversity
on larger, regional scale; assemblages of eutrophic waters are therefore highly heterogeneous (Chase & Leibold 2002).
abiotic variation (disturbance) may overrule biotic influences. On the other hand,
biotic effects may be comparably important
in unstable environments, as biotic units are
frequently replaced; the capacity to switch a
certain function from one to another biotic
unit would require species and functional redundancy, i.e. diversity.
Ecosystem functioning influences. As illustrated in Fig 4, ecosystem functioning in general depends on abiotic and biotic parameters
of the system. Abiotic parameters can either
control properties/process rates by resource
pool size (e.g. nutrient supply for plant primary production), or regulate process efficiency via catalysing effects (e.g. temperature
and pH affecting enzyme activity). Abiotic
effects on ecosystem functioning are often
mediated through the biotic compartments
of the system; i.e. abiotic factors influence
performance of biotic compartments which
in turn regulate ecosystem processes/properties, and ultimatively functioning.
Each biotic unit (species; functional
groups) present in an ecosystem may affect
singular processes/properties/functions (Fig
4); the individual units are strongly affected
by the abiotic parameters of the system, as
well as the interactions with other biotic
units (competition/predation).
Ecosystem functioning of wetlands in particular depends on both the abiotic and biotic prerequisites. The capacity for removing
nutrients (N and P in various forms) is determined by the incoming concentration and
hydraulic loading rate, i.e. the nutrient load
(Kadlec & Knight 1996). However, even the
‘biotic setup’, i.e. the structure, composition,
and extent of the biotic assemblages influence wetland performance capacity (Kadlec
2008; Thullen et al. 2008; Kallner Bastviken
et al. 2009). However, there is need to outline the relative effect of biotic and abiotic
parameters, particularly in aquatic environments (Gamfeldt & Hillebrand 2008). For
example, in highly dynamic wetland systems,
Potential interactions
Interactions between biodiversity and
function as well as among functions are of
importance for understanding freshwater
ecology (Gamfeldt & Hillebrand 2008).
With regard to created wetlands in particular, they are also relevant for developing suitable management strategies.
Biodiversity–function links. Apart from
effects mediated by singular biotic units,
performance of biotic compartments also
depends on the strength, stability, and complexity of biotic interactions between units.
Put simply, the more biotic compartments
present per ecosystem, the more complex the
biotic interactions. This complexity or diversity is considered to be a major driver of ecosystem functioning itself (Loreau et al. 2001,
Hooper et al. 2005). Theories hypothesize
a positive effect of biodiversity on ecosystem functions (Loreau et al. 2001; Hooper
et al. 2005), based on the mechanistic explanations, that more diverse communities
are either (i) more resource-efficient due to
resource complementarity and species facilitation (niche-differentiation effect), or (ii)
have a higher chance of containing species
performing above-average (sampling effect).
Species diversity and functional redundancy
may also provide an insurance against future
environmental change (Hooper et al. 2002;
Loreau et al. 2003).
Consistent with diversity-function predictions, diversity parameters can influence
and enhance wetland functioning. P retention in equally loaded mesocosms increased
Biodiversity and ecosystem functioning in created agricultural wetlands
with submerged plant richness (and associated macroalgae biomass) (Engelhardt &
Ritchie 2001; 2002), N accumulation or
plant uptake increased with plant species
richness (Chabrerie et al. 2001; Callaway et
al. 2003), methane production decreased
with functional richness of wetland plants
(Bouchard et al. 2007), and decomposition
rate increased both with shredder richness
and biomass (Thullen et al. 2008). However,
observations from other studies seem contradictive to the diversity-function hypothesis.
Several studies have shown that monocultures of certain species may exceed diverse assemblages in performing the ecosystem function investigated, particularly if communities
were non-randomly assembled (Smith &
Knapp 2003; Schläpfer et al. 2005; Srivastava
& Velland 2005; Sullivan et al. 2007; Lake et
al. 2007). The created wetlands investigated
in this thesis assemble their biotic communities by natural succession; species assembly
is thus affected by non-random mechanisms
(Weiher & Keddy 1995).
Functional coupling & joint functioning.
Only recently carried out studies outline,
that the more functions considered, the
more apparent the importance of biodiversity for ecosystem functioning (Hector &
Bagchi 2007; Gamfeldt et al. 2008; Gamfeld
& Hillebrand 2008). In contrast, previous
diversity-functioning research commonly assumed that one ecosystem function investigated at a time may serve as an estimate for
overall ecosystem functioning (see Gamfeld
et al. 2008) and diversity mechanisms where
concluded mainly based on studies involving
only one trophic level (often primary producers). These simplifications, however, may
partly be the cause for confounding results
on diversity-function links. Some specific
ecosystems may be sustained without involving any biotic species; or mediated by one socalled ‘keystone’ species alone; in these cases
biodiversity could be of minor importance.
Other functions however, may involve more
complex food web interactions; in these cases
biodiversity effects may become apparent. By
randomly choosing one specific ecosystem
function to represent overall (joint) ecosystem functioning, biodiversity links may or
may not become apparent. Further, single ecosystem processes and properties may affect
each other, and the strength, stability and
complexity of their interaction (functional
coupling) may also affect overall ecosystem
functioning. These aspects need thus further
research attention, particularly in freshwater
habitats (Gamfeldt & Hillebrand 2008).
Created agricultural wetlands
Wetland creation - a potential solution?
As a potential solution to (partially) abate
the habitat/species loss and eutrophication
caused by historic wetland loss, the restoration or creation of wetland areas at large,
watershed scales is suggested (Mitsch et al.
2001; Paludan et al. 2002; Zedler 2004;
Chapman & Reed 2006; Mitsch et al. 2006;
Olde Venterink et al. 2006). The idea is that
an increase of the aquatic habitat in monotonous agricultural landscapes with intensive
production, may benefit species diversity or
retention. Wetland creation with pure biodiversity aims has been successful in the United
States (Galatowitsch & van der Valk 1996;
Seablom et al. 2001; Seablom & van der Valk
2003; Balcombe et al. 2005a, 2005b), with
local and regional diversity of specific wetland habitat types (e.g. prairie potholes) being restored. The creation of small permanent
water bodies may be suitable for sustaining
both the diversity of aquatic as well as transitional wetland species, which has been demonstrated for man-made agricultural ponds
(Declerck et al. 2006; Abellan et al. 2006;
Céréghino et al. 2008).
