A newly developed dispersal capacity metric indicates succession of benthic invertebrates in restored rivers

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DOI: 10.7287/PEERJ.PREPRINTS.911V1
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Abstract
Dispersal capacity plays a fundamental role in riverine benthic invertebrates? colonization of new habitats that emerge following flash floods or restoration. However, an appropriate measure of dispersal capacity for benthic invertebrates is still lacking. Dispersal of benthic invertebrates occurs mainly during the aquatic (larvae) and aerial (adult) life stages, and each can be further subdivided into active and passive modes. Based on these totally four dispersal modes, we firstly developed a simple metric to estimate dispersal capacity for 528 benthic invertebrate taxa by incorporating weight for each mode. Secondly we tested this metric using benthic invertebrate community data from a) 23 restored river sites all involving an improvement of river bottom habitats dating back 1 to 10 years, b) 23 unrestored sites, and c) 298 adjacent surrounding sites in the low mountain and lowland areas of Germany. We hypothesize that our metric will reflect the temporal succession process of benthic invertebrate communities colonizing the restored sites, while no temporal changes are expected in the unrestored and surrounding sites. By applying our metric to these three river treatment categories, we found that the average dispersal capacity of benthic invertebrate communities in the restored sites decreased significantly within the early years following restoration, while there were no changes in both unrestored and surrounding sites. After dividing all taxa into quartiles representing weak to strong dispersers, this pattern became even more obvious; strong dispersers colonized the restored sites during the first year after restoration and then significantly decreased over time, while weak dispersers continuously increased. The successful application of our metric to river restoration might be promising to further apply this metric for example in river assessments or meta-community structure.
1
A newly developed dispersal capacity metric indicates succession of benthic invertebrates in restored
rivers
Fengqing Li1, Andrea Sundermann, Stefan Stoll, Peter Haase
Senckenberg Research Institute and Natural History Museum Frankfurt, Department of River Ecology and
Conservation, Gelnhausen, Germany and Biodiversity and Climate Research Centre (BiK-F), Frankfurt am
Main, Germany
Running head: Dispersal influences community succession
1Corresponding author: Fengqing Li
Postal address: Senckenberg Research Institute, Department of River Ecology and Conservation,
Clamecystrasse 12, Gelnhausen 63571, Germany
E‒mail: Fengqing.Li@senckenberg.de
Tel: +49 (0)6051 61954‒3126
Fax: +49 (0)6051 61954‒3118
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Abstract
Dispersal capacity plays a fundamental role in riverine benthic invertebrates’ colonization of new habitats that
emerge following flash floods or restoration. However, an appropriate measure of dispersal capacity for benthic
invertebrates is still lacking. Dispersal of benthic invertebrates occurs mainly during the aquatic (larvae) and
aerial (adult) life stages, and each can be further subdivided into active and passive modes. Based on these
totally four dispersal modes, we firstly developed a simple metric to estimate dispersal capacity for 528 benthic
invertebrate taxa by incorporating weight for each mode. Secondly we tested this metric using benthic
invertebrate community data from a) 23 restored river sites all involving an improvement of river bottom
habitats dating back 1 to 10 years, b) 23 unrestored sites, and c) 298 adjacent surrounding sites in the low
mountain and lowland areas of Germany. We hypothesize that our metric will reflect the temporal succession
process of benthic invertebrate communities colonizing the restored sites, while no temporal changes are
expected in the unrestored and surrounding sites. By applying our metric to these three river treatment
categories, we found that the average dispersal capacity of benthic invertebrate communities in the restored sites
decreased significantly within the early years following restoration, while there were no changes in both
unrestored and surrounding sites. After dividing all taxa into quartiles representing weak to strong dispersers,
this pattern became even more obvious; strong dispersers colonized the restored sites during the first year after
restoration and then significantly decreased over time, while weak dispersers continuously increased. The
successful application of our metric to river restoration might be promising to further apply this metric for
example in river assessments or meta-community structure.
Key words: integrated dispersal metric, weight approach, macroinvertebrate, community succession, river
restoration.
Highlights
We develop a new dispersal metric for river ecosystems.
We test our metric using 23 restoration projects in Germany.
Our metric successfully elucidates community succession in restored rivers.
Strong and weak dispersers show an inverse succession trend in restored rivers.
Our metric is useful to detect environmental perturbation and community succession.
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Introduction
In the natural state, many ecosystems are characterized by frequent disturbances that result in a constantly
dynamic environmental mosaic. This process is being enhanced by unprecedented global change (e.g. human
disturbance, habitat fragmentation, pollution, and climate warming) on a local, regional and global scale, which
is especially true for river ecosystems (Revenga et al., 2005; Xenopoulos et al., 2005). However, it is currently
unclear whether and how an organisms colonization capacity is able to cope with the new challenges, like
global warming. Colonization is a series of processes that includes population dispersal, establishment, and
reproduction (Wirth et al., 2008). As a key attribute of colonization, dispersal capacity, a measure of frequency
and distance of organisms’ movement, among different habitats can greatly influence community dynamics
(Beisner et al., 2006; Heino, 2013). This topic has been well studied in terrestrial and marine ecosystems
(Grantham et al., 2003; Kinlan and Gaines, 2003; Bowler and Benton, 2005; Lester et al., 2007; Clobert et al.,
2012), but our knowledge of colonization in freshwater ecosystems is still poor. Consequently, any projections
of the impacts of global change on freshwater organisms made or assessments of re-colonization success of
riverine organisms in restored rivers without sufficient information about their dispersal capacities are
inadequate.
