Water availability and proximity to natural areas influence terrestrial plant and macroinvertebrate communities in urban stormwater infrastructures

doi:10.21203/rs.3.rs-3889123/v1

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Water availability and proximity to natural areas influence terrestrial plant and macroinvertebrate communities in urban stormwater infrastructures

https://doi.org/10.21203/rs.3.rs-3889123/v1

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Stormwater infrastructures are primarily used for managing water runoff, but these environments can also foster biodiversity. Despite extensive literature about certain taxa found in these human-made environments, the terrestrial plants and macroinvertebrates present there remain understudied. Here, we compared alpha and beta diversity of plant and macroinvertebrate communities and assessed the influence of landscape characteristics on their composition in different types of urban stormwater infrastructures. Plants and macroinvertebrates were identified at the bottom and on the banks of 54 infrastructures (dry basins, wet basins with and without a water channel and retention ponds) in Quebec City and Trois-Rivieres, in Eastern Canada. Results showed poor and homogenous plant and macroinvertebrate communities in dry basins. Wet basins had the highest plant diversity, with more facultative wetland species. Wet basins with and without water channel had similar plant and macroinvertebrate composition, with the most heterogeneous communities. Retention ponds (with permanent water) had distinct communities with fewer plant species than wet basins. Macroinvertebrate and plant diversity decreased when excluding data from the banks of retention ponds from the analyses. The presence of natural areas around the infrastructures significantly influenced communities within a 2000 m and 500 m radius for plant and macroinvertebrate communities, respectively. Wetland plant species were generally found in infrastructures close to natural areas, whereas generalist species were associated with disturbed environments. Our results suggest that enhancing diversity of the stormwater infrastructure types at the regional and local (microhabitat) scales will maximize diversity of plants and macroinvertebrates.

Beta diversity

urban communities

vegetation

terrestrial macroinvertebrate

water management

Population growth and urban sprawl are continuously in conflict with the conservation of nature and associated biodiversity and ecosystem services (Barlow et al. 2016; McKinney 2002). Although rapid development of impervious surfaces provides immediate benefits to people, enabling them to live in cities, it also compromises the ecological services provided by nature to attenuate adverse effects of urbanization such as heat islands and flooding. In particular, wetland loss across the landscape and the channelling of rivers lead to intense floods in urban environments (Elmqvist et al. 2016; Gulbin et al. 2019; Juan et al. 2020). In synergy with climate change and its associated frequent and heavy precipitation locally (Douville et al. 2021), these losses of natural habitats create challenges for managing stormwater runoff. Cities therefore need to adopt a variety of approaches to increase their ability to respond effectively. For example, best management practices (BMP) such as porous surfaces, vegetated swales, artificial wetlands, and stormwater infrastructures are gradually being adopted to enhance control of rainwater in urban landscapes (Eger et al. 2017).

Stormwater infrastructures are built in cities to ensure quantitative water management during snowmelt or heavy precipitation (Field et al. 2000; Hogan and Walbridge 2007) and trap runoff contaminants (e.g., sediments, nutrients, hydrocarbons, metals) to reduce their discharge into natural aquatic environments (Starzec et al. 2005; Wang et al. 2014). They are engineered on a case-by-case basis to meet a targeted retention capacity that helps ensure flood protection and provide a certain level of control of suspended solids. Stormwater detention basins are specifically designed for temporary water storage and can be defined as either dry or wet basins. According to the stormwater management guide of Quebec province (Rivard 2011), the former typically shows water retention times of less than 24 hours and is frequently mowed, whereas the latter generally has a retention capacity of more than 24 hours, water-saturated soil, and specific wetland vegetation. In addition, wet basins can be channelized to concentrate water or help its evacuation, depending on the slope and the bathymetry of the basin. Stormwater infrastructures can also be designed as retention ponds, retaining water permanently and fostering the establishment of aquatic vegetation. Although stormwater infrastructures are primarily designed and constructed for water management, their abundance in expanding suburbs of North America increases interest for urban planners in their biodiversity and ecosystem services, such as carbon sequestration, and water filtration via uptake of certain contaminants (e.g., nutrients) (Temmink et al. 2022; Wissler et al. 2020).