With regard to eutrophication, nutrient
export from farming areas to aquatic habitats can be abated using different strategies
(Mitsch et al. 2001; Zedler 2004; Olde Ven17
Biodiversity and ecosystem functioning in created agricultural wetlands
biotic uptake
Fig 5. Schematic view of a created wetland and its main ecosystem functions. High concentrations of nutrients enter
at the wetland inlet; nutrients
are then processed via different
pathways: sedimentation of
particle-associated phosphorus, bacterial denitrification of
nitrate-N to N2, biotic uptake
and seasonal storage of both
nitrate-N and phosphate-P,
and delayed nutrient release
via litter decomposition.
terink et al. 2006): (i) a change of nutrient
application and soil preparation practices
may reduce the amount of applied fertilizers
and decrease the risk of leaching and runoff
exports, (ii) the creation of buffer zones (e.g.
floodplain restoration, buffer strips) between
farming areas and streams/ditches may
hinder nutrients from entering the aquatic
system, and (iii) the installation of overflow
areas or permanent pond-like water bodies may slow down the transport velocity
within the watershed, and allow processing
of nutrients which have already entered the
aquatic system. These strategies target a decrease in nutrient (N and P) concentrations
and loads, i.e. aim to increase the retention
capacity of the watershed. The types of measures may differ in efficiency of N or P retention (Mitsch et al. 2001). Created wetlands
with permanent water bodies are considered
particularly suitable to remove N (Fleischer
et al. 1994; Leonardson 1994; Mitsch et al.
2001; Paludan et al. 2002; Hey 2002; Kadlec
2005; Mitsch et al. 2005; Olde Venterink et
al. 2006), by providing required conditions
for denitrification. High nitrate, low oxygen
content (Knowles 1982), and high macrophyte biomass, i.e. litter as carbon sources
and surface for biofilms (Weisner et al. 1994)
facilitate denitrification; these conditions
can be sustained in pond-like wetlands created in agricultural landscapes.
Multiple purposes. This type of created wetlands may have the potential for abating part
of either the biodiversity loss or eutrophication problem, or both simultaneously.
However, the simultaneous targeting of biodiversity and nutrient retention purposes in
created wetlands requires close evaluation, as
high nutrient concentrations may interfere
with biodiversity goals (Hansson et al. 2005).
Wetland creation may also increase the risk
for other undesired environmental consequences, as wetlands sustain a comparably
high risk for climate gas emissions (Mitsch et
al. 2001; Verhoeven et al. 2006; Stadmark &
Leonardson 2005, 2007) or host organisms
potentially hazardous to human health (e.g.
mosquitos, Dale & Knight 2008). On the
other hand, wetlands created for nitrogen
abatement may also sustain several ancillary
beneficial functions (e.g. P retention, Tonderski et al. 2005), and potential links between
functions, and multifunctionality may be of
relevance for restoration management (Findlay et al. 2002; Euliss et al. 2008).
Studied ecosystem functions & services. The
target ecosystem services N retention and
biodiversity (species and functional diversity) of created wetlands are the focus of this
thesis. Nitrogen retention and biodiversity
conservation are stated aims of the Swedish
Biodiversity and ecosystem functioning in created agricultural wetlands
environmental objectives (www.miljomal.
nu), and are functions desired to be achieved
by wetland creation. Further, the ecosystem
functions P retention (ancillary service), litter decomposition and methane emission
(environmental risk), are considered in the
papers of this thesis; firstly to investigate
targeted and ancillary benefits as well as environmental risks of wetland creation, and
secondly to investigate functional coupling
and joint ecosystem functioning.
Ecosystem services and functions present
in created wetlands are summarized in Fig 5.
Certain wetland functions can be passively
facilitated by the physicochemical conditions
(e.g. particle and P sedimentation due to reduced current), others are actively sustained
by biotic compartments (e.g. N retention
due to bacterial denitrification plus plant
nutrient uptake; litter decomposition by
microbes and shredder macroinvertebrates).
Wetland processes/properties depend on the
prevalent abiotic characteristics (e.g. hydraulic turbulence, pH). In addition, the abiotic
environment mediates indirect effects via
the biotic compartments; abiotic conditions influence the composition of biotic assemblages (e.g. absolute amount of nutrient
supply excludes/facilitates species), and the
biotic interactions between biota (e.g. competition of plants and bacteria for nutrients),
and thereby ecosystem rates and properties.
In turn, the ecosystem properties themselves
(e.g. plant biomass) may affect the abiotic
environment (e.g. flow patterns or shading/
UV radiation) or ecosystem processes (e.g.
amount of plant nutrient uptake), thus indirectly or directly affect wetland functions.
Thesis relevance. The effect of created wetlands on regional diversity is still largely
unknown, and with regard to combining
wetlands for several environmental goals,
impacts on diversity need to be considered
(Paper I). Previous diversity investigations in
wetlands focus almost exclusively on higher
organisms; in contrast, information on microbial assemblages is scarce, although major
wetland processes are mediated by bacterial
communities, e.g. denitrification (Paper II).
Despite the multifunctionality of wetlands,
studies that compare (usually very few) created wetlands for their functional capacity are
often limited to one ecosystem function at a
time, or else, the potential for interactions/
relations between ecosystem functions has
not been considered (Papers III & V). Also,
temporal aspects on the functioning and biodiversity of wetlands need to cover longer
time scales, to test if biodiversity-function
interactions interact with time (Paper IV).
Biotic parameters are mostly not considered when wetland functions are assessed/
predicted, and their relative effect in highly
dynamic wetland environments is unknown
(Paper V). Further, research on biodiversity–
function links that considers non-random
species assemblages (Paper IV), and diversity
effects in dynamic environments (Paper V)
are needed.