Due to a lack of information on dispersal capacity for freshwater organisms, the application of community
succession theory to river ecosystems has not been widely addressed until now (Milner et al., 2008).
Nonetheless, dispersal studies have been conducted for some fish taxa in freshwater ecosystems (Pépino, 2012;
Stoll et al., 2013; Radinger and Wolter, 2014), but only a few studies have been carried out to identify the
dispersal capacity of other functionally important groups, such as benthic invertebrates. This is because the
direct measurement of dispersal capacity is notoriously difficult for a large number of small-sized benthic
invertebrates (Hughes, 2007; Brederveld et al., 2011), and which is partly related to their diverse life cycles.
Most benthic invertebrates live at the bottom of a river channel and sometimes on land, such as the adult stages
of most aquatic insects. Consequently, the dispersal of benthic invertebrates is likely to occur in aquatic and
aerial habitats (Bilton et al., 2001; Bohonak and Jenkins, 2003). Within the aquatic dispersal mode, passive
(with the external aid of water flow, wind, or animal vectors) drift and active (self-generated movement)
movement along the river bottom are of particular importance, while within the aerial dispersal mode, the active
(upstream) flight and the passive wind drift of adult aquatic insects predominate (Bilton et al., 2001). It appears
that benthic invertebrates with life cycles restrict to aquatic habitats show lower dispersal capacities whereas
those with a flying adult stage tend to be stronger dispersers (Miller et al., 2002; Hughes, 2007, Kappes et al.
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2012). In addition to life cycle stages, the relative importance of dispersal via active or passive modes also
differs among taxonomic groups. These various mobility and life cycle characteristics make benthic
invertebrates one of ideal model groups for conducting comprehensive ecological studies in river ecosystems.
Although the number of studies of dispersal capacity of freshwater organisms has been built up steadily, the
proportion of the species studied is still low compared to the entire riverine communities. This has resulted in a
gap between the increasing demand for studies of benthic invertebrates’ diversity and the inadequate knowledge
of the species traits. To fill in this gap, a considerable advancement has been made by compiling a large
number of trait attributes into databases, for instance the STAR (Standardization of River Classifications)
project (www.eu-star.at; Furse et al., 2006) or the www.freshwaterecology.info database (Schmidt-Kloiber and
Hering, 2012). The four major dispersal modes, aquatic active, aquatic passive, aerial active, and aerial passive,
are incorporated into those species trait databases (Bis and Usseglio-Polatera, 2004; Furse et al., 2006; Schmidt-
Kloiber and Hering, 2012). However, each individual mode may provide different dispersal aspects in a certain
extent, and a comprehensive measure to quantify an integrated dispersal capacity is still lacking. Therefore, the
main aim of our study is to build a simple metric by incorporating these four dispersal modes to represent an
integrative assessment of dispersal capacity for several hundred riverine benthic invertebrates. This metric will
be beneficial to future freshwater studies investigating for example colonization or meta-community structure.
River restoration provides an opportunity to test the suitability of our metric because restored rivers need to
be (re-)colonized by benthic invertebrates following restoration. This colonization process particularly depends
on the dispersal capacity of benthic invertebrates: species with a high dispersal capacity are expected to colonize
the restored sites first, while species with low dispersal capacities will show up much later. To investigate this
pattern we used riverine benthic invertebrate data from 23 restored sites which have spanned 1 to 10 years after
restoration. All these 23 restoration projects involved significant changes of river bottom sediments including a
removal of specimens. As we neither had data from these restored sites prior to restoration nor from consecutive
yearly monitoring of those restored sites we applied a space-for-time substitution approach, using each
restoration as a temporal replicate. We compared the dispersal capacity values of the 23 restored sites with
dispersal capacity values from 23 unrestored sites, each nearby one of the restored sites. As a second control
group, we calculated the dispersal capacity values of all other available community data from river sites in the
nearby surroundings (< 5 km) of the restoration projects. This 5 km surrounding has been shown to be the
relevant species source pool for the colonization of restored sites (Sundermann et al., 2011a, b; Stoll et al., 2013,
Stoll et al. 2014).
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Based on this study design and using our new developed metric, we calculated an average dispersal capacity
value for the 23 benthic invertebrate communities in the restored sites, the 23 communities in the nearby
unrestored sites, and the 298 communities from the 5 km surroundings. This enables us to test the following
hypotheses: 1) the average dispersal metric of benthic invertebrate communities decreases over time at restored
sites, while no such changes in the average dispersal metric of benthic invertebrate communities can be
observed in the unrestored and surrounding sites, and 2) species that are strong dispersers are expected to
rapidly colonize the restored sites, while weak dispersers need longer time to colonize the restored sites and thus
they are expected to increase continuously in the early post-restoration stage.
