RIPARIAN INVASION by LONICERA MAACKII DRIVES SHIFTS in AQUATIC BIOTA and ECOSYSTEM PROCESSES Di

TERRESTRIAL-AQUATIC CONNECTIONS: RIPARIAN INVASION BY LONICERA

MAACKII DRIVES SHIFTS IN AQUATIC BIOTA AND ECOSYSTEM PROCESSES

Dissertation

Submitted to

The College of Arts and Sciences of the

UNIVERSITY OF DAYTON

In Partial Fulfillment of the Requirements for

The Degree of

Doctor of Philosophy in Biology

Rachel Elizabeth McNeish, M.S.

Dayton, Ohio

May, 2016

TERRESTRIAL-AQUATIC CONNECTIONS: RIPARIAN INVASION BY LONICERA

MAACKII DRIVES SHIFTS IN AQUATIC BIOTA AND ECOSYSTEM PROCESSES

Name: McNeish, Rachel Elizabeth

APPROVED BY:

______Ryan W. McEwan, Ph.D. Faculty Advisor

______M. Eric Benbow, Ph.D. Committee Member

______Alburt J. Burky, Ph.D. Committee Member

______P. Kelly Williams, Ph.D. Committee Member

______Margaret Carreiro, Ph.D. Committee Member

______Karolyn M. Hansen, Ph.D. Committee Member

ABSTRACT

TERRESTRIAL-AQUATIC CONNECTIONS: RIPARIAN INVASION BY LONICERA

MAACKII DRIVES SHIFTS IN AQUATIC BIOTA AND ECOSYSTEM PROCESSES

Name: McNeish, Rachel Elizabeth University of Dayton

Advisor: Dr. Ryan W. McEwan

Invasive species are of global importance due to their impacts on ecological communities, habitat structure, native community dynamics, and ecosystem processes.

Scientists and conservation managers are increasingly focusing on the biological impacts of invasive species and devising management practices that emphasize the health of ecosystems based on measured biological processes. Lonicera maackii is a highly successful invasive shrub in forests of eastern and Midwestern North America. We investigated how riparian invasion of L. maackii influenced (1) the availability of in- stream leaf litter resources, algal growth, above stream canopy cover, and light available to the stream, (2) the functional and taxonomic diversity and community composition of aquatic macroinvertebrate communities, (3) the effects of L. maackii on throughfall chemistry. In summary, the removal of an invasive riparian shrub influenced the timing, deposition, quality, and abundance of leaf litter habitat into a headwater stream, ostensibly driving bottom-up effects on aquatic primary producer biomass and the macroinvertebrate community. Patterns in macroinvertebrate community and functional

iii trait dynamics were influenced by seasons and the L. maackii riparian forest. These findings suggest that functional traits were driven by life history strategies linked with seasonal patterns in temperature and food resources that are also influenced by L. maackii riparian forests. In addition, riparian L. maackii has the potential to alter nutrient subsidies during rain events that enter aquatic systems as throughfall, and suppress stream algal growth early in the growth season, impacting nutrient cross-system subsidies and one of the basal food resources in aquatic systems. Based on these findings we have developed a predictive framework for understanding how this terrestrial invasive shrub influences aquatic ecosystems.

ACKNOWLEDGEMENTS

Special thanks to Dr. Ryan McEwan for his commitment as my advisor and mentor during my dissertation. Additional thanks to all the University of Dayton Biology graduate students, undergraduate community, and the University of Dayton Graduate

School for their support. I would like to further express my appreciate to the Centerville-

Washington Park District, the Bellbrook-Sugarcreek Park District, and the Dayton Five

Rivers MetroParks for allowing me to conduct my research in their park and stream sites.

TABLE OF CONTENTS

ABSTRACT ...... iii ACKNOWLEDGEMENTS ...... v LIST OF ILLUSTRATIONS ...... x LIST OF TABLES ...... xii LIST OF ABBREVIATIONS AND NOTATIONS ...... xiii CHAPTER 1: A REVIEW ON THE INVASION ECOLOGY ...... 1

ABSTRACT ...... 1

INTRODUCTION ...... 2

LONICERA MAACKII ADVANTAGEOUS LIFE HISTORY TRAITS ...... 4 Dispersal Mechanisms ...... 4 Rapid Growth and Environmental Plasticity ...... 5 Phenology ...... 7 Allelopathy and Resistance to Herbivory ...... 8

LONICERA MAACKII INVASION IMPACTS AT VARYING ECOLOGICAL SCALES...... 11 Community-Scale Impacts ...... 11

LONICERA MAACKII INVASION AND LANDSCAPE ECOLOGY ...... 18

LONICERA MAACKII IMPACTS ON ECOSYSTEM PROCESSES ...... 20

MANAGEMENT AND RESTORATION OF LONICERA MAACKII INVADED HABITATS...... 22 Detection and Management ...... 22 Lonicera maackii management impacts on flora and fauna ...... 25

CONNECTIONS TO INVASION THEORY ...... 26

FUTURE DIRECTIONS AND CONSIDERATIONS ...... 29

ACKNOWLEDGEMENTS ...... 31

LITERATURE CITED ...... 32

FIGURE LEGENDS ...... 54

FIGURES ...... 55 CHAPTER 2: REMOVAL OF THE INVASIVE SHRUB LONICERA MACKII ...... 58

ABSTRACT ...... 58

INTRODUCTION ...... 59

METHODS ...... 61 Study Site ...... 61 Experimental Lonicera maackii Removal ...... 62 Leaf Litter Accumulation ...... 62 Benthic Algal Biomass and Above Stream Canopy Cover ...... 63 Macroinvertebrate Community ...... 63 Statistical Analyses ...... 64

RESULTS ...... 65 In-stream Leaf Material ...... 65 Canopy Cover and Benthic Algal Biomass ...... 66 Macroinvertebrate Density ...... 66

DISCUSSION ...... 67 In-stream Organic Matter ...... 67 Macroinvertebrate Density and Benthic Algal Biomass ...... 68

ACKNOWLEDGMENTS ...... 70

REFERENCES ...... 71

TABLES ...... 80

FIGURE LEGENDS ...... 83

FIGURES ...... 85

SUPPLEMENTAL MATERIALS ...... 91 CHAPTER 3: TERRESTRIAL-AQUATIC LINKAGES INFLUENC BENTHIC ...... 92

ABSTRACT ...... 92

INTRODUCTION ...... 93

METHODS ...... 97 Study Site ...... 97 Benthic Macroinvertebrate Sampling ...... 97 Ambient Conditions ...... 98

vii

Statistical Analyses ...... 99

RESULTS ...... 103 Macroinvertebrate Taxonomic and Functional Diversity Metrics ...... 103 Macroinvertebrate Community Structure and Functional Trait responses ...... 103 Ambient Conditions ...... 105

DISCUSSION ...... 106

ACKNOWLEDGEMENTS ...... 109

REFERENCES ...... 110

TABLES ...... 124

FIGURE LEGENDS ...... 134

FIGURES ...... 135

SUPPLEMENTAL MATERIALS ...... 139

SUPPLEMENTAL REFERENCES ...... 143

SUPPLEMENTAL FIGURES ...... 157 CHAPTER 4: TERRESTRIAL-AQUATIC CONNECTIONS: RIPARIAN ...... 166

ABSTRACT ...... 166

INTRODUCTION ...... 167

METHODS ...... 170 Study Sites ...... 170 Throughfall Experimental Design ...... 171 Lonicera maackii Effect on Stream Algal Growth ...... 174 Statistical Analyses ...... 175

RESULTS ...... 177 Throughfall Chemistr: ...... 177 Lonicera maackii Effect on Stream Algal Growth ...... 179

DISCUSSION ...... 180

ACKNOWLEDGEMENTS ...... 184

LITERATURE CITED ...... 186

TABLES ...... 196

FIGURE LEGENDS ...... 198

FIGURES ...... 200

viii

SUPPLEMENTAL MATERIALS ...... 206

SUPPLEMENTAL FIGURES ...... 217

LIST OF ILLUSTRATIONS

Figure 1.1: Lonicera maackii fall fruit production...... 54

Figure 1.2: Scanning electron microscope images ...... 54

Figure 1.3: Predictive framework for Lonicera maackii impacts...... 54

Figure 2.1: Conceptual framework for the effects of Lonicera maackii...... 83

Figure 2.2: Mean (±SE) in-stream leaf litter organic matter ...... 83

Figure 2.3: Mean (±SE) in-stream leaf organic matter for dominant leaf genera ...... 83

Figure 2.4: Mean (±SE) in-stream leaf litter organic matter for the most dominant ...... 84

Figure 2.5: Mean (±SE) monthly percent canopy cover ...... 84

Figure 2.6: Mean (±SE) monthly benthic algal biomass and macroinvertebrate density..84

Figure 3.1: Macroinvertebrate density for L. maackii and removal reaches ...... 134

Figure 3.2: Taxonomic community dynamics ...... 134

Figure 3.3: Functional community dynamics ...... 134

Figure 3.4: Mean macroinvertebrate functional feeding group relative abundance...... 134

Supplemental Figure 3.1: Taxonomic richness and diversity...... 157

Supplemental Figure 3.2: Functional diversity metrics ...... 157

Supplemental Figure 3.3: Taxonomic and functional community dynamics ...... 157

Supplemental Figure 3.4: Mean macroinvertebrate functional feeding group ...... 157

Supplemental Figure 3.5: Mean above stream canopy cover ...... 157

Supplemental Figure 3.6: Mean light at the surface of the stream ...... 158

Supplemental Figure 3.7: Ambient nutrient and total suspended solid dynamics ...... 158

Figure 4.1: Taylorsville MetroPark throughfall field site ...... 198

Figure 4.2: Mean (±SEM) throughfall and rain water volume and pH ...... 198

Figure 4.3: Mean (±SEM) total carbon and organic carbon ...... 198

Figure 4.4: Mean (±SEM) dissolved organic and dissolved inorganic carbon ...... 198

Figure 4.5: Spring mean (±SEM) algal grown (standing stock chlorophyll a) ...... 199

Figure 4.6: Autumn mean (±SEM) algal grown (standing stock chlorophyll a) ...... 199

Supplemental Figure 4.1: Mean (±SEM) total and dissolved nitrogen ...... 217

Supplemental Figure 4.2: Mean (±SEM) nitrite, nitrate, and ammonia ...... 217

Supplemental Figure 4.3: Mean (±SEM) total orthophosphate and SRP ...... 217

Supplemental Figure 4.4: Spring mean (±SEM) algal biomass ...... 217

Supplemental Figure 4.5: Spring mean (±SEM) algal growth to biomass ratios ...... 218

LIST OF TABLES

Table 2.1: Statistical analyses of leaf litter types, canopy cover, algal biomass ...... 80

Table 2.2: Linear regression statistical results for leaf litter accumulation...... 81

Table 2.3: Repeated measures ANOVA for in-stream top leaf litter genera ...... 82

Supplemental Table 2.1: Statistica test for dominant the top five leaf genera ...... 91

Table 3.1: Description of macroinvertebrate functional traits ...... 124

Table 3.2: Community functional richness for both stream reaches ...... 126

Table 3.3: Functionally relevant indicator macroinvertebrate taxa ...... 127

Table 3.4: Functionally relevant indicator taxa within each season ...... 130

Supplemental Table 3.1: Description of functional trait states ...... 139

Supplemental Table 3.2: Macroinvertebrate taxonomic and abundance list ...... 145

Supplemental Table 3.3: Taxonomic and functional diversity metrics...... 150

Supplemental Table 3.4: ADONIS multiple permutation result ...... 152

Supplemental Table 3.5: Functionally relevant macroinvertebrate taxa ...... 153

Table 4.1: Percent difference of nutrient concentration...... 196

Supplemental Table 4.1: Statistical results for throughfall nutrient concentration ...... 206

Supplemental Table 4.2: Statistical results for throughfall nutrient deposition...... 209

Supplemental Table 4.3: Statistical results for stream algal growth respons ...... 212

Supplemental Table 4.4: stream algal growth response when exposed to control ...... 215

xii

LIST OF ABBREVIATIONS AND NOTATIONS

°C Degrees Celcius

× Times

µm Micrometers

> Less Than

< Greater Than

~ Approximately

% Percent

4-D Four Dimensions

AFDM Ash-Free-Dry-Mass

AH Amur Honeysuckle

ANOVA Analysis of Variance

BEL Box Elder Leaves

BO Black Oak Park

CF Collector-Filterer

CG Collector-Gatherer cm centimeter

COM Coarse Organic Matter d day

xiii

D Detritivore

DEB Division of Environmental Biology

ESM Electronic Supplemental Material df Degrees of Freedom

DIC Dissolved Inorganic Carbon

DM Dry Mass

DN Dissolved Nitrogen

DOC Dissolved Organic Carbon

EICA Evolutionary Increased Competitive Ability

EPA Environmental Protection Agency

ERH Enemy Release Hypothesis

ET Evapotranspiration

FD Functional Diversity

FDis Functional Dispersion

FEve Functional Evenness

FFG Functional Feeding Group

FP Fecher Park

FRic Functional Richness

H Herbivore

H1 Hypothesis One

H2 Hypothesis Two

H3 Hypothesis Three

H4 Hypothesis Four

xiv hr Hour

LM Lonicera maackii m2 Concordance or similarity between two ordincations

M Molarity m meter mm millimeter n Sample size

NH3-N Ammonia-Nitrogen

NISC National Invasive Species Council

NMDS Non-Metric Multidimensional Scaling

NO2-N Nitrite-Nitrogen

NO3-N Nitrate-Nitrogen

NSF National Science Foundation

NWH Novel Weapons Hypothesis

OH Ohio

OM Organic Matter

P Predator

PR Photosynthetic Rate

RGR Relative Growth Rate rmANOVA Repeated Measures Analysis of Variance sp. Singular for species unknown spp. Plural for species unknown

SRP Soluble Reactive Phosphorus

Suppl. Supplemental

TC Total Carbon

TM Thematic Mapper

TN Total Nitrogen

TOC Total Organic Carbon

USA United States of America

χ2 Chi-square Value

xvi

CHAPTER 1: A REVIEW ON THE INVASION ECOLOGY OF AMUR

HONEYSUCKLE (LONICERA MAACKII), A CASE STUDY OF ECOLOGICAL

IMPACTS AT MULTIPLE SCALES

ABSTRACT:

Invasive species are of global importance due to their impacts on ecological communities, habitat structure, native community dynamics, and ecosystem processes and function. Scientists and conservation managers are increasingly focusing on the biological impacts of invasive species and devising management practices that emphasize the health of ecosystems based on measured biological processes. Lonicera maackii is a highly successful invasive shrub in forests of eastern North America. The scientific literature surrounding this species has grown in the past several decades as researchers have investigated Lonicera maackii impacts across multiple ecological scales. In this review we synthesized literature on (1) the key traits related to this species’ invasion success, (2) the impacts this invasive has at various ecological scales, (3) the outcomes of restoration efforts, and (4) connections to invasion ecology theories. Lonicera maackii impacts are complex and vary across ecosystems and spatial scales, and we report findings from studies demonstrating a wide range of effects on species composition, community structure, ecosystem function, and successional trajectories. We end by providing a working ecological framework that may help guide future research and conservation efforts.

INTRODUCTION:

Invasive species are considered to be one of the most important threats to biodiversity and ecosystem function across the globe (Ruesink et al. 1995, Wilcove et al.

1998), with the economic impacts estimated as high as ~ $120 US billion/year (Pimentel et al. 2005). Invasive plants have prompted a great deal of interest with particular focus on quantifying the impacts invaders have on plant communities (other articles this special issue, Callaway and Ridenour 2004, Collier et al. 2002, Crooks 2002, Levine 2000).

Effects ascribed to invasive plants include modification of habitat structure, changing ecosystem processes, and decreasing native biodiversity (Hejda et al. 2009). Many states in the USA have rules surrounding invasive plants and efforts to organize activities have become widespread and include cooperative weed management areas and invasive plant boards.

From both a management and scientific perspective there is a clear need for identifying invasive species that are thoroughly studied and have well-established, empirical evidence of ecosystems effects (Gurevitch and Padilla 2004, Sagoff 2005). In one example, Tamarix spp. (Saltcedar; L.) has been classified as an invasive in southeastern USA and previous research suggested this plant reduced water availability through increased evapotranspiration rates (ETR; Thomas 1963). More recent studies have indicated that ETR of riparian invaded Tamarix spp. forests is the same regardless of the density of Tamarix spp., and the impacts of Tamaraix spp. varies between sites

(Stromberg and Chew 2002, Stromberg et al. 2009). Invasion implies high abundance of a particular problematic species, and while this can be easily measured in the field (and is visually obvious), some have argued that exotic species have been too quickly

“demonized” or “vilified” as invasive before their effects on ecosystems are fully understood (Borrell 2009, Davis 2011, Gurevitch and Padilla 2004, Stromberg and Chew

2002, Stromberg et al. 2009). The native vs. exotic paradigm has started shifting, and scientists continue to focus on quantifying the biological impacts of particular species

(Chew and Hamilton 2010, Thompson and Davis 2011) to inform management practices that positively influence the ecosystem’s health based on measured biological processes

(Borrell 2009, Stromberg et al. 2009).

Lonicera maackii (Amur honeysuckle; (Rupr.) Maxim.) is a highly successful invasive shrub in the Eastern Deciduous Forest. Concern over this species is widespread and, in fact, legal limitations exist in some states (USDA 1999). Initial research on this species provided an outstanding background on the invasion biology of L. maackii

(Deering and Vankat 1999, Gould and Gorchov 2000, Hutchinson and Vankat 1998,

Luken and Goessling 1995, Luken 1988, Luken et al. 1995). Since then, an extensive body of empirical evidence has been developed that documents this plant’s (1) suite of invasive traits, (2) success as regulated by landscape characteristics, and (3) significant impacts at multiple ecological scales. This burgeoning scientific literature makes L. maackii an ideal species to serve as a model of invasion impact; however, there has not been a recent consolidation of the empirical evidence.

In this manuscript we synthesize the available literature to provide a framework for understanding the ecological effects of L. maackii and to help direct future research and management practices. We begin by describing how anthropogenic activities and the life-history traits of L. maackii contributed to its invasion success in the USA. Working from this basis, we then summarize findings from empirical studies that identify key traits

3 that regulate the invasion of L. maackii into vulnerable habitats and describe the impacts this invasive has at various ecological scales. We end by connecting L. maackii to invasion ecology theories and identifying directions for future research that will strengthen and advance plant invasion biology.

LONICERA MAACKII ADVANTAGEOUS LIFE HISTORY TRAITS:

Dispersal Mechanisms:

Long-distance dispersal and propagule pressure are key characteristics that contribute to the success of invasive plant populations (Davies and Sheley 2007, Gosper et al. 2005). For example, Microstegium vimineum (Japanese stilt grass; (Trin.) A.

Camus) propagule pressure was found to range from 556.6 to 144.5 per m2 in riparian and upland forests respectively (Eschtruth and Battles 2011). Heterotheca latifolia

(Camphor-weed; Buckey) dispersal was found to be related to wind patterns in the

Georgia piedmont area (USA) with expansion increasing at a rate of 3 miles per year

(Plummer and Keever 1963). Arguably, anthropogenic activities - especially from the industrial revolution and globalization eras - have been the most effective dispersal agents, spreading invasives across oceans and continents (Meyerson and Mooney 2007).

The initial introduction of Phragmites australis (common reed (Cav.) Trin. ex Steudel) to the USA is believed to have occurred at coastal ports along the Atlantic ocean, with dispersal of this invasive promoted by the use of ship ballast to fill marsh sites used for railroad development (Saltonstall 2002). High propagule pressure combined with long distance dispersal vectors and anthropogenic activities are a potent combination for the success of invasive plants.

Lonicera maackii is a good model of long distance dispersal and abundant propagule production in invasive plants. These shrubs have large fruiting events (Fig. 1a) and produce seeds that can germinate in various light, temperature stratification, and soil conditions (Hidayati et al. 2000, Luken and Goessling 1995). We have also documented presence of L. maackii berries submerged in streams with and without L. maackii in the riparian forest (Fig. 1b, pers. observation) – suggesting stream corridors are vectors for the spread of L. maackii propagules. Bartuszevige and Gorchov (2006) found some native birds (e.g. Turdus migratorius; American robin) eat and disperse viable L. maackii seeds via defecation. Castellano and Gorchov (2013) found that 68% of L. maackii seeds were still viable after passing through the intestinal system of Odocoileus virginianus

(white-tailed deer). Although the fruits of this species are a poor source of nutrition for wildlife (Ingold and Craycraft 1983), these animals contribute to the long distance dispersal of L. maackii, and may specifically support dispersal to edge habitats

(Bartuszevige and Gorchov 2006, Castellano and Gorchov 2013).

Rapid Growth and Environmental Plasticity:

Rapid growth and plasticity in response to changing environmental conditions are characteristic traits of successful invasive species. For instance, Polygonum perfliatum

(mile-a-minute; L.) is an invasive Asian vine well known to grow quickly, as indicated by its common name (Oliver 1996). Fogarty and Facelli (1999) demonstrated the invasive

European shrub Cytisus scoparius (scotch broom: (L.) Link) grows faster than, and successfully outcompetes, South Australian natives (Acacia verniciflua A. Cunn, A. myrtifolia (SM.) Wild, and Hakea rostrata F. Muell. Ex Meisn.). A Hawaiian study that surveyed over 60 invasive and native plants found that invasives spent less energy

5 producing leaves and had greater specific leaf area, CO2 assimilation, and N and P levels than natives (Baruch and Goldstein 1999). These results also indicate that invasive plants have phenotypic plasticity, which in turn, may facilitate their success in novel habitats.

Invasive plants are commonly thought to grow faster than natives, this trait is usually accompanied by the plant’s ability to utilize resources well with greater phenotypic and genetic plasticity than natives (Daehler 2003).

Lonicera maackii has a highly competitive growth pattern, which is an important phenotypic characteristic that helps facilitate its success in new habitats. This woody shrub produces numerous stem shoots and grows rapidly as an immature stem, then shifts resource allocation toward height growth and reproduction in its mature stage (Deering and Vankat 1999). When up-right stems are clipped, they re-sprout readily (Deering and

Vankat 1999, Luken and Mattimiro 1991) which adds to this plants already “bushy” appearance and contributes to a dense L. maackii canopy that decreases light availability to the herb layer. This plant is also known to exhibit growth plasticity in different habitats (Luken et al. 1997b, Luken et al. 1995). Compared to Lindera benzoin

(Spicebush; (L.) Blume), L. maackii was found to successfully utilize a range of light levels (1%, 25%, and 100% of full sun photosynthetic photon flux density) more effectively in terms of growth and photosynthesis, and exhibit higher branch plasticity and stomatal density (Luken et al. 1997b). Seedling establishment can occur in a variety of light conditions and shrubs located in open habitats produce significantly greater numbers of fruits than in shaded habitats (Lieurance 2004, Luken and Goessling 1995,

Luken and Thieret 1996). Lonicera maackii’s net primary productivity and aboveground biomass in open, high light habitats is substantially greater compared to shrubs located in

6 low light woodland communities (Lieurance 2004, Luken 1988). The unique growth plasticity of L. maackii leaves and branches across light environments provides this shrub a competitive edge in low light habitats.