The investigated created wetlands as a whole
provided several ecosystem services; Table 1
summarizes ecosystem services studied in the
five papers of this thesis. Please observe that
ecosystem services were measured with different methods, limiting direct comparability. I included Table 1 to provide an overview
of my results, and as a guide for which of the
five papers to consult for details.
Biodiversity results
The requirements for nutrient retention in
created wetlands are not an obstacle for biodiversity.
Incoming nutrient concentrations and hydraulic loading rate (i.e. retention capacity
indicators, Kadlec & Knight 1996; Kadlec
2005) were not associated to composition
or diversity of aquatic macroinvertebrates in
Biodiversity and ecosystem functioning in created agricultural wetlands
Table 1. Ecosystem services related to nutrient retention and biodiversity that are provided by created agricultural wetlands. The table gives an overview on the ecosystem service(s) stucied in each of the five thesis papers,
Plant richness
Macroinv. richness
(Species number)
(Species number)
Paper I
Field, n=36
Vegetation season
Paper II
Field, n=32
Vegetation season
Paper III
Model, n=36
Paper IV
Exp. wetlands, n=18
Functional diversity
Denitrif. richness*
(DGGE band number)
Retention function
(kg ha-1 yr-1)
(kg ha-1 yr-1)
6 – 51
7 – 16
5 – 22
2 – 13
419 – 2135
8505 (15 ha)
3 – 21
6 – 29
675 – 1068
Paper V
Field, n=14
12 – 30
19 – 47
135 – 2156
-77 – 89
Vegetation season
*nir/nos. Bacterial denitrifying enzyme genes nirK+nirS and nosZ coding for enzymes needed for different steps in the denitrification chain (see also Fig 7A). DGGE
(see Paper II). **all years/wetlands. nm not measured.
created wetlands (Table 2; Paper I). Macroinvertebrate and plant assemblages were instead constrained by parameters determining
potential colonization success (wetland age,
distance and connectivity to potential source
habitats, Table 2). In general, large variation
fractions of richness and composition were
unexplained by environmental parameters
(Table 2). Similarly, only low variation fractions could be assigned to environmental factors in previous studies on diversity and assemblage composition of macroinvertebrates
(Lundqvist et al. 2001; Van de Meutter et al.
2008), plants (Edvardsen et al. 2006) and
bacteria (Langenheder & Ragnarsson 2007).
Further, a study comparing lake bacteria,
phytoplankton, zooplankton, and fish assemblages indicated that dispersal predictors are
more important than local environment for
the two latter, larger and less motile groups
(Beisner et al. 2006). The large unexplained
variation may also be due to factors which
where unaccounted for in Papers I and II;
these include for example stochastic effects
(Scheffer et al. 2006), preemption/priority
effects (De Meester et al. 2002; Forbes &
Cole 2002; Chadwell & Engelhardt 2008)
and productivity effects operating on regional scale (Chase & Leibold 2002), which are
particularly relevant for small water bodies.
I also found minor effects of retention capacity indicators on richness and composition of wetland plants (unpublished data).
Plant composition varied with N:P ratio,
and submerged plant richness decreased with
P and suspended solid levels (Table 2). Submerged macrophytes may be effectively suppressed by high P and turbidity levels (Scheffer 1998; Jeppesen et al. 2000). The high
diversity of small isolated ponds is partly assigned to their higher likelihood of being in
a macrophyte-dominated state (compared to
larger lakes with fish; Scheffer et al. 2006). If
high P levels in created wetlands would cause
shifts from a macrophyte to a phytoplankton
state, this may ultimately lead to a decrease
in macroinvertebrate diversity (Declerck
et al. 2005). Very high P levels were associ-
Biodiversity and ecosystem functioning in created agricultural wetlands
ated with low macroinvertebrate richness in
other Swedish created wetlands (Hansson et
al. 2005). The phosphorus levels observed in
created wetlands in Papers I-V were low to
moderate by comparison. Further, the P levels were highly heterogeneous, i.e. negative
P effects are likely to operate on local scale
only. Regarding nitrogen, no adverse effects
on richness of plants or macroinvertebrates
were found, despite the very high ambient
concentrations (3-20 mg l-1), i.e. regionally
elevated N levels.
Composition of eubacterial biofilm communities (Table 2) and overall richness of
bacterial and denitrifying communities (Paper II), were more strongly influenced by the
functional requirements of retention wetlands. Nitrate concentration explained part
of the DGGE (denitrifying gradient gel electrophoresis; see Paper II) band structure of
eubacteria (Table 2). The majority of the ex-
plained variation was accounted for by biotic
parameters, i.e. richness of emergent and submerged plants (Table 2; Paper II). Eubacterial
community structure was earlier shown to be
influenced by vegetation state (Langenheder
& Prosser 2008), water chemistry (Hewson
et al. 2003), and wetland morphology (Hewson et al. 2007), and similar magnitudes of
influence as in Paper II have been reported
for bacterial communities from other aquatic
habitats (Beisner et al. 2006; Langenheder &
Ragnarsson 2007).
In conclusion, environmental parameters
seem to have only minor effects on the assemblages of macroinvertebrates and plants;
accordingly, the prerequisites for a simultaneous retention function of created wetlands
did not hinder the establishment of diverse
local and regional assemblages in these systems. In comparison, the bacterial assemblages seemed to be influenced mainly by
Table 2. Influence of abiotic and biotic wetland parameters on biodiversity of created wetlands.
Wetland characteristics
Species richness
Eubacteria (16S)*
Denitrifying enzyme
Composition Richness
– (Su)
x (nirK)
– (Su)
x (nos)
+ (nos), –
+ (nos)
x (nirS)
+ (nirK)
ns (total)
rS = 0.44
< 10% each 17 – 34%
*Eubacteria were targeted by applying the 16S rDNA primer (Paper II). **The denitrification enzyme genes nirK, nirS and nosZ were sampled from
bacterial biofilm. Wetland characteristics are abbreviated as NO3 nitrate-N and P total phosphorus concentration, N:P total nitrogen:phosphorus ratio,
q hydraulic loading rate, depth mean water depth, area wetland size, T water temperature, TSS total suspended solids, age time since wetland
creation, connect lotic surface water connectivity, agric dominance of agricultural use in direct vicinity, forest distance to forested inland area, north
northing within the region, Em emergent plants, Su submerged plants, Fl floating and floating-leaved plants. x effect present, + – positive/negative
direction. TOTAL EXPL total variation fractions explained by all parameters. ns not significant. rS Spearman Rank correlation.
