Polish Journal of Ecology
Pol. J. Ecol. (2014) 62: 557–575
Regular research paper
Krystian OBOLEWSKI1,2*, Katarzyna GLIŃSKA-LEWCZUK3, Agnieszka STRZELCZAK4,
Paweł BURANDT3
Pomeranian University in Słupsk, Department of Ecology, 76-200 Słupsk, Arciszewskiego 22b,
Poland
2
Kazimierz Wielki University in Bydgoszcz, Faculty of Natural Science, 85-064 Bydgoszcz,
Chodkiewicza 30, Poland,*e-mail: [email protected] (corresponding author)
3
University of Warmia and Mazury in Olsztyn, Department of Land Reclamation and Management,
10-719 Olsztyn-Kortowo, Plac Łódzki 2, Poland
4
West Pomeranian University of Technology, Faculty of Food Sciences and Fisheries, 71-459 Szczecin,
Papieża Pawła VI 3, Poland
1
EFFECTS OF A FLOODPLAIN LAKE RESTORATION ON
MACROINVERTEBRATE ASSEMBLAGES – A CASE STUDY
OF THE LOWLAND RIVER (THE SŁUPIA RIVER, N POLAND)
ABSTRACT: A study on the response of
macroinvertebrate assemblages to the restoration
of hydrological connectivity of an oxbow lake
through the channel excavation is presented. The
study included a five-year environmental monitoring (hydrological, hydro-chemical measurements and invertebrate sampling) carried out in
the years 2008–2009 and 2011–2012 in the floodplain of the Słupia River (N Poland). The results
allowed for assessing ecological effects of the
hydrotechnical treatments (re-opening of an old
river bed, declogging and installation of wooden
deflectors) applied in the oxbow restoration. The
results confirmed the preliminary hypothesis that
the level of hydrological connectivity determines
the dynamics of invertebrate fauna communities in river-floodplain systems. Analysis of the
data revealed that such reconnection considerably influenced the structure of hydrobionts by
altering abiotic habitat conditions. Effects of
radical changes in the habitat morphology and
hydrodynamic conditions and monitoring of
macroinvertebrate assemblages preformed in the
restored lake indicated a significant instability of
the ecosystem soon after the treatment, what was
confirmed by the results of canonical analysis, in
which 50% of the total variance remained unexplained. Among factors analysed, water quality
parameters explained 21.4% of the total variance
in macroinvertebrate communities. However,
the hydrobionts showed a significant instability
with respect to variable hydrological conditions
(flow through the lake) what contributed to a low
share of water flow along the oxbow in explaining the total variance. The analysis of long-term
changes that occurred in the studied ecosystem
showed that the restoration of full hydrological
connectivity brought only a short-term increase
in benthofauna abundance. The most distinct
reorganization in the structure of macroinvertebrate communities was observed in the first year
after the reconnection while the consecutive four
years brought only insignificant changes, mainly
the appearance of a few species, mainly molluscs.
Our investigation suggests that the taxonomic
composition of macroinvertebrates in the studied
oxbow lake having the only one-arm opening,
before the improvement of hydrological connectivity, seemed to be more stable and optimal for
this kind of habitat, than after the hydrotechnical works, when the ecosystem became passable
due to both-arms connections to the river channel. Therefore, semi-lotic oxbows, connected to
parent rivers only with one arm, can be properly
functioning aquatic ecosystems in river floodplains.
KEY WORDS: oxbows, restoration, diversity, macroinvertebrates, river-floodplain system,
Słupia River
Krystian Obolewski et al.
558
1. INTRODUCTION
Decreasing area of wetlands, including
floodplains, is a global phenomenon, which
causes are mainly attributed to anthropogenic
impact (e.g. Mits ch and G oss elin k 2007,
Isl am 2010). The most destructive effects are
brought by hydrological changes occurring in
the result of damming upstream reaches of
rivers and river regulation. River-floodplain
systems (RFS) are particularly strongly affected by such treatments since they initiate
long-term changes such as terrestialization
and fragmentation of wetlands (Ward et al.
1999, Chormański et al. 2011).
Changes of land use in river catchments,
hydrotechnical constructions as well as draining ditch networks have seriously impaired
the functioning of RFS (e.g. Pett s et al. 1989,
Schönbrunner et al. 2012). In order to restore
their (semi-) natural conditions it is important to prepare rehabilitation programmes
(Gi l ler 2005). River hydrodynamics has
turned out to be a driving force significant for
self-sufficient RFS which also contributes to
higher biological diversity (To ck ner et al.
1999, Ga l l ardo et al. 2008, O b ole wsk i et
al. 2014, Wi l k-Woźni a k et al. 2014). Therefore, the improvement and strengthening of
hydrological connectivity between rivers and
floodplains are crucial for the restoration of
the appropriate comprehensive natural functions of those systems (e.g. Buijs e et al. 2005,
Gli ńska-L e wczu k and Bur andt 2011).
This can be achieved by the oxbow lake – river reconnection (O b ole wsk i and Glińsk a Lewczuk 2011). That raises hypothesis that
hydrological parameters determine the structure of invertebrates in RFS. Such an active
approach allows to enhance natural hydraulic
forces preventing oxbow lakes from shallowing and terrestialization and preserves their
role as the centres of biological diversity (“hot
spots”) in river floodplains. This is particularly important because well-retained floodplains mitigate the effects of rapid, damaging
floods occurring in the main river channel
(R e cke ndor fer et al. 2005).
Oxbow lakes can be connected to the
parent river with both arms (lotic water bodies), with one arm (semi-lotic) or completely
isolated (lentic) (Glińska-L e wczu k 2009).