Phenology:

Invasive plant species phenology have been shown to vary compared to natives in the invaded habitat. Berberis thunbergii (Japanese barberry; DC.), an invasive in the

Eastern Deciduous Forest, leafs out nearly one month earlier compared to a native shrub and demonstrated significantly greater photosynthetic capacity compared to native plant species (Xu et al. 2007). An invasive biennial herb Alliaria petiolata (garlic mustard;

(M.Bieb.) Cavara & Grande) has a competitive edge in early spring since its leaves emerge earlier than native herbs and achieve high photosynthetic rates (Myers and

Anderson 2003). Ailanthus altissima (tree-of-heaven (Mill.) Swingle) is another invasive plant that leafs out in early spring, stores high concentrations of photosynthate in leaves and stems, and is efficient at photosynthesis, making this plant highly competitive early in the growing season (Fryer 2010). A combination of early leaf emergence and efficient photosynthetic processes are important characteristics that promote the success and spread of invasive plant species.

Lonicera maackii has an extended growing season in comparison to other plants.

Leaf development and expansion occurs two to three weeks before, and the final leaf abscission is later than, native flora (McEwan et al. 2009a). Another competitive edge is the leaves are freeze resistant and are still present on shrubs during early winter (Fig. 1c;

McEwan et al. 2009a). As stated earlier, L. maackii propagule production is copious in the fall and berry formation begins in early fall and will stay attached well into winter

(pers. observation; Fig 1a). Massive flower production occurs in mid-spring (Luken and

Thieret 1996), and these flowers are a resource for pollinators (Goodell et al. 2010,

McKinney and Goodell 2011). The combination of previously described growth characteristics and phenology gives L. maackii a competitive edge that increases its ability to outcompete native flora.

Allelopathy and Resistance to Herbivory:

Some invasive plant species have been shown to exhibit biochemical effects on predators through the production of allelochemicals (Theoharides and Dukes 2007).

These chemicals are secondary plant compounds that typically suppress the growth, survivorship, and reproductive capabilities of competitors (Hierro and Callaway 2003).

A meta-analysis of common invasive plants in China found 75% of the most noxious invasives displayed evidence of allelopathic effects (Ni et al. 2012). Stinson et al. (2006) found support that the antifungal phytochemistry of A. petiolata can indirectly disrupt the mutualistic relationship between hardwood trees and mycorrhizal fungi, which can result in a reduction of tree seedling regeneration in forest communities. Allelochemicals also impact microbial communities (Callaway and Ridenour 2004), suggesting invasive plants may alter these communities and nutrient cycling in terrestrial and aquatic systems.

A series of studies have been conducted that established evidence of L. maackii allelopathic effects on plants. Cipollini et al. (2008b) identified 13 secondary metabolites present in L. maackii leaves and demonstrated that luteolin and apigenin derivatives were the main allelopathic chemicals present in leaf extracts. Lonicera maackii root and shoot extracts have been shown to reduce germination of several native herbaceous plants including Impatiens capensis (jewelweed Meerb.), and Anemone virginiana (tall

8 anemone L.; Cipollini and Flint 2013, Dorning and Cipollini 2005, McEwan et al. 2010).

McEwan et al. (2010) also demonstrated L. maackii leaf and fruit extracts had differential effects on four grass and forb species. Lonicera maackii fruit extracts suppressed all forb and grass seed germination (Festuca arundinaceae (tall fescue Schreb), Impatiens wallerana (dwarf white baby (Hook.) F.), Coreopsis lanceolata (lanced-leaves coreopsis

L.), and Poa pratensis (Kentucky bluegrass L.)) whereas leaf extracts only suppressed seed germination of I. walleriana. Other experimental studies have focused on the effects of L. maackii extracts on the morphology, fecundity, reproduction, and growth on

Arabidopsis thaliana (Arabidopsis (L.) Heynh.; Cipollini and Dorning 2008, Cipollini et al. 2012). Arabidopsis thaliana grown in L. maackii conditioned soils exhibited decreased survivorship and an 11 day delay in flower production; however, seed production and mature leaves were larger and more abundant compared to those grown in unconditioned soils (Cipollini and Dorning 2008). Schradin and Cipollini (2012) conducted a study to identify positive or negative feedbacks on L. maackii growth compared to the native honeysuckle Diervilla lonicera (Northern bush honeysuckle Mill.) grown in different soil types, and whether growth patterns were due to soil abiotic or biotic conditions. Soil type (sandy and loamy) and conditioning and soil biota influenced the sign and strength plant-soil feedbacks. For example, L. maackii grew ~ 2 × more in its own (L. maackii) conditioned, sandy soil compared to unconditioned soil, resulting in a positive feedback. Soil sterilization resulted in negative feedback, decreasing L. maackii growth in sandy soil. Alternatively, there was negative feedback on D. lonicera growth when grown in its own soil, with feedback effects neutralized when soil was sterilized. Cipollini et al. (2012) also conducted a study to determine if L. maackii and

9 the invasive herb A. petiolata had similar allelopathic effects on A. thaliana. Results indicated that L. maackii leaf extracts reduced reproduction and growth of A. thaliana, whereas there were few effects from A. petiolata. Bauer et al. (2012) cautioned that allelopathic impacts are context dependent based on their field study that suggested soil microorganisms and different native species are important factors that influence net allelopathic effects. These collective findings suggest the chemistry of L. maackii can have an impact on native plant communities and strong evidence suggests allelopathic effects in some experimental settings; however, more research is needed to understand how allelopathic effects are manifested in the invasion biology of L. maackii in the field.

Evidence suggests that L. maackii has some traits that confer resistance to herbivory in its introduced range. McEwan et al. (2009b) investigated the anti-herbivory potential of L. maackii on the invasive generalist caterpillar Lymantria dispar (Gypsy moth; Erebidae). Caterpillar relative consumption and growth rates were significantly reduced when provided with only L. maackii as a food resource. The development time of the caterpillars was also inhibited when fed L. maackii, and all larva died before molting to the next stage. Cipollini et al. (2008b) found indications that the moth generalist Spodoptera exigua (beet armyworm; Noctuidae) had reduced feeding when given food sources made with L. maackii leaf extracts compared to control food.

Lieurance and Cipollini (2013a) conducted a study to identify how juvenile L. maackii shrubs respond to herbivory under environmental stress conditions. They found at low light and nitrogen levels L. maackii tolerance and resistance to herbivory was still high.

These findings suggest L. maackii has strong resistance to natural predators which may be another competitive strategy for this invasive species.

LONICERA MAACKII INVASION IMPACTS AT VARYING ECOLOGICAL SCALES:

Community-Scale Impacts:

Effects on Plant Communities:

As invasive species encroach and proliferate in habitats, they modify substrate, resources, and ecosystem processes, which can result in substantial changes in plant and animal communities taxonomically and functionally (Shea and Chesson 2002, Randall

1996, Vilà et al. 2011). Empirical evidence suggests L. maackii invasion has substantial negative impacts on native plants. Forests with this invasive shrub have significantly lower herb fecundity, fitness, and growth (Gould and Gorchov 2000, Miller and Gorchov

2004). Collier et al. (2002) found that herb abundance and richness significantly decreased under L. maackii shrubs compared to away locations, and with stands that had longer L. maackii residence times. Lonicera maackii exhibited strong above ground competition where removal of its shoots increased seedling and herbaceous growth

(Gorchov and Trisel 2003), survivorship (Cipollini et al. 2008c, Gorchov and Trisel

2003), and species richness (Musson and Mitsch 2003). The presence of L. maackii shrubs may also decrease recruitment of secondary forests because native seedlings experience greater herbivory due to the lack of protective herbaceous cover under L. maackii shrubs (Meiners 2007). White et al. (2014) found when L. maackii abundance increased in riparian zones, native tree seedling and sapling densities decreased. Invaded forests are predicted to experience alterations in species interactions and species composition, ultimately impacting community structure, function, and successional trajectories (Hartman and McCarthy 2008, Hutchinson and Vankat 1997, Luken et al.

1997a).

Animal Community Impact:

Effects of invasive plants on animals is an important motivating factor for natural areas managers and the general public. Pyšek et al. (2012) found nearly 70% of studies reported non-native invasive species vegetation negatively impacted animal communities.

Most invasive plants are pollinated by generalists which can affect native plant-pollinator mutualistic relationships (Richardson et al. 2000, Traveset and Richardson 2006). For example, Impatiens glandulifera (ornamental jewelweed Royle), an Asian invader in central Europe, attracts native bee pollinators due to its more nutrient rich nectar than natives, resulting in decreased fitness and abundance of native flowers (Chittka and

Schürkens 2001). Ballard et al. (2013) found arthropod abundance, biomass, and richness was substantially reduced on non-native plants compared to native plants, suggesting non-native plants may impact food resources (e.g. arthropods) that support higher trophic levels. These studies highlight the importance of research on invasive plant impacts on animal communities.

Lonicera maackii has substantial impacts on food resources for fauna, resulting in alterations in food web dynamics and disease vector population dynamics. Goodell et al.

(2010) found L. maackii serves as a resource for pollinators and was related to increased pollinator visits and pollen deposition of the native herb Hydrophyllum macrophyllum

(largeleaf waterleaf Nutt.) despite L. maackii shading effects (McKinney and Goodell

2011). Alternatively, McKinney and Goodell (2010) found Geranium maculatum

(Spotted geranium L.) pollination visits and seed set were reduced in the presence of L. maackii, suggesting L. maackii has differential impacts on pollinators. Loomis et al.

(2014) found spider taxa and guilds were more abundant in honeysuckle-present plots,

12 with more than double the vertical colonization of spiders compared to honeysuckle- absent plots, suggesting the complex branch architecture of L. maackii is important for spider communities. However, Buddle et al. (2004) reported ground-dwelling predator spider communities were less diverse in narrow riparian forests and hedgerows compared to wider forests buffers due to decreased habitat complexity and ground cover, which the authors suggested may have been due to the presence of L. maackii in these habitats. In a similar study, Loomis et al. (2014) found other arthropod orders (e.g. Diptera,

Hymenoptera, Coleoptera) were more diverse and vertically covered more shrub area in honeysuckle-present plots than honeysuckle-absent plots. Christopher and Cameron

(2012) found L. maackii invasion did not impact arthropod community diversity; however, invaded plots supported greater Acari (mites and ticks) abundance than non- invaded plots. These studies demonstrate L. maackii has differential effects on arthropod communities, influencing resources and habitat substrate for arthropod utilization.

Lonicera maackii can also have consequences on human-related disease vectors.

This shrub has been demonstrated to impact Ochlerotatus triseriatus (known synonym:

Aedes triseriatus; Culidiae) which is the disease vector for the La Crosse encephalitis virus (Conley et al. 2011). In fact, Conley et al. (2011) found the oviposition of this disease vectoring mosquito decreased with L. maackii density, suggesting L. maackii alters the landscape by decreasing the amount of habitat (e.g. tree holes) available for oviposition. A different study found L. maackii may be a facilitator of the West Nile

Virus mosquito vector Culex pipiens (Culicidae; Shewhart et al. 2014). Mosquito eggs exposed to L. maackii leaf and flower leachates had the highest larval survivorship compared to native leaf leachates, and only larvae exposed to L. maackii leachates

13 reached adulthood (Shewhart et al. 2014). In a study observing the effects of L. maackii on tick-borne diseases, it was found white-tailed deer visited L. maackii invaded areas more frequently, which supported a higher number of lone star ticks (Ambylomma americanum) that were infected with a bacterial pathogen from the ehrlichiosis group

(Ehrlichia spp.) compared to areas where L. maackii was removed (Allan et al. 2010).

These studies suggest L. maackii may impact mosquito and tick vector habitats and population dynamics, ultimately affecting human disease instances.

Lonicera maackii invasion across the forest-to-urban gradient has substantially impacted avian species fledgling survivorship and nesting habitat availability (Borgmann and Rodewald 2004, McCusker et al. 2010, Rodewald 2009, Rodewald et al. 2010).

Forests with dense Lonicera spp. invasion had increased densities of understory bird species (e.g. T. migratorius), especially for overwintering birds when Lonicera spp. fruit production was high; however, densities of upper canopy birds (e.g. Contopus virens: eastern wood-pewees) decreased in these forests (McCusker et al. 2010). Nest predation is one of the most important threats to avian fledgling success; therefore, it is important for birds to select appropriate nesting habitats (Martin 1992). Borgmann and Rodewald

(2004) found avian nest success of T. migratorius and Cardinalis cardinalis (Northern cardinal) was lower in invasive Lonicera spp. and Rosa multiflora (Muliflora rose

Thunb.) locations along an increased urban gradient compared to native woody species.

Rodewald (2009) discovered an increase in Empidonax virescens (Acadian flycatcher) brood parasitism was positively related to the number of stems around the nest. Stems were associated with L. maackii invasion suggesting its “bushy” growth pattern creates perching sites for brown-headed cowbirds (Molothrus ater) to view nests, increasing

14 opportunities for brood parasitism to occur. Rodewald et al. (2010) suggest L. maackii presence results in an ecological trap for avian species. Lonicera maackii is suspected to be an ephemeral ecological trap for C. cardinalis due to its unique leaf phenology. Birds may preferentially build nests in L. maackii shrubs in order to hide nests from predators due to early leaf out this plant exhibited in the spring compared to native plants; however,

L. maackii shrubs lack the habitat complexity needed for nesting sites, which in actuality makes nests more susceptible to predation and result in a decrease in overall bird annual production (Rodewald et al. 2010). Research has yet to identify if L. maackii influences bird plumage color, an important trait for mate selection; however, a study conducted by

Witmer (1996) found that E. virescens tail bands change from yellow to orange when

Lonicera morrowii (Morrow’s honeysuckle A. Gray) fruits were consumed. These studies indicate L. maackii shrubs are a poor habitat for avian fauna, creating an ecological trap which reduces avian success.

The invasion of L. maackii can also initiate behavioral changes in animals, resulting in alterations in foraging behavior and predation risk. Peromyscus leucopus

(white-footed mouse) was found to increase in risky behavior (e.g. foraging) in L. maackii stands (contingent upon food availability and moon light), most likely due to the high canopy cover L. maackii provides in the shrub layer (Mattos and Orrock 2010).

Similar findings were observed for other nocturnal mammals (e.g. opossums and raccoons), with mice preferring to forage under L. maackii shrubs on cloudless nights

(Dutra et al. 2011). Rodent granivores also preyed on L. maackii seeds more than native roughleaf dogwood (Cornus drummondii C.A. Mey.) during the spring (Mattos et al.

2013). These studies indicate some granivores and mesopredators may be positively

15 influenced by the presence of L. maackii; however, more research is needed to identify how the presence of this shrub may influence behavior.

There has been some research related to how L. maackii may mediate amphibian communities. Watling et al. (2011c) investigated how invaded L. maackii plots in a deciduous forest altered the understory microclimate (temperature and humidity) and amphibian community. Plots invaded by L. maackii had a lower mean daily temperature, amphibian species richness and evenness, and experienced a shift in the amphibian community composition. In a study focused on the interaction of predators and L. maackii chemistry on amphibian larvae, artificial pools were created in invaded and non- invaded plots (Watling et al. 2011b). Pools were lined with soil that were or were not chemically influenced by L. maackii growth and allowed to fill up with natural rain water. It was found that Anazyrus americans (American toad) larvae development was significantly faster in pools containing leaf litter and soil from L. maackii. Hickman and

Watling (2014) also found A. americans tadpoles exhibited increased risk prone behaviors, such as increased surfacing and swimming behavior in L. maackii leachate regardless of the presence of predator chemical cues. These findings indicate L. maackii chemically alters amphibian habitat, indirectly altering risk-prone behavior and making these animals more susceptible to predators.

Odocoileus virginianus is an important ungulate known to reduce tree seedling and herbaceous regeneration (Rooney and Dress 1997, Tilghman 1989), forest habitat structure (Fuller 2001, McShea and Rappole 1992), and food web interactions (Rooney and Waller 2003). Several studies have indicated there is an important relationship between L. maackii and O. virginianus. Deer are known to eat L. maackii berries and

Castellano and Gorchov (2013) found 68% of seeds were still viable after passing through the deer gut, suggesting this ungulate can be an important dispersal vector of L. maackii. Christopher et al. (2014) demonstrated both L. maackii presence and O. virginianus browsing decreased annual and spring perennial abundance. A L. maackii ×

O. virginianus interaction effect revealed O. virginianus reduced perennial abundance regardless of L. maackii presence or absence. In a study examining the interactive effect of L. maackii and O. virginianus on litter arthropod communities, no interactive effect was found on arthropod diversity and total abundance, but there was a significant interaction effect on the abundance of Acari (Christopher and Cameron 2012). Lonicera maackii led to a decrease in Aranea (orb-weaver siders) abundance while there was an increase in Acari abundance, suggesting L. maackii and deer may have indirectly impacted arthropod communities (Christopher and Cameron 2012). Future research is needed to understand how the interactive role between L. maackii and O. virginianus can impact higher trophic levels and ecosystem processes.

Microbial Community Impacts:

Invasive species impact soil and aquatic microbial communities that are crucial to organic matter processing and nutrient cycling, resulting in alterations in ecosystem processes and function (Claeson et al. 2014, Gessner et al. 2007, Hawkes et al. 2005,

Kourtev et al. 2002). Invasive plants tend to support different microbial communities compared to native plants and coupled with unique leaf characteristics (e.g. increased nitrogen), result in alterations in decomposition and nutrient transformation. Arthur et al.

(2012) reported L. maackii leaf litter breakdown was 5 × faster, had higher nitrogen, and lower lignin concentrations when compared to native Fraxinus americana (white ash L.)

17 and Carya spp. (hickory Nutt.). Through the decomposition process, L. maackii leaf litter maintained microbial communities that were distinguishable from the community present on native species (Arthur et al. 2012). Ali et al. (2015) found when L. maackii leachate was added to sterilized soil there was a 1.5 fold increase in mycorrhizal infection of I. capensis compared to sterilized soil, generally increasing I. capensis growth; however, when leaf extracts were added to live soil, there was a decrease in mycorrhizal infection of I. capensis and overall growth. In a plant-soil feedback study examining changes in arbuscular mycorrhizal fungi abundance associated with invasive species, it was found that L. maackii reduced arbuscular mycorrhizal fungi on native plant roots indirectly via soil legacy effects and directly when grown in conjunction with L. maackii (Shannon et al. 2014). These findings suggesting L. maackii leaves support unique microbial communities and leachate alters soil communities, impacting ecosystem function and processes. Further research is needed to fully understand how these microbial effects may manifest as alterations of ecosystem function.

LONICERA MAACKII INVASION AND LANDSCAPE ECOLOGY:

Plant species invasions have been strongly linked to land use and characteristics of the landscape matrix in which the potentially invadable habitat is embedded

(Bartuszevige et al. 2006, Hutchinson and Vankat 1998, Johnson et al. 2006). For instance, Johnson et al. (2006) reported a relationship between invasive species colonization and site features such as soil pH. Edge effects can also strongly influence resource availability within a given site. Bartuszevige et al. (2006) and Yates et al.

(2004) found the amount of edge in the landscape surrounding habitat patches was a strong determinant of invasion and small forest patches are susceptible to edge effects

18 well into their interiors. Intensive land use is well-known to have a lasting influence on the vegetation composition of forests (Bellemare et al. 2002, Foster et al. 2003), including the facilitation of invasive species establishment (Johnson et al. 2006).

Moreover, if the historical land-use creates an “extinction debt” of native species

(Vellend et al. 2006), invasive species may exploit the resources that were made available in this “empty niche” (Hierro et al. 2005).

Lonicera maackii is a good model species for understanding and demonstrating the landscape ecology of plant invasion. Landscape features have been shown to influence the invasion biology of this species and anthropogenic features are particularly important. For instance, Bartuszevige and Gorchov (2006) and Hutchinson and Vankat

(1997) both demonstrated this species’ presence in forest patches was positively related to distance to the nearest town. White et al. (2014) found L. maackii was indicative of areas that were more urbanized and less associated with forested areas. Borgmann and

Rodewald (2005) found that L. maackii cover was best explained by an increase in urban land cover, and that L. maackii was more pervasive in urban forests compared to rural ones. Flory and Clay (2005 & 2009) demonstrated the density and germination success of L. maackii increased closer to roadways and in forests at early and mid-successional stages in central and southern Indiana, USA. In a study conducted in Louisville, KY

USA, L. maackii was shown to be less successful in terms of stem density within forest patches located outside 10 km from the center of the city compared to those within that distance (Trammell and Carreiro 2011). Pennington et al. (2010) found L. maackii was a dominant species in urban riparian systems and McNeish et al. (2012, 2015) found that L. maackii dominated riparian zones impacted aquatic ecosystems, suggesting accumulated

19 impacts at the watershed scale may be more severe.

LONICERA MAACKII IMPACTS ON ECOSYSTEM PROCESSES:

Ecosystems are open systems susceptible to subsidies and allochthonous flows of resources from adjacent habitats (Baxter et al. 2005, Leroux and Loreau 2008). Invasive species can substantially alter ecosystem processes such as nutrient cycling, decomposition, and energy transformation within and across ecosystems. Pueraria montana (kudzu (Lour.) Merr.) is an invasive nitrogen fixing legume found in the southeastern USA that significantly increased net N mineralization, nitrification, and nitric oxide emissions from invaded soils by more than 100% (Hickman et al. 2010).

Mineau et al. (2012) found the riparian invasive tree Elaeagnus angustifolia (Russian olive L.) substantially increased terrestrial organic matter subsidies and retention of leaf organic matter in a stream system resulting in an estimated 14% decrease in stream ecosystem efficiency (ratio of ecosystem respiration to organic matter input). In a meta- analysis of 199 papers presenting data related to invasive species impacts, it was found that although presence of invasive species tends to result in a decline of local native species, many plant invasions result in increased ecosystem function (Vilà et al. 2011).

Several studies have demonstrated L. maackii can impact a variety of terrestrial and aquatic ecosystem processes and function. For example, Trammell et al. (2012) found total foliar biomass was 1.5 × lower in invaded L. maackii forests; however, L. maackii foliar biomass was 16 × greater in these plots compared to low invaded forests.

This study suggests L. maackii may negatively impact native tree and shrub species production. In a tree-ring study, Hartman and McCarthy (2007) found upper canopy trees had reduced radial growth in forested sites with dense L. maackii compared to non-

20 invaded sites. Lonicera maackii impacts organic matter processing and availability in terrestrial and aquatic habitats as described via five leaf breakdown experiments where it was found that L. maackii leaf breakdown was up to ~ 4 × faster compared to several native leaf species (Arthur et al. 2012, Fargen et al. 2015, McNeish et al. 2012, Poulette and Arthur 2012, Trammell et al. 2012). In an aquatic leaf pack study, it was found that

L. maackii leaves supported greater and a more complex microbial growth, possibly due to the increased number of trichomes which created a more complex leaf surface topography at the micro scale compared to Prunus serotina leaves (cherry Ehrh.; Fig. 2;

R. W. McEwan, unpublished data). Lonicera maackii leaf deposition into streams is quite high (Fig. 1d; McNeish et al. 2015); therefore, there may be impacts on aquatic leaf processing scaling from the reach to the watershed level.