Stormwater infrastructures hold great potential for the establishment of fauna and flora such as amphibians (Birx-Raybuck et al. 2010; Hamer and Parris 2011), fishes (Battisti et al. 2020; Schaeffer et al. 2012) and aquatic macroinvertebrates (Le Viol et al. 2009; Mackintosh et al. 2015; Meland et al. 2020). Aquatic plant diversity in retention ponds is well-known; macrophyte cover and diversity have been shown to decrease with turbidity, nutrient concentration, and proportion of urban areas (Akasaka et al. 2010; Mackintosh et al. 2015). In turn, the presence of macrophytes in urban stormwater infrastructures was found to promote the establishment of amphibians (Bounas et al., 2020) and macroinvertebrates (Holtmann et al. 2018; Perron and Pick 2020; Sinclair et al. 2021). Several studies found similar macroinvertebrate diversity between natural/unmanaged ponds and retention ponds (Hassall and Anderson 2015; Le Viol et al. 2009). However, Meland et al. (2020) observed different communities in retention and natural ponds with higher richness depending on conductivity, surface area and pond connectivity. Overall, these studies focused on the wet part of these infrastructures, leaving the surrounding banks unaccounted for. Despite their potential to host communities that are markedly different than those found in the water body, the banks contribute to homogenizing the vegetation between retention ponds (lower beta-diversity among sites), reducing similarities with natural wetlands and promoting exotic plant establishment (Bergeron d'Aoust 2021). Moreover, while retention ponds have been extensively studied for aquatic fauna and flora in urban environments (Hassall 2014), studies focusing on wet detention basins (channelized or not) remain scarce.

Landscape composition influences biodiversity of stormwater infrastructures. For instance, the occurrence and abundance of macroinvertebrates and amphibians decrease with larger impervious surface area (Mackintosh et al. 2015; Simon et al. 2009). Perron et al. (2021) also observed an association between Odonata community structure in urban ponds and landscape factors such as wetland area and wood cover. Urban fragmentation and densification can lower beta diversity and change plant community structure (Blouin et al. 2018). However, few studies have investigated landscape influence on vegetation in stormwater infrastructures, which could serve as habitat stepping stones toward establishment.

In this study, we examined plant and macroinvertebrate biodiversity in stormwater infrastructures to assess differences in community between types of infrastructures. Selected basins were categorized as dry detention basins, wet detention basins, channelized wet detention basins, or retention ponds (hereafter, the terms detention and retention will be omitted for facilitating the reading). We compared the specific richness, community composition, and heterogeneity (beta diversity) of plants and macroinvertebrates between these four infrastructure types. We hypothesized that (1) plant and macroinvertebrate communities differ between dry basins, wet basins (channelized or not) and ponds due to the difference of habitats; (2) macroinvertebrate community structure is associated with plant community structure; (3) the banks of the infrastructure contribute to plant and macroinvertebrate diversity; and (4) plant and macroinvertebrate communities in wet basins (channelized or not) are significantly affected by the proportion of natural areas in the landscape matrix since their conditions should allow colonization by plants and insects found in riparian zones, wetlands or wet prairies located close-by.

2.1. Study Sites

A survey of stormwater infrastructures was undertaken in Quebec City and Trois-Rivieres (Quebec, Canada) using Google Earth software (Google Earth Pro 2023) and information provided by the two municipalities. Quebec City covers a total surface area of about 485 km² and has a population of 549 459 inhabitants (Statistics Canada 2021). This area includes natural habitats (46%), built areas (40%), agricultural zones (12%) as well as vacant lots (2%). Trois-Rivieres is a smaller city, with a surface area of 288 km² and 137 188 inhabitants. This city includes more natural habitats (58%), less built areas (25%) and agricultural zones (14%), and a similar percentage of vacant lots (3%). The climate in the province of Quebec is characterized by a high thermal amplitude, with strong temperature differences between summer and winter and between the south and the north. The winter is characterized by a snow cover of 140 days, a three-meter snowfall and a mean temperature of -11°C in January. The growing period is approximately 165 days, and the mean summer temperature is 20°C in July (Gérardin and McKenney 2001; Environnement Canada 2012). Both cities are in a humid continental climate zone.

A total of 54 stormwater infrastructures were selected (Fig. 1). All the selected infrastructures were at least 10 years old, to ensure that biodiversity was well established and had lived through maintenance constraints. The surface area of the sites ranged from 650 m² to 16 000 m² (except for one 63 000 m² basin), with an average of 5 366 m². The selected infrastructures were categorized based on the presence of a permanent pool of water or not. The seven dry basins were frequently mowed. The twenty wet basins had at least temporarily water-saturated soil. The fifteen channelized wet basins had a channel often saturated in water. Finally, each of the twelve ponds held a permanent body of water.