Materials and methods
Study system and data collection
Twenty-three larger restoration projects (Fig. 1) in the low mountain and lowland areas (26268 m above sea
level) of Germany were selected all aiming at an improvement of habitats, hydrological conditions, and species
diversities (Sundermann et al., 2011a, b; Stoll et al., 2013). At all sites the applied restoration measures initially
led to a significant disturbance of the riverbed, following a removal of specimens and opening new habitat for
colonization. Benthic invertebrate community data were compiled from these 23 restored sites for our analyses.
Beside, benthic invertebrate community data from two control groups, unrestored and surrounding, were used to
differentiate the temporal colonization patterns of benthic invertebrate communities among these three river
treatment categories. The unrestored site was located upstream of the restored site to avoid the influence of
organisms drifting from the restored site, and the mean distance between the paired restored and unrestored sites
was 1 km. The unrestored control site was selected because it could represent the conditions of the restored site
prior to the restoration action. In each river, both restored and unrestored sites were similar in terms of geology,
adjacent land-use, river type, and catchment area. The surrounding sites were 05 km away from the restored
sites. Only river sites within the same catchment where the restoration project was conducted were considered
for the surrounding site datasets, which resulted into 298 surrounding sites. The distance between the
surrounding site and the restored site was calculated using the network analyst tool in ArcMap 10.0 (ESRI Inc.,
Redlands, USA). Because data from consecutive yearly monitoring for those river sites were not available, the
space-for-time substitution approach was used to represent the riverine biological conditions during the 110
years. This is generally not the best choice, and repeated sampling at the same restored site over several years
would be more valuable in carrying out dispersal studies. However, the substitution approach is still valuable in
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the context of limited human and financial resources, and this approach has been widely used in previous studies
(Sundermann et al., 2011b; Lorenz et al., 2012; Haase et al., 2013; Lorenz et al., 2013; Stoll et al., 2013; Blois et
al., 2014).
Benthic invertebrates were collected from March to July in 2007 and 2008 in the restored and unrestored sites
following the EU Water Framework Directive (WFD) sampling protocol (Haase et al., 2004a, b). Twenty
multiple habitat samples were taken in each site within 200 m river reaches using a shovel sampler (25 × 25 cm
sampling area and 500 μm mesh size). All benthic invertebrates were preserved in 70% ethanol and identified in
the laboratory following the protocol of Haase et al. (2004a, b). Data from the surrounding sites were collected
by governmental environmental agencies of the states of Hesse and North Rhine-Westphalia during 20042008
and followed the same sampling protocol as for the above restored and unrestored sites. All analyses of benthic
invertebrates in our study were based on both quantitative (abundance of a given species in one sampling site)
and qualitative (presence/absence of a given species in one sampling site) data.
Dispersal capacity
We selected four dispersal modes, i.e. aquatic active, aquatic passive, aerial active, and aerial passive to
determine an integrative measure of dispersal capacity for each benthic invertebrate. Aquatic dispersal is
dominant in the larval stages of aquatic insects and for many other taxa which spend their entire life in the water.
However, most aquatic insects are not solely restricted to instream dispersal but are potentially capable of aerial
(over-land) dispersal in their winged adult stage. Dissimilar to the common aquatic and aerial dispersal ways of
benthic invertebrates, some individuals may disperse by walking through the ground, for instance Gammarus,
thus exhibiting terrestrial dispersal. For most riverine benthic invertebrates, however, there are only a few
studies quantifying the minimum terrestrial dispersal distances (Flecker and Allan 1988, Hershey et al. 1993)
and there is no database available to quantify terrestrial dispersal. We assume that terrestrial dispersal does not
greatly contribute to the overall dispersal capacity of benthic invertebrates. Based on the STAR database, the
four selected dispersal modes were available for 528 out of 641 of the studied taxa (Table S2). In the STAR
database, a positive integer, ranges from 0 (no affinity) to 3 (high affinity), is assigned to each taxon, which
describes its affinity to each dispersal mode. For example, Haplotaxis gordioides (Oligochaeta) has 1 point for
aquatic active and passive modes, respectively, but 0 point for both aerial active and passive, resulting in a low
value of standardized dispersal metric (standDM), i.e. 0.08 (Table S2) and an assignment to the weak disperser
group; while Hydropsyche saxonica (Trichoptera) has 2, 3, 3, and 1 points for aquatic active, aquatic passive,
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aerial active, and aerial passive, respectively, leading it to a high value of standDM, i.e. 1.00 (Table S2) and an
assignment to the strong disperser group.