The process of their formation is mostly in-
fluenced by geographical factors, i.e. geological structure, river hydrological regime,
geomorphometric parameters and soil characteristics as well as land use in the catchment
area (Gli ńska- L e wcz u k and Burandt
2011). Partial or total separation of meanders
from the main river channel causes their shallowing, overgrowing and increases the trophy
(Mits ch and G oss eli n k 2007, Gli ńskaL e wc z u k and Bu r andt 2011) which,
in turn, deteriorates habitat conditions for
invertebrate fauna (G a l l ard o et al. 2008,
Jurk i e w i cz -Kar nowska 2011, O b o le wsk i et al. 2013). That particularly concerns some of benthofauna representatives
very sensitive to unfavourable environmental
conditions. Therefore, qualitative and quantitative structure of benthic fauna can reflect
and indicate changes occurring in the aquatic environment (e.g. Z i mmer et al. 2000,
Whi les and G old ow itz 2005, Ga l l ard o
et al. 2008, Ju rk i e w i c z - Kar now sk a 2011,
O b ole wsk i et al. 2014). It has been reported
that the functioning of benthic communities
in floodplain water bodies depends on habitat
conditions regulated by hydrological regime
being a key factor responsible for nutrient
and energy cycling as well as regulating water
temperature (Bu rg me r et al. 2007) and aeration (O b ol e w sk i et al. 2009). Differences
in the connection of floodplain lakes with the
main channel determine the availability of
nutrients and the degree to which processes
such as primary productivity and decomposition are controlled by the system (Wang et
al. 2007). Authors (Ward et al. 1999) indicated the intensive development of the isolated water bodies, controlled mainly by autogenic processes, leading to eutrophication
and then to terrestrialization (Pa lme r et al.
2005, O b ole wsk i and Gli ńska- L e wcz u k
2013). Consequently, the lack of regeneration
by flow dynamics will lead to the disappearance of lentic ecosystems and other wetland
areas that contribute to the loss of biodiversity.
Effective restoration of oxbow lakes requires comprehensive analysis of relationships between hydrological connectivity
and floodplain ecosystems. To ck ne r et al.
(1999) and Ward et al. (1999) underline that
restoration should base on the diversification of water flow dynamics and graduation
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
of the connection between river channel and
floodplain water bodies. That can bring the
presence of varied habitats for organisms of
different habitat requirements.
A set of hydrotechnical works can be applied in the activation of oxbow lakes such as
dredging, excavation of ditches or planting.
In Europe the large-scale restoration works
have been performed on the floodplain lakes
of the Danube river (e.g. To ck ne r et al. 1999,
R e ckendor fer et al. 2005), the Rhine and
the Meuse rivers (Nien huis et al. 2002). In
Poland, the active protection of oxbow lakes
against shallowing has been implemented
in the Słupia River floodplain (O b ol e w sk i
et al. 2014). The abundance and diversity of
macroinvertebrates inhabiting oxbow lake reconnected to the Słupia with pipes have been
reported so far (O b ole wsk i and GlińskaL e wczu k 2011).
In spite of increasing interest in the restoration of floodplain ecosystems there is still
the lack of studies on the influence of restored
connectivity between river channels and
semi-lotic floodplain lakes on invertebrates
inhabiting those water bodies. Such knowledge is crucial for the assessment of restoration results (P y wel l et al. 2003).
To cover the need of such an investigation, the main goal of this study was to analyse the reaction of benthic macroinvertebrate
communities to changes triggered by the
restoration of surface connectivity between
a semi-lotic oxbow lake and its parent river.
The research aimed at answering the following questions: (I) To what extend changes in
the level of hydrological connectivity in RFS
determine the structure of invertebrate fauna
communities? (II) What is the short- and
long-term influence of restored connectivity
559
on the qualitative and quantitative structure
of macrozoobenthos? (III) What guidelines
for the future RFS restoration measures result
from this study?
2. STUDY AREA: FLOODPLAIN OF THE
SŁUPIA RIVER
The Słupia River is a lowland watercourse
which flows into the Baltic Sea. Its length
amounts to 138.6 km while the catchment
area covers 1620 km2. Hydrographic network
in the Słupia River catchment started to form
around 15 000–10 000 years BP soon after last
glaciation (late vistulian glaciation, pomeranian phase) (O b ole wsk i et al. 2009). The
width of the riverbed ranges from 7 m in the
upper part to 40 m in the lower part and the
average river gradient is 1.3‰. Water flow in
the investigated section of the Słupia River
ranged from 10 to 22 m3 s–1 and the average
flow was similar to the multiannual value and
amounted to 16 m3 s–1 (Hydrographic Map of
Poland 2010). However, large-scale regulation
works had been conducted since the middle of
XVIII century. Drainage was first performed
in 1890 and the scope of regulation works
included riverbed deepening accompanied
by the removal of boulders, trunks and sandbanks. Channelizations along the river section between the city of Słupsk and Krzynia in
the years of 1915–1919 and in 1922 substituted bends with straight cuts, river banks were
reinforced and numerous channels, weirs,
dams and water reservoirs were constructed.
As a result of the technical regulation, the
length of the Słupia channel was considerably reduced and 50 former river bends were
cut off and formed oxbow lakes (O b ole wsk i
and Glińska-L e wczu k 2011). Similarly,
Table 1. Morphological characteristic of the studied OLS3 oxbow lake before and after the restoration of
full hydrological connectivity with the Słupia River (SLO – semi-lotic, LO – lotic).
DisDistance from
VolDepth
Sinu- tance
the river
ume
GeographiArea osity
beLength Width
Downcal coordi- Time Type
A S=D/a- tween UpVav
D (m) S (m)
stream
stream Max. hav hmax
nates
(ha)
c
arms
(thou
arm
arm
(m)
(m)
(m)
(-)
a-c
m3)
(m)
(m)
(m)
BeN
SLO
143
10.5 0.14
1.1
131
124
0
118 0.6 1.30 0.82
54o23’25.62” fore
E
17o01’59.16” After LO
249
10.5 0.21
2.0
125
0
0
118 0.4 1.15* 0.84
* – depth in 2012
Krystian Obolewski et al.