2.2. Vegetation Characterization

Each selected infrastructure was first visited to identify and map the different types of plant communities at the bottom and on the banks of the infrastructure (e.g., reed beds, shrubs, wet meadows, lawn, etc.), excluding submerged macrophyte communities. On a second visit (June 25 to July 29, 2019), three quadrats (1 m x 2.5 m) per type of plant community were selected randomly and surveyed to identify herbaceous and shrub species. In rare cases, particularly small homogenous communities (e.g., aquatic communities) were sampled with two quadrats and large heterogenous communities (e.g., wet meadows) with four. Species cover was estimated using seven classes (0–1%, 1–10%, 10–25%, 25–50%, 50–75%, 75–90% and 90–100%). Species typical of wetlands were categorized as facultative wetland species (FACW) or obligate wetland species (OBL) according to Bazoge et al. (2014) and the PLANTS Database (USDA 2019) (see Supplementary Information S1).

2.3. Macroinvertebrate Sampling

Sampling of terrestrial macroinvertebrates was carried out from August 1 to 23 (2019) in 28 of the 54 stormwater infrastructures, corresponding to seven sites for each of the four types. The macroinvertebrates were sampled between 10 a.m. and 4 p.m. under sunny or partial cloudy skies and low winds. A sweep net (30 cm diameter; 0.35 mm mesh size) was used to capture macroinvertebrates that were on the vegetation. Three sessions of 20 seconds (one second per meter) were randomly carried out in each type of vegetation community. As the dry basins had homogeneous plant communities resulting in only three sessions of netting, they were sampled a second time in July 2021 to increase the number of samples after verifying that species richness and community composition were not different between the two sampling years. This enabled better comparison of the macroinvertebrate communities in the dry basins with those in the other infrastructure types which had more netting sessions. Macroinvertebrates were transferred into Falcon tubes (50 mL) with 90% ethanol and stored at 4°C. The specimens were then identified and counted at family level, when possible, as suggested in several studies (e.g., Rosenheim and Ward 2020; Sáenz-Romo et al. 2019). Diptera and Hymenoptera were mostly identified at order level, except for some targeted groups such as Nematocera and Brachycera (suborder level), as well as Apidae, Halictidae, and Dolichopodidae (family level).

2.4. Environmental Data Collection

Landscape variables were acquired using QGIS (QGIS 2023). Land use (urban, natural or agricultural) was characterized for each site within a radius of 500, 1000 and 2000 meters based on maps from the Canadian government (Canada Government 2020). Natural areas combine forests, grasslands, wetlands and rivers. The proportion of vegetation within the three radii was also estimated with the Normalized Difference Vegetation Index (NDVI), which is a good estimation of plant biomass on a finer scale than the government maps of land uses (Liang and Wang 2019). NDVI values were calculated from spectral data obtained from Landsat 8 satellite imagery (Libra 2015), according to Parviainen et al. (2010), with the following formula: NDVI = (NIR - R) / (NIR + R) (NIR: near infrared, R: red). We measured distances between the sampled stormwater infrastructures and the closest major roads and highways, as well as to the closest river and forest. These metrics make it possible to determine the proximity between each infrastructure and natural areas and their potential disconnection by major roads and highways. Nearby major roads and highways may be a corridor for plants like the exotic invasive Phragmites australis (Jodoin et al. 2008) or a barrier for flying insects (Andersson et al. 2017). Distance from the nearest stormwater infrastructure was used to analyze the proximity between them. The different distances were calculated using the NNJoin tool in QGIS. Local variables included a dummy variable corresponding to the city within which the sampled stormwater infrastructures are situated, as well as their respective surface area, calculated by GIS. All the variables are listed with the respective code names and units in Supplementary Information S2.

2.5. Statistical Analysis

All statistical analyses were carried out in the R environment (R Core Team 2023). For macroinvertebrates, family richness in the dry basins was first compared between the two sampling years using an ANOVA. Community composition was also compared with a multilevel pairwise comparison of permutational multivariate analyses of variance (9 999 permutations). We found no difference between years and therefore pooled the data for the other analyses. Family richness of macroinvertebrates as well as plant species richness and proportion of wetland plants were compared between infrastructure types with Generalized Linear Models. We used a Poisson distribution for the cumulative species richness according to the quadrat number and a quasibinomial distribution for the proportion of wetland plants. These models were selected according to the ratio between the residual deviance and the residual degree of freedom (ratio lower than 4). Type II analysis of variance was then used to estimate the difference in richness or proportion between the types of infrastructures. Post hoc multiple comparisons were performed using least-squares means with a Tukey adjustment. These analyses were illustrated by line graphs representing species richness and proportion of wetland plants according to quadrat number. We compared these results including and excluding quadrats sampled on the banks of the infrastructures to determine the influence of this structure on diversity.