Biodiversity and ecosystem functioning in created agricultural wetlands
Created wetlands are equally valuable habitats as natural ponds and large-scale wetland
creation has the potential to enhance regional
Created agricultural wetlands sustained
similar local and regional macroinvertebrate
richness and overall species pools when compared to natural ponds in the same region
(Paper I) and in agricultural landscapes
elsewhere (Williams et al. 2004; Robson &
Clay 2005; Biggs et al. 2007; Davies et al.
2008). In created wetlands, insects clearly
dominated local assemblages as well as the
regional pool; the most diverse orders being
aquatic beetles, dragon-and damselflies, caddisflies, and water bugs (Paper I). Compared
to natural ponds in the same region, created
wetlands hosted more lotic groups (mayflies,
stoneflies), probably as a result of higher connectivity to the watersheds.
Created wetlands sustained highly individual plant and macroinvertebrate assemblages, i.e. their spatial heterogeneity and
β diversity was high (Paper I; unpublished
data). Wetlands located in landscapes with
extensive wetland creation (hence, high total aquatic habitat density; Paper I) hosted
richer local plant and macroinvertebrate assemblages (Fig 6A; Paper I). With regard to
macroinvertebrates, a positive density effect
prevailed even on regional scale: the cumulative macroinvertebrate species pool sustained
in the high density landscape was greater (110
compared to about 90 species, Fig 6B) than
that of the other two regions. In small isolated
ponds, the richness of some organism groups
(including macroinvertebrates and plants)
may be promoted by second order effects of
habitat size (Scheffer et al. 2006), implying
that several small wetland sites likely sustain
more species than one large site of equal area
(Oertli et al. 2002).
Predicted diversity (n=8)
vegetation and water quality, i.e. factors that
may also affect the N retention capacity of
wetlands (Kadlec 2005; Kadlec 2008).
Habitat density
Fig 6. Effects of habitat density on local (A) and regional (B) species diversity of plants and macroinvertebrates. Density of aquatic habitats per landscape
(Low, Mod, High) varied due to differences in wetland creation efforts (0.18, 0.24, and 0.35 wetlands
per km-2); 15% (i.e. n=13, 8, and 15 wetlands) of all
created wetlands were investigated per landscape (Paper I).
In conclusion, created wetlands serving
simultaneous diversity and N retention purposes have similar capacities for biodiversity
conservation than natural agricultural ponds.
The creation of many small created wetlands
may promote both local and regional diversity of macroinvertebrates, particularly if wetland creation efforts raise total aquatic habitat densities by more than 30% (Paper I).
Biodiversity and ecosystem functioning in created agricultural wetlands
NO3- NO2- NO N2O N2
DGGE band richness
Created wetlands sustain functionally diverse
denitrifier communities; the local environment is closely linked to bacterial denitrifier
The three studied denitrifying enzyme
genes (Paper II) were present in all wetlands,
indicating that denitrifier communities with
the potential for processing early and late
denitrification steps (Fig 7A) were ubiquitously established (Table 1). Complexity and
diversity of denitrifying enzyme gene composition of created wetlands were also comparable to that of agricultural soils (Throbäck
et al. 2004). Further results from DNA sequencing of the denitrifying communities
from the same created wetlands confirm that
these systems host highly diverse denitrifier
communities (Milenkovski 2009).
In created wetlands, the nirK and nirS
enzyme genes decoding for an early denitrification step were more diverse on local and
regional scales, than the nosZ enzyme gene
coding for the last denitrification step (Table 1). The finding that diversity was greatest
for the nir genes and lowest for nosZ is supported by a general relationship of lower diversity for nosZ compared to both nirK and
nirS (Wallenstein et al. 2006). Generally, the
band structure of DGGE patterns for the
three enzyme genes was only weakly related
to environmental conditions of created wetlands (Table 2) and unrelated to eubacterial
diversity (Paper II), however similarly low
explanation fractions have been reported
from other systems (Langenheder & Ragnarsson 2007).
Interestingly, the spatial variation of denitrifying enzyme gene richness among wetlands was linked to factors which are known
(Kadlec & Knight 1996; Kadlec 2005) to
affect N retention, i.e. inlet nitrate concentration and hydraulic loading rate. Further,
the enzyme gene richness of the early (nir)
and last denitrification step (nos) responded
differently to nitrate concentration; while
Nitrogen retention g m-2 d-1
Functional diversity results
5 - 10
> 10
Nitrate-N concentration mg l-1
Spearman Rank
rS = 0.605***
Shannon Index functional diversity
Fig 7. Denitrifying functional diversity. (A) Denitrification chain: reaction steps catalyzed by the nir
and nos enzyme types are highlighted; the community
composition of the enzyme genes nirK, nirS and nosZ
of 32 created wetlands was investigated in Paper II.
(B) Richness of bacterial denitrification enzyme genes
coding for the second (nir) and last (nos) step in the
denitrification chain was affected by nitrate concentration. At high nitrate concentrations, both enzyme
genes became equally rich, otherwise the nir type
(nirK + nirS) enzyme genes dominate. (C) Shannon
functional diversity (based on DGGE band numbers
of the nir and nos enzyme genes, data from Paper II)
correlated (p<0.0005) with nitrogen retention of created wetlands (annual predictions from Paper III).
Biodiversity and ecosystem functioning in created agricultural wetlands
the nirK+nirS band richness decreased with
nitrate concentration, the nosZ richness increased (Fig 7B). Similar to our findings,
Kjellin et al. (2007) found that the nosZ enzyme gene composition varied with nitrogen
and hydraulic loading rates. In contrast however, the nosZ diversity was lowest at high
nitrogen and hydraulic loads (Kjellin et al.