560
Table 2. The range of restoration works performed within the semi-lotic oxbow lake (OLS3) in the
Słupia River floodplain. Denotations: Numbers related to Fig. 1
No
1
Location along the river
channel (km+m)
0+000 – 0+020
(km 46+030 Słupia River)
2
0+055 – 0+145
3
4
5
7
0+150
0+145 – 0+242
0+200
0+218 – 0+242
(km 45+640 Słupia River)
0+219
8
0+240
6
Range of works
Declogging and enabling water flow from oxbow to the Kwacza River, mechanical excavation
Fragment of natural bed with yellow water lily and reed field designated for
protection
Wooden deflector on the excavated right bank
Excavation (re-opening of old bed) leading water to oxbow lake
Wooden deflector on the excavated left bank
Double sided reinforcement with live vascines applied to canal and Słupia
River banks, stockade in the river
Wooden footbridge
Removal of willow thicket from the left bank of the river at the place of
oxbow-river connection
Fig. 1. Location of the studied oxbow lake (OLS3) in the Słupia River floodplain (N Poland) and the
distribution of sampling sites (Upstream – A, Middle zone – B, Downstream – C). 1–8 hydrotechnical
works performed within the restoration project in 2008.
688 ha of riverside meadows were drained.
The Słupia River is a watercourse with altered
hydrological activity by hydroelectric power
stations situated ca. 2 km above the studied
oxbow lake.
2.1. Oxbow lake restoration project
The restored oxbow lake is located at
46+030 km of the course of the Słupia River
within the “Słupia River Valley” Landscape
Park (Fig. 1). It is a shallow and small water
body which was connected to the river with
its downstream arm (semi-lotic oxbow lake)
before the restoration (Table 1). In 2008,
within the scope of the project, a ditch connecting the upstream arm with the Słupia was
excavated and the lower connection was widened. Additionally, a wooden deflector was
installed and willow thickets were removed
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
from the left bank of the river at the place of
oxbow-river connection. The range of works
performed is given in Table 2.
3. METHODS
3.1. Sampling
Water and invertebrate samples were collected in the years 2008–2009 and 2011–2012
every three months. The sampling was performed during the pre-restoration period
(from January to November 2008, n=12) and
the period of full hydrological connectivity
(from January to November 2009, n=12; from
April 2011 to November 2012, n=18). Three
sampling sites were selected: in the closed arm
(A), in the middle zone (B) and in the open
arm (C) (Fig. 1). Soon after the hydrotechnical
works completed in 2009, site A was moved to
the excavated upstream arm (~50m).
3.2. Hydrological measurements
Discharges of water at the same sampling
sites were measured simultaneously with the
time of sampling and with the same number
of replications. The methods depended on
the volume of outflowing water. In most cases
an electromagnetic velocity sensor FlowSensmodel 801 (Valeport, UK) was used and
the discharge was calculated using the standard velocity-area method. Low discharges
(Q<5.0 L s–1) were determined by the standard volumetric or velocity-area methods.
Additionally, Mini Diver data loggers (Van
Essen Instr.) were installed to monitor continuously the water stages. Stage-discharge
rating curves were plotted for 3 cross-sections to obtain a relationship between flow
and measured stage height. This relationship
allowed discharge to be estimated from registered measurements alone and allowed for
adjustment of error in individual discharge
measurements. For the need of the study we
narrow the hydrological analyses to the periods of observations April–November.
561
monium (NH4+-N), nitrate (NO3--N) and
nitrite nitrogen (NO2--N) were measured
in situ using calibrated multiparametric
probe YSI Professional Plus (YSI, USA) and
HQD40 (Hach-Lange, USA). Simultaneously, water samples of 2.5 L volume each were
taken from the depth of ca. 20 cm and placed
in polyethylene bottles. Within 24 h samples
were filtered through Whatman® GF/F glass
fibre filters (pre-combusted at 450oC for 4 h)
to determine the amount of dissolved solids
(TDS) (APHA, 1989). The concentrations of
cations and anions (Cl-, SO42-, K+, Ca2+, Mg2+
and Na+) were determined with ionic chromatography. The FLOWSYS-SYSTEA® analyser was used to indicate the concentrations of
total phosphorus (TP) and orthophosphates
(PO43--P) (APHA, 1989). After the incubation at 450°C for 4 h the quantity of mineral
matter (MM) was determined. Chemical oxygen demand (COD) and the concentration of
HCO3- were measured with the help of DR2800 spectrophotometer (Hach-Lange, USA)
and the cuvette method.
3.4. Biological analyses
Invertebrates were sampled with the Ekman’s grab sampler (225 cm2 surface) three
times at each zone of the oxbow lakes. Those
sites differed by the thickness of bottom sediments and their composition (e.g. the presence of leaves, branches, and submerged
plants). The sediments were sieved through a
300 mm mesh size sieve, placed in containers
and fixed in 5% formalin. In the laboratory
benthic invertebrates were identified to the
possibly lowest systematic level, except for
Oligochaeta (class). Results of that identification were considered in two ways: separately
and altogether. Additionally, the following
zoocenotic indices were used: Shannon diversity (H’), Pielou’s evenness (J’), domination
(D), number of taxa and the total abundance.
The diversity indices (H’ and J’) were calculated with the help of Past v.2.17c software
(Hammer et al. 2001).
3.3. Physico-chemical analyses of water
3.5. Statistical analyses
Water temperature, specific electrical
conductivity (SEC) and pH as well as the
concentration of dissolved oxygen (DO), am-
To assess the statistical differences among
years for macroinvertebrate data (grouped
into classes/orders) and hydro-chemical
562
Krystian Obolewski et al.
data, the non-parametric analysis of variance (Kruskal-Wallis and Dunn’s tests, P
≤0.05) were performed. Canonical Correspondence Analysis (CCA) was applied to
obtain a synthesized profile of macrozoobenthos taxa against the background of environmental conditions (time since opening)
in the studied oxbow lake and to evaluate
similarities in composition of benthic fauna
between pre- and post- treatment phases. It
is a highly useful ordination method which
supports the multivariate analysis of variance
(MANOVA) among selected environmental factors, water quality data and macroinvertebrate assemblages (e.g. Ga l l ard o et al.
2008, O b ol e w sk i 2011). The method also
facilitated the interpretation of complex correlations and a synthetic presentation of the
obtained results.
CCA was used between 19 environmental variables (hydrological and water quality parameters, time since the reconnection)
and 14 groups of macroinvertebrates with the
use of the CANOCO 4.5 software package
Fig. 2. (A) Water stages and discharges in the Słupia River (N Poland). In the hydrograph are presented
average stages and discharges from the period April–November calculated for years 2007–2012 based
on measuremts at the water gauge in Słupsk. (B) Box (first and third quartiles) and whiskers (maximum,
minimum, median) plots of discharges recorded for upstream, downstream arms and middle zone of
the studied oxbow lake during summer half-years over the study period.