Before evaluating beta diversity, species-by-site and family-by-site matrices with Hellinger transformation were computed for mean plant cover and relative abundance of macroinvertebrates, respectively. Species or families that had a frequency of occurrence lower than 5% were excluded. A site-by-site matrix using the Gower distance was calculated to compute the group centroid of each type of infrastructure. The difference between these group centroids (representing community differences between the types of infrastructures) was tested with a multivariate analysis of variance, and post-hoc tests were conducted with a multilevel pairwise comparison of permutational multivariate analyses of variance (9 999 permutations). The distance of each site to its associated group centroid, determining whether the dispersion of one or more groups differed, was calculated and tested with a permutation-based test of multivariate homogeneity of variance (9 999 permutations). Differences in multivariate dispersion and composition between infrastructure types were illustrated in a Principal Coordinate Analysis. They were compared including and excluding the banks to determine the involvement of this structure in the composition of the communities (PCoA).

The relative abundance of plant species and families of macroinvertebrates in each infrastructure type was investigated using a Principal Component Analysis (PCA). This analysis provides more information than a PCoA on specific species or families that characterize the different types of infrastructures. We excluded the banks of the infrastructures to better capture the differences between the type of infrastructure within the main portion of the basin. A Mantel test was conducted to investigate the correlation between vegetation and macroinvertebrate communities. Borcard and Legendre (2012) highlighted the efficiency of this test for ecological analyses. Percent cover and relative abundance of targeted plants and macroinvertebrates were illustrated by boxplots according to the type of infrastructure and tested with Kruskal-Wallis tests with Dunn’s multiple comparison test and a Bonferroni adjustment.

Each environmental variable was standardized prior to statistical analyses. Both plant and macroinvertebrate community differences were first assessed between Quebec City and Trois-Rivieres by a PERMANOVA on an RDA with the two cities as dummy variables to test for a difference in plant and macroinvertebrate communities between the two locations. Then, a partial RDA was conducted for plant communities (with ‘city’ used as a co-variable), as they were significantly different according to the city, whereas a classic RDA was performed for macroinvertebrates. A forward selection model using ordiR2step function in the “vegan” R package was performed to avoid multicollinearity of the variables and to test for the most significant variables influencing communities (α = 0.05).

3.1. Species Richness

A total of 291 plant species were identified within the 54 sites surveyed. For the dry basins, plant species with a frequency of occurrence of more than 70% were Festuca rubra, Lathyrus pratensis, Vicia cracca, Plantago major, Taraxacum officinale, Trifolium pratense and Trifolium repens,which are mostlygrass and melliferous species (Supplementary Information S3). For wet basins, the wetland species Carex scoparia, Galium palustre, Juncus vaseyi, Scirpus atrovirens, Typha sp., Juncus effusus, Lythrum salicaria and Equisetum arvensis were the most frequent (more than 70%). For channelized wet basins, plant species with a frequency of occurrence of more than 70% were Solidago canadensis, Vicia cracca, Juncus vaseyi and Symphyotrichum lanceolatum, as well as Equisetum arvensis, Lythrum salicaria and Typha sp., that were found in all these sites. Few species were frequent in ponds, with only Lythrum salicaria and Typha sp. showing a frequency of occurrence of more than 70% of the sites.

We identified 84 families within the subset of 28 sites selected for the macroinvertebrate survey. For dry basins, a total of 20 families were found at a frequency of occurrence of more than 70%, among them Apidae as well as 14 families found on all sites, such as Brachycera, Hymenoptera and Nematocera (see Supplementary Information S4). For wet basins, a total of 11 families were captured on more than 70% of the sites, including Aphididae, Brachycera, Cicadellidae and Heteroptera, which were found on all sites. For channelized wet basins, 25 families were sampled at a frequency of occurrence of more than 70%, including 12 families found on all sites, such as Brachycera, Hymenoptera, Nematocera and Lepidoptera. Ponds had only seven families frequently encountered on more than 70% of the sites. Only Aphididae, Brachycera and Nematocera were captured at all ponds.

Differences in specific richness between infrastructure types were assessed using species richness accumulation curves with a significant increase in richness as the number of quadrats sampled increases (Fig.2). Total plant richness was the lowest for dry basins (Fig.2a). Wet basins had greater diversity than ponds, with channelized wet basins showing intermediate values (Fig 2a). The higher specific richness in wet basins was associated with a greater FACW species richness (Fig.2c). When the banks were excluded, ponds had significantly fewer species than all other types of basins (Fig.2b), with the lowest FACW species richness as well (Fig.2d). Wet basins and channelized wet basins were similar in terms of plant richness whether the banks of the sites were included or not (Fig.2a, 2b). The same analyses of OBL species richness showed no significant differences between wet basins, channelized wet basins and ponds, whether including or excluding the banks (results not shown). A comparison between ponds, wet basins and channelized wet basins focusing on OBL and FACW species proportion of the total pool of sampled species, excluding the banks of the infrastructure, showed a greater proportion of FACW species in wet basins than in ponds (Fig.2g) and inversely, a greater proportion of OBL species in ponds than in wet basins (Fig.2h), with channelized wet basins showing intermediate proportion values for both. Overall, proportion of OBL species decreased with the number of quadrats for the three infrastructure types.