Lonicera maackii may also have substantial impacts on water and nutrient transformation and availability. In a terrestrial leaf pack study, L. maackii released nitrogen faster compared to Acer saccharum leaf litter (sugar maple Marshall; Trammell et al. 2012). Poulette and Arthur (2012) found L. maackii increased N loss in mixed honeysuckle-hickory leaf packs up to ~2 × more than any other honeysuckle-native leaf pack combination. A throughfall study demonstrated there was decreased throughfall volume under L. maackii shrubs and cation concentrations increased up to 3 × (McEwan et al. 2012). Lonicera maackii has been shown to have a higher transpiration rate compared to other shrubs, using an estimated 10% of ground and surface water (Boyce et al. 2011). Rapid uptake of surface water, coupled with high leaf N concentration and loss during decomposition, may support a positive nutrient feedback loop for L. maackii

21 shrubs, resulting in an impact on forest production and water and nutrient transformation and availability in terrestrial and aquatic habitats.

MANAGEMENT AND RESTORATION OF LONICERA MAACKII INVADED HABITATS:

Detection and Management:

Early detection of invasive species is an important proactive approach for management (Moody and Mack 1988). The USA National Invasive Species Council

(NISC) has developed guidelines for early detection and rapid response of invasive species (USDOI 2015). Early detection allows for preemptive eradication of nascent foci before these small populations combine into larger populations and become more difficult to control (Moody and Mack 1988). Simberloff (2003) discusses the importance of early detection and rapid response of invasive species, highlighting several successes in the eradication of invasives such as Vachellia karroo (acacia karroo thorn (Hayne)

Banfi & Galasso) in Western Australia and Senecio jacobaea (ragwort L.) in New

Zealand. For example, Caulerpa taxifolia is a tropical algae that has invaded coasts along several countries including France, Italy, and Spain. This invasive algae was first sighted in a small coastal area outside Monaco in 1984; however, no management efforts were made until the population spread along thousands of hectares (Simberloff 2003).

Wilfong et al. (2009) demonstrated that the satellite Landsat 5 TM and Landsat 7 ETM+ imagery coupled with the unique, seasonal timing of leaf abscission of many invasive plants compared to native species, can be used to facilitate detection of invasive species.

These studies highlight early detection methods and preemptive efforts are pivotal in the quest for managing plant invaders.

A suite of studies have identified practical detection and treatment methods that are useful for management of L. maackii. Environmental remote sensing images can be utilized to detect and map locations of invasive species in a cost effective manner (Huang and Asner 2009). Lonicera maackii’s extended leaf phenology allows for detection in early spring and late fall using image differencing of satellite photos (Wilfong et al.

2009). A combination of images with fine spatial resolution, data from Thematic Mapper

(TM) satellites, and ground truthing (field observations) can potentially be one of the most effective methods for early L. maackii detection (Johnston et al. 2012, Shouse et al.

2013).

Several eradication studies have been conducted in order to identify the most effective way to manage L. maackii. Schulz et al. (2012) found seasonal stem cutting followed by an application of 18% glyphosate resulted in a 75-85% reduction of L. maackii individuals, while stem cutting followed by spraying of regrowth shoots with 1% glyphosate ~ 40 days later was less effective, resulting in up to 56% of L. maackii individuals killed. Rathfon and Ruble (2007) tested four removal treatment methods: foliar application, streamline basal bark application, full basal bark application, and stump cutting with a chemical application. Foliar applications were effective in controlling 70-90% whereas basal bark applications did not give consistent results, and were thus considered unreliable. Stump cutting coupled with chemical application was most effective against larger L. maackii shrubs which resulted in over 90% control of L. maackii shrubs. In a study examining the rationale of one year versus annual treatment of

L. maackii, it was found that annual treatment of L. maackii via stump cutting treated with glyphosate coupled with pulling plants was up to ~ 6× more effective than a one

23 event treatment (Loeb et al. 2010). Hartman and McCarthy (2004) found that stump cutting with a treatment of glyphosate and stem injection with EZ-Ject lance were both highly effective at killing L. maackii shrubs (> 94%). These collective studies suggest best management practice for L. maackii is to apply a chemical herbicide immediately following stump cutting and to repeat the process on an annual or semi-annual basis; although, a more cost effective method may be to use foliar application.

In addition to human induced efforts, there has been some natural dieback of invasive honeysuckle species reported in several states throughout the Midwest USA

(Boyce et al. 2014). In recent years some L. maackii shrubs have exhibited dieback due to the presence of a honeysuckle leaf blight fungus (Insolibasidium deformans:

Basidiomycete; Boyce et al. 2014). This fungus affects individuals of the Lonicera genus and is widespread throughout northcentral and northeastern USA and the UK (Beales et al. 2004, Riffle and Watkins 1986). Symptoms of I. deformans start in spring when lesions on new leaves develop, which result in leaves eventually browning and senescing prematurely (Beales et al. 2004, Riffle and Watkins 1986). In a recent survey of L. maackii shrubs conducted around Cincinnati, OH USA, it was found 61.8% of L. maackii stems were dead which was ~ 58% more compared to reports from 1989 (3.2%; Boyce et al. 2014). Thus far there has been no studies on biocontrol agents for L. maackii; however, a study by Waipara et al. (2007) demonstrated herbivore and pathogen damage was low on L. japonica in New Zealand and suggested further research on natural enemies in L. japonica’s native range is necessary for the development of biocontrol agents.

Lonicera maackii populations may be susceptible to forest litter and allelopathic effects from native plant species. Wilson et al. (2013) conducted a study to identify forest characteristics related to L. maackii invasion. Findings indicated increased leaf litter depth, with increasing presence of oak leaf litter, were negatively associated with the presence of L. maackii. Rietveld (1983) found the allelochemical juglone, commonly found in the walnut family, had negative impacts on L. maackii. In this study a variety of herbaceous and woody plant species, including L. maackii, were exposed to different juglone concentrations (0-10-3M). There was a significant decrease in L. maackii seed germination (92-43%), radical growth (8-0 mm), and shoot elongation (20-1cm) when exposed to increasing juglone concentrations. These studies are important because they identified forest characteristics that may increase forest community resistance to L. maackii invasion.

Lonicera maackii management impacts on flora and fauna:

Several studies have documented the effects on removal and managing L. maackii on plant and animal communities. Runkle et al. (2007) found that 7-8 years after L. maackii removal plant cover, tree seedling density, and species richness increased – suggesting removal of L. maackii can enhance plant ground cover and impact ecosystem productivity and function. Hartman and McCarthy (2004) conducted a removal and non- removal study to identify how removal of L. maackii impacted native seedling survivorship. They found three years post L. maackii removal seedling survivorship was higher compared to plot locations with L. maackii present; however, seedling survivorship varied among genera. Removal of L. maackii shrubs also resulted in an increase in the abundance of P. leucopus an important generalist rodent (Shields et al.

2014). McNeish et al. (2015) found removal of L. maackii from the riparian zone of a stream significantly increased in-stream light availability, terrestrial organic matter contribution to the stream, and aquatic macroinvertebrate density compared to a non- removal stream reach, indicating management of L. maackii in terrestrial habitats can impact adjacent aquatic systems. In a similar study Fargen et al. (2015) found removal of riparian L. maackii did not influence in-stream leaf litter decomposition; however, L. maackii leaf litter packs supported lower macroinvertebrate abundance compared to native A. saccharum Marsh. (sugar maple) litter packs.

CONNECTIONS TO INVASION THEORY:

Several theories have been developed to address how and why some non-native species become overabundant, successfully outcompete natives, and eventually become classified as invasive. For example, the Empty Niche Hypothesis presumes a non-native can become successful when there is open habitat for the organism to colonize, proliferate, and spread outside its native range (Shea and Chesson 2002). The Enemy

Release Hypothesis details the idea a non-native becomes invasive when a natural enemy

(e.g. specialist consumer) is not present in the new range, and thus, the non-native has essentially escaped from its natural enemy (Colautti et al. 2004, Keane and Crawley

2002). The Novel Weapons Hypothesis states invasive plants may bring with them into

“recipient” communities new biochemistry that provides an advantage and, through generations, this alteration may become selected for and increase in strength (Callaway and Ridenour 2004). Some invasive plants are known to exhibit increased allelochemical production outside their native range which may serve as a “novel weapon” providing a competitive edge against native species (Ni et al. 2012, Vivanco et al. 2004). The

Evolution of Increased Competitive Ability theory postulates that with the removal of natural enemies, an invasive will shift resources from defense to growth to improve its competitive ability (Blossey and Notzold 1995). Another major ecological theory,

Invasional Meltdown, predicts the success and establishment of one invasive species may facilitate and/or increase the establishment of other invasive species (Simberloff and

Holle 1999, Simberloff 2006). For example, the Rhamnus cathartica (European buckthorn L.) and the Lumbricus terrestricus (European earthworm) may have co- facilitated one another in North America, and their co-invasion may have facilitated several other invasive pests and insects (Heimpel et al. 2010). These theories are often studied and considered separately; however, there is evidence that many of these theories may apply to single species, and the theories themselves are conceptually linked , which highlights the importance of exploring these theories in tandem (Hierro et al. 2005, Joshi and Vrieling 2005).

The invasion biology of L. maackii is a good model for understanding the overlap and interconnection of some important invasion theories. For instance, the success of L. maackii has been linked to both the Novel Weapons Hypothesis (NWH) and the Enemy

Release Hypothesis (ERH). Several studies presented above (Forest Community Impacts

& Resistance to Herbivory) suggested L. maackii may have allelopathic effects, inhibiting native plant germination, growth, and development and suppressing insect survivorship (Cipollini et al. 2008a, Cipollini et al. 2008b, Dorning and Cipollini 2005,

McEwan et al. 2010, McEwan et al. 2009b). Lonicera maackii generally experiences low levels of arthropod herbivory (Lieurance and Cipollini 2012) but has been commonly observed to be browsed by O. virginianus (Castellano and Gorchov 2013). Cipollini et

27 al. (2008b) identified secondary compounds in L. maackii leachate that decreased insect herbivore consumption and native seed germination, suggesting allelochemistry may in part explain why L. maackii is so successful and lending support to the NWH. Lieurance and Cipollini (2013b) also studied herbivory effects of the specialist North American honeysuckle sawfly (Zaraea inflate: Cimibicidae) and the generalist caterpillar

Spodoptera frugiperda (fall armyworm: Noctuidae) on L. maackii, L. reticulate (native honeysuckle Raf.), Viburnum prunifolium (native confamilial L.), and L. japonica

(Japanese honeysuckle Thunb.) in field and laboratory settings. Lonicera maackii had significantly less foliage damage and was not impacted by the sawfly specialist, which preferred L. reticulate over L. maackii when given a choice (Lieurance and Cipollini

2013b). The generalist caterpillar fed equally on all Lonicera species, but effects on L. maackii were more strongly observed in laboratory assays (Lieurance and Cipollini

2013b). These findings lend support to the Novel Weapons and Enemy Release hypotheses in that (1) L. maackii produces allelochemicals that negatively impact native plants and arthropods, (2) herbivory is quite low on L. maackii, and (3) this invasive can escape from specialist and generalist insect herbivores which hinder native honeysuckle and confamiliars. These findings also suggest strong potential for Evolution of Increased

Competitive Ability hypothesis, particular given L. maackii’s rapid growth rate and phenotypic plasticity. Observations by the authors from forests in Ohio and Kentucky

USA suggest Invasional Meltdown may also be occurring with L. maackii invasion as control sites often display a profusion of A. petiolata growth, and some areas of heavy invasion by L. maackii are underlain by an invasion of Euonymus fortuneii (winter creeper (Turcs.) Hand.-Maz.).

FUTURE DIRECTIONS AND CONSIDERATIONS:

In this review we have synthesized empirical literature that demonstrated L. maackii’s effects across various ecological scales, ecosystem processes and function, and restoration efforts. Lonicera maackii has been shown to (1) suppress local plant species survivorship, growth, and reproduction, (2) decrease primary productivity and enhance decomposition and nutrient turnover ecosystem processes, (3) increase risk prone behavior and decrease reproductive success of certain animals, (4) provide needed protection from predators and supply seeds as a food resource for grainivores, (5) provide support for human disease vectors, and (6) support high vertical diversity of certain arthropod communities. These findings lend support to the hypothesis that L. maackii effects are complex and vary across ecosystems and multiple ecological scales (Fig. 3), suggesting this invasive has a diversity of impacts on species interactions and composition, community structure and successional trajectories, and ecosystem function and processes.

Lonicera maackii transforms aquatic and terrestrial ecosystems via alterations of terrestrial subsidies, habitat structure, community composition, and ecosystem function

(Fig. 3). The presence of L. maackii along stream habitats (or potentially along other water bodies) resulted in a substantial change in the pool of resources in the aquatic system that support aquatic food webs and ecosystem processes (McNeish et al. In

Press). Many aquatic macroinvertebrates have a terrestrial adult phase (e.g. mosquitos and blackflies) and impacts on the population dynamics of aquatic insect life stages may result in a bottom-up effect on secondary production that is present in the terrestrial habitat and serves as an important food resource for terrestrial vertebrates (Baxter et al.

2005; Burdon and Harding 2008, Lounibos 2002). Co-occurring with impacts in aquatic systems, are similar effects of L. maackii in terrestrial systems. Lonicera maackii alters plant community composition and microbial communities, impacting the pool of resources available in terrestrial systems, ultimately influencing amphibian, avian, and arthropod communities (Gould and Gorchov 2000, Loomis et al. 2014, Miller and

Gorchov 2004, Watling et al. 2011a; Watling et al. 2011b). Due to the presence of L. maackii along water bodies, we expect impacts of this invasive plant crosses the terrestrial-aquatic interface and will have major implications at the watershed scale (Fig.

3). This framework of L. maackii impact should be considered when managing for this invasive species so that adjacent habitats and communities are not adversely impacted.

The research synthesized in this review has assisted in developing a broad framework to explain L. maackii effects across ecological scales (Fig. 3) and elucidated several avenues for future research. Several studies have suggested an allelopathic mechanism is in part linked to the success of this invasive; however, more work is necessary to tease out direct and indirect effects on native plant and herbivore communities and provide more evidence for the Novel Weapons and Enemy Release hypotheses. Restoration efforts have identified removal of L. maackii positively impacts native plant communities, but it would be interesting to expand research efforts to understand how various management efforts (e.g. cutting and removal versus defoliation) impact resources that support animal communities. For example, McNeish et al. (2015) suggested L. maackii branch architecture is responsible for delayed and reduced availability of in-stream leaf litter resources that serve as aquatic food and habitat substrates, which suggests cut and removal efforts may be more beneficial compared to

30 defoliation methods for aquatic ecosystem health. Finally, a link has been identified between L. maackii and human disease vectors; however, very few studies have been conducted in this area. Continued emphasis on synthesizing effects of widespread species such as L. maackii is critical for building generalizable frameworks for invasion impacts and guiding research and management efforts.

ACKNOWLEDGEMENTS:

Appreciation to Anastasia Stolz for her confocal photos of L. maackii and P. seratona images and to Meghan Maloney and Erin Rowekamp for their photos of A. americans and T. migratorius photos. This work was also supported in part by the

University of Dayton Office for Graduate Academic Affairs through the Graduate

Student Summer Fellowship Program and by the National Science Foundation (DEB

1352995).

LITERATURE CITED:

ALI, J., D. LIEURANCE, AND D. CIPOLLINI. 2015. Soil biota affect mycorrhizal infection

and growth of Impatiens capensis and alter the effects of Lonicera maackii

rhizosphere extracts. J. Torrey Bot. Soc. 142:1–11.

ALLAN, B. F., H. P. DUTRA, L. S. GOESSLING, K. BARNETT, J. M. CHASE, R. J. MARQUIS,

G. PANG, G. A STORCH, R. E. THACH, AND J. L. ORROCK. 2010. Invasive honeysuckle

eradication reduces tick-borne disease risk by altering host dynamics. Proc. Natl.

Acad. Sci. U. S. A. 107:18523–18527.

ARTHUR, M. A., S. R. BRAY, C. R. KUCHLE, AND R. W. MCEWAN. 2012. The influence of

the invasive shrub, Lonicera maackii, on leaf decomposition and microbial

community dynamics. Plant Ecol. 213:1571–1582.

BALLARD, M., J. HOUGH-GOLDSTEIN, AND D. TALLAMY. 2013. Arthropod communities

on native and nonnative early successional plants. Environ. Entomol. 42:851–859.

BARTUSZEVIGE, A. M., AND D. L. GORCHOV. 2006. Avian seed dispersal of an invasive

shrub. Biol. Invasions 8:1013–1022.

BARTUSZEVIGE, A. M., D. L. GORCHOV, AND L. RAAB. 2006. The relative importance of

landscape and community features in the invasion of an exotic shrub in a fragmented

landscape. Ecography (Cop.). 29:213–222.

BARUCH, Z., AND G. GOLDSTEIN. 1999. Leaf construction cost, nutrient concentration,

and net CO 2 assimilation of native and invasive species in Hawaii. Oecologia

121:183–192.

BAUER, J. T., S. M. SHANNON, R. E. STOOPS, AND H. L. REYNOLDS. 2012. Context

dependency of the allelopathic effects of Lonicera maackii on seed germination.

Plant Ecol. 213:1907–1916.

BAXTER, C. V., K. D. FAUSCH, AND W. CARL SAUNDERS. 2005. Tangled webs: reciprocal

flows of invertebrate prey link streams and riparian zones. Freshw. Biol. 50:201–

220.

BEALES, P. A., J. SCRACE, R. T. A. COOK, A. V. BARNES, AND C. R. LANE. 2004. First

report of honeysuckle leaf blight (Insolibasidium deformans) on honeysuckle

(Lonicera spp.) in the UK. Plant Pathol. 53:536–536.

BELLEMARE, J., G. MOTZKIN, AND D. R. FOSTER. 2002. Legacies of the agricultural past in

the forested present: An assessment of historical land-use effects on rich mesic

forests. J. Biogeogr. 29:1401–1420.

BLOSSEY, B., AND R. NOTZOLD. 1995. Evolution of increased in invasive competitive

ability nonindigenous a hypothesis plants : A Hypothesis. J. Ecol. 83:887–889.

BORGMANN, K. L., AND A. D. RODEWALD. 2004. Nest predation in an urbanizing

landscape: The role of exotic shrubs. Ecol. Appl. 14:1757–1765.

BORGMANN, K. L., AND A. D. RODEWALD. 2005. Forest restoration in urbanizing

landscapes: interactions between land uses and exotic shrubs. Restor. Ecol. 13:334–

340.

BORRELL, B. 2009. Alien invasion ? An ecologist doubts the impact of exotic species. Sci.

Am.

exotic/>.

BOYCE, R. L., S. N. BROSSART, L. A. BRYANT, L. A. FEHRENBACH, R. HETZER, J. E. HOLT,

B. PARR, Z. POYNTER, C. SCHUMACHER, S. N. STONEBRAKER, M. D. THATCHER, AND

M. VATER. 2014. The beginning of the end? Extensive dieback of an open-grown

Amur honeysuckle stand in northern Kentucky, USA. Biol. Invasions 16:2017–

2023.

BOYCE, R. L., R. D. DURTSCHE, AND S. L. FUGAL. 2011. Impact of the invasive shrub

Lonicera maackii on stand transpiration and ecosystem hydrology in a wetland

forest. Biol. Invasions 14:671–680.

BUDDLE, C. M., S. HIGGINS, AND L. RYPSTRA. 2004. Ground-dwelling spider assemblages

inhabiting riparian forests and hedgerows in an agricultural landscape. Am. Midl.

Nat. 151:15–26.

BURDON, F. J., AND J. S. HARDING. 2008. The linkage between riparian predators and

aquatic insects across a stream-resource spectrum. Freshw. Biol. 53:330–346.

CALLAWAY, R. M., AND W. M. RIDENOUR. 2004. Novel Weapons: Invasive success and

the evolution of increased competitive ability. Front. Ecol. Environ. 2:436–443.

CASTELLANO, S. M., AND D. L. GORCHOV. 2013. White-tailed deer (Odocoileus

virginianus) disperse seeds of the invasive shrub, Amur honeysuckle (Lonicera

maackii). Nat. Areas J. 33:78–80.

CHEW, M. K., AND A. L. HAMILTON. 2010. The rise and fall of biotic nativeness: a

historical perspective. Pages 35–47 In D. M. Richardson [ed.], Fifty Years of

Invasion Ecology: The Legacy of Charles Elton, 1st edition. Blackwell Publishing

Ltd, Oxford, UK.

CHITTKA, L., AND S. SCHÜRKENS. 2001. Successful invasion of a floral market. Nature

411:653.

CHRISTOPHER, C. C., AND G. N. CAMERON. 2012. Effects of invasive Amur honeysuckle

(Lonicera maackii) and white-tailed deer (Odocoileus virginianus) on litter-dwelling

arthropod communities. Am. Midl. Nat. 167:256–272.

CHRISTOPHER, C. C., S. F. MATTER, AND G. N. CAMERON. 2014. Individual and interactive

effects of Amur honeysuckle (Lonicera maackii) and white-tailed deer (Odocoileus

virginianus) on herbs in a deciduous forest in the eastern United States. Biol.

Invasions 16:2247–2261.

CIPOLLINI, D., AND M. DORNING. 2008. Direct and indirect effects of conditioned soils

and tissue extracts of the invasive shrub, Lonicera maackii, on target plant

performance. Castanea 73:166–176.

CIPOLLINI, D., R. STEVENSON, AND K. CIPOLLINI. 2008a. Contrasting effects of

allelochemicals from two invasive plants on the performance of a nonmycorrhizal

plant. Int. J. Plant Sci. 169:371–375.

CIPOLLINI, D., R. STEVENSON, S. ENRIGHT, A. EYLES, AND P. BONELLO. 2008b. Phenolic

metabolites in leaves of the invasive shrub, Lonicera maackii, and their potential

phytotoxic and anti-herbivore effects. J. Chem. Ecol. 34:144–152.

CIPOLLINI, K. A., G. Y. MCCLAIN, AND D. CIPOLLINI. 2008c. Separating above- and

belowground effects of Alliaria petiolata and Lonicera maackii on the performance

of Impatiens capensis. Am. Midl. Nat. 160:117–128.

CIPOLLINI, K., AND W. FLINT. 2013. Comparing allelopathic effects of root and leaf

extracts of invasive Alliaria petiolata, Lonicera maackii, and Ranunculus ficaria on

germination of three native woodland. Ohio J. Sci. 112:37–43.

