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2019 THE GREAT LAKES BOTANIST

DIFFERENTIAL PERSISTENCE AMONG NATIVE SPECIES PLANTED IN A STORMWATER RETENTION POND

Bradley M. Herrick1 Elizabeth Buschert University of Wisconsin-Madison Arboretum University of Wisconsin-Madison Arboretum 1207 Seminole Highway 1207 Seminole Highway Madison, WI 53711 Madison, WI 53711

James M. Doherty Erik R. Olson Stanford Online High School Northland College 220 Panama St. 1411 Ellis Ave. Stanford, CA 94305 Ashland, WI 54806

ABSTRACT

Retention ponds are a commonly implemented best management practice to treat urban stormwater. Newly constructed ponds offer an opportunity to increase native plant diversity in these artificial structures. However, plantings are rarely monitored, and shallow water areas often become colonized by non-native, invasive Typha spp. We documented the occurrence and abundance of eight planted native macrophytes in a newly constructed retention pond after one growing season and again after six growing seasons. Typha species were found in almost all plots after only one growing season and had completely colonized the planted emergent zone after six years. All native planted species, except one, were observed after six growing seasons, albeit at relatively low frequency and low cover. However, Pontederia cordata L. and Scirpus spp. were observed in over half of the plots, which suggests that they are able to coexist with Typha spp. even at low abundances. We suggest that P. cordata and Scirpus spp. should be considered in plantings in urban retention ponds to increase native species diversity.

KEYWORDS: Macrophytes, Native plants, Restoration, UW-Madison Arboretum

INTRODUCTION

Retention (wet-detention) ponds are a commonly implemented best management practice (BMP) to treat stormwater runoff in urban areas and to protect downstream aquatic systems from nutrient enrichment, sedimentation, toxicity, and reduced biodiversity (Dunn et al. 1995; Marsalek 1998; Tenenbaum 2004; Chen and Adams 2006; Wadzuk et al. 2010; Wang and Sample 2014). In addition, retention ponds can reduce peak flows (Hunt et al. 2008) and can mitigate flooding risk in urban areas (Dunn et al. 1995; Kentula at al. 1992). While they are primarily designed to improve water quality downstream, retention ponds themselves are subject to flashy hydroperiods (Bonilla-Warford and Zedler 2002), turbid water, high levels of soil phosphorus (Hogan and Walbridge 2007), contaminated sediments (Marsalek and Marsalek 1997), and invasion by non

1 Author for correspondence (bradley.herrick@wisc.edu)

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native plant species—conditions that are often inhospitable for native plant establishment and use by some wildlife (Bishop et al. 2000a, 2000b).

Retention ponds are designed to store water at stable depths, which reduces the energy of incoming flows and increases the accumulation of sediment-bound phosphorus and toxins. Non-vegetated areas of stormwater ponds are often colonized by non-native invasive plants such as Typha spp. (cattail), Phragmites australis (common reed), and Lythrum salicaria (purple loosestrife) (Schueler 1994; Moore et al. 1999). In particular, the invasive Typha × glauca, an F1 hybrid between the non-native and invasive Typha angustifolia L. (narrow-leaved cattail) and the native Typha latifolia L. (broad-leaved cattail) (Smith 1987), has been shown to expand under stable water levels and high nutrient inputs (e.g., Woo and Zedler 2002; Boers and Zedler 2008) making it difficult to increase native plant diversity (Boers et al. 2007).

Despite these disturbances, retention ponds are often planted with native wetland species in an effort to create or restore habitat in these artificial systems. Plant diversity can influence ecosystem properties such as productivity, decomposition rates, nutrient cycling (Engelhardt and Ritchie 2001), and resistance and resilience to disturbance (Hopple and Craft 2013). In addition, plant diversity is positively related to mycorrhizal and insect diversity (Knops et al. 1999). Pier et al. (2015) found that the diversity of the macroinvertebrate community increased after the introduction of 85 native plants to a newly created system of stormwater ponds. However, since constructed wetlands that include native plantings are rarely monitored (Balcombe et al. 2005), this study was undertaken to investigate the disturbance tolerance and competitive ability of eight native macrophytes planted in a newly constructed stormwater retention pond.