Macroinvertebrate richness showed no difference among the four types of stormwater infrastructures (Fig.2e). However, when the banks were excluded, dry basins showed greater family richness than ponds (Fig.2f), indicating that the banks attracted more macroinvertebrate families compared to the bottom of the ponds.

3.2. Beta Diversity and Community Composition

Beta diversity was used to assess the variability in community composition among the studied sites. Dry basins had significantly lower beta diversity than the other types of infrastructure when data from the banks were included (Table 1; Fig.3a). When the banks were excluded, channelized wet basins had the greatest beta diversity, followed by wet basins, whereas dry basins and ponds had the lowest (Table 1; Fig.3b). Ponds had a high variability in beta diversity. Indeed, four ponds were markedly different in community composition from the others and caused greater variability, even though average beta diversity was low. Plant communities differed between types, except for wet basins and channelized wet basins, which had a similar composition (Table 1; Fig.3a, 3b; centroid position) whether the data from the banks were considered or not.

Beta diversity analyses for macroinvertebrates yielded patterns like those observed for the vegetation. Dry basins had significantly lower beta diversity than the other infrastructure types when data from the banks were included (Table 1; Fig.3c). When excluding the banks, channelized wet basins had the highest beta diversity, followed by wet basins and dry basins, while ponds showed an intermediate beta diversity between that of dry and wet basins (Table 1; Fig.3d). Dry basins, channelized wet basins and ponds each hosted different communities, whether considering banks data or not (Table 1; Fig.3c, 3d, centroid position). Nevertheless, wet basins showed comparable communities with both ponds and channelized wet basins. The relationship between plant and macroinvertebrate communities was supported by a significant Mantel correlation between the two taxa groups (observation: 0.58, p-value < 0.001).

The differences in vegetation community composition between infrastructure types are shown in a PCA (Fig.4a). The first two axes explain 29.4 % of the variation. The results revealed the importance of grasses as well as melliferous plants such as Trifolium repens, Prunella vulgaris and Taraxacum officinale in dry basins. OBL species characterizing ponds included Typha sp., Schoenoplectus tabernaemontani and the invasive FACW species Phragmites australis. Typha sp. and Phragmites australis tended to cover higher portions of the communities in channelized wet basins and ponds compared to the other types of infrastructures (Fig.5). Frequency of occurrence of Typha sp. was 0%, 70%, 100%, 100% for dry basins, wet basins, channelized wet basins and ponds, respectively, and for Phragmites australis, 0%, 20%, 53%, 50%. Similar communities in wet basins and channelized wet basins were represented by FACW species such as Juncus vaseyi, Scirpus rubrocinctus, S. atrovirens, Juncus effusus, Phalaris arundinacea, Carex scoparia, Symphyotrichum lanceolatum and Galium palustris. Generalist species such as Equisetum arvense, Trifolium hybridum and Lathyrus pratensis were also commonly present in these types of infrastructures (Fig.4a).

Macroinvertebrate community composition was also represented using a PCA, with differences in families plotted according to the type of infrastructure (Fig.4b). The first two axes explain 40.5 % of the variation. Hymenoptera and Apidae tended to be associated with dry basins, whereas Nematocera, Brachycera and Coenagriidae, which represent mosquitoes, flies and dragonflies respectively, were more commonly found in ponds. Heteroptera, Cicadellidae, Membracidae and Miridae were well represented in wet basins and channelized wet basins. Pollinators were supplemented by higher abundances of Apidae and Hymenoptera (including several species of pollinators) in dry basins compared to ponds (Fig.5). However, the high variability in abundance among dry basins must be considered. Frequency of occurrence of Apidae and Hymenoptera was 71%, 28%, 28%, 14% and 100%, 86%, 100%, 86% for dry basins, wet basins, channelized wet basins and ponds, respectively.