CIPOLLINI, K., K. TITUS, AND C. WAGNER. 2012. Allelopathic effects of invasive species

(Alliaria petiolata, Lonicera maackii, Ranunculus ficaria) in the Midwestern United

States. Allelopath. J. 29:63–76.

CLAESON, S. M., C. J. LEROY, J. R. BARRY, AND K. A. KUEHN. 2014. Impacts of invasive

riparian knotweed on litter decomposition, aquatic fungi, and macroinvertebrates.

Biol. Invasions 16:1531–1544.

COLAUTTI, R. I., A. RICCIARDI, I. A. GRIGOROVICH, AND H. J. MACISAAC. 2004. Is

invasion success explained by the enemy release hypothesis? Ecol. Lett. 7:721–733.

COLLIER, M. H., J. L. VANKAT, AND M. R. HUGHES. 2002. Diminished plant richness and

abundance below Lonicera maackii, an invasive shrub. Am. Midl. Nat. 147:60–71.

CONLEY, A. K., J. I. WATLING, AND J. L. ORROCK. 2011. Invasive plant alters ability to

predict disease vector distribution. Ecol. Appl. 21:329–334.

CROOKS, J. A. 2002. Characterizing ecosystem-level consequences of biological

invasions: the role of ecosystem engineers. Oikos 97:153–166.

DAEHLER, C. C. 2003. Performance comparisons of co-occurring native and alien

invasive plants: Implications for conservation and restoration. Annu. Rev. Ecol.

Evol. Syst. 34:183–211.

DAVIES, K. W., AND R. L. SHELEY. 2007. A conceptual framework for preventing the

spatial dispersal of invasive plants. Weed Sci. 55:178–184.

DAVIS, M. 2011. Don’t judge species on their origins. Nature 474:153–154.

DEERING, R. H., AND J. L. VANKAT. 1999. Forest colonization and developmental growth

of the invasive shrub Lonicera maackii. Am. Midl. Nat. 141:43–50.

DORNING, M., AND D. CIPOLLINI. 2005. Leaf and root extracts of the invasive shrub,

Lonicera maackii, inhibit seed germination of three herbs with no autotoxic effects.

Plant Ecol. 184:287–296.

DUTRA, H. P., K. BARNETT, J. R. REINHARDT, R. J. MARQUIS, AND J. L. ORROCK. 2011.

Invasive plant species alters consumer behavior by providing refuge from predation.

Oecologia 166:649–657.

ESCHTRUTH, A. K., AND J. J. BATTLES. 2011. The importance of quantifying propagule

pressure to understand invasion: An examination of riparian forest invasibility.

Ecology 92:1314–1322.

FARGEN, C., S. M. EMERY, AND M. M. CARREIRO. 2015. Influence of Lonicera maackii

invasion on leaf litter decomposition and macroinvertebrate communities in an

urban stream. Nat. Areas J. 35:392–403.

FLORY, S. L., AND K. CLAY. 2005. Invasive shrub distribution varies with distance to

roads and stand age in eastern deciduous forests in Indiana, USA. Plant Ecol.

184:131–141.

FLORY, S. L., AND K. CLAY. 2009. Effects of roads and forest successional age on

experimental plant invasions. Biol. Conserv. 142:2531–2537.

FOGARTY, G., AND M. FACELLI. 1999. Growth and competition of Cytisus scoparius, an

invasive shrub, and Australian native shrubs. Plant Ecol. 144:27–35.

FOSTER, D., F. SWANSON, J. ABER, I. BURKE, N. BROKAW, D. TILMAN, AND A. KNAPP.

2003. The importance of land-use legacies to ecology and conservation. Bioscience

53:77–88.

FRYER, J. L. 2010. Ailanthus altissima. In: Effects information system. U.S. Department

of Agriculture, Forest Service, Rocky Mountain Research.

FULLER, R. J. 2001. Responses of woodland birds to increasing numbers of deer: A

review of evidence and mechanisms. Forestry 74:289–298.

GESSNER, M. O., V. GULIS, K. A. KUEHN, E. CHAUVET, AND K. F. SUBERKROPP. 2007. The

Mycota: Environmental and microbial relationships. Pages 301–324 In C. P.

Kubicek and I. S. Druzhinia [eds.], The Mycota: Environmental and microbial

relationships, 2nd edition. Springer Berlin / Heidelberg, New York, NY.

GOODELL, K., A. M. MCKINNEY, AND C. LIN. 2010. Pollen limitation and local habitat-

dependent pollinator interactions in the invasive shrub Lonicera maackii. Int. J.

Plant Sci. 171:63–72.

GORCHOV, D. L., AND D. E. TRISEL. 2003. Competitive effects of the invasive shrub,

Lonicera maackii (Rupr.) Herder (Caprifoliaceae), on the growth and survival of

native tree seedlings. Plant Ecol. 166:13–24.

GOSPER, C. R., C. D. STANSBURY, AND G. VIVIAN-SMITH. 2005. Seed dispersal of fleshy-

fruited invasive plants by birds: contributing factors and management options.

Divers. Distrib. 11:549–558.

GOULD, A. M. A., AND D. L. GORCHOV. 2000. Effects of the exotic invasive shrub

Lonicera maackii on the survival and fecundity of three species of native annuals.

Ammerican Midl. Nat. 144:36–50.

GUREVITCH, J., AND D. K. PADILLA. 2004. Are invasive species a major cause of

extinctions? Trends Ecol. Evol. 19:470–474.

HARTMAN, K. M., AND B. C. MCCARTHY. 2004. Restoration of a forest understory after

the removal of an invasive shrub, Amur honeysuckle (Lonicera maackii). Restor.

Ecol. 12:154–165.

HARTMAN, K. M., AND B. C. MCCARTHY. 2007. A dendro-ecological study of forest

overstorey productivity following the invasion of the non-indigenous shrub Lonicera

maackii. Appl. Veg. Sci. 10:3–14.

HARTMAN, K. M., AND B. C. MCCARTHY. 2008. Changes in forest structure and species

composition following invasion by a non-indigenous shrub, Amur honeysuckle

(Lonicera maackii). J. Torrey Bot. Soc. 135:245–259.

HAWKES, C. V, I. F. WREN, D. J. HERMAN, AND M. K. FIRESTONE. 2005. Plant invasion

alters nitrogen cycling by modifying the soil nitrifying community. Ecol. Lett.

8:976–985.

HEIMPEL, G. E., L. E. FRELICH, D. A. LANDIS, K. R. HOPPER, K. A. HOELMER, Z. SEZEN, M.

K. ASPLEN, AND K. WU. 2010. European buckthorn and Asian soybean aphid as

components of an extensive invasional meltdown in North America. Biol. Invasions

12:2913–2931.

HEJDA, M., P. PYSEK, AND V. JAROSIK. 2009. Impact of invasive plants on the species

richness, diversity and composition of invaded communities. J. Ecol. 97:393–403.

HICKMAN, C. R., AND J. I. WATLING. 2014. Leachates from an invasive shrub causes risk-

prone behavior in a larval amphibian. Behav. Ecol. 25:300–305.

HICKMAN, J. E., S. WU, L. J. MICKLEY, AND M. T. LERDAU. 2010. Kudzu (Pueraria

montana) invasion doubles emissions of nitric oxide and increases ozone pollution.

Proc. Natl. Acad. Sci. U. S. A. 107:10115–10119.

HIDAYATI, S. N., J. M. BASKIN, AND C. C. BASKIN. 2000. Dormancy-breaking and

germination requirements of seeds of four Lonicera species (Caprifoliaceae) with

underdeveloped spatulate embryos. Seed Sci. Res. 10:459–469.

HIERRO, J. L., AND R. M. CALLAWAY. 2003. Allelopathy and exotic plant invasion. Plant

Soil 256:29–39.

HIERRO, J. L., J. L. MARON, AND R. M. CALLAWAY. 2005. A biogeographical approach to

plant invasions: the importance of studying exotics in their introduced and native

range. J. Ecol. 93:5–15.

HUANG, C., AND G. P. ASNER. 2009. Applications of remote sensing to alien invasive

plant studies. Sensors 9:4869–4889.

HUTCHINSON, T. F., AND J. L. VANKAT. 1997. Invasibility and effects of Amur

honeysuckle in southwestern Ohio forests. Conserv. Biol. 11:1117–1124.

HUTCHINSON, T. F., AND J. L. VANKAT. 1998. Landscape structure and spread of the

exotic shrub Lonicera maackii (Amur honeysuckle) in southwestern Ohio forests.

Am. Midl. Nat. 139:383–390.

INGOLD, J. L., AND M. J. O. CRAYCRAFT. 1983. Avian frugivory on honeysuckle

(Lonicera) in southwestern Ohio in fall. Ohio J. Sci. 83:256–258.

JOHNSON, V. S., J. A. LITVAITIS, T. D. LEE, AND S. D. FREY. 2006. The role of spatial and

temporal scale in colonization and spread of invasive shrubs in early successional

habitats. For. Ecol. Manage. 228:124–134.

JOHNSTON, S. E., M. C. HENRY, AND D. L. GORCHOV. 2012. Using Advanced Land Imager

(ALI) and Landsat Thematic Mapper (TM) for the detection of the invasive shrub

Lonicera maackii in southwestern Ohio forests. GIScience Remote Sens. 49:450–

462.

JOSHI, J., AND K. VRIELING. 2005. The enemy release and EICA hypothesis revisited:

incorporating the fundamental difference between specialist and generalist

herbivores. Ecol. Lett. 8:704–714.

KEANE, R. M., AND M. J. CRAWLEY. 2002. Exotic plant invasions and the enemy release

hypothesis. Trends Ecol. Evol. 17:164–170.

KOURTEV, P. S., J. G. EHRENFELD, M. HÃ¤GGBLOM, AND M. HÄGGBLOM. 2002. Exotic

plant species alter the microbial community structure and function in the soil.

Ecology 83:3152–3166.

LEROUX, S. J., AND M. LOREAU. 2008. Subsidy hypothesis and strength of trophic

cascades across ecosystems. Ecol. Lett. 11:1147–1156.

LEVINE, J. M. 2000. Species diversity and biological invasions: Relating local process to

community pattern. Science 288:852–854.

LIEURANCE, D., AND D. CIPOLLINI. 2012. Damage levels from arthropod herbivores on

Lonicera maackii suggest enemy release in its introduced range. Biol. Invasions

14:863–873.

LIEURANCE, D. AND. D. CIPOLLINI. 2013a. Environmental influences on growth and

defense responses of the invasive shrub, Lonicera maackii, to simulated and real

herbifory in the juvenile stage. Annals of Botany 112:741-749

LIEURANCE, D., AND D. CIPOLLINI. 2013b. Exotic Lonicera species both escape and resist

specialist and generalist herbivores in the introduced range in North America. Biol.

Invasions 15:1713–1724.

LIEURANCE, D. M. 2004. Leaf phenology, fecundity, and biomass allocation of the

invasive shrub Lonicera maackii (Rupr.) Maxim in contrasting light environments.

M.S. thesis. Ohio University, Athens, OH.

LOEB, R. E., J. GERMERAAD, T. TREECE, D. WAKEFIELD, AND S. WARD. 2010. Effects of

1-year vs. annual treatment of Amur honeysuckle (Lonicera maackii) in forests.

Invasive Plant Sci. Manag. 3:334–339.

LOOMIS, J. D., G. N. CAMERON, AND G. W. UETZ. 2014. Impact of the invasive shrub

Lonicera maackii on shrub-dwelling Araneae in a deciduous forest in eastern North

America. Am. Midl. Nat. 171:204–218.

LOUNIBOS, L. P. 2002. Invasions by insect vectors of human disease. Annu. Rev.

Entomol. 47:233–266.

LUKEN, J. O. 1988. Population structure and biomass allocation of the naturalized shrub

Lonicera maackii (Rupr.) Maxim. in forest and open habitats. Am. Midl. Nat.

119:258–267.

LUKEN, J. O., AND N. GOESSLING. 1995. Seedling distribution and potential persistence of

the exotic shrub Lonicera maackii in fragmented forests. Am. Midl. Nat. 133:124–

130.

LUKEN, J. O., L. M. KUDDES, AND T. C. THOLEMEIER. 1997a. Response of understory

species to gap formation and soil disturbance in Lonicera maackii thickets. Restor.

Ecol. 5:229–235.

LUKEN, J. O., L. M. KUDDES, T. C. THOLEMEIER, AND D. M. HALLER. 1997b. Comparative

responses of Lonicera maackii (Amur honeysuckle) and Lindera benzoin

(Spicebush) to increased light. Am. Midl. Nat. 138:331–343.

LUKEN, J. O., AND D. T. MATTIMIRO. 1991. Habitat-specific resilience of the invasive

shrub Amur honeysuckle (Lonicera maackii) during repeated clipping. Ecol. Appl.

1:104–109.

LUKEN, J. O., T. C. THOLEMEIER, B. A. KUNKEL, AND L. M. KUDDES. 1995. Branch

architecture plasticity of Amur honeysuckle (Lonicera maackii (Rupr.) Herder):

initial response in extreme light environments. Bull. Torrey Bot. Club 122:190–195.

LUKEN, J., AND J. W. THIERET. 1996. Amur Honeysuckle, its fall from grace. Bioscience

46:18–24.

MARTIN, T. E. 1992. Breeding productivity considerations: What are the appropriate

habitat features for management? Pages 455–473 In J. M. Hagan III and D. W.

Johnston [eds.], Ecology and Conservation of Neotropical Migrant Landbirds.

Smithsonian Institution Press, Washington, DC, USA.

MATTOS, K. J., AND J. L. ORROCK. 2010. Behavioral consequences of plant invasion: an

invasive plant alters rodent antipredator behavior. Behav. Ecol. 21:556–561.

MATTOS, K. J., J. L. ORROCK, AND J. I. WATLING. 2013. Rodent granivores generate

context-specific seed removal in invaded and uninvaded habitats. Am. Midl. Nat.

169:168–178.

MCCUSKER, C. E., M. P. WARD, AND J. D. BRAWN. 2010. Seasonal responses of avian

communities to invasive bush honeysuckles (Lonicera spp.). Biol. Invasions

12:2459–2470.

MCEWAN, R. W., M. A. ARTHUR, AND S. E. ALVERSON. 2012. Throughfall chemistry and

soil nutrient effects of the invasive shrub Lonicera maackii in deciduous forests.

Am. Midl. Nat. 168:43–55.

MCEWAN, R. W., L. G. ARTHUR-PARATLEY, L. K. RIESKE, AND M. A. ARTHUR. 2010. A

multi-assay comparison of seed germination inhibition by Lonicera maackii and co-

occurring native shrubs. Flora - Morphol. Distrib. Funct. Ecol. Plants 205:475–483.

Elsevier.

MCEWAN, R. W., M. K. BIRCHFIELD, A. SCHOERGENDORFER, AND M. A. ARTHUR. 2009a.

Leaf phenology and freeze tolerance of the invasive shrub Amur honeysuckle and

potential native competitors. J. Torrey Bot. Soc. 136:212–220.

MCEWAN, R. W., L. K. RIESKE, AND M. A. ARTHUR. 2009b. Potential interactions

between invasive woody shrubs and the gypsy moth (Lymantria dispar), an invasive

insect herbivore. Biol. Invasions 11:1053–1058.

MCKINNEY, A. M., AND K. GOODELL. 2010. Shading by invasive shrub reduces seed

production and pollinator services in a native herb. Biol. Invasions 12:2751–2763.

MCKINNEY, A. M., AND K. GOODELL. 2011. Plant–pollinator interactions between an

invasive and native plant vary between sites with different flowering phenology.

Plant Ecol. 212:1025–1035.

MCNEISH, R. E., M. E. BENBOW, AND R. W. MCEWAN. 2012. Riparian forest invasion by

a terrestrial shrub (Lonicera maackii) impacts aquatic biota and organic matter

processing in headwater streams. Biol. Invasions 14:1881–1893.

MCNEISH, R. E., E. M. MOORE, M. E. BENBOW, AND R. W. MCEWAN. 2015. Removal of

the invasive shrub, Lonicera maackii, from riparian forests influences headwater

stream biota and ecosystem function. River Res. Appl. 31:1131–1139.

MCSHEA, W. J., AND J. H. RAPPOLE. 1992. White-tailed deer as keystone species within

forest habitats of Virginia. Va. J. Sci. 43:177–186.

MEINERS, S. J. 2007. Apparent competition: an impact of exotic shrub invasion on tree

regeneration. Biol. Invasions 9:849–855.

MEYERSON, L. A., AND H. A. MOONEY. 2007. Invasive alien species in an era of

globalization. Front. Ecol. Environ. 5:199–208.

MILLER, K., AND D. GORCHOV. 2004. The invasive shrub, Lonicera maackii, reduces

growth and fecundity of perennial forest herbs. Oecologia 139:359–375.

MINEAU, M. M., C. V BAXTER, A. M. MARCARELLI, AND G. W. MINSHALL. 2012. An

invasive riparian tree reduces stream ecosystem efficiency via a recalcitrant organic

matter subsidy. Ecology 93:1501–1508.

MOODY, M. E., AND R. N. MACK. 1988. Controlling the spread of plant invasions: the

importance of nascent foci. J. Appl. Ecol. 25:1009–1021.

MUSSON, J., AND W. J. MITSCH. 2003. The effects of the invasive shrub Lonicera maackii

on species richness and soil moisture in the bottomland hardwood forest at the

ORWRP.

MYERS, C. V., AND R. C. ANDERSON. 2003. Seasonal variation in photosynthetic rates

influences success of an invasive plant, Garlic mustard (Alliaria petiolata). Am.

Midl. Nat. 150:231–245.

NI, G. Y., P. ZHAO, Q. Q. HUANG, Y. P. HOU, C. M. ZHOU, Q. P. CAO, AND S. L. PENG.

2012. Exploring the Novel Weapons Hypothesis with invasive plant species in

China. Allelopath. J. 29:199–214.

OLIVER, J. D. 1996. Mile-a-minute weed, (Polygonum perfoliatum L .), an invasive vine

in natural and disturbed sites. Castanea 61:244–251.

PENNINGTON, D. N., J. R. HANSEL, AND D. L. GORCHOV. 2010. Urbanization and riparian

forest woody communities: diversity, composition, and structure within a

metropolitan landscape. Biol. Conserv. 143:182–194.

PIMENTEL, D., R. ZUNIGA, AND D. MORRISON. 2005. Update on the environmental and

economic costs associated with alien-invasive species in the United States. Ecol.

Econ. 52:273–288.

PLUMMER, G. L., AND C. KEEVER. 1963. Autumnal daylight weather and Camphor-weed

cispersal in the Georgia Piedmont region. Bot. Gaz. 124:283–289.

POULETTE, M. M., AND M. A. ARTHUR. 2012. The impact of the invasive shrub Lonicera

maackii on the decomposition dynamics of a native plant community. Ecol. Appl.

22:412–424.

PYŠEK, P., V. JAROŠÍK, P. E. HULME, J. PERGL, M. HEJDA, U. SCHAFFNER, AND M. VILÀ.

2012. A global assessment of invasive plant impacts on resident species,

communities and ecosystems: the interaction of impact measures, invading species’

traits and environment. Glob. Chang. Biol. 18:1725–1737.

RANDALL, J. M. 1996. Weed control for the preservation of biological diversity. Weed

Technol. 10:370–383.

RATHFON, R., AND K. RUBLE. 2007. Herbicide treatments for controlling invasive Bush

honeysuckle in a mature hardwood forest in West-Central Indiana. Pages 187–197

Proceedings of the 15th Central Hardwood Forest Conference.

RICHARDSON, D. M., N. ALLSOPP, C. M. D’ANTONIO, S. J. MILTON, AND M. REJMÁNEK.

2000. Plant invasions--the role of mutualisms. Biol. Rev. Camb. Philos. Soc. 75:65–

93.

RIETVELD, W. J. 1983. Allelopathic effects of Juglone on germination and growth of

several herbaceous and woody species. J. Chem. Ecol. 9:295–308.

RIFFLE, J. W., AND J. F. WATKINS. 1986. Honeysuckle leaf blight: general technical report

RM-129. (J. W. Riffle and G. W. Peterson, Eds.) Diseases of Trees in the Great

Plains.

RODEWALD, A. D. 2009. Urban-associated habitat alteration promotes brood parasitism of

Acadian Flycatchers. J. F. Ornithol. 80:234–241.

RODEWALD, A. D., D. P. SHUSTACK, AND L. E. HITCHCOCK. 2010. Exotic shrubs as

ephemeral ecological traps for nesting birds. Biol. Invasions 12:33–39.

ROONEY, T. P., AND W. J. DRESS. 1997. Species loss over sixty-six years in the ground-

layer vegetation of Heart’s Content, an old-growth forest in Pennsylvania USA. Nat.

Areas J. 17:297–305.

ROONEY, T. P., AND D. M. WALLER. 2003. Direct and indirect effects of white-tailed deer

in forest ecosystems. For. Ecol. Manage. 181:165–176.

RUESINK, J. L., I. M. PARKER, M. J. GROOM, AND P. M. KAREIVA. 1995. Reducing the

risks of reducing nonindigenous species introductions. Bioscience 45:465–477.

RUNKLE, J. R., A. DISALVO, Y. GRAHAM-GIBSON, AND M. DORNING. 2007. Vegetation

release eight years after removal of Lonicera maackii in West-Central Ohio. Ohio J.

Sci. 107:125–129.

SAGOFF, M. 2005. Do invasive species threaten the environment? J. Agric. Environ.

Ethics 18:215–236.

SALTONSTALL, K. 2002. Cryptic invasion by a non-native genotype of the common reed,

Phragmites australis, into North America. Proc. Natl. Acad. Sci. U. S. A. 99:2445–

2449.

SCHRADIN, K., AND D. CIPOLLINI. 2012. The sign and strength of plant-soil feedback for

the invasive shrub, Lonicera maackii, varies in different soils. Forests 3:903–922.

SCHULZ, K. E., J. WRIGHT, AND S. ASHBAKER. 2012. Comparison of invasive shrub

honeysuckle eradication tactics for amateurs: stump treatment versus regrowth

spraying of Lonicera maackii. Restor. Ecol. 20:788–793.

SHANNON, S. M., J. T. BAUER, W. E. ANDERSON, AND H. L. REYNOLDS. 2014. Plant-soil

feedbacks between invasive shrubs and native forest understory species lead to shifts

in the abundance of mycorrhizal fungi. Plant Soil 382:317–328.

SHEA, K., AND P. CHESSON. 2002. Community ecology theory as a framework for

biological invasions. Trends Ecol. Evol. 17:170–176.

SHEWHART, L., R. W. MCEWAN, AND M. E. BENBOW. 2014. Evidence for facilitation of

Culex pipiens (Diptera: Culicidae) life history traits by the nonnative invasive shrub

Amur honeysuckle (Lonicera maackii). Entomol. Soc. Am. 43:1584–1593.

SHIELDS, J. M., M. A. JENKINS, P. A. ZOLLNER, AND M. R. SAUNDERS. 2014. Effects of

Amur honeysuckle invasion and removal on white-footed mice. J. Wildl. Manage.