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
563
Table 3. Changes in physico-chemical parameters of water (mean ± SD) in the studied oxbow lake before and after the restoration of full hydrological connectivity with the Słupia River. Denotations: SEC
= Specific Electrical Conductivity, DO = Dissolved Oxygen, COD = Chemical Oxygen Demand, TP =
Total Phosphorus, TDS = Total Dissolved Solids, MM = Mineral Matter.
Significant differences are marked with bold (nonparametric Kruskal-Wallis Test, *P <0.05, ** P <0.01,
*** P <0.001); a–d = significant differences between consecutive years (nonparametric Dunn’s Test, *P
<0.05, ** P <0.01, *** P <0.001); a – 2008, b – 2009, c – 2011, d – 2012.
Parameter
T
Before
re-opening
2008
(n = 12)
average
± SD
a
10.90
5.49
unit
C
o
pH**
8.05
0.34
SEC
μS cm
339
DO
mg L
**
-1
*
-1
After re-opening
2009
(n = 12)
average
± SD
b
7.21
11.00
2011
(n = 9)
average
± SD
c
12.45
4.86
7.44c**
7.72
8.38b**
363
260
0.25
8.16
0.46
27
304
82
13
444
7.62
b*
4.66
10.54
10.69
8.29
1.12
7.57
3.44
72.3b*
48.5
87.7a,d*
95.7
77.1
4.6
68.1b*
29.9
15.9
4.3
3.3
15.8
c**
a*
b**
DO%*
%
COD
mg L
NO2--N**
mg L-1
0.0088c**
0.0042
0.0100
0.0112
0.0203a**
0.0062
mg L-1
0.368
0.178
0.317c**
0.462
0.157b**
NH4+-N*
mg L-1
0.113b*
0.050
0.015a*
0.060
0.068
PO43--P
***
mg L
0.140
0.038
0.132
0.135
TP
mg L
0.351
0.575
TDS***
mg L-1
234d**
27
245d***
MM
-1
mg L
166
10
Mg2+
mg L-1
7.4
1.7
Ca2+
mg L-1
48.8
Na+
mg L-1
10.1
K
-1
mg L
Cl-
mg L-1
***
NO3--N**
+
HCO3
-*
-1
-1
-1
mg L
-1
2012
(n = 9)
average
± SD
d
12.46
5.60
b**
9.4
a**,c***,d*
8.4
18.8
b***
b*
4.5
0.0227
0.0225
0.053
0.307
0.147
0.014
0.093
0.060
0.162
0.064
0.192
0.067
0.514
0.385
0.152
232
207
158
163
5.0
6.2
4.8
41.1
1.5
7.8
2.0
0.7
19
17
14.5
0.832
c*,d***
119.8
d**
a*
0.339
a***
0.082
19
188a**,b***
169
19
162
13
5.0
1.7
5.4
1.7
49.6
51.2
1.3
57.0
10.2
9.3
9.0
0.5
23.4
26.7
2.3
2.1
1.7
0.2
2.5
0.9
12
15
12
3
30
113.3
126.0
126.2
7.8
(Ter Braa k and Šmi l auer 2002). In order
to identify a minimum subset of variables that
significantly explain variation in the chemical data, redundant variables were removed
through a form of step-wise regression (forward selection) together with Monte Carlo
permutation tests. Because environmental
gradients had not previously been evaluated
in the study area, we ran a manual, forwardselection procedure, which included variables
that had a conditional effect significant at the
5% level (P ≤0.05). Using partial canonical
correspondence analysis (pCCA) the variance
was partitioned into three variable groups (hydrological, trophic and physico-chemical water quality parameters), variance shared and
17
39
173.6
a**
92.4
unexplained. The data have been transformed
to logarithms log (x+1) and centered by taxa.
4. RESULTS
4.1. Hydrological situation
The key factor direcly influencing the habitat conditions of the studied oxbow is the Słupia
river. The section of the Słupia river within
the area under this study is characterized by a
relatively low amplitude of water levels over a
year (in the studied period it was only 1.1 m)
but very frequent fluctuations (Fig. 2A). No extreme hydrological phenomena, like low flows
or high floods, occurred during the research pe-
564
Krystian Obolewski et al.
riod and the studied oxbow lake was surficially
recharged with the river water through its arms.
Hydrological situation in the studied
oxbow lake has changed distinctly since the
restoration works (Fig. 2B). Hydrotechnical
treatments applied in November 2008 enabled free water flow through the excavated
upstream arm and contributed to the higher
water exchange rate in the reservoir.
Since the end of hydrotechnical works,
significant spatial and temporal changes in
hydrological pattern of discharges were observed. Based on the flow measurements
conducted in the 3 crossections (A, B, C)
permanent but varied flow rates have been
observed along the oxbow (Fig. 2B). Water
movement through the oxbow was maintained and observed at each cross-section
even at low water stages. Reconnection of the
oxbow with the Słupia River triggered significantly higher water flow rates in the newly
formed upstream arm (Kruskal-Wallis test,
P = 0.015) and in the middle zone (KruskalWallis test, P=0.032), particularly in the first
year after reconnection (Fig. 2B). No significant differences in water flows were observed
in the downstream arm (Kruskal-Wallis test,
P = 0.160). The activity of the water flow also
differed among seasons. The most active
water flows took place during spring thawings (out of the sampling season), when the
discharges at the inflow achieved 1.89 m3 s–1
(not shown in the figs.). In the period of biological observations the highest water flows
appeared in July 2009 and 2010 (0.67 m3 s–1
and 0.58 m3 s–1, respectively) as an effect of
heavy rainfalls and water released from hydropowers located upstream the study area.