3.3. Landscape Variable Analyses

Out of the 18 variables tested (Supplementary Information S2), only the presence of natural areas within a 2000 m radius (Nat2000) was found to be associated with plant composition in stormwater infrastructures and accounted for 15.2% of the variation (Fig.6; first axis). Dry basins were excluded from this analysis because of the particularly homogenous and poorly diversified plant composition in these highly disturbed sites (frequent maintenance). A higher proportion of natural areas favoured FACW species such as Scirpus rubrotinctus, S. atrocinctus, Carex echinacea, C. garberi, C. crinita, whereas generalist species such as Lathyrus pratensis, Phleum pratense, Agrostis stolonifera, Vicia sepium, Solidago canadensis, and Lotus corniculatus, were noticeably present with less natural areas around. The second axis seemed to reflect the water availability gradient (8.7% of the variability explained) with the OBL species Schoenoplectus tabernaemontani and Typha sp. associated with ponds and a higher number of FACW species such as Carex scoparia, C. vulpinoidea and Phalaris arundinacea occurring in wet basins, whether in the presence of a channel or not. The same analysis was conducted including only wet basins and channelized wet basins sites, and Nat2000 was again the only significant variable (15.5% of variation explained, plot not shown). For ordination analyses on ponds alone, no variable was found to be significantly associated with plant composition. The same analyses were conducted on macroinvertebrate data and showed a significant influence of natural areas as well as vegetation cover mapped from NDVI index on a finer scale within a 500 m radius. The two variables explained 41% of the variation (plot not shown). Neither local variables such as site surface area or distance to main roads, nor the distance between one site and other sites were significantly associated with plant and macroinvertebrate community structures.

Vegetation and macroinvertebrate communities were greatly influenced by the infrastructure type in urban stormwater infrastructures, and the composition of the banks tended to increase diversity as their conditions contrasted with those found in the bottom of the infrastructure. Macroinvertebrate communities were driven by the vegetation composition, showing the same responses in terms of differences in community composition and heterogeneity as vegetation according to infrastructure type. A scarcity of natural areas around the stormwater infrastructures led to a shift of community composition in both taxa compared to infrastructures surrounded by more natural areas.

4.1. Water and Bank Area Drive Local Vegetation Diversity and Composition

Wet basins enhanced vegetation diversity compared to dry basins and ponds, providing a habitat for a greater number of facultative wetland species. This is consistent with findings in the literature, highlighting that wet basins, or more generally constructed wetlands, increase biodiversity (De Martis et al. 2016; Zhang et al. 2020). Moreover, channelization of wet basins had no effect on vegetation communities, probably because there was an important humidity gradient in both types of wet basins. However, channelized wet basins had the most heterogenous communities, as shown by a higher beta-diversity between sites. This could be explained by a high variability of standing water in channelized wet basins caused by well-structured and regularly dredged channels or variation in sediment accumulation depending on maintenance frequency. Surprisingly, obligate wetland species richness (OBL) did not differ between wet basins and ponds, even though the area permanently submerged in ponds is much greater. Stormwater infrastructures in urban areas may not be appropriate for attracting a high richness of OBL species because of urban pressures, such as high turbidity and nutrient concentration, or the high pond water levels that leaves little space for the establishment of wetland plants (Akasaka et al. 2010; Lewerentz et al. 2021; Mackintosh et al. 2015). However, a higher proportion of OBL was found in ponds compared to wet basins, certainly due to the permanent presence of water, favorable to the establishment of these species while limiting others. Moreover, the proportion of OBL species decreased as the number of quadrats sampled increased. This may be explained by the replacement of OBL species by more diverse communities characterized by FACW species, both present in the surveyed habitat.

As expected, dry basins had the lowest diversity of all stormwater infrastructure types. A substantial body of literature has highlighted the consequences of intensive mowing of urban lawns and vacant lots. Mowed vegetation is usually represented by a few dominant species, thus resulting in reduced biodiversity (Lerman et al. 2018; Norton et al. 2019; Sehrt et al. 2020). The upper portion of the banks in wet basins, channelized wet basins and ponds may be compared to dry basins because they are dry and likely to favor generalist and tolerant species in urban areas. However, the types of maintenance and their frequency differ: Dry basins were frequently mowed, contrary to the other types of infrastructures in the studied cities. For this reason, the vegetation found in the upper portion of the banks in wet basins, channelized wet basins and ponds may be compared to urban spontaneous vegetation and to plant communities colonizing vacant lots (Blouin et al. 2018).