78:867–880.

SHOUSE, M., L. LIANG, AND S. FEI. 2013. Identification of understory invasive exotic

plants with remote sensing. Int. J. Appl. Earth Obs. Geoinf. 21:525–534.

SIMBERLOFF, D. 2003. Eradication - preventing invasions at the outset. Weed Sci.

51:247-253.

SIMBERLOFF, D. 2006. Invasional meltdown 6 years later: important phenomenon,

unfortunate metaphor, or both? Ecol. Lett. 9:912–919.

SIMBERLOFF, D., AND B. VON HOLLE. 1999. Positive interactions of nonindigenous

species: invasional meltdown? Biol. Invasions 1:21–32.

STINSON, K. A., S. A. CAMPBELL, J. R. POWELL, B. E. WOLFE, R. M. CALLAWAY, G. C.

THELEN, S. G. HALLETT, D. PRATI, AND J. N. KLIRONOMOS. 2006. Invasive plant

suppresses the growth of native tree seedlings by disrupting belowground

mutualisms. PLoS Biol 4:727–731.

STROMBERG, J. C., AND M. K. CHEW. 2002. Chapter 11: Foreign visitors in riparian

corridors of the American Southwest. Pages 195–227 In B. Tellman [ed.], Invasive

Exotic Species in the Sonoran Region. Tucson, AZ.

STROMBERG, J. C., M. K. CHEW, P. L. NAGLER, AND E. P. GLENN. 2009. Changing

perceptions of change: the role of scientists in Tamarix and river management.

Restor. Ecol. 17:177–186.

THEOHARIDES, K. A, AND J. S. DUKES. 2007. Plant invasion across space and time: factors

affecting nonindigenous species success during four stages of invasion. New Phytol.

176:256–273.

THOMAS, H. E. 1963. Causes of depletion of the Pecos River in New Mexico. U.S.

Geological Survey Water-Supply.

THOMPSON, K., AND M. A. DAVIS. 2011. Why research on traits of invasive plants tells us

very little. Trends Ecol. Evol. 26:155–1556.

TILGHMAN, N. G. 1989. Impacts of white-tailed deer on forest regeneration in

Northwestern Pennsylvania. J. Wildl. Manage. 53:524–532.

TRAMMELL, T. L. E., AND M. M. CARREIRO. 2011. Vegetation composition and structure

of woody plant communities along urban interstate corridors in Louisville, KY,

U.S.A. Urban Ecosyst. 14:501–524.

TRAMMELL, T. L. E., H. A. RALSTON, S. A. SCROGGINS, AND M. M. CARREIRO. 2012.

Foliar production and decomposition rates in urban forests invaded by the exotic

invasive shrub, Lonicera maackii. Biol. Invasions 14:529–545.

TRAVESET, A., AND D. M. RICHARDSON. 2006. Biological invasions as disruptors of plant

reproductive mutualisms. Trends Ecol. Evol. 21:208–216.

USDA. 1999. Lonicera maackii ( Rupr.) Herder. Retrieved January 1, 2014, from USDA.

USDOI. 2015. National invasive species council. Retrieved August 1, 2015, from

USDOI. < https://www.doi.gov/invasivespecies>.

VELLEND, M., K. VERHEYEN, H. JACQUEMYN, A. KOLB, H. VAN CALSTER, G. PETERKEN,

AND M. HERMY. 2006. Extinction debt of forest plants persists for more than a

century following habitat fragmentation. Ecology 87:542–548.

VILÀ, M., J. L. ESPINAR, M. HEJDA, P. E. HULME, V. JAROŠÍK, J. L. MARON, J. PERGL, U.

SCHAFFNER, Y. SUN, AND P. PYŠEK. 2011. Ecological impacts of invasive alien

plants: a meta-analysis of their effects on species, communities, and ecosystems.

Ecol. Lett. 14:702–708.

VIVANCO, J. M., H. P. BAIS, F. R. STERMITZ, G. C. THELEN, AND R. M. CALLAWAY. 2004.

Biogeographical variation in community response to root allelochemistry: novel

weapons and exotic invasion. Ecol. Lett. 7:285–292.

WAIPARA, N. W., C. J. WINKS, L. A. SMITH, AND J. P. WILKIE. 2007. Natural enemies of

Japanese honeysuckle, Lonicera japonica, in New Zealand. New Zeal. Plant Prot.

60:158–163.

WATLING, J. I., C. R. HICKMAN, E. LEE, K. WANG, AND J. L. ORROCK. 2011a. Extracts of

the invasive shrub Lonicera maackii increase mortality and alter behavior of

amphibian larvae. Oecologia 165:153–159.

WATLING, J. I., C. R. HICKMAN, AND J. L. ORROCK. 2011b. Predators and invasive plants

affect performance of amphibian larvae. Oikos 120:735–739.

WATLING, J. I., C. R. HICKMAN, AND J. L. ORROCK. 2011c. Invasive shrub alters native

forest amphibian communities. Biol. Conserv. 144:2597–2601.

WHITE, R. J., M. M. CARREIRO, AND W. C. ZIPPERER. 2014. Woody plant communities

along urban, suburban, and rural streams in Louisville, Kentucky, USA. Urban

Ecosyst. 17:1061–1094.

WILCOVE, D. S., D. ROTHSTEIN, J. DUBOW, A. PHILLIPS, AND E. LOSOS. 1998. Threats to

imperiled quantifying species in the United States. Bioscience 48:607–615.

WILFONG, B. N., D. L. GORCHOV, AND M. C. HENRY. 2009. Detecting an invasive shrub in

deciduous forest understories using remote sensing. Weed Sci. 57:512–520.

WILSON, H. N., M. A. ARTHUR, A. SCHÖRGENDORFER, R. D. PARATLEY, B. D. LEE, AND R.

W. MCEWAN. 2013. Site Characteristics as Predictors of Lonicera maackii in

second-growth forests of Central Kentucky , USA. Nat. Areas J. 33:189–198.

WITMER, M. C. 1996. Consequences of an alien shrub on the plumage coloration and

ecology of Cedar Waxwings. Auk 113:735–743.

XU, C. Y., K. L. GRIFFIN, AND W. S. F. SCHUSTER. 2007. Leaf phenology and seasonal

variation of photosynthesis of invasive Berberis thunbergii (Japanese barberry) and

two co-occurring native understory shrubs in a northeastern United States deciduous

forest. Oecologia 154:11–21.

YATES, E. D., D. F. LEVIA, AND C. L. WILLIAMS. 2004. Recruitment of three non-native

invasive plants into a fragmented forest in southern Illinois. For. Ecol. Manage.

190:119–130.

FIGURE LEGENDS:

Figure 1.1: Lonicera maackii fall fruit production in (a) terrestrial and (b) stream run habitats, (c) freeze resistant leaves in mid-December 2009, and (d) senesced leaf litter in a headwater stream in southwestern Ohio, USA. Photos were taken by Rachel E.

McNeish.

Figure 1.2: Scanning electron microscope images of fresh L. maackii (a) and P. serotina

(b) leaves depicting trichome density and aquatic microbial and fungal growth on L. maackii (c) and P. serotina (d) leaves anchored in a headwater stream for three days.

Pilot study was conducted summer of 2010 with L. maackii and P. serotina leaf packs anchored in a small headwater stream for 7 days. Photos were taken by Anastasia Stolz.

Figure 1.3: Predictive framework for Lonicera maackii impacts across ecosystems at multiple ecological scales.

FIGURES:

Figure 1.1:

Figure 1.2:

Figure 1.3:

CHAPTER 2: REMOVAL OF THE INVASIVE SHRUB, LONICERA MAACKII,

FROM RIPARIAN FORESTS INFLUENCES HEADWATER STREAM BIOTA AND

ECOSYSTEM FUNCTION

ABSTRACT:

Riparian forests and streams are interlinked by cross-system subsidies and alterations of the terrestrial environment can have substantial effects on aquatic biota and ecosystem function. In the Midwestern USA the exotic shrub Lonicera maackii (Amur honeysuckle) has successfully invaded many riparian habitats, creating near- monocultures in some locations. This terrestrial invasion has strong potential to modify cross-system subsidies and impact stream ecosystems. We removed L. maackii from a riparian forest to assess impacts on the aquatic environment. In August 2010, removal occurred along a 150m stream reach, 10m downstream of a non-removal reach, before natural leaf senescence. Over 74 d, in-stream leaf litter (organic matter: OM) was collected weekly from plots located in riffles (5/reach). Benthic algal biomass, above stream canopy cover, and macroinvertebrate density were measured for 18 months.

Lonicera maackii removal was associated with decreased canopy cover and a significant increase in total in-stream leaf OM in early autumn (P < 0.05). Removal also differentially influenced the timing and abundance of specific leaf litter genera within the stream (P < 0.05). Macroinvertebrate density was significantly higher in the removal reach, especially during autumn one year after removal (P = 0.0294). In both reaches,

macroinvertebrate density peaks lagged behind benthic algal biomass peaks. In summary, the removal of an invasive riparian shrub influenced the timing, deposition, quality, and abundance of leaf litter habitat into a headwater stream, ostensibly driving bottom-up effects on aquatic primary producer biomass and the macroinvertebrate community.

INTRODUCTION:

Riparian forests and streams are interlinked by the transfer of materials and energy between these habitats. Subsides that enter the stream from the terrestrial environment (allochthonous inputs) provide a critical nutrient and energy resource and act as habitat for aquatic biota (Vannote et al., 1980; Webster et al.,1995; Polis et al.,

1997, Paetzold et al., 2011). Riparian forests are highly susceptible to invasion by exotic plants, some of which have strong ecological effects (Futoshi et al., 2000; Nakamura et al., 2000; Daehler, 2003; Richardson et al., 2007). For example, deposition of leaf litter from invasive plants can change the rate of organic matter processing, alter nutrient availability, and modify habitat heterogeneity within the aquatic system (Kennedy and

Hobbie, 2004; Swan and Palmer, 2004; Swan et al., 2008; McNeish et al., 2012).

Riparian invasive shrubs and trees known to alter aquatic ecosystems include Tamarix spp. (Salt cedar), which lowers groundwater and depth in lentic and lotic habitats due to extremely high evapotranspiration rates (Brotherson et al., 1984; Brotherson and Winkel,

1986; Di Tomaso, 1998; Stromberg et al., 2007) and Elaeagnus angustifolia (Russian olive), which has been shown to contribute nitrogen to streams (Mineau et al., 2011).

The non-native, invasive shrub Lonicera maackii (Rupr.) Maxim. (Amur honeysuckle) is a terrestrial deciduous shrub that can impact aquatic ecosystem processes

(McNeish et al., 2012). Dense invasion by this species has provoked removal and restoration efforts in many natural areas in eastern North America which provides a scientific opportunity for understanding cross-systems subsidy effects of an exotic plant

(Luken and Therit, 1996; USDA, 2011). Lonicera maackii negatively impacts the survivorship, reproduction, and growth of neighboring plants through competition and allelopathy (Gould and Gorchov, 2000; Gorchov and Trisel, 2003; Hartman and

McCarthy, 2004; Cipollini et al., 2008a, Cipollini et al., 2008b; McEwan et al., 2009b).

McEwan et al. (2009b) demonstrated that L. maackii leaf litter had adverse effects on larvae survivorship of the polyphagous moth Lymantria dispar (European gypsy moth).

Extracts from L. maackii leaves inhibited amphibian larval survivorship and growth and increased surfacing behavior and susceptibility to predation (Walting et al., 2011).

Lonicera maackii has higher evapotranspiration rates in wetlands, which may further negatively impact amphibians that rely on the availability of ephemeral ponds and small streams to complete their life cycles (Boyce et al., 2012). Lonicera maackii litter decomposes 5 × faster than native species, such as Fraxinus americana (White ash) and

Carya spp. (Hickory), in terrestrial forests and exhibits a unique community of microbial decomposers (Arthur et al., 2012). Furthermore, dense stands of L. maackii have been reported to mediate vector and host dynamics of tick-borne disease (Allan et al., 2010).

Since this invasive species supplants natives in riparian corridors, creating near- monocultures and a dense canopy overarching headwater streams, there is potential for L. maackii to influence cross-system subsidies (e.g., nutrients, allochthonous materials) with ultimate effects on aquatic trophic interactions and ecosystem processes (Fig. 1).

In a previous study, we found that L. maackii foliage breakdown in streams was 4

× faster than native species with differences in macroinvertebrate functional feeding group dominance in L. maackii leaf packs compared to native species (McNeish et al.,

2012). In this study, we assessed larger-scale effects by removing L. maackii in the riparian zone of a headwater stream and measuring terrestrial inputs and the response of the aquatic biota. Lonicera maackii has a long leaf duration (McEwan et al., 2009a) and creates dense canopies overarching headwater streams. Therefore, we hypothesized (H1) that the stream reach with a L. maackii dominated riparian zone would have a greater total volume of leaf material entering the stream, as compared to the L. maackii removal reach, due to copious leaf deposition by the invasive shrub. We also hypothesized (H2) that riparian removal of L. maackii would result in an increase in chlorophyll a from benthic biofilm communities due to increased light conditions associated with riparian canopy removal. Finally, we hypothesized (H3) that L. maackii riparian forests would result in a decrease in macroinvertebrate density due to bottom-up effects on the temporal dynamics of in-stream organic matter and chlorophyll a standing stock biomass (Fig. 1).

METHODS:

Study Site:

The study site was located in the upper reaches of a 3rd order headwater stream in the Little Miami watershed within Black Oak Park of the Centerville-Washington Park

District in southwestern Ohio, USA (84.12°W; 36.63°N; hereafter referred to as Black

Oak (BO). The streambed was 1.5 – 4.0 m wide, and the benthic substrata consisted of rocks, sand, and clay and was formed on limestone geology (Schneider, 1957). This

tributary was chosen because it was typical of L. maackii dominated headwater streams in the region, and the site had an abundance of riffle, run, and pool habitats.

Experimental Lonicera maackii Removal:

The riparian flora of BO was surveyed July 2010 within a 160 x 5m reach along both stream banks due to accessibility of the riparian zone and the size of the park.

Removal of aboveground L. maackii biomass (approx. 63% of riparian basal area), and all other minor (< 2%) woody invasive flora (Ligustrum sp. (Privet) and Elaeagnus umbellata (Autumn olive)), was completed within the designated removal area in August

– September 2010 to create a 150m removal reach and an up-stream non-removal reach.

The buffer between these two reaches was 10m. All debris from the invasive species was disposed of offsite. AquaNeat® Aquatic Herbicide, an Ohio EPA approved aquatic herbicide (EPA regulation number: 228-365; Nufarm Manufacturer; active ingredient

Glyphosate N-glycine), was applied to cut stumps of invasive species within 48h to prevent re-growth in future growing seasons. There was no rainfall after AquaNeat herbicide application. All native plants were left uncut, and all native species coarse organic matter (COM; e.g., leaves, snags) was left in place.

Leaf Litter Accumulation:

In-stream leaf litter collection plots (3 per riffle) were installed haphazardly in the thalweg (deepest channel) of riffles (n = 5) within the removal and non-removal reaches on 31 October 2010 to assess benthic leaf accumulation. Natural leaf buildup was collected by the placement of a PVC frame (12.5 × 12.5 cm) above each plot and removing any leaf material present within the frame. Accumulated leaf litter within each plot was collected weekly from 7 November 2010 to 13 January 2011 and preserved in

70% ethanol until processed in the lab. Leaves were rinsed, identified to genus when possible or classified as unknown (13%), and dried at 50º C for 48h. Lonicera maackii leaf litter was obvious regardless of decomposition stage, and so, all unknown leaf litter was classified as native. Sub-samples of dried leaves were then combusted at 550º C for ash-free-dry-mass (AFDM; Benfield, 2007) estimates. To determine OM for individual leaf genera, leaf collection sub-plots were averaged at the riffle scale to create a riffle average, which allowed riffles to be used as the unit of replication for each stream reach to calculate total organic matter available at each time point and the rate of organic matter accumulation.

Benthic Algal Biomass and Above Stream Canopy Cover:

Benthic algal biomass was assessed monthly (July 2010 – May 2012) from a small rock that was haphazardly collected from each riffle when water was present (n = 5 per reach), stored in a foil covered canister, and frozen at 0° C until processed within three months of collection. Rock samples were thawed, scrubbed, and rinsed to dislodge collected algal growth. Each algal sample was brought up to 500mL with DI water, homogenized by stirring, and then 50 mL were filtered through a GC-50 25mm glass fiber membrane filter. Chlorophyll a pigment estimates were made using methods outlined by APHA (1999) and Steinman et al. (2007) and used to calculate benthic algal biomass. Above stream canopy cover was recorded monthly for each riffle in the North,

East, South, and West cardinal directions with a spherical densiometer (Lemmon 1956).

Macroinvertebrate Community:

Monthly Surber samples (n = 5 per reach) were collected when water was present in the stream (generally September – July). The Surber sampler was haphazardly placed

in a riffle, and all benthic substrata were then scrubbed with a bristle brush to dislodge macroinvertebrates. Displaced macroinvertebrates were caught in the sampler net and preserved in 70% ethanol. Samples were sorted for macroinvertebrates in the lab to determine macroinvertebrate density.

Statistical Analyses:

Organic matter (i.e. leaves) inputs to the stream were compared between the L. maackii removal and non-removal reaches to assess the effect of riparian L. maackii on subsides entering the stream from 7 November 2010 to 13 January 2011. Within each sampling date, we compared the mean organic matter (AFDM) from total, native, and

Lonicera leaf litter in the removal and non-removal reaches. The data were first screened for normality using the Shapiro-Wilk normality test, and normally distributed data were compared with paired t-tests. Those data without a normal distribution were compared using Wilcoxon matched-pairs tests (Sokal and Rohlf, 1981; Zar, 1999). These approaches were conducted for all comparisons between L. maackii removal and non- removal reaches.

We next compared organic matter abundance of the most dominant leaf genera making up the streambed organic matter within the L. maackii removal and non-removal reaches. Five leaf litter genera composed ~78% of all leaf organic matter: Lonicera,

Platanus spp. (Sycamore), Acer spp. (Maple), Quercus spp. (Oak), and Fraxinus spp.

(Ash). We compared those genera and combined all other genera into an “other” category for analyses. To compare leaf litter dominance within each treatment reach, we used repeated measures ANOVA (rmANOVA) to analyze mean organic matter of the top five most dominant species, and, where significant model differences were detected, pair-

wise comparisons were made with Bonferroni post-tests. To compare leaf litter dominance between treatment reaches, the mean organic matter for the top five leaf genera were compared within sampling date with paired t-tests.

Leaf OM accumulation rates for total, native, and L. maackii leaf litter were calculated over 35 d and statistically compared between L. maackii removal and non- removal reaches using linear regression models. Canopy cover, benthic algal biomass, and macroinvertebrate density comparisons were made within each sampling point between L. maackii removal and non-removal reaches with Wilcoxon matched pairs test or paired t-tests per Shapiro-Wilk normality tests (Sokal and Rohlf, 1981; Zar, 1999). All statistical analyses were conducted with GraphPad Prism version 5.0 (GraphPad

Software, San Diego CA, USA, www.graphpad.com).

RESULTS:

In-stream Leaf Material:

We found significantly more total leaf material in the streambed of the removal reach than in the non-removal in the early autumn (Fig 2a). This difference was driven by an increase in native inputs (Fig. 2b). Lonicera leaf OM was greater in the non- removal reach, and this difference was statistically significant from 7-14 November (all P

< 0.05; Table 1; Fig. 2c). Total and native leaf OM in-stream accumulation rates were not significantly different between L. maackii removal and non-removal reaches (all P >

0.05; Table 2). Lonicera leaf OM accumulation rate was significantly higher in the non- removal reach, nearly doubling the rate in the removal reach (Table 2).

The number of leaf taxa entering the stream during the study was identical between reaches (24), but the mean leaf OM from individual leaf genera varied between 65

reaches and through time (Table 3; Fig. 3). In the non-removal reach, L. maackii contributed ~25% of leaf OM between 7-21 November (Fig. 3a). In contrast, when L. maackii was removed from the riparian zone, Platanus spp. significantly dominated the in-stream leaf litter community (~33%) on 7-14 November (P <0.01; Fig. 3b). Of all the dominant leaf genera, only Platanus spp. and Lonicera (ESM Table 1) were different between stream reaches. Platanus spp. was significantly greater in the removal reach from 7-21 November, while Lonicera was significantly greater in the non-removal reach from 7-14 November (all P < 0.05). All other leaf genera were statistically indistinguishable between reaches for all sampling dates (P > 0.05; Fig. 4).

Canopy Cover and Benthic Algal Biomass:

As expected, canopy cover above the stream was reduced by the removal of L. maackii (Table 1; Fig. 5), and this difference was statistically significant for several months after removal (P <0.05; Fig. 5). Canopy cover ranged from ~70-100% in the non-removal reach and ~20-80% in the L. maackii removal reach. While there were some monthly differences in benthic chlorophyll a, there were no overall differences between the reaches throughout the study (Table 1; Fig. 6a). Chlorophyll a peaked twice during the study (Fig 6a). The first production peak was one month earlier and 31% higher in the L. maackii removal reach than the non-removal reach (Fig. 6a). During the second peak, chlorophyll a concentration were erratic in the non-removal reach, spiking to levels much higher than in the removal reach (Fig. 6a).

Macroinvertebrate Density:

Macroinvertebrate density was significantly greater in the stream reach associated with L. maackii removal, especially during the autumn of 2011 (Table 1; Fig. 6b). We

detected two peaks in macroinvertebrate density with the greatest densities in both reaches during April 2012 (Fig. 6b). Within the non-removal reach, macroinvertebrate density peaked one month after the first benthic chlorophyll a standing stock biomass peak and three months after the second peak. In the L. maackii removal reach, macroinvertebrate density peaks lagged four months behind chlorophyll a standing stock biomass each year.

DISCUSSION:

In-stream Organic Matter:

The presence of L. maackii in the riparian community substantially altered the timing and composition of terrestrial organic matter subsidies entering the stream.

Removal of this invasive species led to an increase in the amount of leaf litter present within the stream, which was a surprising result that contradicted our original hypothesis

(H1). Lonicera maackii’s dense canopy across the stream may have acted as a physical barrier or 'filter' preventing or delaying the entrance of senesced native leaves from the upper riparian canopy (personal observations). Lonicera maackii retains most of its leaves until late November – early December (McEwan et al., 2009a, Arthur et al., 2012;

Fig. 4); therefore, this dense canopy may persist throughout the fall season. We postulate that when L. maackii was removed, so was this barrier, which allowed increased volumes of native material to drop into the stream during autumn 2010. Many studies have examined the effects of riparian vegetation removal on aquatic ecosystems (Hetrick et al.,

1998; Sabater et al., 2000; MacKenzie, 2008; Jowett et al., 2009; Reid et al., 2010), but to our knowledge, this data set is the first to demonstrate an increase in allochthonous inputs (leaf litter) associated with the removal of a dominant riparian invasive plant. We

found significant variation in in-stream leaf genera COM between reaches (H1).