However, this “washing-out” effect was shortterm – the water flow considerably decreased
in 2012. Significant differences in the flow
rates were only observed between years 2008
and 2009 (Dunn’s test, P <0.001) as well as between 2008 and 2011 (Dunn’s test, P <0.05).
typical of hypertrophic habitats (according to
the Nürnberg’s classification (2001)). Water
reaction and specific electrical conductivity
(SEC) differed significantly between years
2009 and 2011 (Kruskal-Wallis test, P = 0.007,
Table 3). Water aeration changed distinctly
only in the first year after the reconnection
(Kruskal-Wallis test, P <0.05 for DO and
DO%, P <0.01 for COD). The concentrations
of mineral nitrogen forms varied between
consecutive years. After the reconnection,
contents of nitrate and ammonium nitrogen
decreased while the concentration of nitrites
gradually increased and a significant difference was observed between years 2009 and
2011 (Dunn’s test, P <0.01). Increased water
flow might have contributed to the decrease
in the concentration of total phosphorus and
TDS (2.5-fold and 1.2-fold, respectively). The
values of TDS differed significantly between
the pre-restoration period and the first year
after the reconnection (Dunn’s test, P <0.01)
as well as between the first year and the last
year since restoration (Dunn’s test, P <0.001).
Among the remaining parameters, the concentration of bicarbonates was the only one
4.2. Environmental conditions
Restoration of surficial connectivity of the
studied RFS significantly influenced physicochemical parameters of oxbow lake water
(Table 3). In general, significant differences
were observed between 2009 and the other
years. The water quality parameters were
Fig. 3. Results of Canonical Correspondence
Analysis (biplot) performed with hydrochemical
parameters and sampling sites located in OLS3
oxbow lake using forward selection of variables (P
≤ 0.05). No of sampling sites: 1–12 – 2008 year;
13–24 – 2009 year; 25–30 – 2011 year; 31–39 –
2012 year. Explanation of abbreviations in Table 3.
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
that increased significantly as a result of restoration (Dunn’s test, P <0.001).
565
4.4. Patterns of macroinvertebrate abundance
Restored
hydrological
connectivity
through the opening of oxbow lake did not
influenced drastically macrozoobenthos in
a long term perspective (Table 4). Soon after the reconnection the average density of
macroinvertebrate assemblages increased
almost twofold but insignificantly (KruskalWallis test, P = 0.099) but then decreased.
The number of taxa fluctuated among the
years but no statistically significant effect was
noted (Kruskal-Wallis test, P = 0.077). Similar situation was observed for the Shannon’s
and Pielou’s indices (Kruskal-Wallis test,
P = 0.735 and P = 0.086, respectively).
4.3. Patterns of physico-chemical
parameters of water
CCA model for physico-chemical parameters of water and sampling sites (Fig. 3) explained a considerable part of total variance
– the first axis 30% and the second axis 19%.
Water quality varied over the study period but
seasonal conditions in the arms and the centre
of the floodplain lake were similar among years.
Considerable changes have been observed until
2012 and concerned main cations (Na+, Ca2+)
and ions (Cl-, NO3-) as well as COD.
Table 4. Benthic invertebrate composition (mean ± SD) in the studied oxbow lake before and after the
restoration of full hydrological connectivity with the Słupia River.
Significant differences are marked with bold (nonparametric Kruskal-Wallis Test, *P <0.05, ** P <0.01,
*** P <0.001); a–d = significant differences between consecutive years (nonparametric Dunn’s Test,
*P <0.05, ** P <0.01, *** P <0.001); a – 2008, b – 2009, c – 2011, d – 2012; + - total abundance and difference among groups is considered.
Parameter
Abundance
+
Before
re-opening
2008
(n = 12)
average
± SD
2009
(n = 12)
average
± SD
2011
(n = 9)
average
± SD
2012
(n = 9)
average
± SD
a
b
c
d
938.5
No of taxa
1515.7
After re-opening
3714.2
20
3985.5
339.2
28
324.3
805.3
11
1033.1
22
Shannon index H‘
1.124
0.478
0.943
0.748
0.842
0.704
1.209
0.394
Pielou‘s evenness
index J’
0.695
0.265
0.398
0.299
0.611
0.374
0.693
0.235
Oligochaeta
288.9
569.3
1007.8
2245.3
42.0
66.5
452.1
1085.9
Hirudinea
94.7
97.8
122.4
145.0
83.9
120.2
132.8
186.8
Malacostraca
383.7
680.9
2025.6
3361.4
4.9
12.1
25.8
27.8
Megaloptera
30.9
47.5
57.2
103.4
9.9
17.9
1.6
4.9
Odonata
0.0
0.0
3.7
9.2
0.0
0.0
1.6
4.9
Ephemeroptera
3.7
9.2
117.5
155.5
0.0
0.0
11.0
29.4
Plecoptera
0.0
0.0
0.4
1.4
0.0
0.0
34.6
103.7
Trichoptera
10.4
25.4
108.2
0.0
0.0
14.8
39.2
Diptera
101.5
192.9
51.7
66.0
58.3
135.7
4.9
10.5
Coleoptera
8.7
21.4
6.6
21.3
0.0
0.0
0.0
0.0
Hemiptera
3.7
9.2
12.3
20.5
0.0
0.0
0.0
0.0
0.0
0.0
0.4
0.0
0.0
66.7
75.7
*
Arachnida
Gastropoda
Bivalvia
**
b*
0.0
b*,d**
12.3
71.6
a,c,d*
b*
1.4
0.0
0.0
0.0
a*
97.3
242.5
89.8
141.9
25.9
139.6
419.2
50.4
90.1
b*
a**
59.3
70.5
566
Krystian Obolewski et al.
Fig. 4. Density and percentage contribution of macrozoobenthos taxa in consecutive zones of OLS3
oxbow lake before (A–B) and after (C–D) restoration of full hydrological connectivity with the river
channel.
Altogether 4 250 macroinvertebrate representatives were identified over the study period which belonged to 22 species, 19 genera
and 4 higher taxonomic units (see: Appendix). Crustaceans constituted the most common group (40–42%) both before and after
the restoration works and were accompanied
mainly by Oligochaeta but the differences
in their densities between years were insignificant (Kruskal-Wallis test, P = 0.069 and
P = 0.154, respectively). Only Trichoptera
and Gastropoda were significantly affected
by the reconnection (Kruskal-Wallis test,
P <0.05 and P <0.01, respectively).