For the majority of benthic invertebrates, it is supposed that aerial dispersal distance is greater than aquatic
dispersal distance (Minshall and Petersen, 1985). For example, when water velocity was approximately 50 cm s-
1, nymphs of Hydropsyche spp. could drift 11.5 m on average while Baetis rhodani travelled 4.4 m (Elliott,
1971). The average flight distances of adult hydropsychids along the Detroit River and Lake St. Clair in Canada
were 1.8 km, 5 km in some cases with the light trap collection (Kovats et al., 1996), and half of the emerging
Baetis in an Arctic stream might travel at least 1.6 km upstream from their emergence sites (Hershey et al.,
1993). Therefore, it is necessary to assign more weight to aerial dispersal mode to increase the accuracy of our
metric. In our study, we estimated the suitable weight using the following approaches: First, 14 weights of aerial
dispersal modes (110, 15, 20, 25, and 30) were selected as candidate weights where the value of 1 referred to
equal weights for aquatic and aerial modes and the value of 30 referred to the highest candidate weight of aerial
mode. We then applied minimum-maximum normalization approach to those 14 candidate dispersal metrics to
produce standDMs with the range between 0 and 1 (equation 1). Second, we made a series of regressions of
years after restoration against the candidate standDM with average values of all individuals (abundance) and
species presence (Fig. 2A, C). Third, the most suitable weight was selected as the one with the highest
explanatory power and the lowest P value. The statistical results showed that the highest explanatory power and
lowest P value occurred when the weight was 2 using abundance data (Fig. 2 B), while the increasing of
explanatory power and decreasing of P values were keeping until the weight reached 10 using presence/absence
data (Fig. 2D). Although we used 10 as the weight for presence/absence data, we should note that the significant
tests have no changes incorporating different weights (Table S3). Consequently, we presented our metrics for
abundance (equation 2) and presence/absence data (equation 3) with the following equations:
(1)
(2)
(3)
Where standDM refers to standardized dispersal metric; DM refers to dispersal metric of species i; AP refers to
abundance or presence of species i at site j; DM_abund refers to dispersal metric with abundance data;
DM_pres/abs refers to dispersal metric with presence/absence data; aqa refers to dispersal capacity of species i
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via aquatic active mode; aqp refers to aquatic passive mode; aea refers to aerial active mode; and aep refers to
aerial passive mode.
Statistics
To test the first hypothesis, a significant temporal change in the standDM of benthic invertebrate communities
only occurs in the restored sites, the standDM of all taxa at a given site was plotted as a function of time. Non-
linear regressions (inverse first order, equation 4) were used to extract the temporal trends of the standDM of
benthic invertebrate communities in the restored, unrestored, and surrounding sites. Inverse first order
regression was used because the standDM of benthic invertebrate communities changed significantly in the first
half decade and then reached an ecologically dynamic equilibrium. Similarly, inverse regressions were broadly
used, e.g., to estimate decomposition rate of leaf litter over time in river systems (Austin and Vitousek, 2000;
Cusack et al., 2009).
(4)
where y refers to dependence (standDM), y0 refers to intercept, a refers to corelation coefficient, and x refers to
independence (years after restoration).
To test the second hypothesis, strong dispersers rapidly colonize the restored sites while the colonization of
weak dispersers is slow, we arranged all taxa in an ascending order according to their dispersal metrics and then
allocated them to four dispersal groups using a quartile approach. Taxa in the 1st quartile (Q1) were defined as
weak dispersers, and taxa in the 4th quartile (Q4) were strong dispersers. Taxa in the 2nd and 3rd quartiles were
categorized as weak-medium dispersers (Q2) and strong-medium dispersers (Q3), respectively (Table S2; Fig.
3A). For all three river treatment categories, the temporal changes of four dispersal groups in proportion were
then made using the inverse first order regressions.
Results
Dispersal metrics of various taxonomic groups
Similar results of estimated standDMs were observed using abundance and presence/absence data, but only
results evaluated with abundance data were shown here. Values of standDMs were lower for Oligochaeta and
Turbellaria and higher for Ephemeroptera and Trichoptera (Fig. 3B). After splitting all taxa into four dispersal
groups representing weak to strong dispersers, the value of standDM for each taxanomic group became more
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obvious; all taxa of Oligochaeta, Turbellaria, Hirudinea, Gastropoda, Crustacea, and Megaloptera were weak
dispersers, whereas most of Ephemeroptera and Trichoptera were strong dispersers (Fig. 4).
Ecological application of the dispersal metric
Overall, the standDM of the benthic invertebrate communities in the restored sites decreased significantly
(abundance: F1, 21 = 5.37, R2 = 0.20, P = 0.03; presence/absence: F1, 21 = 11.65, R2 = 0.36, P < 0.01) during the
110 years after restoration (Fig. 5), while there were no significant trends in the unrestored and surrounding
sites using both qualitative and quantitative data (Fig. 5).
Succession of benthic invertebrate communities was observed in the restored sites over the 10year period
with weak and strong dispersers responding contrarily in the first half decade and later reaching dynamic
equilibrium (Fig. 6A, B). Specifically, the strong dispersers rapidly colonized the restored sites in the first year
after restoration, and their proportional species richness dramatically decreased from the second year following
restoration onwards (F1, 21 = 9.00, R2 = 0.30, P < 0.01; Fig. 6B). The proportion of weak dispersers in the
communities significantly increased over the 10year period (abundance: F1, 21 = 4.78, R2 = 0.19, P = 0.04;
species richness: F1, 21 = 4.35, R2 = 0.17, P = 0.05; Fig. 6A, B). However, there were no significant trends in the
relative abundance of strong dispersers (Fig. 6A). For weak-medium and strong-medium dispersers, there were
no significant trends neither when using both quantitative and qualitative data (Fig. 6A, B). As expected, no
changes were observed for the four types of dispersers in the unrestored and surrounding sites.