Lonicera maackii was the dominant contributor to the in-stream leaf litter within the non- removal reach. In contrast, the leaf litter community within the removal reach was dominated by native taxa such as Platanus spp. The large size of Platanus spp. leaves may have caused this species, in particular, to be prone to being trapped and held within the L. maackii shrub canopy, preventing or delaying its entry to the aquatic substrate

(Henry and Flood, 1920). These results suggest that (1) the physical presence of riparian zone invaders can create a significant shift in the timing, quantity, and quality of native allochthonous materials entering streams and (2) the in-stream benthic habitat heterogeneity (as derived from leaf COM) can be substantially altered by terrestrial invasion.

Macroinvertebrate Density and Benthic Algal Biomass:

Aquatic macroinvertebrate density may in part be regulated by the availability of benthic algal biomass, which can result in a bottom-up effects on trophic interactions.

We found that algal biomass increased immediately after autumnal leaf senescence; however, the overall availability was the same between stream reaches, indicating that riparian removal may have influenced the timing of algal biomass (H2).

Macroinvertebrate peaks followed chlorophyll a, suggesting macroinvertebrates were partially dependent on this resource (H3; Fig. 1). These results suggest a link between senescence events and increased epilithic biofilm production, a common food resource for certain benthic macroinvertebrates (Cummins and Klug, 1979; Feminella and

Hawkins, 1995; Steinman, 1996; Carrick et al., 2012). Magoba and Samways (2010) reported that aquatic macroinvertebrate taxon richness increased following large-scale

removal of invasive trees in riparian forests in Africa, suggesting macroinvertebrate assemblages can recover post riparian invasive removal. Gratton and Denno (2005) also found an increase in arthropod communities in salt marshes once invasive Phragmites australis (Phragmites) was removed. Several studies have demonstrated leaf litter composition influences macroinvertebrate community structure (Abelho and Graca,

1995; Kominoski et al., 20011; McNeish et al., 2012); and our study suggests that the removal of L. maackii, which had such a substantial contribution to the in-stream litter community, had a strong impact on aquatic macroinvertebrates. Monitoring at the site is ongoing to assess long-term effects on light availability, benthic algal biomass, and macroinvertebrate life histories, secondary production, and community dynamics.

In summary, our data (1) suggests that invasive plants moving into the riparian zone of headwater streams can alter the aquatic food web and (2) provide initial support for a mechanistic explanation of this effect (Fig. 1). Removal of the invasive species resulted in a temporal shift and increase in the abundance of native leaf litter within the stream (H1). This alteration of allochthonous inputs, which serve as habitat and food resources for aquatic organisms, can drive alterations in aquatic macroinvertebrates (H3).

We postulate that filtering of organic matter by the shrub canopy is an important mechanism effect. From an applied perspective, our results suggest that removing invasive species from the riparian zone of streams, even over relatively short reaches

(150 m in our case) of otherwise heavily invaded streams, can have a substantial local influence on aquatic biota and in-stream ecosystem function.

ACKNOWLEDGMENTS:

We would like to thank the Centerville-Washington Park District for access to the site and Lucy Siefker who helped coordinate our project with park district. Assistance in the field and lab was provided by dozens of McEwan lab members including Amy

Hruska, Charlie Jackson, Simon McClung, and Joe Branner. Helpful comments on previous versions of the manuscript were provided by Caitlin Bojanowski, Jessica Davis, and Casey Hanley. The project was partially supported in part by a University of Dayton

Learn, Lead, and Serve grant awarded to Eryn Moore. This work was also supported in part by the University of Dayton Office for Graduate Academic Affairs through the

Graduate Student Summer Fellowship Program and by the National Science Foundation

(DEB 1352995).

REFERENCES:

Abelho M, Graca MA 1996. Effects of eucalyptus afforestation on leaf litter

dynamics and macroinvertebrate community structure of streams in Central

Portugal. Hydropbiologia 324: 195-204.

Allan B F, Dutra HP, Goessling LS, Barnett D, Chase JM, Marquis RJ, Pang G, Storch

GA, Thach RE, Orrock JL. 2010. Invasive honeysuckle eradication reduces tick-

borne disease risk by altering host dynamics. Proceedings of the National

Academy of Science 107: 18523-18527.

American Public Health Association. 1999. Standard methods for the examination of

water and Wastewater. (20th ed). Washington D.C.: American Public Health

Association.

Arthur MA, Bray SR, Kuchle C, McEwan RW. 2012. The influence of the

invasive shrub, Lonicera maackii, on leaf decomposition and microbial

community dynamics. Plant Ecology 213: 1571-1582.

Benfield EF. 2007. Decomposition of Leaf Material. In: Hauer FR, Lamberti GA (Eds).

Methods in stream ecology (2nd ed). (pp. 711-720). San Diego: Academic Press.

Boyce RL, Durtsche RD, Fugal SL. 2012. Impact of the invasive shrub Lonicera

maackii on stand transpiration and ecosystem hydrology in a wetland forest.

Biological Invasions 14: 671-680.

Brotherson JD, Carman JG, Szyska LS. 1984. Stem-diameter age relationships

of Tamarix ramosissima in central Utah. Journal of Range

Management 37: 362- 364

Brotherson JD, Winkel V. 1986. Habitat relationships of saltcedar (Tamarix

ramosissima) in central Utah. Great Basin Naturalist 46: 535-541.

Carrick HJ, Dananay KL, Eckert RA, Price KJ. 2012. Decomposition during

autumn foliage leaf-fall in wetlands situated along a biogeochemical gradient in

Pennsylvania, USA. Journal of Freshwater Ecology 27: 1-17.

Cipollini D, Stevenson R, Cipollini D. 2008a. Contrasting effects of

allelochemicals from two invasive plants on the performance of a nonmycorrhizal

plant. International Journal of Plant Science 169: 371-375.

Cipollini D, Stevensen R, Enright S, Eyles A, Bonello P. 2008b. Phenolic metabolites in

leaves of the invasive shrub, Lonicera maackii, and their potential phytotoxic and

anti-herbivore effects. Journal of Chemical Ecology 34: 144-152.

Cummins KW, Klug MJ 1979. Feeding ecology of stream invertebrates. Annual Review

of Ecology and Systematics 10: 147-172.

Daehler CC. 2003. Performance comparisons of co-occurring native and alien invasive

plants: implications for conservation and restoration. Annual Review of Ecology,

Evolution and Systems 34: 183-211.

Di Tomaso JM. 1998. Impact, biology, and ecology of saltcear (Tamarix spp.) in the

southwestern United States. Weed Technology 12: 326-336.

Feminella JW, Hawkins CP. 1995. Interactions between stream herbivores and

periphyton: a quantitative analysis of past experiments. Journal of the North

American Benthological Society 14: 465-509.

Futoshi N, Swanson FJ, Wondzell SM. 2000. Disturbance regimes of stream and

riparian systems – a disturbance-cascade perspective. Hydrological Processes 14:

2849–2860.

Gorchov DL, Trisel DE. 2003. Competitive effects of the invasive shrub, Lonicera

maackii (Rupr.) Herder (Caprifoliaceae), on the growth and survival of native tree

seedlings. Plant Ecology 166: 13-24.

Gould AMA, Gorchov DL. 2000. Effects of the exotic invasive shrub Lonicera

maackii on the survival and fecundity of three species of native annuals.

American Midland Naturalist 144: 36-50.

GraphPad Prism, URL http://www.graphpad.com/support/faqid/248/ [accessed on 6

August 2013].

Gratton C, Denno RF. 2005. Restoration of arthropod assemblages in a Spartina

salt marsh following removal of the invasive plant Phragmites australis.

Restoration Ecology 13: 358-372.

Hartman KM, McCarthy BC. 2004. Restoration of a forest understory after the

removal of an invasive shrub, Amur honeysuckle (Lonicera maackii). Restoration

Ecology 12: 154-165.

Henry A, Flood MG. 1920. The history of the London Plane, Platanus acerifolia,

with notes on the genus Platanus. Proceedings of the Royal Irish Academy 35: 9-

28.

Hetrick NJ, Brusven MA, Bjornn TC, Keith RM, Meehan WR. 1998. Effect of canopy

removal on invertebrates and diet of juvenile Coho salmon in a small stream

in southeast Alaska. Transactions of the American Fisheries Society 127: 876-

888.

Jowett IG, Richardson J, Boubee JAT. 2009. Effects of riparian manipulation on

stream communities in small streams: Two case studies. New Zealand Journal of

Marine and Freshwater Research 43: 763-774.

Kennedy TA, Hobbie SE. 2004. Saltcedar (Tamarix ramosissima) invasion alters

organic matter dynamics in a desert stream. Freshwater Biology 49: 65-76.

Kominoski JS, Marczak LB, Richardson JS. 2011. Riparian forest composition

affects stream litter decomposition despite similar microbial and invertebrate

communities. Ecology 92: 151-159.

Lemmon PE. 1956. A spherical densiometer for estimating forest overstory density.

Forest Science 2: 314–320

Luken JO, Thieret JW. 1996. Amur honeysuckle, its fall from grace. BioScience 46: 18-

24.

MacKenzie RA. 2008. Impacts of riparian forest removal on Palauan streams.

Biotropica 40: 666-675.

Magoba RN, Samways MJ. 2010. Recovery of benthic macroinvertebrate and

adult dragonfly assemblages in response to large scale removal of riparian

invasive alien trees. Journal of Insect Conservation 14: 627-636.

McEwan RW, Birchfield MK, Schoergendorfer A, Arthur MA. 2009a. Leaf

phenology and freeze tolerance of the invasive shrub Amur honeysuckle and

potential native competitors. Journal of the Torrey Botanical Society 136: 212-

220.

McEwan RW, Rieske LK, Arthur MA. 2009b. Potential interaction between

invasive woody shrubs and the gypsy moth (Lymantria dispar), an invasive insect

herbivore. Biological Invasions 11: 1053-1058.

McEwan RW, Arthur-Paratley LG, Rieske LK, Arthur MA. 2010. A multi-

assay comparison of seed germination inhibition by Lonicera maackii and co-

occurring native shrubs. Flora 205: 475-483.

McNeish RE, Benbow ME, McEwan RW. 2012. Riparian forest invasion by a

terrestrial shrub (Lonicera maackii) impacts aquatic biota and organic matter

processing in headwater streams. Biological Invasions 14: 1881-1893.

Mineau MM, Baxter CV, Marcarelli AM. 2011. A non-native riparian tree

(Elaeagnus angustifolia) changes nutrient dynamics in streams. Ecosytems 14:

353-365.

Nakamura F, Swanson FJ, Wondzell SM. 2000. Disturbance regimes of stream

and riparian systems – a disturbance-cascade perspective. Hydrological

Processes 14: 2849-2860.

Paetzold A, Smith M, Warrant P, Maltby L. 2011 Environmental impact propagated by

cross-system subsity Chronic stream pollution controls riparian spider

populations. Ecology 92: 1711-1716.

Polis GA, Anderson WB, Holt RD. 1997. Toward an integration of landscape and food

web ecology: the dynamics of spatially subsidized food webs. Annual Review of

Ecology and Systematics 28: 289-316.

Reid DJ, Quinn JM, Wright-Stow AE. 2010. Responses of stream

macroinvertebrate communities to progressive forest harvesting: influences of

harvest intensity, stream size and riparian buffers. Forest Ecology and

Management 260: 1804-1815.

Richardson DM, Holmes PM, Esler KJ, Galatowitsch SM, Strogerg JC,

Kirkman SP, Pysek P, Hobbs RJ. 2007. Riparian vegetation: degradation,

alien plant invasions, and restoration prospects. Diversity and Distributions 13:

126-139.

Sabater F, Butturini A, Marti E, Munoz I, Romani A, Wray J, Sabater S.

2000. Effects of riparian vegetation removal on nutrient retention in a

Mediterranean stream. Journal of the North American Benthological Society 19:

609-620.

Schneider WJ. 1957. Relation of geology to stream flow in the upper little Miami

basin. The Ohio Journal of Science 57: 11-14.

Sokal RR, Rohlf FJ 1981. Biometery: The principles and practice of statistics in

biological research. (2nd Ed). New York: W.H. Freeman and Company.

Steinman AD, Lamberti GA, Leavitt PR. 2007. Biomass and pigments of

benthic algae. In: Hauer FR, Lamberti GA. (ed). Methods in stream ecology

(2nd ed). (pp. 357-379). San Diego: Academic Press.

Steinman AD. 1996. Effects of grazers on freshwater benthic algae. In: Stevensen JR,

Bothwell ML, Lowe RL. (eds). Algal ecology: freshwater benthic ecosystems

(pp. 341-374). New York: Academic Press.

Stromberg JC, Lite SJ, Marler R, Paradzick C, Shafroth PB, Shorrock D, White

JM, White MS. 2007. Altered stream-flow regimes and invasive plant

species: the Tamarix case. Global Ecology and Biogeography 16: 381-393.

Swan CM, Palmer MA. 2004. Leaf diversity alters litter breakdown in a Piedmont

stream. Journal of the North American Benthological Society 23: 15-28.

Swan CM, Healey B, Richardson DC. 2008. The role of native riparian tree

species in decomposition of invasive tree of heaves (Ailanthus altissima) leaf

litter in an urban stream. Ecoscience 15: 27-35.

United States Department of Agriculture, Plants Profile: L. maackii (Rupr.) Herder Amur

L. maackii. URL http://plants.usda.gov/java/profile?symbol=LOMA6 [accessed

on 4 October 2011].

Vannote RL, Minshall GW, Cummins KW, Sedell JR, Cushing CE. 1980. The river

continuum concept. Canadian Journal of Fisheries and Aquatic Science 37:

130-137.

Walting JI, Hickman CR, Lee E, Want K, Orrock LL. 2011. Extracts of the invasive

shrub Lonicera maackii increase mortality and later behavior of amphibian larvae.

Oecologia 165: 153-159.

Webster, JR, Wallace JB, Benfield EF. 1995. Organic processes in streams of the

eastern United States. Pages 117-187 in Cushing EE, Cummins KW, Minshall

GW. (eds). River and stream ecosystems. Elsevier, Amsterdam.

Zar JH. 1999. Biostatistical analysis 4th Edition. Pages 261-263. Prentice-Hall, New

Jersey.

TABLES:

Table 2.1: Two-tailed paired t-test and Wilcoxon matched pairs test for organic

matter types, canopy cover, benthic algal biomass, and total macroinvertebrate

density. Mean comparisons were conducted between L. maackii removal and

non-removal reaches.

P-

Parameter df t-value P-value W-value value

Total Organic Matter 3 0.798 0.4835

Native Organic Matter 3 1.025 0.3803

Lonicera Organic Matter 3 16.970 0.0004

Canopy Cover 10 6.448 < 0.0001

Benthic Algal Biomass -70 0.1650

Macroinvertebrate Density 101 0.0294

Table 2.2: Linear regression statistical results for leaf litter accumulation over 35

days for total, native, and Lonicera coarse organic matter accumulation rates (slope

± standard error) between L. maackii removal and non-removal reaches. The slope

P values indicates if leaf litter accumulation rates were significantly different

between reaches.

Organic Matter Reach Slope R2 F-Statistic P-Value

Total Non-removal 1.320 ± 0.234 0.58 0.175 0.677 Removal 1.450 ± 0.204 0.69

Native Non-removal 1.071 ± 0.199 0.56 0.781 0.382

Removal 1.309 ± 0.181 0.70

Lonicera Non-removal 0.248 ± 0.043 0.60 4.555 0.038

Removal 0.141 ± 0.027 0.55

Table 2.3: Repeated measures ANOVA for in-stream top leaf litter genera collected. Comparisons were conducted within a sampling reach between leaf genera for each sampling point.

Percent Stream Variation Reach Source of Variation df (%) F-Value P-Value

Non- Leaf Genera 5 16.99 5.693 0.0013 Removal

Time 5 3.32 1.477 0.2022

Interaction 25 11.38 1.011 0.4583

Subjects (matching) 24 14.3247 1.327 0.1616

Lonicera Leaf Genera 5 14.45 10.21 < 0.0001 Removal

Time 5 11.63 6.908 < 0.0001

Interaction 25 26.71 3.173 < 0.0001

Subjects (matching) 24 6.7923 0.8403 0.6796

FIGURE LEGENDS:

Figure 2.1: Conceptual framework for the effects of Lonicera maackii, an invasive woody shrubs that dominates riparian corridors, on cross-system subsidies and food webs.

Figure 2.2: Mean (±SE) in-stream leaf litter organic matter for total, native, and invasive

L. maackii leaf litter in L. maackii removal and non-removal reaches for 35 days in Black

Oak stream, an upper 3rd order headwater stream. Significant differences between reaches on a sampling day were determined by Bonferroni pairwise post-tests comparison.

Figure 2.3: Mean (±SE) in-stream leaf organic matter for dominant leaf genera within L. maackii removal and non-removal reaches in a headwater stream. Significant Bonferroni pairwise post-test comparisons between leaf genera on the same date after two-way

ANOVA. Symbols: (a) Indicates that Plantanus significantly contributed the most OM than all other leaf genera community (P < 0.01). * Indicates that Acer significantly contributed more OM than Fraxinus and Other (P Indicates that Lonicera significantly contributed more OM than Quercus and Other (P < 0.05). (b) Indicates that

Lonicera significantly contributed more OM than Platanus and Other (P < 0.05).

Indicates that Acer significantly contributed more OM than all other leaf genera except for Lonicera (P < 0.001).

Fig 2.4: Mean (±SE) in-stream leaf litter organic matter for the most dominant leaf genera in Black Oak stream, an upper 3rd order headwater stream. Significant differences between L. maackii removal and non-removal reaches on a sampling day were determined by Bonferroni pairwise post-tests comparison.

Figure 2.5: Mean (±SE) monthly percent canopy cover for L. maackii removal and non- removal reaches from March 2011 – June 2012 after a two-tailed paired t-test. Paired values were between reaches for each sampling date.

Figure 2.6: Mean (±SE) monthly (a) benthic algal biomass and (b) total macroinvertebrate density in L. maackii removal and non-removal reaches from July

2010 – May 2012. Samples were not collected during the month of January because the stream was frozen. The hashed line indicates the beginning of L. maackii removal. An asterisk symbol indicates significant difference between reaches at P < 0.05.

FIGURES:

Figure 2.1:

Figure 2.2:

Figure 2.3:

Figure 2.4:

Figure 2.5:

Figure 2.6:

SUPPLEMENTAL MATERIALS:

Supplemental Table 2.1: Two-tailed paired t-test and Wilcoxon matched pairs test for dominant the top five leaf genera. Mean comparisons were conducted between L. maackii removal and non-removal reaches.

Paired t-Test

Leaf Genus df t-value P -value

Platanus 5 2.455 0.0043

Acer 5 0.451 0.6712

Quercus 5 1.010 0.3589

Fraxinus 5 0.657 0.5403

Lonicera 5 16.970 0.0004

CHAPTER 3: TERRESTRIAL-AQUATIC LINKAGES INFLUENCE BENTHIC

MACROINVERTEBRATE FUNCTIONAL DIVERSITY

ABSTRACT:

Developing a predictive framework to understand how communities assemble is linked to an organism’s functional traits that are shaped in response to biotic and abiotic conditions. Riparian zones are crucial interfaces that support terrestrial-aquatic linkages; therefore, alterations in the riparian zone can result in fluctuations in terrestrial-aquatic connections. Invasion of Lonicera maackii throughout the Midwest has led to riparian near-monocultures along river systems, influencing terrestrial-aquatic linkages. We investigated how riparian invasion of L. maackii influenced the functional and taxonomic diversity and community composition of aquatic macroinvertebrate communities in a headwater stream. We expected that L. maackii invaded riparian forests would have lower aquatic macroinvertebrate taxonomic and functional diversity, resulting in a reduction in macroinvertebrate functional richness. Aquatic macroinvertebrates were sampled monthly from autumn 2010 to winter 2013 in L. maackii invaded and removal stream reaches (n = 5 riffles/reach). We found that macroinvertebrate density was significantly reduced in the L. maackii stream reach (P < 0.05); however, there was no effect on taxonomic and functional diversity metrics. Macroinvertebrate taxonomic community structure and functional trait presence was distinct between stream reaches and were different across seasons (P < 0.05). Winter and spring seasons were

characterized primarily by stenothermic organisms while summer and autumn seasons were characterize by eurythermic organisms. Collector-gatherers were primarily present during winter and spring seasons while herbivores and detritivores characterized summer an autumn seasons. Patterns in macroinvertebrate community and functional trait dynamics were influenced by seasons and the L. maackii riparian forest. These findings suggest that functional traits were driven by life history strategies linked with seasonal patterns in temperature and food resources that are also influenced by L. maackii riparian forests.

INTRODUCTION:

Community composition and functional diversity are shaped by biotic and abiotic conditions and have influence on broader ecosystem processes (Allison 2012, Webb et al.

2010, Xiao et al. 2012). Developing a predictive framework to understand how communities assemble is an important goal in community ecology and is linked to understanding how organismal traits are linked to coexistence (Keddy 1992, Weiher and

Keddy 2001). Community assembly is a combined result of environmental “filter effects” (e.g., hydrologic and temperature gradients) and biotic interactions (e.g., competition, facilitation) that influence community composition and trait presence from plot-to-global spatial scales (Gross et al. 2009, Lamouroux et al. 2002, McGill et al.

2006, Olden and Kennard 2010, Reich and Oleksyn 2004). Trait-based approaches are becoming a common approach for understanding community assembly (Ackerly and

Cornwell 2007, Allison 2012, Case 1981, Gross et al. 2009, McGill et al. 2006). Species inherently have functional traits that may determine the species abundance in communities and are linked with changes in environmental conditions (Funk et al. 2008,

Lamouroux et al. 2002), impacting the resulting composition and functional diversity of communities and ecosystems (Webb et al. 2010).

Riparian zones are crucial interfaces that support terrestrial-aquatic linkages and are often habitats of disturbance events that influence biodiversity and ecosystem processes (Baxter et al. 2005, Gregory et al. 1991, Naiman and Decamps 1997). These zones function as nutrient sinks, prevent bank erosion, and support the transfer of ecosystem subsidies between terrestrial and aquatic systems (Baxter et al. 2005, Gregory et al. 1991). Alterations in the riparian zone via natural disturbance events (e.g. fire and flood incidents) and anthropogenic activities (e.g. stream channelization) can result in fluctuations in terrestrial-aquatic connections and can influence riparian zone function

(Dwire and Kauffman 2003, Garssen et al. 2015, Greene 2014, Jacobs et al. 2007, Walsh et al. 2005).