4.2. Macroinvertebrates Directly Depend on Vegetation Composition

Macroinvertebrate communities were strongly associated with vegetation composition, having similar patterns of differences in community composition between the types of infrastructure and community heterogeneity among sites (beta diversity). The higher beta diversity in macroinvertebrate communities in wet basins and channelized wet basins may be associated with the heterogeneity in vegetation communities. This concurs with findings by Sartori et al. (2015), who demonstrated that heterogeneity of macroinvertebrate communities was related to vegetation heterogeneity in constructed wetlands. Florencio et al. (2014) also observed a direct relationship between macroinvertebrate communities and aquatic plant richness and composition in temporary wet ponds. Odonata are particularly influenced by both obligate wetland and upland plant species (Goertzen and Suhling 2013; Johansson et al. 2019; Perron and Pick 2020) due to emergent vegetation offering perching and sheltering locations as well as oviposition substrate (Corbet 1980; Remsburg et al. 2008). Moreover, a high number of Nematocera, essentially represented by mosquitoes, were present in ponds, which contrasts with the Yadav et al. (2012) who has demonstrated lower larval density in deepwater ponds compared to stormwater wetlands. A longer hydroperiod can facilitate predator establishment, which decrease larval density (Holmes and Cáceres 2020). Moreover, the occurrence of mosquitoes is higher with mixed vegetation communities, fringing trees and cattails in stormwater infrastructures (Hunt et al. 2006; Yadav et al. 2012). This is consistent with our results since we observed more cattails in ponds. Other studies have shown an increase in abundance, diversity and biomass of aquatic macroinvertebrates related to the presence of macrophytes in natural/constructed wetlands and ponds (Gleason et al. 2018; Hsu et al. 2011; Theel et al. 2008). However, we did not observe a similar pattern between plant specific richness and macroinvertebrate family richness when comparing the infrastructure types. Even though some studies on aquatic macroinvertebrates identified at the family level obtained reliable results (Heino and Soininen 2007; Marshall et al. 2006), this taxonomic resolution certainly has limitations that may partly explain the similarity in macroinvertebrate richness values we observed among infrastructure types. Finally, we chose to investigate many macroinvertebrate taxonomic groups using one protocol to determine overall diversity, whereas Gál et al. (2019) advised on targeting only certain taxonomic groups to better assess the effect of urbanization.

Macroinvertebrate diversity in dry basins, whose most frequently mowed grassy vegetation lies at the bottom and on the banks, was similar to the diversity found in the other infrastructure types when banks were included in the analyses. Its grassy vegetation is known to attract specific macroinvertebrates, such as Hymenoptera and pollinators, because melliferous plant species predominate. However, mowed lawns seem to cause a shift in pollinator communities, with more dominant species and less supporting pollinators than meadows (Cloutier et al., unpublished results). This is consistent with our results that showed dry basin communities were particularly different and homogenous compared to the other infrastructure types. Spontaneous vegetation in urban areas offers better quality habitats for macroinvertebrate communities such as pollinators compared to regularly mowed lawns (Robinson and Lundholm 2012), and several studies have proven the benefits of reducing mowing frequency to allow for macroinvertebrate establishment, especially flying insects, and pollinators, in urban meadows (Norton et al. 2019).

4.3. Turnover of Species Composition with Landscape Proximity

Vegetation composition of both wet basins and channelized wet basins changed with the proportion of natural areas within a 2000 m radius. Generalist and tolerant species were dominant when natural areas were less present around the infrastructures, whereas their proximity favored the establishment of wetland species in wet basins and channelized wet basins. Urban fragmentation has a marked effect on vegetation and macroinvertebrate biodiversity (Faeth et al. 2011; Theodorou et al. 2020). Within our study area (Quebec City and Trois-Rivieres), riparian zones and woodlands are rich in wetlands, which could explain the diversity and abundance of FACW species found on our sites that were close to natural environments. Proximity of natural areas had no influence on plant communities in ponds. This could be explained by the scarcity of permanent water wetlands in the urban landscape, which limits the establishment of specific aquatic plants that rely on connectivity for dispersion.

The presence of permanent water in channelized wet basins and ponds seems to favor the establishment of P. australis in our study sites, although its coverage remained below 30% in these infrastructures. In contrast, P. australis was rarely present in the wet basins except at four sites. These results are surprising, at least for ponds, as P. australis tends to establish better in shallow water (Coops et al. 1996; Vretare et al. 2001). However, P. australis has high phenotypic plasticity and is thus able to acclimatize to deep water by allocating more resources to stem weight, and fewer but taller stems (Vretare et al. 2001). Further studies are required to investigate the colonization potential of P. australis and Typha sp. when in competition with different native wetland species to determine the factors of its establishment success in different urban stormwater infrastructure types.

Two proxies of natural area cover within a 500 m radius influenced terrestrial macroinvertebrate composition. The first proxy is based on the government map of land uses and roughly defines the natural areas in the city and its surroundings, whereas the second proxy with NDVI calculation is more precise and provides information about the small, vegetated areas in residential or highly urbanized areas such as vacant lots. Surprisingly, these two variables were not multicollinear, meaning that macroinvertebrates may potentially have different sensitivity to natural cover at two different scales. Certain macroinvertebrate families may have been more directly influenced by the connectivity to small natural areas, while others were more directly affected by the proximity to vast green spaces. Hyseni et al. (2021) insisted on the importance of green and blue landscape proximity for macroinvertebrate biodiversity in urban ponds, showing that specific richness and community composition are correlated with functional connectivity. Overall, our findings showed that landscape influenced macroinvertebrates within a smaller radius than for vegetation, which may reflect the need for closer proximity with green spaces to allow dispersion.