Discussion
Strengths, weaknesses and challenges in the dispersal traits of benthic invertebrates
In our study, we developed a simple dispersal metric based on the widely-used dispersal modes from the STAR
database. Our metric provides a first estimate of dispersal capacity of benthic invertebrates, and it is exceedingly
valuable since our metric leads to the proper interpretation of community succession in the restored habitats. An
investigation of literatures indicates that dispersal traits are available for some benthic invertebrates, but there
are still a large number of species for which dispersal information is lacking (Bilton et al., 2001; Bohonak and
Jenkins, 2003; Brederveld et al., 2011). Based on the available knowledge of dispersal from various sources,
each species’ dispersal mode is assigned a positive integer between 0 and 3 in the STAR database (Furse et al.,
2006). However, those values are expert judgements, and as such prone to misjudgements. To fill in these
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knowledge gaps, further real data on species dispersal capacities are needed so that more comprehensive
analyses can be carried out in future studies.
Dispersal capacity of most benthic invertebrates is constrained in comparison to terrestrial organisms due to
both the distinct boundaries of freshwater ecosystems and the short-lived flying adult stages (Bohonak and
Jenkins, 2003; Tonkin et al., 2014). In contrast, some recent genetic studies indicate that aerial dispersal over
long distances within and across catchments may be common (Miller et al., 2002; Hughes, 2007), most likely by
means of passive dispersal modes (Bohonak and Jenkins, 2003) or because the distance between two adjacent
headwaters is within the dispersal range of some flying adults (Griffith et al., 1998; Geismar et al., 2015).
Nevertheless, these various studies are in line with the conclusion that the dispersal capacity of benthic
invertebrates is remarkably stronger via air than via water. In our study, we selected the suitable weight based
on the statistical results of 14 regression models. A higher weight with presence/absence data indicated that
species’ flying capacity is considerable more important than their aquatic drifting capacity. This stems from the
assumption of the regression model that a good model is with higher dispersal capacity in the early stage
following restoration activities and lower values in the latter stage. The increasing of weight of aerial dispersal
mode can therefore considerably increase the proportion of strong dispersers and further enlarge the value of the
metric in the early stage, which results into a higher explanatory power. In some extreme cases, it seems that the
contribution of the aquatic mode to the metric can be ignored if the aerial weight is large enough (e.g. > 30).
However, the clear bell curves indicated a relative lower weight using abundance data. This suggests the
importance of aquatic mode cannot be ignored because most benthic invertebrates spend their lives under the
water for a considerable long time (Hughes, 2007; Brederveld et al., 2011).
Dispersal in restored rivers
Our approach is based on the assumption that in undisturbed rivers, the standDM of benthic invertebrate
communities are in stable status and should not change over time, while in the recently disturbed rivers, strong
dispersers have higher probabilities to arrive earlier than weak dispersers and thus the standDM of benthic
invertebrate communities should change over time. This was reflected in our study design that covered restored,
unrestored, and surrounding river treatment categories. By applying our metric to these three river treatment
categories, a significant decrease in standDM of benthic invertebrate communities was observed in the restored
sites, particularly in the first 35 years, while there were no significant trends in either unrestored or surrounding
sites (Fig. 5), which supports our first hypothesis. In addition, non-significant trend in the unrestored sites
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indicated that the standDM of benthic invertebrate communities have no remarkable differences among the
restored sites prior to the restoration activities.
Yet, this raises another question: why was there a decrease in standDM of benthic invertebrate communities
in the restored sites over time? To answer this question, we investigated the community succession of benthic
invertebrates in the restored sites. Communities in the newly restored habitats were rapidly assembled by strong
dispersers. Species with low dispersal capacity needed longer time to arrive at the restored sites. But generally
species that are poor at dispersing tend to be better competitors once habitats have stabilized, and hence replace
the early arriving but less competitive strong dispersers. Taking the Simulium spp. as an example, they possess
strong dispersal capacities, but other freshwater species outcompete and displace the Simulium spp. in the
ongoing process of succession, which results in their absence or low abundance after a certain period of time
(Downes and Lake, 1991). An increase in Simulium spp. following disturbance was also reported by Milner et al.
(2008) who investigated the Glacial Wolf Point Creek in Alaska between 1977 and 2005. Taxa with good
dispersal capacity but poor competitive ability are defined as fugitive species (Horn and MacArthur, 1972;
Milner et al., 2008). Beside Simulium spp., many other taxa also belong to fugitive species, such as chironomid
Cricotopus intersectus (Milner et al., 2008) and Baetis spp. (Minakawa and Gara, 2003). Therefore, the non-
random establishment and persistence of strong and weak dispersers throughout the succession process of
communities is the answer to the above question and it also supports our second hypothesis.