Lonicera maackii (Rupr.) Maxim. (Amur honeysuckle) is considered an invasive shrub in forests throughout much of the contiguous USA and this species has densely invaded the riparian zones surrounding headwater stream systems in the Midwest

(McNeish et al. 2014, USDA 1999). This invasive plant is known to substantially suppress native plant species survivorship, reproduction, and understory recruitment, impacting forest community structure, primary productivity, and nutrient dynamics

(Arthur et al. 2012, Collier et al. 2002, Deering and Vankat 1999, Gorchov and Trisel

2003, Luken and Thieret 1996, McEwan et al. 2010, McNeish et al. 2014, Poulette and

Arthur 2012, Trammell et al. 2012). Lonicera maackii has also influenced terrestrial arthropod diversity, survivorship, and community dynamics (Buddle et al. 2004,

Christopher and Cameron 2012, Cipollini et al. 2008, Conley et al. 2011, Loomis et al.

2014, McEwan et al. 2009, McNeish et al. 2012, Shewhart et al. 2014). For example,

Shewhart et al. (2014) and Allan et al. (2010) found L. maackii may facilitate disease vectors such as the West Nile Virus vector mosquito Culex pipiens L. (Dipteran:

Culicidae) and lone star ticks (Ixodida: Ambylomma americanum L.) that are vectors for the bacterial pathogen ehrlichiosis. Lonicera maackii leaf litter is high in nitrogen, supports a unique microbial community, and leaf breakdown is up to 5 × faster than native leaves (Arthur et al. 2012, McNeish et al. 2015, Poulette and Arthur 2012,

Trammell et al. 2012). Lonicera maackii shrubs have been estimated to utilize ~10% of available water resources in wetland forests and reduce throughfall volume available to the forest floor in second-growth forests (Boyce et al. 2011, McEwan et al. 2012), which suggest this invasive plant may have substantial impacts on water resources.

Collectively, these studies identify L. maackii effects that span multiple ecological scales and suggest this species has similar impacts to that of other invasive plants (e.g.,

McNeish & McEwan In Press, Ballard et al. 2013, Chittka and Schürkens 2001, Myers and Anderson 2003, Oliver 1996), making L. maackii a model species to study how riparian invasion influences terrestrial-aquatic linkages.

Our previous work found that L. maackii had substantial effects on headwater stream systems, which may be directly connected to the in-stream availability of resources that support aquatic community assembly. In an aquatic leaf breakdown study, we found that L. maackii leaf pack breakdown was ~ 4 × faster and supported different macroinvertebrate taxonomic and functional feeding group abundances compared to native and mixed (i.e, native mixed with L. maackii leaves) leaf packs (McNeish et al.

2012). In another study, we reported significantly lower autumnal in-stream leaf litter,

decreased macroinvertebrate densities, and reduced light available to the aquatic system when L. maackii dominated stream reaches when compared to reaches where L. maackii was absent because of restorative removal (McNeish et al. 2015). Our data suggest that terrestrial invasion alters basal resources in these streams, potentially resulting in bottom- up effects on the macroinvertebrate community, influencing taxonomic and functional macroinvertebrate community composition and associated ecosystem processes

(McNeish & McEwan In Press, McNeish et al. 2015). Although this past work highlighted how L. maackii riparian invasion altered aquatic habitat and food resources, little is understood about the impact of this invasive riparian species and others on the long-term responses of entire macroinvertebrate community assemblages in headwater streams. In this study, our goal was to investigate how the invasion of L. maackii into riparian forests influenced benthic macroinvertebrate community dynamics in a headwater stream. We hypothesized that in stream reaches where the riparian forests are dominated by L. maackii (H1) that the macroinvertebrate taxonomic and functional diversity will be reduced; this reduction would result (H2) in a contraction and convergence in trait distribution (reduced functional richness); and that (H3) based on the results of our previous studies that L. maackii invaded stream reaches would support a distinct taxonomic and functional macroinvertebrate community. Lastly, we tested the hypothesis (H4) that riparian forests dominated by L. maackii would result in increased above-stream canopy cover resulting in significant light attenuation and increased in- stream nitrogen and phosphorus levels compared a stream reach with no riparian L. maackii.

METHODS:

Study Site:

Black Oak Park stream (hereafter referred to as BO) is an un-named tributary within the Little Miami watershed located in southwestern OH, USA (84.12°W;

36.63°N). The stream site is a 3rd order headwater stream with a benthic substrata of sand, clay, and rocks, a streambed that was 1.5 – 5.0 m wide underlain by limestone geology (Schneider 1957). As part of an ecological restoration project, L. maackii was removed from a portion of the riparian forest from August – September 2010 as described in McNeish et al. (2014). All woody invasive plants were removed from a 160 m stream reach with a 5 m buffer on each stream side to create a L. maackii removal reach and an upstream L. maackii (non-removal) reach. The in-stream buffer between these two reaches was 10 m. All native plant species were left intact, and all native species coarse organic matter (COM; e.g., leaves, snags) were left in place. AquaNeat® Aquatic

Herbicide, an Ohio EPA approved aquatic herbicide (EPA regulation number: 228-365;

Nufarm Manufacturer; active ingredient Glyphosate N-glycine), was applied to cut stumps of invasive species within 48 h to prevent re-growth in future growing seasons.

Maintenance removal took place twice after L. maackii removal to prevent re-growth in the removal zone. The experimental stream reach is located within the Centerville-

Washington Park district and all aspects of the project were undertaken in cooperation with park district land managers.

Benthic Macroinvertebrate Sampling:

The benthic macroinvertebrate community was sampled monthly from September

2010 – December 2013 via a Surber sampler (n = 5 per reach) as long as water was

flowing and the stream surface was not frozen. The Surber sampler was haphazardly placed in each riffle, and all benthic substrata were scrubbed with a brush to dislodge aquatic macroinvertebrates and coarse particulate organic matter (COM). All dislodged materials were captured in the sampler net and preserved in 70% ethanol on site. Surber samples were brought to the lab and macroinvertebrates were identified to genus when possible using Merritt et al. (2008), Peckarsky et al. (1990), and Thorp and Covich

(2001).

Ambient Conditions:

Several nutrient parameters were measured monthly from October 2012 –

December 2013 in both L. maackii and removal stream reaches. One L water samples were collected at the upstream beginning of each riffle along the width of the stream in an acid washed amber Nalgene container. Samples were brought back to the lab and processed within 24 h and analyzed for nutrient content via colorimetric methods. Total orthophosphate (PO4-3) was assessed using the malachite green method (D’Angelo et al.

2001). Five mL samples were acidified with a 1.75% (w/v) ammonium heptamolybdate tetrahydrate solution in 6.3N sulfuric acid and shaken for 10 min. with a benchtop shaker.

A 0.035% (w/v) solution of Malachite Green carbinol hydrochloride in 0.35% (w/v) aqueous polyvinyl alcohol solution was added to the sample and then shaken for 20 min.

The absorbance was read at 630 nm and used to calculate orthophosphate concentrations

(hereafter referred as P). Nitrogen was measured via nitrite (NO2-N), nitrate (NO3-N), and ammonia (NH3-N) using the DREL 2800 water quality kit from Hach Company.

Nitrite was determined via the diazotization method and read at 507 nm. The cadmium reduction method was used to identify nitrate concentrations colorimetrically at 500 nm.

Ammonia was quantified using the Nessler method and absorbance was read at 524 nm.

Total suspended solids (TSS) was measured photometrically at 810 nm after the water sample was blended for 2 min.

Above stream canopy cover and light availability were recorded monthly and randomly within each riffle site to represent the difference in forest canopy between stream reaches throughout the study period. Canopy cover was measured with a spherical densiometer in the North, South, East, and West cardinal directions (Lemon

1959). Light was measured at the surface of the water with a waterproof Milwaukee

MW700 standard Portable Lux Meter.

Statistical Analyses:

Macroinvertebrate total density, taxon richness, and taxonomic diversity were calculated for each sample. Total density represented the number of macroinvertebrate individuals per m2, taxon richness was the sum of the number of taxa present in each sample, and diversity was calculated using Hill’s numbers (effective taxon numbers) as presented as in Jost (2006). Macroinvertebrate density, richness, and diversity failed the

Shapiro-Wilk normality test. Thus, these metrics were analyzed with the Wilcoxon matched-pairs test to identify differences in these metrics between the L. maackii removal and non-removal reaches over the sampling period (Sokal and Rohlf 1981, Zar 1999).

Taxonomic macroinvertebrate community dynamics were visualized with non-metric multidimensional scaling (NMDS). A total of three NMDS tests were conducted using a sample × taxon abundance matrix with metaNMS() (max try of 100 iterations, Bray-

Curtis similarity distance) using the ‘vegan’ package in R (McCune and Grace 2002,

Oksanen et al. 2015). The second and third NMDS ordinations were conducted from the

previous best NDMS result. Procrustes and Protest analyses were conducted between the second and third NMDS solutions to ensure the final NMDS result was a stable solution. The Procrustes test can be used to resize and rotate two ordinations in order to match them for a best fit solution between the ordinations (Jackson 1995). The difference between each ordination point and its corresponding partner is then calculated and summed, resulting in the residual sum squared (m2) that indicates the concordance or similarity between the two ordinations (Jackson 1995). A small m2 indicates the two ordination solutions are very similar. All m2 values were < 0.0001 (protest() using

‘vegan’), indicating the final NMDS result was a stable solution relative to previous

NMDS solutions. The effect of stream reach (L. maackii removal and non-removal) and sampling season (spring, summer, autumn, and winter) and the interaction of these two factors on community dynamics were analyzed using ADONIS (Adonis()) in the ‘vegan’ package.

We took a trait based approach to understand how L. maackii impacted macroinvertebrate community composition not only taxonomically but also functionally responded to removal. Seven ecological, life history, and morphological functional traits

(resulting in 26 trait states; Table 1) were linked with each aquatic macroinvertebrate taxon (ESM Table 1). Traits were identified from those described in Poff et al. (2006) that were considered to be most likely influenced by alterations in terrestrial-aquatic connections. For taxa that were not identified to genus, the most prevalent trait of the broader (e.g. family) taxonomic assignment was chosen or a representative genus was selected based on most complete trait information available (Beche et al. 2006). These traits (Table 1) were used to calculate macroinvertebrate functional diversity (FD)

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indices. Eleven taxa were removed from FD analyses due to insufficient information in the literature to confidently assign a trait state to each trait for these taxa.

Macroinvertebrate FD was calculated via functional richness (FRic), functional evenness (FEve), and functional dispersion (FDis) for each sample. Functional richness represents the volume of trait space utilized in n-dimensional space, which is the degree of functional convergence or divergence of a community (Mason et al. 2005, Petchey and

Gaston 2006, Villéger et al. 2008). Larger FRic values for one treatment indicate more trait space is utilized or filled compared to a treatment with small FRic values (functional divergence; Boersma et al. 2015). Functional evenness represents how even the distribution of traits are in trait space (Mason et al. 2005, Villéger et al. 2008).

Functional dispersion detects trait abundance shifts via differences in the relative abundance of traits in each community (Boersma et al. 2015, Laliberté and Legendre

2010). Functional diversity metrics were calculated with a sample × trait value presence and absence matrix and a sample × species abundance matrix using dbFD() from the

‘FD’ package in R (Laliberté et al. 2015). The cailliez method was used to calculate distance matrices since matrix distance could not be represented in Euclidean space

(Laliberté et al. 2015). Functional diversity metrics were checked for normality via the

Shapiro-Wilk test and then analyzed with the Wilcoxon matched-pairs test to test temporal changes in FD metrics between stream reaches. Overall functional trait community composition was visualized with NMDS using a sample × trait value presence and absence matrix (max. try of 100 iterations with the Jaccard similarity index) in R using ‘vegan’. Functional community trait dynamics were visualized and statistically analyzed as previously explained for taxonomic macroinvertebrate

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community dynamics with NMDS. Functional indicator species were identified as presented by Ricotta et al. (2015). Indicator species were first identified for each grouping factor and then the functional association was identified for each indicator species. Functional association was determined by measuring the functional distance between the centroids of all samples for each grouping factor and the indicator species

(Ricotta et al. 2015). Overall FRic at the community scale was calculated for removal and L. maackii stream reaches and across seasons by calculating the volume of each community’s convex hull (as identified from NMDS results) using convhulln() from the

‘geometry’ package in R (Barber et al. 2015).

Macroinvertebrate functional feeding groups (FFG; trophic habitat trait) patterns within and between stream reaches were identified by first calculating the relative abundance of each FFG per sample. All FFG data were non-normal; therefore, differences in FFG relative abundances within and between stream reaches (i.e., between

FFG within each reach) were determined using Wilcoxon matched-pairs test and

Friendman’s test respectively (Sokal and Rohlf 1981, Zar 1999). FFG pairwise comparisons within each reach were conducted using the post.hoc.friedman.nemenyi.test() with the ‘PMCMR’ package in R (Pohlert 2015).

Differences in ambient conditions were identified between stream reaches for above stream canopy cover, light availability, and nutrient dynamics. All conditions were non normal and analyzed with the Wilcoxon matched-pairs test within and between stream reaches (Sokal and Rohlf 1981, Zar 1999).

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RESULTS:

Macroinvertebrate Taxonomic and Functional Diversity Metrics:

The macroinvertebrate community in the stream reach with a riparian zone dominated by L. maackii (L. maackii reach) differed from the experimental restoration reach where L. maackii had been removed (removal reach). A total of fifty taxa were identified in the entire study system, 33 of which were observed in both stream reaches, seven were unique to the L. maackii reach, and ten were unique to the removal reach

(ESM Table. 2). Chironomidae (Diptera) and Naididae (Oligochaeta) both comprised over 80% of the individuals present in both removal and L. maackii reaches (ESM Table

2). The removal reach supported significantly higher macroinvertebrate density and taxonomic richness compared to the L. maackii reach for most sampling events (Figs. 1;

ESM 1a & ESM Table 3). Macroinvertebrate densities peaked during spring months, with densities generally greater in the removal reach one year post L. maackii removal

(September 2011; Fig.1). Taxonomic richness generally increased earlier in the L. maackii reach but peaks lasted longer in the removal reach (ESM Fig. 1a). The L. maackii riparian forest did not significantly influence taxonomic diversity, trait space occupied (FRic), trait evenness distribution (FEve), and relative abundance of trait combinations over time (trait shift: FDis; ESM Figs 1b & 2).

Macroinvertebrate Community Structure and Functional Trait responses:

Functional and taxonomic macroinvertebrate assemblages were significantly different between stream reaches and seasons (ESM Table 4; ESM Fig. 3).

Macroinvertebrate communities were taxonomically different between reaches during spring and autumn seasons but only functionally different during the summer season

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(ESM Table 4; Figs 2 & 3). The functional traits of communities were similar between stream reaches during spring and autumn seasons; however, these communities were taxonomically different, indicating the functional traits and not taxa associations of these communities were maintained during these seasons. However, summer communities were taxonomically similar between reaches but were functionally different, suggesting even with a small change in taxonomic composition there could be major functional changes. Overall macroinvertebrate community functional richness (as calculated by volume of the convex hull) occupied 86% of trait space in the L. maackii reach whereas the functional richness of removal communities occupied 49% (Table 2). Assemblages in the removal reach had a community functional richness that was ~ 1.7 × greater than the L. maackii reach during spring and summer seasons. Community functional richness in the L. maackii reach was ~ 1.4 and ~ 2.8 × greater than the removal reach during autumn and winter seasons, respectively (Table 2).

There were a diversity of functionally relevant taxa associated with specific stream reaches and seasons (Tables 3 & 4; ESM Table 5). Indicator taxa were uniquely identified between stream reaches (Table 3). Interestingly, many of the indicator taxa in the L. maackii reach were classified as medium-large body sized, whereas taxa indicating the communities of the removal reach were mainly small bodied (ESM Table 1). Winter and spring seasons had the most functionally relevant taxa in common whereas summer and autumn had the most relevant taxa in common (ESM Table 5), suggesting functionally relevant taxa were similar between winter and spring compared to summer and autumn seasons. Winter and spring seasons were characterized by predators/collector-gatherers, uni- and multivoltine life history strategies, and small body

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size traits. Summer and autumn seasons were characterized by taxa that were primarily herbivores, predators, and detritivores, eurythermals, multivoltine, and medium body size traits. During spring and winter seasons most functionally relevant taxa in the removal reach were also present in the L. maackii reach (Table 4). The removal reach had more unique functionally relevant taxa (20% and 85%) compared to the L. maackii reach (5% and 75%) during summer and autumn seasons, respectively (Table 4). Overall, L. maackii invaded stream reach supported a macroinvertebrate community that was both taxonomically and functionally distinct as compared to the removal reach, with effects influenced be stream seasonality.

Macroinvertebrate FFG abundance was significantly affected by intact L. maackii riparian forest (L. maackii reach: X2 = 25.804, df = 4, P < 0.0001; removal reach: X2 =

36.927, df = 4, P < 0.0001). Collector-gatherer relative abundance was the dominant

FFG within both stream reaches (all P < 0.05; Fig. 4& b). Herbivores were greater than predators within the L. maackii reach (P < 0.05; Fig. 4a). Within the removal reach, predator and herbivore relative abundance was less than all other FFG (all P < 0.05; Fig.

4b). Herbivore relative abundance was significantly greater in the L. maackii reach compared to the removal reach (ESM Table 3; ESM Fig. 4b). All other FFG relative abundances were similar between reaches (ESM Fig. 4); however, collector-gatherer relative abundance was generally greater in the removal reach (ESM Fig. 4c).

Ambient Conditions: Lonicera maackii riparian forests had differential effects on ambient stream corridor conditions. Above stream canopy cover was significantly less in the L. maackii reach, resulting in a substantial increase in light availability (ESM Table 3; EMS Figs. 5

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& 6). Canopy cover and light availability followed typical seasonal patterns with canopy cover peaking during growing seasons and light availability peaking during winter months. Nutrient dynamics were similar between removal and L. maackii reaches (ESM

Table 3; ESM Fig. 7). Phosphorous peaked earlier and higher in the removal reach compared to the L. maackii reach; although, this pattern was not significant (ESM Fig.

5b). There were two substantial spikes in nitrate and ammonia in the L. maackii reach during autumn and winter months (ESM Fig. 5c,e) and a large spike in TSS for both stream reaches during May (ESM Fig. 5d).

DISCUSSION:

The inherent seasonality of streams has important repercussions on aquatic organism life history strategies, the availability of resources, and serves as a temporal environmental filter (Lytle and Poff 2004, Poff et al. 2006, Verberk et al. 2008). Aquatic macroinvertebrate functional and taxonomic diversity has been directly related to stream hydroperiod (Schriever et al. 2015), which influences the evolutionary strategies of macroinvertebrates (Lytle and Poff 2004). Shifts in temperature can also drive macroinvertebrate thermal guilds, which is linked to climate change and anthropogenic activities (Heino et al. 2009, Lake et al. 2000). Our study identified strong seasonal effects on the taxonomic and functional composition of the macroinvertebrate assemblages, which has been well documented in the literature (Beche et al. 2006,

Hawkins and Sedell 1981, Murphy and Giller 2000). Winter and spring seasons were characterized primarily by stenothermic organisms while summer and autumn seasons were characterize by eurythermic organisms, which is likely attributed to seasonal changes in stream water temperature (Allan and Castillo 2007, Cummins 1974).

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Collector-gatherers were predominately present during winter and spring seasons while herbivores and detritivores characterized summer an autumn seasons, potentially reflecting seasonal patterns and the availability of food resources (Cummins and Klug

1979, Richardson 1991, Thompson and Townsend 1999, Wallace and Webster 1996).

Interestingly, detritivore (i.e., shredder) abundance was overall low in this stream site, which contradicted with the expectation that this FFG would be a dominant FFG in headwater systems (Vannote et al. 1980). Black Oak stream is located in a developed area in the Miami Valley, OH and was dominated by Chironomidae and Oligochaeta but

Plecoptera (stoneflies) were missing, which are pollution sensitive taxa that serve as detritivores found commonly in headwater streams (Vannote et al. 1980). In a study measuring macroinvertebrate diversity within along the Little Miami River continuum in the same region as our study, Hanley (2008) found only two Plecoptera taxa across 13 sampling sites over two years; this may be associated with the fact that Ohio has lower taxonomic richness of Plecoptera relative to some surrounding states (Edward deWalt et al. 2012).

Riparian invasive plants can have substantial impacts on terrestrial-aquatic linkages via the transfer of ecosystem subsidies between terrestrial and aquatic habitats

(Greene 2014). For example, Elaeagnus angustifolia (Russian olive L.) is a riparian invasive tree in western USA known to increase in-stream terrestrial organic matter subsidies and organic nitrogen, altering stream ecosystem efficiency and biogeochemistry

(Mineau et al. 2011, 2012). Differences in macroinvertebrate taxonomic and functional community structure between stream reaches may in part be linked to the availability of in-stream leaf litter subsidies. Our previous work demonstrated L. maackii riparian

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forests substantially decreased the abundance and altered the temporal deposition of in- stream leaf litter (McNeish et al. 2015), an important subsidy from riparian forests that is utilized by aquatic macroinvertebrates as a habitat and food resource (Minshall 1967,

Petersen et al. 1989, Vannote et al. 1980). McNeish et al. (2012) and Fargen et al. (2015) demonstrated aquatic L. maackii leaf litter breakdown was faster and supported different macroinvertebrate FFG composition in comparison to native litter, influencing leaf litter processing efficiency even over a year time period.

The presence of L. maackii in the riparian forest had substantial effects on the macroinvertebrate community and ambient conditions. Removal of L. maackii resulted in increased macroinvertebrate density and taxonomic richness, supporting our hypotheses (H1 & H2). Presence of riparian L. maackii resulted in a macroinvertebrate community composition that was both taxonomically and functionally distinct compared to the removal reach (supporting H3). Macroinvertebrate community functional richness within the L. maackii reach was greater – or more divergent – compared to communities in the removal reach; however, strong seasonal patterns resulted in functional convergence in the L. maackii reach during spring and summer seasons (supporting H2).

Above stream canopy cover was greater in the L. maackii reach, resulting in a shading effect; however, ambient nutrient conditions were similar between stream reaches. Our study is the first to show invasion of L. maackii in riparian forests is linked to the entire taxonomic and functional trait community composition (i.e. not only FFG composition) of aquatic macroinvertebrate communities (as compared to other studies e.g. Fargen et al.

2015, McNeish et al. 2012).

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In summary, our data (1) suggests riparian invasive species are linked to aquatic macroinvertebrate taxonomic community composition and functional trait diversity and

(2) lends support to our predictive framework that riparian invasive plants can have a bottom-up effect on aquatic ecosystems, impacting aquatic food web dynamics and ecosystem function and processes (McNeish & McEwan In Press, McNeish et al. 2015).

Our study suggests there is a link between changes in macroinvertebrate community structure to temperature and food resources; however, further research is needed to clearly identify these links. This study highlights that changes in the riparian zone can have substantial implications on aquatic macroinvertebrate community structure and functional diversity.