METHODS AND MATERIALS

In 2009, a 1.6 ha retention pond was constructed at the University of Wisconsin-Madison Arboretum in Madison, Wisconsin (Figure 1) to treat an average of one hundred and sixty-six million liters (135 acre-feet) of stormwater annually. The pond receives water from a 111 ha urban watershed comprised primarily of residential, commercial, and industrial land uses (University of Wisconsin- Madison Arboretum 2006). The pond edges were graded to insure a relatively consistent water level in the emergent zone (Figure 1). This design allowed the installation of native plants that are appropriate to shallow-water conditions. Plants were purchased from Wildlife Nurseries, Inc. in Oshkosh, Wisconsin approximately 87 miles north-northeast of the study site. All plants were grown from locally collected seed or root stock. In late April and early May of 2010, we planted ninety-six 6 ¥ 1.5 m plots within the pond’s emergent zone (Figure 2) with root stocks and bulbs of eight native wetland species: Acorus calamus L. (sweet flag), Bolboschoenus fluviatilis (Torr) Soják (river bulrush), Juncus effusus L. (common rush), Pontederia cordata (pickerel-weed), Sagittaria rigida Pursh (sessile- fruited arrow-head), Schoenoplectus acutus (C.C.Gmel.) Palla (hard-stem bulrush), Schoenoplectus tabernaemontani (C.C.Gmel.) Palla (soft-stem bulrush), and Sparganium eurycarpum Engelm. (common bur-reed). These species were selected because they are readily available, commonly found in natural and restored wetlands and are often included in species lists for constructed stormwater retention ponds. We randomly assigned each species to one-, three-, or six-species assemblages within plots at a density of 2.7 individuals per m2. Each unique monotype or assemblage was assigned to four replicate plots, which in turn were assigned at random within four nearly equal sections of the pond perimeter. We erected chicken-wire fence around the plots in the summer of 2010 to minimize goose herbivory.

In the fall of 2010, and again in 2016, we recorded species presence and percentage cover in all

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FIGURE 1. Retention pond study site at the University of Wisconsin-Madison Arboretum, Madison, Wisconsin. Emergent zone and submergent benches are shown in white. Eight native macrophyte species were planted randomly in ninety-six 6.Â¥ 1.5 m plots within the emergent zone. Image credit: Mark Wegener.

plots. Species percentage cover was recorded in six cover classes, adapted from the Daubenmire scale (Mueller-Dombois and Ellenberg 1974): 1 (0–5%), 2 (6–25%), 3 (26–50%), 4 (51–75%), 5 (76–95%), and 6 (96–100%). For analysis, the categorical cover classes were converted to continuous data by calculating the mid-point of each class. Means are based on the mid-point data. In 2015, the plantings were surveyed using a larger plot size of approximately 3x6 m, which provided a broader picture of the in situ emergent plant community that was not focused solely on the original planting area and allowed for an examination of where the emergent plant community became established over time. We present these results separately to provide additional context. For all years, we lumped species of bulrush—Bulboschoenus fluviatilis, Scirpus acutus, and Schoenoplectus tabernaemontani— as “Scirpus spp.†because it was difficult to discern between species. Scirpus spp. is treated as a single species for analysis purposes. We used paired t-tests to compare the percentage cover of each species between sampling years. Since the ecological management goal of the plantings was to increase native wetland plant diversity, no plots were left as unplanted controls due to the risk of colonization by invasive plants.

In addition to the eight macrophytes, five submergent species were planted in the spring of 2010 on three submerged benches at different water depths at each of six locations around the pond edge (Figure 1). The benches were two, three, and four feet below water level and consisted of a combined total area of 200.94 m2, 251.67 m2, and 299.43 m2, respectively. The five submergent species were: Elodea canadensis Michx. (elodea), Stukenia pectinata (L.) Börner (sago pondweed), Vallisneria americana Michx. (water celery), Nymphaea odorata Aiton (white water-lily), and Nelumbo lutea Willd. (American lotus). In order to provide good soil contact and encourage rooting into the substrate (Eleuterius 1975), tubers or seeds were put into a cotton mesh bag with several small rocks and tossed from shore into the appropriate water depth. The bags sank to the pond bottom.