Although clear temporal trends of the entire benthic invertebrate communities in the restored sites were
observed, the temporal trend of strong dispersers using abundance data was not significant. This is most likely
because a few strongly dispersing individuals can colonize the restored sites in the early post-restoration stage,
but they may not establish substantial populations in a short term. Such an effect can greatly influence the
responses of communities to environmental changes, and which leads to a relatively low proportional abundance
of strong dispersers in the early stage and a non-significant trend in reduction over time.
Milner et al. (2008) pointed out that the dispersal constraints largely influenced the community succession, as
non-insect taxa required at least 20 years to colonize. In our case, the colonization speed of non-insect taxa (e.g.
Oligochaeta, Turbellaria, Hirudinea, Gastropoda, and Crustacea) was slower than of insect taxa (e.g.
Ephemeroptera and Trichoptera). However, in comparison to the study conducted by Milner et al. (2008), the
colonization speed was relatively high in our case, and it took around 35 years for those non-insect taxa to
colonize the restored sites in this temperate climatic region. Minshall et al. (1983) also reported that it took three
years to get the fully colonization of the original taxa in the Teton River (Idaho) following a major flash flood.
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We are fully aware that no single mechanism can completely describe community succession. Beside
dispersal capacity, extrinsic (e.g., competition, landscape barriers) and intrinsic drivers (e.g., species’ life cycles,
parasite load) are also of the utmost importance. For example, landscape surface roughness or differences in
parasite loads may also influence succession (Grabner et al., 2014). Overall, our study provides a simple
dispersal metric which has proofed to be a useful tool to assess riverine organisms’ colonization of new habitats
emerging after dramatically anthropogenic disturbances. By means of this metric, our study demonstrates that
benthic invertebrate communities in new river habitats can develop rapidly, and the non-random succession of
benthic invertebrate communities indicates that a period of 35 years after restoration is needed to reach an
equilibrium in terms of community dispersal capacity. To further improve our metric, direct measurements of
dispersal frequency and distance for individual benthic invertebrates will be of the utmost importance. Beyond
stimulating work to refine taxon-specific estimates of dispersal capacity, the incorporation of our metric into
conventional bioassessment indices may improve the sensitivity of assessment indices to detect perturbations
and increase the ability of assessment indices to explore changes in river benthic invertebrate communities.
Acknowledgements
This study was supported by Deutsche Bundesstiftung Umwelt (FK 25032-33 ⁄ 2), Hessisches Ministerium für
Umwelt, ländlichen Raum und Verbraucherschutz (FK III 2-79i 02) and the research funding programme
‘LOEWE –Landes-Offensive zur Entwicklung Wissenschaftlich-ökonomischer Exzellenz’ of Hesse’s Ministry
of Higher Education, Research, and the Arts. Data from the surrounding sites were kindly provided by
Hessisches Landesamt für Umwelt und Geologie (HLUG) and Landesamt für Natur, Umwelt und
Verbraucherschutz NRW (LANUV).
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Figure legends
Fig 1. Geographic locations of the restoration projects and their surrounding sites in the low mountain and
lowland areas of Germany. Unrestored sites were not shown in the figure because the mean distance between
the paired restored and unrestored sites were 1 km, and all restored and unrestored sites were overlapped at the
defined spatial scale. Full name of the restored site is given in Table S1.
Fig 2. The 10-year trends in average standardized dispersal metric of benthic invertebrate communities in 23
restored sites. In total, 14 curves are presented, referring to 14 weights of the standardized dispersal metric: 1
10, 15, 20, 25, and 30 with (A) abundance and (C) presence/absence data. The R2 and P value of each regression
model are presented in (B, D).
Fig 3. Summary plots of standardized dispersal metrics for (A) 528 species and (B) 15 taxanomic groups from
low to high dispersal capacity. The classification of four dispersal groups is based on the weight calculated with
abundance data, i.e. 2 in (A). Four dispersal groups defined by a quartile approach are: weak dispersers = 025th;
weak-medium dispersers = 25th50th; strong-medium dispersers = 50th75th; and strong dispersers = 75th1. The
dot refers to mean value, the whisker refers to standard error, and the number above/below the whiskers refers to
the number of species on which the calculation is based in (B). Full name of each taxanomic group is given in
Fig. 4.
Fig. 4. The proportion of species richness among four dispersal groups for each taxanomic group.
Fig 5. The 10-year trends in average standardized dispersal metric of benthic invertebrate communities in the
restored, unrestored, and surrounding sites using (A) abundance and (B) presence/absence data. The weights of
abundance and presence/absence data are 2 and 10 for the standardized dispersal metric, respectively.
Fig 6. The 10-year trends in proportion of abundance and species richness of four dispersal groups of benthic
invertebrates in the restored sites using (A) abundance and (B) presence/absence data. The trends in the
unrestored and surrounding sites are not presented because of the missing of significant trends. Four dispersal
groups defined by a quartile approach are: weak dispersers = 025th; weak-medium dispersers = 25th50th;
strong-medium dispersers = 50th75th; and strong dispersers = 75th1.
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Fig. 1.
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Fig. 2.