Previous research (Wallenstein et al. 2006;
Kandeler et al. 2006) suggests that denitrifier
community structure and abundance in soils
is primarily controlled by factors other than
nitrate supply (e.g. carbon content, degree of
water saturation, pH), despite the fact that
the denitrification rate is stimulated by N/
nitrate (Kjellin et al. 2007). Nitrate did not
affect DGGE band pattern composition or
richness within single enzyme genes in Paper
II (Table 2); however, the overall complexity
of bacterial denitrifier assemblages was influenced. Earlier studies of environmental effects
on bacterial denitrifier communities have
not compared functionally different enzyme
genes (Braker et al. 1998; Hallin & Lindgren
1999; Braker & Tiedje 2003; Hannig et al.
2006; Bremer et al. 2007), and largely assumed that effects on one gene can represent
effects on the overall denitrifier community.
The results from Paper II strongly suggest
that environmental effects (particularly nitrate supply) on denitrifier richness, and to
a lesser extent composition, differ for functionally different enzyme genes. Hence, the
use of single enzyme genes will not suffice to
characterize the environmental influence on
the overall bacterial denitrifier community in
created wetlands.
Ecosystem functioning results
Ecosystem functions measured in this
thesis included the (target and ancillary)
ecosystem services N retention and P retention (Tables 1 & 3) and further, the wetland
functions methane production and litter decomposition (Table 3, Papers III & V). Also
studied were the abiotic and biotic controls
(Table 3), potential links between functions
(Papers III & V), and temporal changes of N
retention (Fig 8). In the final Paper V, joint
ecosystem functioning (i.e. simultaneous
performance of N retention, P retention and
litter decomposition, Fig 9) was studied.
Biotic factors affect the functioning of highly
dynamic created wetlands and biotic influences
partly differ between functions.
In Paper V, ecosystem functions in created wetlands with variable abiotic and biotic characteristics (dynamic environments)
were investigated. Variation occured spatially
(among wetlands; controlled for by parallel
investigations in 14 wetlands) and temporally (over seasons and years; controlled for
by repeated sampling). This setup allowed to
test if biotic variables explained differences
in three ecosystem functions (N retention,
P retention and litter decomposition) additionally to abiotic factors, in dynamic environments.
Biotic parameters affected the processes/
properties underlying the three functions
(Table 3), and models containing both abiotic and biotic factors, explained more functional variation than abiotic factors alone (Paper V). Similarly, in Paper III, the variation
in methane production was best explained
by including biotic and abiotic explanatory
variables in the model (Table 3). Prediction
models for wetland functions (e.g. N and P
retention, Kadlec & Knight 1996; methane
production, Bastviken et al. 2004) are commonly based on the abiotic, physicochemical
dynamics. The inclusion of biotic parameters
may significantly improve predictability.
Nevertheless, abiotic factors explained larger
variation fractions for most functions; the
abiotic factors which were found to influence
the wetland functions (Table 3) agreed with
previous studies (see detailed discussions in
Papers III & V).
The biotic parameters in Paper V covered
influences of two biotic groups (plants or
macroinvertebates), and distinguished be-
Biodiversity and ecosystem functioning in created agricultural wetlands
Table 3. Influence of abiotic and biotic wetland parameters on ecosystem functions of created wetlands.
Nitrogen retention
(m d-1)
PO4 load
Morphohydrology q
Physicochemistry pH
Pl rich
Pl fuDiv
M div
M shr rich
Pl cover
Pl biom
M shr abu
N rate
(g m-2 d-1)
Phosphorus retention
kNO3* NO3 rate
(m d-1)
(g m-2 d-1)
(m d-1)
P rate
(g m-2 d-1)
Litter Decomposition
kPO4 PO4 rate
(m d-1)
(g m-2 d-1)
kD 1mm
kD 5mm
CH4 *
(g m-3)
*at reference temperature. Ecosystem processes/properties include the removal rate coefficients (kN, kNO3, kP, kPO4) and areal removal rates (N
rate, NO3 rate, P rate, PO4 rate) for total nitrogen, nitrate-N, total phosphorus, and phosphate-P; the litter decomposition rate coefficient in 1 and 5
mm mesh size bags (kD1mm, kD5mm); the dissolved concentration of methane (CH4). Wetland characteristics are abbreviated as NO3 nitrate-N, P
total phosphorus, and PO4 phosphate-P concentration, PO4 load phosphate-P load, q hydraulic loading rate, depth mean water depth, area wetland
size, T water temperature, age time since construction, surface inlet type: surface-fed. Pl rich Plant richness, Pl fuDiv Plant functional diversity, M
shr rich Macroinvertebrate shredder richness, Pl cover Plant cover, Pl biom Plant biomass, M shr abu macroinvertebrate shredder abundance.
TOTAL EXPL total variation fraction explained by all parameters. + – positive/negative influence. ns not significant.
tween (i) diversity and (ii) abundance effects. The significant diversity effects that
were found, were positive (Table 3). At least
one process involved in N retention, P retention and litter decomposition, respectively,
increased with diversity of either plants (retention) or macroinvertebrates (decomposition). The abundance (biomass, cover) of
plants on the other hand, seemed to influence ecosystem functions differently (Table
3; Papers V & III). Plant biomass was linked
to increasing N retention, plant cover to decreasing P retention, litter decomposition,
and methane production.
In conclusion, biotic parameters need to
be accounted for in order to predict wetland
functioning. Management of vegetation succession in created wetlands needs to balance
between strategies facilitating plant richness
and biomass to sustain desired functions;
alternatively, one wetland function could
be prioritized above others, and vegetation
management adopted accordingly.
Biodiversity - function links
Functional diversity of denitrifiers correlates to
N retention capacity of created wetlands.