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FIGURE 2. Study site at the time of planting in spring 2010. Typha spp. is absent. Photo by Brad Herrick.

RESULTS

Within one year after planting, Typha spp. was found in 93% of plots, and by 2016 it had invaded all plots (Figures 3 and 4). After six years, all planted species decreased in frequency except Pontederia cordata, which showed slight increases between the original planting frequency and subsequent survey events (Figure 4). In 2016, plants with the two highest frequencies were P. cordata and Scirpus spp., at 66% and 53%, respectively. Juncus effusus had the lowest plot frequency, being found in only 2% of plots. Sparganium eurycarpum was not detected in surveys after the original planting.

By visual estimation, the clear majority of Typha spp. in our plots consisted of T.× glauca, although T. angustifolia was also present (Galen Smith, personal correspondence). We saw no individuals of T. latifolia in our plots or anywhere within the retention pond. Hereafter, Typha spp. refers only to T.× glauca and

T. angustifolia. Between 2010 and 2016, the percentage cover of all planted species except Pontederia cordata significantly decreased (p < 0.05, Figure 4). Pontederia cordata decreased slightly from 8% to 7% but this decrease was not statistically significant (p = 0.49). Typha spp. significantly increased from 26% to 96% (p < 0.0001) over the same time period.

By 2016, 80% of the plots had only one or two planted species present, and

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FIGURE 3. Study site at the time of sampling in the fall of 2016. Typha spp. have invaded the emergent zone. Photo by Brad Herrick.

an additional 14% of plots were devoid of any planted species. Furthermore, there were no plots by this time with more than three planted species.

Pontederia cordata was the only planted species to increase in plot frequency during the study period; all others decreased (Figure 4). However, all species were present in at least a few plots with the exception of Sparganium eurycarpum (Figure 4). In addition, the abundance of P. cordata remained relatively stable while all others decreased (Table 1). None of the planted species achieved greater than 9% cover during the study period, likely due to the rapid establishment and increase in cover by Typha spp. (Table 1). Planting a higher

TABLE 1. Mean percentage cover of planted native species and invasive Typha spp. in 2010 and 2016 (± standard deviation) per 6 ¥ 1.5 m plots. Asterisks represent significant difference between years at a􀀀=0.05.

Mean Percentage Cover Species 2010 2016 Acorus calamus 3.85 ± 0.92 0.45 ± 0.17* Juncus effusus 0.23 ± 0.07 0.05 ± 0.04* Pontederia cordata 8.05 ± 1.53 6.8 ± 1.25 Sagittaria rigida 0.94 ± 0.34 0.1 ± 0.05* Scirpus spp. 5.05 ± 0.9 2.21 ± 0.48* Sparganium eurycarpum 0 0 Typha spp. 26.48 ± 3.00 96.46 ± 0.35*

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FIGURE 4. Frequency of planted species and Typha spp. in 6 x 1.5 m plots at initial planting in 2010 and 2016.

density of native plants, especially P.cordata and Scirpus spp., may have improved diversity long-term.

2015 data

All planted native species except Sparganium eurycarpum were still present in 2015 (Figure 5). Acorus calamus was found in 17% of the plots in 2015, as

FIGURE 5. Frequency of planted species and Typha spp. at the time of planting and in 2015 (6 x 3 m plots).

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compared with 14% in 2016. Juncus effusus and S. rigida were found in 13% and 14% of plots, respectively, in 2015, as compared with 2% and 4% in 2016. Although these three species exhibited increased frequency with a slightly larger sampling unit, percentage cover was similar to that recorded in 2016. Finally, in 2015, Pontederia cordata and Scirpus spp. were found in 92% and 82% of the plots, respectively, as compared to 66% and 53% in 2016, but with mean cover of just 12% and 7%, respectively.

Planted Submergent Species

Likely due to poor water clarity, none of the planted submergent species were observed in 2016, except for a few scattered patches of Nymphaea odorata.