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Fig. 3.
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Fig. 4.
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Fig. 5.
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Fig. 6.
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Supplementary
Table S1. Characteristics of 23 restored sites in the low mountain and lowland areas of Germany.
Site
Code
Parameter
Years after
restoration
Stream
order
Elevation
(m)
Catchment
area (km2)
Birkigsbach
A
5
4
120
25
Dill
B
2
6
217
314
Fallbach
C
4
2
127
30
Fauerbach
D
2
3
205
14
Fulda_Mecklar
E
2
7
196
2375
Fulda_Niederaula
F
2
6
207
1290
Gersprenz
G
1
6
162
154
Horloff
H
4
4
121
173
Josbach
I
5
3
268
29
Kinzig
J
7
6
134
885
Lache
K
3
3
104
11
Lahn
L
7
6
188
650
Nidda Bad Vilbel
M
6
6
96
1200
Nidda Ilbenstadt
N
1
6
104
1168
Nidda Ranstadt
O
3
5
129
226
Nidder
P
5
5
112
153
Niers
Q
7
5
26
658
Rodau
R
5
4
111
71
Ruhr
S
3
7
159
1000
Salz
T
3
5
144
83
Schwalm
U
10
4
28
250
Sulzbach
V
8
4
96
33
Ulster
W
1
5
242
384
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Table S2. Values of four dispersal modes and their standardized dispersal metric based on the weight of aerial
dispersal mode calculated with abundance data, i.e. 2, as well as the group information for 528 species of
benthic invertebrates using a quartile approach. A positive integer, ranges from 0 (no affinity) to 3 (high
affinity), is assigned to each taxon, and which describes the affinity to each dispersal mode. The standardized
dispersal metric ranges between 0 and 1, and which is produced using the minimum-maximum normalization
approach. Gen. refers to general family group; Ad. Refers to adult; and Lv. refers to larvae. All information is
available in the attached Adobe Acrobat file.
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Fig. S1. Sketch map of the relative localities of the restored, unrestored, and surrounding sites.
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  • Chapter
    The Teton River is located in southeastern Idaho and drains the west slope of the Rocky Mountains southwest of Yellowstone National Park (Fig. la). In the mid-1970’s a dam was constructed near the lower end of the river. The dam stood 95 m high and, during its initial filling, retained a volume of 3.10 × 108 m3. One June 5, 1976 the dam failed and released 70% of its volume within 2½h. The resultant wall of water reached 23 m in height near the dam and further downstream, where the water left the canyon in which the dam was located and spread out across the flood plain, it was 15 m high. Discharges near our eventual study area jumped from 30 to 3500 m3/s and velocities exceeded 12 m/s (Ray and Kjelstrom, 1978). About 9.4 km below the dam site, the Teton River splits into North and South Forks (Fig. la). The North Fork was severely damaged by the flood but the adjacent South Fork sustained much less structural upheaval. In order to facilitate channel restoration and bridge replacement, water was diverted from the North Fork in August and the streambed remained dry until mid-December.
  • Article
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    The colonization cycle hypothesis states that stream ecosystems would become depleted of insects if flying adults did not compensate for drifting immatures. Using long-term drift and benthic abundance data, we show that a Baetis mayfly nymph population moves downstream during development in the Kuparuk River in arctic Alaska. Baetis relative benthic abundance decreased from early to late season in an upstream unfertilized river section, while simultaneously increasing in the downstream fertilized section. Baetis nymphs drifted significantly more in the upstream unfertilized section, compared to the downstream fertilized section where food was more abundant. Approximately one-third to one-half of the nymph population drifted at least 2.1 km downstream during the arctic summer. A stable isotope tracer experiment and mathematical models show that about one-third to one-half of the adult Baetis population flew 1.6-1.9 km upstream from where they emerged. These results provide a quantitative test of the colonization cycle for the dominant grazer/collector in the Kuparuk River. Quantifying the colonization cycle is essential to understanding stream ecosystem function because offspring of downstream insects are needed for nutrient cycling and carbon processing upstream. Since downstream drift and upstream flight are important components in recovery of streams from disturbances, our results provide a quantitative method for predicting recolonization rates from downstream, essential to estimating recovery.
  • Article
    Streams and rivers represent a special and complex case of dendritic networks, where comparable stretches of river habitat are isolated from each other by river stretches with different abiotic conditions as well as the terrestrial matrix. The capacity of stream insects for ‘within-network’ (i.e. in-stream) and ‘out-of-network’ (i.e. lateral overland) dispersal varies dramatically, making them good models for studying how these different dispersal modes affect community or population structure, and how landscape structure affects dispersal.In this study, we genotyped 1129 individuals of the montane caddisfly species Drusus discolor from 44 sampling sites for 20 microsatellite loci to assess local population structure and the importance of overland dispersal and gene flow in two central German highlands. Using species distribution models, we assessed whether suitability of the terrestrial habitat between streams influenced genetic structuring among sampling sites.Our results generally indicate high levels of overland dispersal at small and medium distances and show that the surrounding landscapes or catchment boundaries do not drive population structure in the study species. The data further show a significant change in spatial autocorrelation at about 20 km distance in both regions, indicating that dispersal is distance-limited and reduced at distances above 20 km.