Denitrification is the major pathway of N
removal in nitrate-rich environments (Seitzinger et al. 2006; Beaulieu et al. 2008), and
has the potential to limit the N retention
function of created wetlands. The denitrification chain is a series of reaction steps, in
which bacterial denitrifiers play a crucial role
in expressing the enzymes needed to catalyze
the three reductions from nitrite to dinitro25
Biodiversity and ecosystem functioning in created agricultural wetlands
gen gas (Zumft 1997; Fig 7A). Richness and
structure of the bacterial enzyme gene assemblage may affect these steps, as denitrifier
populations differ in physiological properties,
e.g. their affinity for electron acceptors and
donors or the relative reaction rates of denitrification steps (Phillipot & Hallin 2005).
Results from previous studies (Cavigelli &
Robertson 2000, 2001; Holtan-Hartwig
et al. 2001, 2002; Rich et al. 2003) suggest
that differences in the community composition of soil-denitrifying bacteria may explain
differences in denitrification rates. A skewed
community composition of the enzyme
genes (nir compared to nos) may potentially
limit one of the denitrification reactions, i.e.
form a bottleneck in the denitrification process. Based on data from Paper II, I calculated
the Shannon index for richness of the functionally different enzyme genes nir and nos
per wetland (i.e. assuming equal importance
of band richness for early and late denitrification steps), to assess denitrifier functional diversity. I then compared functional diversity
with the predicted annual nitrogen retention (case study predictions from Paper III)
for the 32 wetlands, which were included in
both studies. I found that denitrifier functional diversity was positively correlated to
predicted annual N retention (Fig 7C). Bell
et al. (2005) suggest that bacterial community structure affects ecosystem functioning,
and that species richness has positive effects
on bacterial ecosystem functioning; Jayakumar et al. (2004) linked nirS diversity to
high denitrification rates. In contrast to my
results, Kjellin et al. (2007) observed highest
denitrification rates at sites with lowest nosZ
diversity (compared to other sites in the same
wetland). Rich et al. (2004) found differences
in denitrification rates between wetland and
upland soils, but no structural or diversity
differences in the nosZ gene. However, earlier
studies on enzyme genes did not cover more
than one denitrification step and focussed on
richness within single enzyme genes; my results indicate that N retention may be highest if functional diversity across genes is high.
Vegetation type rather than diversity affects N
In Paper IV, biodiversity and N retention
were investigated in experimental wetlands
with controlled abiotic conditions. This
allowed to investigate the effect of vegetation state and diversity-function links (Fig
8). Regarding N retention, wetlands with a
vegetation state of high biomass (tall emergent plants) exhibited continuously higher
N removal during four years (Fig 8E), although plant Shannon diversity in these systems decreased over time and in relation to
other vegetation states (submerged or freelydeveloped vegetation; Fig 8B). Thus, the N
retention function in experimental wetlands
seems closely linked to effects mediated by
plant identity or vegetation state, while it was
indifferent to vegetation diversity. Latest diversity-function research suggests that diversity effects are often associated to a ‘sampling
effect’, i.e. ecosystem function is facilitated if
species with high performance capacity are
present (e.g. Bracken & Stachowicz 2006;
Cardinale et al. 2006).
In Paper V, three parallel ecosystem functions (N and P retention and litter decomposition) and biodiversity were investigated
in highly dynamic, full-scale created wetlands. Biodiversity influences on ecosystem
functioning, particularly regarding plant
diversity, became apparent after accounting
for abiotic factors (Table 3). Plant diversity
parameters affected both P retention (plant
richness) and N retention (functional plant
diversity) positively (Paper V). This suggests that in highly dynamic created wetland environments, plant diversity may be
important to assure functioning over time.
Biodiversity is hypothesized to serve as an insurance against disturbance and to stabilize
ecosystem functioning (Hooper et al. 2002,
Loreau et al. 2003). Also, while N retention
seemed indifferent to species loss within
functionally uniform plant assemblages (e.g.
emergent plants, Paper IV), freely assembled
vegetation in agricultural created wetlands
(Paper V) could consist of up to five different functional groups (submerged, rooted
Biodiversity and ecosystem functioning in created agricultural wetlands
floating-leaved, free-floating, emergent, and
woody wetland species). Created wetlands
did not maintain extensive cover of emergent
vegetation; large parts exceeded depth limits
tolerable by emergent plants (average depth
1 m, Paper I). These parts of the water body
were instead vegetation-free or colonized by
aquatic (obligate hydrophytic) vegetation.
The positive influence of functional plant
diversity on N retention found in paper V
may therefore rather be interpreted as a positive effect of vegetation as such (compared
to vegetation-free), or as an effect of an even
distribution among all the functional plant
groups, ensuring that emergent plants are
abundant (among others).
Temporal trends are apparent for biodiversity,
but not for N retention; time effects depend on
vegetation type.
In Paper IV, biodiversity and N retention
were investigated over a four year period (Fig
8) to investigate effects of time and succession.
Plant richness
N retention varied among years (likely due
to temperature effects), but no trend across
years was apparent (Fig 8E). Only in the first
year after creation, wetlands without planted
vegetation performed somewhat lower N retention than in the following years.
Diversity of macroinvertebrates and plants
undergoes temporal changes as created wetlands mature (Paper IV; age effect in Paper
I). Planted wetlands had higher initial plant
diversity than unplanted wetlands, but in all
vegetation states plant and macroinvertebrate
species numbers increased over time (Fig 8A,
C). Over a four-year period, however, plant
Shannon index decreased in planted emergent wetlands, as a few plant species became
highly dominant (Fig 8B). Hence, there were
significant interactions of time and vegetation state. In contrast, the initial differences
in macroinvertebrate diversity between vegetation states leveled out over time (Fig 8D).
These results suggest that time effects
on biodiversity differ between (i) diversity
measures (species richness/Shannon index),
Macroinvert. richness
N retention
Vegetation state (treatment)
Plant Shannon index
Macroinvert. Shannon index
Planted with emergent veg.
Planted with submerged veg.
Freely-developed veg.