DISCUSSION

We expected Sparganium eurycarpum to perform well in a retention pond setting, since it is a robust plant and is often found with Scirpus spp. and Typha spp. under flooded conditions. Additionally, it has been shown to produce a high amount of above-ground biomass and to reduce competition when planted with other native species (Chapman et al. 2013). However, in our study it was never recorded in surveys after the initial planting. While our data suggest that S. eurycarpum may not do well in stormwater retention ponds, we believe that the complete failure of S. eurycarpum to establish anywhere within the entire pond as opposed to other planted species obtained from the same nursery and geographic region of Wisconsin, suggests another alternative: that we may have received bad root stock for S. eurycarpum or that we planted it incorrectly.

While the invasion of Typha × glauca was not surprising, due to its establishment throughout the watershed, the likely high levels of nutrients in the pond water and open, shallowly flooded conditions along the pond edge created ideal conditions for invasion and likely contributed to the high abundance after only one year. Typha angustifolia increased over time in five constructed wetlands in Connecticut (Moore et al. 1999). In a study of a Typha invasion in a Lake Michigan coastal wetland, Mitchell et al. (2011) found that Typha density (primarily

T.× glauca and T. angustifolia) increased significantly during the first ten years of invasion. Additionally, experiments show that T.× glauca can reduce native macrophyte biomass and diversity (e.g., Woo and Zedler 2002; Hall and Zedler 2010; Larkin et al. 2012), and Typha litter alone can reduce the abundance and diversity of native plants after only one year (Farrer and Goldberg 2009). In this study, assemblages of one, three, and six species did not provide any short (one year) or longer-term (six years) competitive advantage. Additionally, the competitive dominance of Typha spp. does not seem to have been hindered by the native plantings. Pontederia cordata and Scirpus spp. were more abundant outside of the plots in slightly deeper water. In addition, many individuals of Scirpus spp. (likely Bolboschoenus fluviatilis) were found outside of the plots in shallower water,

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closer to the shoreline. Although these two groups of species were not represented in the 2016 plot data, it seems that they were able to coexist with Typha spp. at different water depths. The competitive advantage of many emergent species has been shown to vary substantially with subtle changes in water depth (e.g., Olson and Doherty 2012). Whether these species were excluded from other areas due to competition with Typha spp. or gradually shifted to more optimal sites is difficult to determine. Pontederia cordata can grow well in deeper water compared to most Typha spp., (Nichols 1999) and B. fluviatilis has been shown to establish under a wide range of water depths (Galatowitsch and van der Valk 1996) but produce the most biomass under shallowly flooded conditions (Hudon 1997). These observations also highlight one of the limitations of a priori study designs in dynamic systems. As the system changed over time, species were able to establish outside of our study plots, and that shift was difficult to capture within our predefined study plots. This is not surprising as stormwater retention ponds have been shown to have flashy hydroperiods (Bonilla-Warford and Zedler 2002), and it is difficult to predict such dynamics during the planting phase. Therefore, we recommend that future efforts to establish native plant communities within stormwater retention ponds consider planting native species over a broader range of water depths and planting species adapted to different water depths.

CONCLUSION

Despite the overwhelming invasion by Typha spp. shortly after the planting of the native species, all but one of the native species persisted, albeit most of them at low frequency. Pontederia cordata and Scirpus spp. showed some resistance to invasion, and the cover of P. cordata, although low, was stable over six years. This study shows that while Typha spp. may readily invade the emergent zone of artificial retention ponds, planting native species during pond creation may increase diversity, even without follow-up management. We recommend the use of

P. cordata and Scirpus spp. as good options to increase native plant richness in stormwater retention ponds. ACKNOWLEDGMENTS

We thank Dr. Joy Zedler for the idea to incorporate research plots into the redesign of the retention pond as well as for guidance in designing the study. Numerous volunteers assisted in planting, and Charlie Tucker assisted with erecting the goose fence. Thanks to Mark Wegener for the map design and Jessica Ross for data analysis. Comments and suggestions by Dr. Michael Huft and two anonymous reviewers greatly improved the manuscript.

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