The abundance and composition of macroinvertebrates varied between the arms and
the central zone of the lake (see: Appendix).
Before the reconnection the highest density
was observed in the middle part of the water
body (Fig. 4A). Each zone was predominated
by Oligochaeta and Crustacea. The share of
Oligochaeta decreased and more taxonomic
groups appeared (Fig. 4B), particularly in
the arms. After the reconnection, the highest benthofauna density was observed in the
middle zone and was significantly different
from other sites (Fig. 4C), (Kruskal-Wallis
test, P = 0.032). The river flow in the newly-formed upstream arm induced the 4-fold
increase in macroinvertebrate density (Kruskal-Wallis test, P = 0.015) while in the downstream arm the changes were insignificant
(Kruskal-Wallis test, P = 0.160).
Surface connectivity strongly interrelated
with the density of molluscs, i.e. Pisidium
amnicum Müller, Lymnaea sp. and Viviparus contectus Millet (see: Appendix), share
of which increased mainly in both arms
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
(Fig. 4D). In turn, the share of Diptera larvae decreased, which was mainly observed
in the lower arm (Fig. 4D). Sergentia sp. and
Procladius sp. (Diptera, Chironomidae), numerous in the semi-lotic oxbow lake, after reconnection only sporadically occurred in the
lotic floodplain lake or was not observed at all
(see: APPENDIX).
567
4.5. Primary gradients affecting
aquatic community structure
CCA analysis revealed the relationships
between the structure of macroinvertebrate
assemblages and environmental conditions
(Fig. 5A). Among 20 explanatory parameters,
the forward selection indicated 19 variables
Fig. 5. Results of Canonical Correspondence Analysis performed with invertebrate and environmental
data from the reconnected river-floodplain lake system using forward selection of variables (P ≤0.05).
(A) Biplot of significant environmental variables and invertebrate data of axis 1 and 2; (B) Biplot showing significant invertebrate data and time scores; (C) Biplot showing significant invertebrate data and
heterogenic habitats; (D) Biplot shows the share of significant invertebrate communities in seasons; (E)
Biplot showing significant invertebrate data and intensity of flow.
Krystian Obolewski et al.
568
which significantly contributed to the model performance (Table 3). Dissolved oxygen
(DO%) and HCO3- were excluded from the
data set since they were redundant. The final
model explained 49% of the total variance
of benthic fauna communities and both two
canonical axes were statistically significant
(Monte Carlo test, P = 0.002). The first axis
revealed the trend from high eutrophication to high concentrations of ions (Fig. 5A).
Variance partitioning indicated, that hydrological parameters (discharge and time since
re-opening) explained only 0.7% of the total
variance while physico-chemical variables
(temperature, SEC, pH, TDS, DO, COD, K+,
Na+, Cl-, MM, Ca2+ and Mg2+) – 18.3%. Parameters related to the level of trophy (NO3-N, NO2--N, NH4+-N, PO43--P, TP) contributed with 3.1% of variance shared among
those three explanatory subsets amounted to
27 and 51% of the total variability remained
unexplained.
CCA analysis generally indicated moderate changes in the structure of macroinvertebrates triggered by altered hydrological
conditions after restoration works (Fig. 5A).
Distinct differences in the structure of benthofauna were observed only in the first year
after reconnection (Fig. 5B). However, the
insight into pies classes diagrams, focused on
consecutive years revealed much more information (Fig. 5B). During the first year after
the reconnection of RFS the share of most of
the invertebrate groups increased and preferred the middle zone and the downstream
oxbow arm (Fig. 5C). Summer seems to be
the most favourable season for macroinvertebrates since they reached high abundance
(Fig. 5D). Water flow influenced the structure
of macrofauna in a diverse way – most of the
invertebrates preferred a moderate water flow
rate and only Plecoptera occurred abundantly
at higher discharges (Fig. 5E).
5. DISCUSSION
Floodplain lakes form a different type
of aquatic biotopes comparing to rivers they
origin from. They constitute habitats for biocenoses transitional between stagnant and
flowing waters. Most of a year they are at least
partly cut off from their rivers and except
for the duration of high floods, the exchange
of water in oxbow lakes is highly limited
(Mit s ch and G o ss el i n k 2007).
The assemblages of aquatic invertebrates
in the studied floodplain lake included insects, worms, molluscs and crustaceans,
what is in accordance with the results of
studies performed on other European rivers
(Whi les and G old ow itz 2005, Ga l l ard o
et al. 2008, Ja kubi k 2012, O b ol e w sk i et
al. 2014). In partly isolated oxbow lakes, the
relatively long periods of stabilization and
low water exchange in their central zones are
favourable to the taxa with longer life cycles
but less effective colonization strategies, e.g.
isopods (G as it h and R esh 1999). Similarly,
insects with short life cycles like Chironomids reach the highest abundance in the same
parts of semi-lotic oxbow lakes (Fig. 4). They
colonize the water body mostly from the air
(Ma l lor y et al. 1994). After the restoration,
the observed high abundance of benthic invertebrates in the middle part of the oxbow
lake spread around the whole water body
(see: APPENDIX).
Preliminary we hypothesed that, similarly to water bodies of natural origin, a level of
hydrological connectivity determines the dynamics of invertebrate fauna communities in
river-floodplain systems. The results showed
that hydrotechnical works performed in the
studied oxbow lake changed the morphology
of its bed which probably limited the sedimentation of mineral and organic matter carried by river water as well as the development
of macrophytes. Such a trend is unfavourable
for benthofauna diversity. The following decrease in the flow rate of river water intensified the deposition of small-grained material.
That resulted in lake shallowing, increased
development of vegetation, higher evaporation in summer what caused the oxbow lake
discharge dropped to 0.1–0.3 m3 s–1 (Fig. 2).
Hydrological measurements indicated
gradual decline in water inflow through the
upstream arm and the formation of a levee
built mainly of mineral fraction. That implies
the limited occurrence of backwater and the
presence of groundwater recharge which was
partly identified by the changes in water temperature and conductivity (Table 3).