Fig 8. Temporal variation in biodiversity and nitrogen retention of experimental wetlands with differing vegetation states (n=6 each; Paper IV). Temporal trends (across years) for biodiversity (A-D) depend on vegetation
state (legend), biotic group (plants or macroinvertebrates), and applied diversity measure (richness or Shannon
index); time and vegetation state interactions were significant (Paper IV). Nitrogen retention (E) function
differs between years and between vegetation states, but no continuous trend across years and no interaction
were observed.
Biodiversity and ecosystem functioning in created agricultural wetlands
(ii) biotic groups (plants/macroinvertebrates), and (iii) vegetation states (emergent/
In conclusion, for an assessment of created wetland services, across-year trends need
to be regarded, particularly for biodiversity.
Species accumulation of both plants and
macroinvertebrates is prevailing during at
least 4 years after establishment, indicating
ongoing external recruitment. Relative species abundance is changing faster in planted
than unplanted wetlands, but differs between
emergent and submerged vegetation states,
suggesting that competition impacts on di-
Plant richness
Hydr. load. rate
N:P ratio
Plant biomass
Fig 9. Joint ecosystem functioning in created wetlands. Three ecosystem functions, nitrogen retention,
phosphorus retention, and litter decomposition, described by the process rate coefficients kN, kP and
kD (black arrows), respectively, were simultaneously
employed as multiple response variables in redundancy analysis RDA (Paper V); a set of 10 environmental parameters was tested for how much variation
in joint functioning was explained among 14 created
wetlands (black circles). The graph shows the final
ordination diagram. Significant environmental gradients (grey arrows) delineate strength and direction of
environmental influence by arrow length and direction (analogous interpretation for response variables;
black arrows). Proximity between circles delineates
similarity of environmental characteristics and ecosystem function of wetlands. In total, 67% (p=0.002)
of the functional variation was explained.
versity are prevailing already after two years
if emergent vegetation dominates.
Function - function links
The risk for methane emission is independent
from the N retention capacity of created wetlands.
Methane emission is an example for an
environmental risk (climate gas emission)
which may increase due to large-scale wetland creation (Mitsch et al. 2001; Verhoeven
et al. 2006; Stadmark & Leonardson 2005,
2007). N retention is often the primary environmental goal targeted by large-scale wetland creation (Fleischer et al. 1994; Leonardson 1994; Mitsch et al. 2001; Paludan et al.
2002; Hey 2002; Kadlec 2005).
Methane emission (diffusional flux) from
created wetlands was found to be generally
low, although all wetlands had quantifiable
methane production (Paper III). The extent
of production could be predicted by wetland
characteristics (Table 3; Paper III), of which
some (nitrate concentration, plant cover)
were shown to affect N retention processes of
created wetlands (Table 3; Paper V). While
nitrate concentration and plant cover can be
expected to support N retention (Weisner et
al. 1994; Kadlec 2005; Paper V), they tend to
seem to suppress methane production (Paper
III). Methane emission was not correlated to
the N retention predictions (Table 1) for created wetlands. In conclusion, the investigated environmental risk and benefit of created
wetlands can be managed independently and
there is potential to optimize N retention.
Certain wetland functions are coupled: N
retention increases with fast litter decomposition.
In accordance to the literature, the regression models for N retention processes (Table
3) indeed predicted 40-70% of the functional
variation based on parameters known to affect
denitrifier activity or composition: hydraulic
load (Phipps & Crumpton 1994; Kjellin et
al. 2007; Paper II), nitrate concentration (Ka-
Biodiversity and ecosystem functioning in created agricultural wetlands
dlec 2005; Seitzinger et al 2006; Beaulieu et
al. 2008; Paper II), plant biomass (Weisner
et al. 1994; Eriksson & Weisner 1997, 1999;
Lin et al. 2002) and plant diversity (RuizRueda et al. 2008). However, in addition to
these factors, the model fit and explanatory
power improved further (60-90%; Paper V),
if the decomposition rate coefficient kD was
included as an independent variable. Litter
decomposition rate limits the turnover of assimilated carbon (i.e. plant biomass) for heterotrophic consumption (Webster & Benfield 1986; McKie et al. 2006). For created
wetlands, not only carbon amount (resource
pool size: plant biomass), but also efficiency
of carbon processing (litter decomposition)
facilitate N retention. Hence, different process rates may be coupled in created wetlands.
In conclusion, if N retention and decomposition rate are linked, wetland management needs to develop strategies to optimize
several functions simultaneously.
N & P retention
P retention
N retention
No substantial
retention service
Fig 10. Multiple ecosystem services in created wetlands. Most wetlands (circles as in Fig 9) have suboptimal capacity of either N or P retention. Some wetlands, however, are capable of performing substantial
N and P retention simultaneously (upper left corner).
Furthermore, plant richness (an ancillary ecosysystem
service) could also be sustained parallelly in these
systems, considering the environmental parameters
distinguishing joint ecosystem functioning among
created wetlands (grey arrows in Fig 9).
Joint ecosystem functioning: Created wetlands can sustain multiple ecosystem services.
In Paper V, joint ecosystem functioning,
i.e. the simultaneous performance of the wetland functions N retention, P retention and
litter decomposition, was investigated (Fig
9) to assess the factors which distinguish the
capacity for multifunctionality.
Variation in joint ecosystem functioning
was explained by abiotic and biotic wetland
characteristics, i.e. N:P ratio and hydraulic
loading rate, as well as plant richness and
plant biomass. Two thirds of the total functional variation were explained; biotic factors accounted for more than one third of the
explainable variation.
The functions responded differently to environmental parameters (Fig 9). P retention
increased with plant richness and decreased
with plant biomass. N retention and litter
decomposition varied independently from
plant richness and biomass; these two functions were constrained by hydraulic load (+)
and increasing N:P ratio (-) instead (Fig 9).
These results indicate that only some wetlands were capable of performing the ecosystem services N and P retention simultaneously and substantially (Fig 10), while most
wetlands performed one service suboptimally. Wetlands with high simultaneous N and
P retention tended to have higher hydraulic
loading rate, lower N:P ratio, higher plant
richness and lower plant biomass, compared
to the average over all 14 wetlands. Suboptimal capacities of either service were distinguished by very high hydraulic load in
combination with high plant biomass (suboptimal P retention); alternatively (suboptimal N retention) by high N:P ratio, thus
slow litter decomposition, and low hydraulic
loads. However, with constellations of low
hydraulic load and plant richness, combined
with high N:P ratio and plant biomass, created wetlands may run the risk of providing
no substantial nutrient retention service.