This study highlighted the benefits that different types of stormwater infrastructures can provide in terms of vegetation and macroinvertebrate diversity. Comparison between different infrastructure types showed that conditions in wet basins enhance wetland vegetation biodiversity and habitat heterogeneity. Dry basins attracted more generalist plant species and pollinators, whereas ponds favored few aquatic plant species, Odonata and other water-dependent species. The macroinvertebrate communities were directly associated with vegetation communities. The banks were particularly important for the establishment of plant and macroinvertebrate diversity in ponds. We also provided evidence that natural areas around these urban infrastructures influenced the community composition of vegetation and macroinvertebrates. Although managers generally perceive stormwater infrastructures as constructions with the primary function of regulating urban runoff, these human-made ecosystems have the potential to offer several other services if biodiversity considerations are integrated in the design (Saulnier-Talbot and Lavoie 2018). Pollution retention, organic matter decomposition and air temperature regulation are other examples of services provided by urban constructed wetlands. These infrastructures also offer cultural services, such as adding aesthetic value or natural spaces to the urban landscape. In the context of current efforts to increase urban biodiversity, we recommend creating microhabitat heterogeneity at the local scale with varying topography in wet basins and constructed ponds, as well as a mosaic of habitats at the regional scale that will maximize biodiversity in stormwater infrastructures. Better knowledge of urban ecological restoration is essential to adapt green infrastructures to the current management paradigm, even if these novel ecosystems do not provide the same conditions and biodiversity as natural lands (Saulnier-Talbot and Lavoie 2018; Zedler et al. 2012). Indeed, novel ecosystems such as stormwater infrastructures have the potential to generate new services with a biodiversity specific to urban areas. Improving their ecological value should therefore be a focus in the context of urban planning.

This work was supported financially by the Natural Sciences and Engineering Research Council of Canada (NSERC; grant number CRDPJ 530548-18), Ville de Québec, Ville de Trois-Rivieres and the Groupe de Recherche Interuniversitaire en Limnologie (GRIL).

The authors have no relevant financial or non-financial interests to disclose.

Author Contribution

M.P.P., M.P. and I.L. conceptualized the project. M.P.P. and M.P. worked on the methodology. M.P.P. wrote the main manuscript and original draft as well as data curation. M.P.P. and A.D. realized formal analysis. M.P., A.D. and I.L. supervised the study and validated all the steps. M.P., A.D. and I.L. reviewed and edited the manuscript. M.P. and I.L. provided access to the resources required for the study. I.L. contributed to funding acquisition.

Acknowledgements

We would like to thank our funding agencies, the Natural Sciences and Engineering Research Council of Canada (NSERC) and the Groupe de Recherche Interuniversitaire en Limnologie, for their financial support as well as the municipalities of La Ville de Québec and La Ville de Trois-Rivieres for their collaboration. Our thanks to Karen Grislis for linguistic revision. We are grateful to all the interns for assisting with fieldwork. We thank Maxime Tisserant, Léo-Janne Paquin and Steven L’Heureux-Le Page for their help with taxonomic identifications.

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Table 1. Differences in species beta diversity and species composition between the four stormwater infrastructure types including or excluding the data of the banks for vegetation and macroinvertebrates. Difference in beta diversity was tested with ANOVA by permutations on site distances to centroid (a,b) and difference in species composition was tested with PERMANOVA (c,d). DF: degree of freedom

Area of site	DF	F	P
a. Vegetation beta diversity
Including banks	3	30.11	0.001
Excluding banks	3	15.09	0.001
b. Macroinvertebrate beta diversity
Including banks	3	10.78	0.001
Excluding banks	3	3.81	0.014
c. Vegetation composition
Including banks	3	2.05	0.001
Excluding banks	3	2.6	0.001
d. Macroinvertebrate composition
Including banks	3	4.18	0.001
Excluding banks	3	3.27	0.001

No competing interests reported.

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Water availability and proximity to natural areas influence terrestrial plant and macroinvertebrate communities in urban stormwater infrastructures

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Abstract

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1. Introduction

2. Materials and Methods

2.1. Study Sites

2.2. Vegetation Characterization

2.3. Macroinvertebrate Sampling

2.4. Environmental Data Collection

2.5. Statistical Analysis

3. Results

4. Discussion

4.1. Water and Bank Area Drive Local Vegetation Diversity and Composition

4.2. Macroinvertebrates Directly Depend on Vegetation Composition

4.3. Turnover of Species Composition with Landscape Proximity

5. Conclusion

Declarations

Author Contribution

Acknowledgements

References

Table

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Supplementary Files

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