This study shows that the restored hydrological connectivity in RFS is not always
the only factor for the development of ben-
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
thofauna assemblages (hydrological variables
explained less than 1% of the total variance;
Fig. 5A). There are also biotic drivers of
macrozoobenthos variability. Complex and
strong biological relationships within macrofauna communities seem to be the reason of
such situation (Heino 2000, Gr if f it h et al.
2001). Other studies indicate that the qualitative and quantitative structure of benthofauna communities is mainly influenced by such
environmental parameters as nutrient availability and temperature (e.g. Gr i f f it h et al.
2001, Mur phy and D av y-B ow ke r 2005).
This confirms our results since the most favourable conditions for macrozoobenthos
development were stated in summer (Fig. 5D)
and water temperature strongly correlates
with other environmental variables (Z i mmer et al. 2000). Moreover, if we assume that
hydrological connectivity determines environmental variables, then the influence of hydrological parameters on invertebrate assemblages is indirect (Peres-Neto et al. 2006).
Most of reports on biodiversity indicate
that the abundance of invertebrates is the
highest in semi-open floodplain lakes (To ckne r et al. 1999, Amoros and B or ne tte
2002, Whi les and G oldow itz 2005). Considerable diversity of habitats in such water
bodies (favourable to a wide range of hydrobionts) as well as better environmental conditions (improved by hydrological connectivity)
seem to be the main reasons of such situation
(Glińska-L e wczu k 2009). That was also
confirmed by our study since a relatively high
invertebrate diversity was observed in the
semi-lotic oxbow lake and the restoration of
full hydrological connectivity did not bring
a longlasting increase in the benthofauna diversity and abundance (Table 4). However,
the shift from environmental conditions in
the lotic floodplain lake enabled the appearance of organisms sensitive to water quality
such as mayflies (Caenis macrura Stephens,
Heptagenia sp.), stoneflies (Leuctra sp., Chloroperla sp.) and molluscs (see: APPENDIX).
Restored hydrological connectivity in RFS
favoured the colonization of floodplain lakes
particularly by that last group, probably due
to the increase in Ca2+ concentrations in water (r = 0.765, P <0.05).
It has been proved that physico-chemical
and hydrological factors influence directly or
569
indirectly hydrobionts’ communities (To ck ner et al. 1999, Je pp e s e n et al. 2003, Wang
et al. 2007). There is common conclusion that
concentrations of dissolved oxygen and calcium ions are by some authors considered as
crucial for macrozoobenthos abundance and
diversity (e.g. Ga l l ard o et al. 2008, O b ole wsk i et al. 2009). Such parameters as the
contents of magnesium, orthophosphates,
total phosphorus, nitrates, chlorides, sodium and organic substances as well as pH
and conductivity we proved to have meaningful significances as environmental factors, regardless on their origin (e.g. Hei no
2000, Z i m me r et al. 2000). Other factors
that could explain the equivocation of our
study results are: morphometry (He i no
2000, Je pp e s e n et al. 2003), the presence
of macrophytes (P y wel l et al. 2003), helophytes (Gr i f f it h et al. 2001, Mu r phy and
D av y - B ow ke r 2005), trophic interactions
(Jepp es en et al. 2003, G a l l ard o et al. 2008,
Jurk i e w i cz - Kar nowska 2011) and pollution (Wo o d c o ck and Hu r y n 2007).
6. CONCLUSIONS
Five-year investigation of changes occurring in benthic communities inhabiting the
oxbow lake due to hydrological connectivity
gained with the Słupia River revealed that
such reconnection considerably influenced
the structure of hydrobionts by altering abiotic habitat conditions. Changes in water quality resulted from the restoration works should
be considered as factors that more strongly
influence invertebrate assemblages than hydrological parameters. Changes in hydrodynamic conditions disturbed biogeochemical
processes which drastically affected all organisms in the oxbow lake. Increased values
of all the biocenotic indices in the first year
after reconnection indicate the importance
of hydrological connectivity in RFSs. Results
of this study show that habitat conditions of
oxbow lakes connected with their rivers only
by one arm can be optimal for high biodiversities and should be carefully managed in the
river floodplains.
ACKNOWLEDGEMENTS: We would like
to thank Szymon Kobus for their help in morphometric and hydrochemical studies (Dept. of
Krystian Obolewski et al.
570
Land Reclamation and Environmental Management, University of Warmia and Mazury) as well
as Łukasz Maksymowski and Natalia Jarząb for
their help in collecting and preparation of biological material (Dept. of Ecology, Pomeranian
University in Słupsk). This study was supported
financially by the National Science Centre, grant
no. NN305 143240. Reprint was financed by Polish National Fund for Environmental Protection
and Water Management.
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572
APPENDIX. Mean density (ind. m–2), constancy of occurrence (C, %), dominance index (D, %) and
index of ecological significance (Q) of benthofauna inhabiting the studied oxbow lake before and after
the restoration.
Taxa/parameter
Oligochaeta
Hirudinea
Glossiphonia
complanata L.
Hirudo medicinalis L.
Erpobdella octoculata L.
Helobdella stagnalis L.
Piscicola geometra L.
Malacostraca
Asellus aquaticus L.
Gammarus fossarum Koch.