Biodiversity and ecosystem functioning in created agricultural wetlands
What explains diversity and ecosystem function of created wetlands?
Both diversity and functions of created
wetlands were influenced by abiotic, vegetation, spatial (regional), and temporal parameters, and the magnitude and direction of the
influence differed. Further, two types of interactions were observed, between diversity
and function and among functions.
Local environment. Generally, the explained variation fractions for composition
and diversity of any of the three investigated
biotic groups (plants, macroinvertebrates,
bacteria) were low, usually below 30% (Table 2), whereof minor proportions were accounted for by abiotic, water quality parameters (Papers I & II; unpublished data).
In contrast, major variation fractions of
ecosystem functioning (single and joint functioning) were explained by the local environment (40% to 70%), with abiotic parameters
exceeding the proportions explained by biotic ones, for most functions (Table 3).
Vegetation. Plant diversity influenced bacterial diversity and composition (Paper I),
and more so than abiotic parameters. Vegetation states with high biomass affected plant
diversity and evenness negatively in the longterm (Paper IV).
Vegetation effects on ecosystem functioning were related to diversity and abundance
(Table 3) as well as vegetation state (Fig 8).
Vegetation diversity had positive or no effects
on functions (Table 3; Paper IV), while vegetation abundance was positively related to
some functions (N retention) and negatively
to others (P retention, litter decomposition,
methane production). Vegetation effects
were of subordinate importance compared to
abiotic parameters in dynamic wetland environments for all functions but P retention. In
more stable abiotic environments, N retention capacity differed with vegetation state,
being highest in highly productive states.
Regional watershed/landscape. Spatial factors (including regional and location parameters) were rather important for biodiversity,
explaining more than the local abiotic/biotic
environment, at least for larger biota (plants
and macroinvertebrates). For these groups,
local assemblage establishment seemed to
depend on the landscape context (connectivity, distance to source habitats) and the
total regional habitat pool (habitat density).
Similar patterns have been shown for lake
assemblages; small organisms depending on
environmental, larger organisms on dispersal
conditions (Beisner et al. 2006).
Spatial parameters were not of directly
related to ecosystem functioning. However,
the location of created wetlands in a watershed is a management decision; spatial location determines the magnitude of load and
concentration (which are important for N
retention) received by a given wetland; thus
spatial factors may indirectly influence function.
Time. Biodiversity was also affected by
time, i.e. ongoing succession/aging over
the first few years after establishment. The
number of species increased (at least over 4
years), while the compositional diversity and
evenness developed depending on the initial
species constellations. If planted wetlands
serve as models for ‘late succession stages’,
diversity of plants is likely to decrease, when
created wetlands reach the later succession
Temporal succession trends seemed less
important for the ecosystem service N retention, however daily and seasonal variations
were observed (Papers IV & V). Retention
function (N) has been reported to be annually variable (Kadlec & Knight 1996; Kadlec
2005), but variation is often related to seasonal/interannual variation in flow, concentration, and temperature, rather than to aging/
succession effects. Wetland age could partly
be relevant for methane production, i.e. the
risk for climate gas emission may increase in
older wetlands. Although not investigated
Biodiversity and ecosystem functioning in created agricultural wetlands
here, aging effects may also be relevant for
litter decomposition and P retention. Earlier studies showed that litter decomposition increases with wetland age (Atkinson &
Cairns 2001; Spieles & Mora 2007) mainly
as a result of litter shortage during the first
years, and that P retention decreases in older
systems (Kadlec & Knight 1996; Braskerud
et al. 2005), mainly as a result of saturated P
binding capacities.
Interactions. Diversity-function links were
shown to be particularly relevant for highly
dynamic environments, where positive effects of diversity parameters on process rates
involved in N retention, P retention and litter
decomposition were observed after accounting for abiotic variation (Table 3). Functional
diversity among bacterial denitrifying genes
was positively correlated to N retention (Fig
7). N retention in experimental wetlands
with low environmental variation, seemed
indifferent to succeeding biodiversity loss
(Fig 8). Function-function links were found
for N retention and litter decomposition; N
retention processes were facilitated when litter decomposition was fast (Paper V).
(1) The ecosystem services provided by created agricultural wetlands are comparable
to natural systems or to other constructed wetland types; created wetlands may
thus contribute to biodiversity conservation and eutrophication abatement on
watershed scales.
(2) Ecosystem functioning of created wetlands was more clearly linked to environmental conditions than biodiversity.
Functioning may thus be managed primarily by wetland design and location in
the watershed. Biodiversity management
seems most efficient on regional scale;
benefits due to increasing total habitat
densities per watershed by more than
30% seem very likely. Optimizing the
abiotic prerequisites for N retention (N
concentration and hydraulic load) does
not seem to be contradictory to biodiversity aims.
Apart from wetland design/placement,
management of vegetation type and
extent is important for ecosystem functioning and biodiversity. Vegetation
management affects ecosystem services
differently: a high plant biomass favors
N retention, but may inhibit plant diversity, and also P retention.
Diversity parameters were positively related to ecosystem functions; a given
function was enhanced by diversity of a
particular biotic group or by functional
diversity. This suggests, that if created
wetlands are aimed at sustaining several
parallel ecosystem functions (and services), biodiversity should be promoted.
Functional diversity and composition
of denitrifiers is influenced by retention
prerequisites (N concentration and hydraulic load); high functional diversity
may be linked to higher N retention capacity of created wetlands.
Wetland creation contributes to (yet
suboptimal) N retention, while the risk
for simultaneous methane emission is
low. An optimization of the N retention
function is unlikely to increase methane
Rates of N retention and litter decomposition were functionally coupled. Created wetlands may require management
for multifunctionality, in order to sustain a specific ecosystem service.
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