  • Book
    http://ukcatalogue.oup.com/product/9780199608904.do
  • Article
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    Restoration is an increasingly central theme in river ecology. Recent studies have highlighted the importance of the species pool in the surrounding river network for determining colonisation of restored river reaches by both invertebrates and fish. Using a comprehensive data set of 21 river restoration sites and 292 sites in the immediate surroundings, we tested the influence of distance to nearest colonist source on invertebrate colonisation based on a comparison of river network distances and Euclidean distances, expecting river network distances would better align with colonisation rates. We then assessed the importance of dispersal distance in relation to several other parameters, such as the number and intensity of barriers along the river network, surrounding taxon pool occupancy rate, physical characteristics of the restored sites and restoration techniques used in determining colonisation of commonly occurring benthic invertebrates. We hypothesised that (i) distance would be critical, with colonisation of restored sites declining with increasing distance; (ii) barriers between these sites would be a minor, but taxon-specific, influence on the colonisation; and (iii) the higher the regional pool occupancy rate of a certain taxon, the higher its probability of presence at a restored site. Overall, taxon pool occupancy rate was the most important driver of colonisation likelihood, followed by distance to nearest source, with the first kilometre particularly important. The effect of barriers was minor but significant, and taxon identity had no effect on the predictive ability of the model. Factors associated with the restoration projects such as techniques used and physical characteristics had minor influences, being completely outweighed by taxon pool and dispersal-related factors. To gauge the likelihood of successful outcomes of habitat restoration projects, we suggest it is important to assess regional taxon pools and ensure distances between healthy populations are minimised. These results clearly emphasise the importance of spatial planning for restoration projects.
  • Article
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    Increasing temperatures can be a significant stressor for aquatic organisms. Amphipods are one of the most abundant and functionally important groups of freshwater macroinvertebrates. Therefore, we conducted a laboratory experiment with Gammarus pulex, naturally infected with microsporidians. In each group, 42 gammarids were exposed to 15[degree sign]C and 25[degree sign]C for 24 h. Sex of gammarids was determined and microsporidian infections were detected by specific PCR. To quantify stress levels of the amphipods, the 70 kDa heat shock proteins (hsp70) were analyzed by western blot. More males than females were detected in the randomized population sample (ratio of females/males: 0.87). No mortality occurred at 15[degree sign]C, while 42.9% of gammarids died at 25[degree sign]C. Sequences of three microsporidians (M1, M2, M3) were detected in this G. pulex population (99.7%-100% sequence identity to Microsporidium spp. from GenBank). Previous studies showed that M3 is vertically transmitted, while M1 and M2 are presumably horizontally transmitted. Prevalences, according to PCR, were 27.0%, 37.8% and 64.9% for Microsporidium sp. M1, M2 and M3, respectively. Cumulative prevalence was 82.4%. Multiple infections with all three microsporidians in single gammarids were detected with a prevalence of 8.1%, and bi-infections ranged between 12.2% and 25.7%. In dead gammarids, comparatively low prevalences were noted for M1 (males and females: 11.1%) and M2 (females: 11.1%; males 0%), while prevalence of M3 was higher (females: 66.7%; males: 88.9%). No significant effect of host sex on microsporidian infection was found.Significant effects of temperature and bi-infection with Microsporidium spp. M2 + M3 on hsp70 response were detected by analysis of the whole sample (15[degree sign]C and 25[degree sign]C group) and of M2 + M3 bi-infection and gammarid weight when analyzing the 25[degree sign]C group separately. None of the parameters had a significant effect on hsp70 levels in the 15[degree sign]C group. This study shows that some microsporidian infections in amphipods can cause an increase in stress protein level, in addition to other stressors. Although more harmful effects of combined stressors can be expected, experimental evidence suggests that such an increase might possibly have a protective effect for the host against acute temperature stress.
  • Article
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    It is commonly assumed that the colonization of restored river reaches by fish depends on the regional species pools; however, quantifications of the relationship between the composition of the regional species pool and restoration outcome are lacking. We analyzed data from 18 German river restoration projects and adjacent river reaches constituting the regional species pools of the restored reaches. We found that the ability of statistical models to describe the fish assemblages established in the restored reaches was greater when these models were based on 'biotic' variables relating to the regional species pool and the ecological traits of species rather than on 'abiotic' variables relating to the hydromorphological habitat structure of the restored habitats and descriptors of the restoration projects. For species presence in restored reaches, 'biotic' variables explained 34% of variability, with the occurrence rate of a species in the regional species pool being the most important variable, while 'abiotic' variables explained only the negligible amount of 2% of variability. For fish density in restored reaches, about twice the amount of variability was explained by 'biotic' (38%) compared to 'abiotic' (21%) variables, with species density in the regional species pool being most important. These results indicate that the colonization of restored river reaches by fish is largely determined by the assemblages in the surrounding species pool. Knowledge of species presence and abundance in the regional species pool can be used to estimate the likelihood of fish species becoming established in restored reaches.