Insecta
Before re-opening
(n = 12)
Mean density
Mean C D Q Up- Mid- Downdle
stream
stream
zone
340.7 44.4 32.4 37.9 276.5 646.9
98.8
26.3 44.4 2.5 10.5
After re-opening
(n = 30)
Mean density
Mean C D Q Up- Mid- Downdle
stream
stream
zone
533.3 27.8 70.0 44.1 271.3 888.7 320.0
9.9
14.8
54.3
17.1
0.9 50.0 6.7
16.0
28.1
19.3
0
0
0
0.2
0.0 3.3 0.2
0.5
0.0
0.0
28.0 55.6 2.7 12.2
0.0
34.6
49.4
91.5
4.8 76.7 19.1 49.0
125.9
81.5
24.7 44.4 2.3 10.2
4.9
34.6
34.6
6.1
0.3 10.0 1.8
7.0
13.3
10.4
0.0
0.0
0.0
0.0
0.4
0.0 6.7 0.4
1.0
1.5
0.0
0.0
0.0 0.0 0.0
0.0 0.0 0.0
436.2 55.6 41.5 48.0 197.5 869.1
4.9
242.0
865.5 45.2 53.3 49.1 455.0 939.3 1005.9
22.2 0.5 3.2
9.9
4.9
0.0
44.5
2.3 23.3 7.4
50.5
62.2
7.4
19.8 22.2 1.9 6.5
0.0
39.5
19.8
28.2
1.5 40.0 7.7
3.5
63.7
28.1
Lestes sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
1.1
0.1 6.7 0.6
0.0
1.5
1.5
Ischnura sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
1.1
0.1 3.3 0.4
0.0
0.0
3.0
1.6
11.1 0.2 1.3
0.0
4.9
0.0
23.2
1.2 13.3 4.0
0.5
0.0
62.2
3.3
11.1 0.3 1.9
0.0
9.9
0.0
3.8
0.2 10.0 1.4
1.5
8.9
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
6.6
0.3 3.3 1.1
0.0
0.0
17.8
0.0
0.0 0.0 0.0
0.0
0.0
0.0
22.2
1.2 10.0 3.4
24.5
13.3
22.2
Leuctra sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
11.5
0.6 3.3 1.4
0.0
31.1
0.0
Chloroperla sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.2
0.0 3.3 0.2
0.5
0.0
0.0
Limnephilus sp.
4.9
11.1 0.5 2.3
0.0
14.8
0.0
29.4
1.5 23.3 6.0
43.0
32.6
11.8
Phryganea sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
4.4
0.2 3.3 0.9
0.0
11.9
0.0
Psychomyidae
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.5
0.0 3.3 0.3
0.0
0.0
1.5
Cyrnus sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
2.2
0.1 6.7 0.9
0.0
4.4
1.5
Tabanus sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
3.5
0.2 13.3 1.6
3.5
0.0
5.9
Procladius sp.
32.9 66.7 3.1 14.4 19.8
69.1
9.9
0.5
0.1 6.7 0.4
0.0
4.4
0.0
Megaloptera
Sialis lutaria L.
Odonata
Ephemeroptera
Cloeon sp.
Ephemerella
ignite (Poda)
Caenis macrura
Stephens
Heptagenia sp.
Plecoptera
Trichoptera
Diptera
Macroinvertebrates vs. RFS (river floodplain system) hydrological connectivity
Taxa/parameter
Chironomus sp.
Chironomidae
pupa
Chaoborus sp.
Before re-opening
(n = 12)
Mean density
Mean C D Q Up- Mid- Downdle
stream
stream
zone
46.1 44.4 4.4 14.0 19.8 108.6
9.9
573
After re-opening
(n = 30)
Mean density
Mean C D Q Up- Mid- Downdle
stream
stream
zone
33.6 1.8 33.3 7.6 47.5 28.1 15.0
1.6
11.1 0.2 1.3
4.9
0.0
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
1.6
11.1 0.2 1.3
4.9
0.0
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
Sergentia sp.
49.4 33.3 4.7 12.5
4.9
118.5
24.7
0.0
0.0 0.0 0.0
0.0
0.0
0.0
Coleoptera
Hyphydrus ovatus L.
Platambus sp.
8.2
11.1 0.8 2.9
0.0
24.7
0.0
2.7
0.1 3.3 0.7
0.0
7.4
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.2
0.0 6.7 0.3
2.0
0.0
0.0
Dytiscidae larvae
1.6
11.1 0.2 1.3
0.0
0.0
4.9
0.0
0.0 0.0 0.0
0.0
0.0
0.0
Hydrophilus sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.2
0.0 3.3 0.2
0.5
0.0
0.0
Hemiptera
Hesperocorixa sp.
0.0
0.0 0.0 0.0
0.0
0.0
0.0
2.2
0.1 10.0 1.1
1.0
4.4
0.4
Corixa sp.
Notonecta glauca
L.
Nepa cinera L.
1.6
11.1 0.2 1.3
0.0
4.9
0.0
1.6
0.1 6.7 0.8
0.0
1.5
4.4
0.0
0.0 0.0 0.0
0.0
0.0
0.0
1.6
0.1 3.3 0.5
0.0
0.0
4.4
1.6
11.1 0.2 1.3
0.0
4.9
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.2
0.0 3.3 0.2
0.5
0.0
0.0
0.0
0.0 0.0 0.0
0.0
0.0
0.0
74.6
3.9 26.7 10.2
7.4
44.4
149.6
16.5 44.4 1.6 8.3
9.9
34.6
4.9
10.7
0.6 20.0 3.3
18.5
3.0
7.4
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.5
0.0 3.3 0.3
0.0
1.5
0.0
0.0 0.0 0.0
0.0
0.0
0.0
7.3
0.4 16.7 2.5
2.0
4.4
13.3
0.0 0.0 0.0
0.0
0.0
0.0
48.6
2.5 23.3 7.7
36.4
8.9
85.9
0.0 0.0 0.0
0.0
0.0
0.0
3.8
0.2 3.3 0.8
0.0
10.4
0.0
0.0 0.0 0.0
0.0
0.0
0.0
0.5
0.0 3.3 0.3
0.0
1.5
0.0
0.0 0.0 0.0
0.0
0.0
0.0
29.3
1.6 36.7 7.6
9.0
35.6
34.6
0.0 0.0 0.0
0.0
0.0
0.0
0.7
0.0 6.7 0.5
1.0
0.0
1.0
553.1
1915.5
11
40
1.420 1.898
1.898
3.433
2.886 2.125
2.403
0.554 0.597
0.701
0.649
0.689 0.506
0.576
Arachnida
Argyroneta
aquatic Clerck
Bivalvia
Pisidium amnicum Müller
Sphaerium corneum L.
Unio pictorum L.
Gastropoda
Valvata piscinalis
0.0
Müller
Lymnaea sp.
0.0
Bithynia tentacu0.0
lata L.
Planorbarius
0.0
corneus L.
Viviparus contec0.0
tus Millet
Viviparus vivipa0.0
rus L.
Total
1051.8
No of taxa
20
Shannon index
2.708
H‘
Pielou‘s evenness
0.685
index J’
563.0 2039.5
11
16
1043.0 2382.0 1916.1
26
28
26