ACKNOWLEDGEMENTS:

We would like to thank the Centerville-Washington Park District, OH for use of the stream field site, Julia Chapman and Tiffany Schriever for assistance with R programming, Jim Crutchfield for assistance in nutrient analyses, and Casey Hanley for use of laboratory equipment and space. Special thanks to Eryn Moore, Courtney

Dvorsky, Ryan Reihart, Dani Theimann, Patrick Vrablik, Michael Ruddy and all the undergraduate students at the University of Dayton who contributed time and effort to field and laboratory work. This work was supported by the National Science Foundation

(NSF: DEB 1352995) and in part by the University of Dayton Office for Graduate

Academic Affairs through the Graduate Student Summer Fellowship Program. Any opinions, findings, and conclusions or recommendations expressed are those of the author and do not necessarily reflect the views of the National Science Foundation.

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REFERENCES:

BENFIELD, E.F. 2007. Decomposition of Leaf Material. Pages 711-720 in F.R. Hauer and

G.A. Lambertic, editors. Methods in Stream Ecology, 2nd edition. Academic Press,

San Diego.

ACKERLY, D. D., AND W. K. CORNWELL. 2007. A trait-based approach to community

assembly: partitioning of species trait values into within- and among-community

components. Ecol. Lett. 10:135–45. Retrieved May 21, 2013, from Ecology letters.

ALLAN, B. F., H. P. DUTRA, L. S. GOESSLING, K. BARNETT, J. M. CHASE, R. J. MARQUIS,

G. PANG, G. A STORCH, R. E. THACH, AND J. L. ORROCK. 2010. Invasive honeysuckle

eradication reduces tick-borne disease risk by altering host dynamics. Proc. Natl.

Acad. Sci. U. S. A. 107:18523–18527. Retrieved June 4, 2013, from Proceedings of

the National Academy of Sciences of the United States of America.

rez&rendertype=abstract>.

ALLAN, J. D., AND M. M. CASTILLO. 2007. Stream ecology: Structure and function of

running waters, 2nd edition. Springer, Dordrecht, The Netherlands.

ALLISON, S. D. 2012. A trait-based approach for modelling microbial litter

decomposition. Ecol. Lett. 15:1058–70. Retrieved May 27, 2013, from Ecology

letters. .

ARTHUR, M. A., S. R. BRAY, C. R. KUCHLE, AND R. W. MCEWAN. 2012. The influence of

the invasive shrub, Lonicera maackii, on leaf decomposition and microbial

110

community dynamics. Plant Ecol. 213:1571–1582. Retrieved June 13, 2013, from

Plant Ecology. .

BALLARD, M., J. HOUGH-GOLDSTEIN, AND D. TALLAMY. 2013. Arthropod Communities

on Native and Nonnative Early Successional Plants. Environ. Entomol. 42:851–859.

Retrieved from Environmental Entomology.

BARBER, C., K. HABEL, R. GRASMAN, R. GRAMACY, A. STAHEL, AND D. C. STERRATT.

2015. geometry: Mesh generation and surface tesselation. 0.3-6. Retrieved from 0.3-

6. .

BAXTER, C. V., K. D. FAUSCH, AND W. CARL SAUNDERS. 2005. Tangled webs: reciprocal

flows of invertebrate prey link streams and riparian zones. Freshw. Biol. 50:201–

220. Retrieved May 25, 2013, from Freshwater Biology.

BECHE, L. A., E. P. MCELRAVY, AND V. H. RESH. 2006. Long-term seasonal variation in

the biological traits of benthic-macroinvertebrates in two Mediterranean-climate

streams in California, U.S.A. Freshw. Biol. 51:56–75. Retrieved from Freshwater

Biology. .

BOERSMA, K., L. DEE, S. MILLER, M. BOGAN, D. LYTLE, AND A. GITELMAN. 2015.

Linking multidimensional functional diversity to quantitative methods: A graphical

hypothesis-evaluation framework. Ecol. Soc. Am.:81–87. Retrieved from Ecological

Society of America. .

BOYCE, R. L., R. D. DURTSCHE, AND S. L. FUGAL. 2011. Impact of the invasive shrub

111

Lonicera maackii on stand transpiration and ecosystem hydrology in a wetland

forest. Biol. Invasions 14:671–680. Retrieved May 27, 2013, from Biological

Invasions. .

BUDDLE, C. M., S. HIGGINS, AND L. RYPSTRA. 2004. Ground-dwelling Spider

Assemblages Inhabiting Riparian Forests and Hedgerows in an Agricultural

Landscape. Am. Midl. Nat. 151:15–26.

CASE, T. J. 1981. Niche packing and coevolution in competition communities. Proc. Natl.

Acad. Sci. U. S. A. 78:5021–5. Retrieved from Proceedings of the National

Academy of Sciences of the United States of America.

ez&rendertype=abstract>.

CHITTKA, L., AND S. SCHÜRKENS. 2001. Successful invasion of a floral market. Nature

411:653.

CHRISTOPHER, C. C., AND G. N. CAMERON. 2012. Effects of Invasive Amur Honeysuckle

(Lonicera maackii) and White-Tailed Deer (Odocoileus virginianus) on Litter-

Dwelling Arthropod Communities. Am. Midl. Nat. 167:256–272.

CIPOLLINI, D., R. STEVENSON, AND K. CIPOLLINI. 2008. Contrasting effects of

allelochemicals from two invasive plants on the performance of a nonmycorrhizal

plant. Int. J. Plant Sci. 169:371–375. Retrieved June 13, 2013, from International

Journal of Plant Sciences. .

COLLIER, M. H., J. L. VANKAT, AND M. R. HUGHES. 2002. Diminished plant richness and

abundance below Lonicera maackii, an invasive shrub. Am. Midl. Nat. 147:60–71.

112

CONLEY, A. K., J. I. WATLING, AND J. L. ORROCK. 2011. Invasive plant alters ability to

predict disease vector distribution. Ecol. Appl. 21:329–334.

CUMMINS, K. W. 1974. Structure and function of stream ecosystems. Bioscience 24:631–

641.

CUMMINS, K. W., AND M. J. KLUG. 1979. Feeding ecology of stream invertebrates. Annu.

Rev. Ecol. Syst. 10:147–172.

D’ANGELO, E., J. CRUTCHFIELD, AND M. VANDIVIERE. 2001. Rapid, sensitive, microscale

determination of phosphate in water and soil. J. Environ. Qual. 30:2206–9.

Retrieved from Journal of environmental quality.

DEERING, R. H., AND J. L. VANKAT. 1999. Forest colonization and developmental growth

of the invasive shrub Lonicera maackii. Am. Midl. Nat. 141:43–50.

DWIRE, K. A., AND J. B. KAUFFMAN. 2003. Fire and riparian ecosystems in landscapes of

the western USA. For. Ecol. Manage. 178:61–74. Retrieved from Forest Ecology

and Management.

EDWARD DEWALT, R., Y. CAO, T. TWEDDALE, S. A. GRUBBS, L. HINZ, M. PESSINO, AND J.

L. ROBINSON. 2012. Ohio USA stoneflies (Insecta, Plecoptera): Species richness

estimation, distribution of functional niche traits, drainage affiliations, and

relationships to other states. Zookeys 178:1–26.

FARGEN, C., S. M. EMERY, AND M. M. CARREIRO. 2015. Influence of Lonicera maackii

Invasion on Leaf Litter Decomposition and Macroinvertebrate Communities in an

113

Urban Stream. Nat. Areas J. 35:392–403.

FUNK, J. L., E. E. CLELAND, K. N. SUDING, AND E. S. ZAVALETA. 2008. Restoration

through reassembly: plant traits and invasion resistance. Trends Ecol. Evol. 23:695–

703. Retrieved May 23, 2013, from Trends in ecology & evolution.

GARSSEN, A. G., A. BAATTRUP-PEDERSEN, L. A. C. J. VOESENEK, J. T. A. VERHOEVEN,

AND M. B. SOONS. 2015. Riparian plant community responses to increased flooding:

a meta-analysis. Glob. Chang. Biol. 21:2881–2890. Retrieved from Global Change

Biology. .

GORCHOV, D. L., AND D. E. TRISEL. 2003. Competitive effects of the invasive shrub,

Lonicera maackii (Rupr.) Herder (Caprifoliaceae), on the growth and survival of

native tree seedlings. Plant Ecol. 166:13–24.

GREENE, S. L. 2014. A roadmap for riparian invasion research. River Res. Appl. 30:663–

669.

GREGORY, S. V, F. J. SWANSON, W. A. MCKEE, AND K. W. CUMMINS. 1991. An ecosystem

perspective of riparian zones. Bioscience 41:540–551.

GROSS, N., G. KUNSTLER, P. LIANCOURT, F. DE BELLO, K. N. SUDING, AND S. LAVOREL.

2009. Linking individual response to biotic interactions with community structure: a

trait-based framework. Funct. Ecol. 23:1167–1178. Retrieved from Functional

Ecology. <://WOS:000271632200017>.

HANLEY, C. 2008. A river continuum analysis of relationships between bioassessment,

land use, and macroinvertebrate assemblages. University of Dayton, OH.

114

HAWKINS, C., AND J. SEDELL. 1981. Longitudinal and seasonal changes in functional

organization of macroinvertebrate communities in four Oregon streams. Ecology

62:387–397. Retrieved from Ecology. .

HEINO, J., R. VIRKKALA, AND H. TOIVONEN. 2009. Climate change and freshwater

biodiversity: Detected patterns, future trends and adaptations in northern regions.

Biol. Rev. 84:39–54.

JACKSON, D. A. 1995. A PROcrustean Randomization TEST of community environment

concordance. Ecoscience 2:297–303.

JACOBS, S. M., J. S. BECHTOLD, H. C. BIGGS, N. B. GRIMM, S. LORENTZ, M. E. MCCLAIN,

R. J. NAIMAN, S. S. PERAKIS, G. PINAY, AND M. C. SCHOLES. 2007. Nutrient vectors

and riparian processing: A review with special Reference to African semiarid

savanna ecosystems. Ecosystems 10:1231–1249. Retrieved from Ecosystems.

JOST, L. 2006. Entropy and diversity. Oikos 113:363–375.

KEDDY, P. A. 1992. Assembly and response rules: Two goals for predictive community

ecology. J. Veg. Sci. 3:157–164. Retrieved from Journal of Vegetation Science.

eralSearch&qid=2&SID=W1pPLzAZjdAHkE5rN3O&page=1&doc=1>.

LAKE, P. S., M. A. PALMER, P. BIRO, J. COLE, A. P. COVICH, C. DAHM, J. GIBERT, W.

GOEDKOOP, K. MARTENS, AND J. VERHOEVEN. 2000. Global change and the

biodiversity of freshwater ecosystems: Impacts on linkages between above-sediment

and sediment biota. Bioscience 50:1099–1107.

115

LALIBERTÉ, E., AND P. LEGENDRE. 2010. A distance-based framework for measuring

functional diversity from multiple traits. Ecology 91:299–305. Retrieved from

Ecology. .

LALIBERTÉ, E., P. LEGENDRE, AND B. SHIPLEY. 2015. FD: Measuring functional diversity

(FD) from multiple traits, and other tools for functional ecology. 1.0-12.

LAMOUROUX, N., N. L. POFF, AND P. L. ANGERMEIER. 2002. Intercontinental convergence

of stream fish community traits along geomorphic and hydraulic gradients. Ecology

83:1792–1807.

LOOMIS, J. D., G. N. CAMERON, AND G. W. UETZ. 2014. Impact of the invasive shrub

Lonicera maackii on shrub-dwelling Araneae in a deciduous forest in eastern North

America. Am. Midl. Nat. 171:204–218.

LUKEN, J., AND J. W. THIERET. 1996. Amur Honeysuckle, its fall from grace. Bioscience

46:18–24.

LYTLE, D. A., AND N. L. POFF. 2004. Adaptation to natural flow regimes. Trends Ecol.

Evol. 19:94–100.

MARCARELLI, A. M., C. V BAXTER, M. M. MINEAU, AND R. O. HALL. 2011. Quantity and

quality: unifying food web and ecosystem perspectives on the role of resource

subsidies in freshwaters. Ecology 92:1215–1225. Retrieved from Ecology. <

ISI>://WOS:000292633900004>.

MASON, N. W. H., D. MOUILLOT, W. G. LEE, AND J. B. WILSON. 2005. Functional

richness, functional and functional evenness divergence: the primary of functional

components diversity. Oikos 111:112–118.

116

MCCUNE, B., AND J. B. GRACE. 2002. Analysis of Ecological Communities. MjM

Software, Gleneden Beach, Oregon, USA. Retrieved from .

MCEWAN, R. W., M. A. ARTHUR, AND S. E. ALVERSON. 2012. Throughfall chemistry and

soil nutrient effects of the invasive shrub Lonicera maackii in deciduous forests.

Am. Midl. Nat. 168:43–55.

MCEWAN, R. W., L. G. ARTHUR-PARATLEY, L. K. RIESKE, AND M. A. ARTHUR. 2010. A

multi-assay comparison of seed germination inhibition by Lonicera maackii and co-

occurring native shrubs. Flora - Morphol. Distrib. Funct. Ecol. Plants 205:475–483.

Elsevier. Retrieved June 13, 2013, from Flora - Morphology, Distribution,

Functional Ecology of Plants.

MCEWAN, R. W., L. K. RIESKE, AND M. A. ARTHUR. 2009. Potential interactions between

invasive woody shrubs and the gypsy moth (Lymantria dispar), an invasive insect

herbivore. Biol. Invasions 11:1053–1058. Retrieved May 23, 2013, from Biological

Invasions. .

MCGILL, B. J., B. J. ENQUIST, E. WEIHER, AND M. WESTOBY. 2006. Rebuilding

community ecology from functional traits. Trends Ecol. & Evol. 21:178–185.

Retrieved from Trends in Ecology & Evolution.

MCNEISH, R. E., M. E. BENBOW, AND R. W. MCEWAN. 2012. Riparian forest invasion by

a terrestrial shrub (Lonicera maackii) impacts aquatic biota and organic matter

processing in headwater streams. Biol. Invasions 14:1881–1893. Retrieved June 13,

117

2013, from Biological Invasions.

0199-8>.

MCNEISH, R. E., E. M. MOORE, M. E. BENBOW, AND R. W. MCEWAN. 2014. Editorial.

River Res. Appl. 7.

MCNEISH, R. E., E. M. MOORE, M. E. BENBOW, AND R. W. MCEWAN. 2015. Removal of

the invasive shrub, Lonicera maackii, from riparian forests influences headwater

stream biota and ecosystem function. River Res. Appl. 31:1131–1139.

MERRITT, R., K. W. CUMMINS, AND M. BERG. 2008. An introduction to the aquatic insects

of North America, 4th edition. Kendall/Hunt, Iowa.

MINEAU, M. M., C. V. BAXTER, AND A. M. MARCARELLI. 2011. A Non-Native Riparian

Tree (Elaeagnus angustifolia) Changes Nutrient Dynamics in Streams. Ecosystems

14:353–365. Retrieved June 13, 2013, from Ecosystems.

MINEAU, M. M., C. V. BAXTER, A. M. MARCARELLI, AND G. W. MINSHALL. 2012. An

invasive riparian tree reduces stream ecosystem efficiency via a recalcitrant organic

matter subsidy. Ecology 93:1501–1508. Retrieved from Ecology. <

ISI>://WOS:000306829300001>.

MINSHALL, G. W. 1967. Role of allochthonous detritus in the trophic structure of a

woodland springbrook community. Ecology 48:139–149.

MURPHY, J. F., AND P. S. GILLER. 2000. Seasonal dynamics of macroinvertebrate

assemblages in the benthos and associated with detritus packs in two low-order

streams with different riparian vegetation. Freshw. Biol. 43:617–631.

118

MYERS, C. V., AND R. C. ANDERSON. 2003. Seasonal variation in photosynthetic rates

influences success of an invasive plant, Garlic mustard (Alliaria petiolata). Am.

Midl. Nat. 150:231–245.

NAIMAN, J., AND H. DECAMPS. 1997. The ecology of interfaces: Riparian Zones. Annu.

Rev. Ecol. Syst. 28:621–658.

OKSANEN, A. J., F. G. BLANCHET, R. KINDT, P. R. MINCHIN, R. B. O. HARA, G. L.

SIMPSON, P. SOLY-, M. H. H. STEVENS, AND H. WAGNER. 2015. vegan: Community

Ecology Package. R 562 package version 2.2-1. Retrieved from

project.org/package=vegan>.

OLDEN, J. D., AND M. J. KENNARD. 2010. Intercontinental comparison of fish life history

strategies along a gradient of hydrologic variability. Am. Fish. Soc. Symp. 73:83–

107. Retrieved from American Fisheries Society Symposium.

content/uploads/2013/03/AFS_Chapter_2010.pdf>.

OLIVER, J. D. 1996. Mile-a-minute weed, (Polygonum perfoliatum L .), an invasive vine

in natural and disturbed sites. Castanea 61:244–251.

PECKARSKY, B. L., P. FRAISSINET, M. PENTON, AND D. CONKLIN. 1990. Freshwater

macroinvertebrates of Northeastern North America. Cornell University Press, Ithica

and London.

PETCHEY, O. L., AND K. J. GASTON. 2006. Functional diversity: back to basics and

looking forward. Ecol. Lett. 9:741–758. Retrieved from Ecology Letters.

119

PETERSEN, R. C., K. W. CUMMINS, AND G. M. WARD. 1989. Microbial and Animal

Processing of Detritus in a Woodland. Ecol. Monogr. 59:21–39.

POFF, N. L., J. D. OLDEN, N. K. M. VIEIRA, D. S. FINN, M. P. SIMMONS, AND B. C.

KONDRATIEFF. 2006. Functional trait niches of North American lotic insects: traits-

based ecological applications in light of phylogenetic relationships. J. North Am.

Benthol. Soc. 25:730–755. The Society for Freshwater Science. Retrieved from

Journal of the North American Benthological Society.

3593(2006)025[0730:FTNONA]2.0.CO;2>.

POHLERT, A. 2015. PMCMR: Calculate multiple comparisons of mean rank sums. 4.0.

Retrieved from 4.0.

project.org/web/packages/PMCMR/PMCMR.pdf>.

POULETTE, M. M., AND M. A. ARTHUR. 2012. The impact of the invasive shrub Lonicera

maackii on the decomposition dynamics of a native plant community. Ecol. Appl.

22:412–424.

REICH, P. B., AND J. OLEKSYN. 2004. Global patterns of plant leaf N and P in relation to

temperature and latitude. Proc. Natl. Acad. Sci. U. S. A. 101:11001–11006.

RICHARDSON, J. S. 1991. Seasonal food limitation of detritivores in a montane stream: An

experimental test. Ecology 72:873–887.

RICOTTA, C., M. CARBONI, AND A. T. R. ACOSTA. 2015. Let the concept of indicator

species be functional! J. Veg. Sci. Retrieved from Journal of Vegetation Science.

120

SCHNEIDER, W. J. 1957. Relation of geology to stream flow in the Upper Little Miami

Basin. Ohio J. Sci. 57:11–14. Retrieved from The Ohio Journal of Science.

SCHRIEVER, T. A., M. T. BOGAN, K. S. BOERSMA, M. CAÑEDO-ARGÜELLES, K. L. JAEGER,

J. D. OLDEN, AND D. A. LYTLE. 2015. Hydrology shapes taxonomic and functional

structure of desert stream invertebrate communities. Freshw. Sci. 34:399–409.

Retrieved from Freshwater Science.

SHEWHART, L., R. W. MCEWAN, AND M. E. BENBOW. 2014. Evidence for facilitation of

Culex pipiens (Diptera: Culicidae) life history traits by the nonnative invasive shrub

Amur Honeysuckle (Lonicera maackii). Entomol. Soc. Am. 43:1584–1593.

SOKAL, R., AND F. ROHLF. 1981. Biometery: the Principles and Practice of Statistics in

Biological Research, 2nd edition. W.H. Freeman and Company, NewYork.

THOMPSON, R. M., AND C. R. TOWNSEND. 1999. The effect of seasonal variation on the

community structure and food-web attributes of two streams: Implications for dood-

web science. Oikos 87:75–88. Retrieved from Oikos.

pdf?acceptTC=true>.

THORP, J., AND A. P. COVICH. 2001. Ecology and classification of North American

freshwater invertebratess. Academic Press, San Diego.

TRAMMELL, T. L. E., H. A. RALSTON, S. A. SCROGGINS, AND M. M. CARREIRO. 2012.

Foliar production and decomposition rates in urban forests invaded by the exotic

121

invasive shrub, Lonicera maackii. Biol. Invasions 14:529–545.

USDA. 1999. Lonicera maackii ( Rupr .) Herder. Retrieved January 1, 2014, from

USDA. .

VANNOTE, R. L., G. W. MINSHALL, K. W. CUMMINS, J. R. SEDELL, AND C. E. CUSHING.

1980. The River Continuum Concept. Can. J. Fish. Aquat. Sci. 37:130–137.

Retrieved from Canadian Journal of Fisheries and Aquatic Sciences.

VERBERK, W. C. E. P., H. SIEPEL, AND H. ESSELINK. 2008. Life-history strategies in

freshwater macroinvertebrates. Freshw. Biol. 53:1722–1738.

VILLÉGER, S., N. W. H. MASON, AND D. MOUILLOT. 2008. New multidimensionale

functional diversity indices for a multifaceted framework in functional ecology.

Ecology 89:2290–2301. Retrieved from Ecology.

1206.1>.

WALLACE, J. B., AND J. R. WEBSTER. 1996. The role of macroinvertebrates in stream

ecosystem function. Annu. Rev. Entomol. 41:115–139.

WALSH, C. J., A. H. ROY, J. W. FEMINELLA, P. D. COTTINGHAM, P. M. GROFFMAN, AND R.

P. MORGAN. 2005. The urban stream syndrome: current knowledge and the search

for a cure. J. North Am. Benthol. Soc. 24:706–723.

WEBB, C. T., J. A HOETING, G. M. AMES, M. I. PYNE, AND N. LEROY POFF. 2010. A

structured and dynamic framework to advance traits-based theory and prediction in

ecology. Ecol. Lett. 13:267–83. Retrieved May 23, 2013, from Ecology letters.

122

WEIHER, E., AND P. KEDDY. 2001. Ecological assembly rules: Perspectives, advances,

rereats. Cambridge University Press, Cambridge, UK.

XIAO, Y., G.-D. XIE, K. AN, AND C.-X. LU. 2012. A research framework of ecosystem

services based on functional traits. Chinese J. Plant Ecol. 36:353–362. Retrieved

from Chinese Journal of Plant Ecology. <://BCI:BCI201200379218>.

ZAR, J. 1999. Biostatistical Analysis, 4th edition. Prentice-Hall, New Jersey.

123

TABLES:

Table 3.1: Description of macroinvertebrate functional traits (7) and their trait states (26) used for functional diversity calculations and NMDS community dynamics.

Category Trait Trait States

Life History Voltinism Semivoltine, Univoltine,

Bi- or Multivoltine

Development Slow season

Fast seasonal

Nonseasonal

Morphology Respiration Tegument, Gills, Aerial

Size at Maturity Small (<9 mm), Medium

(9-16 mm), Large

(>16mm)

Ecology Trophic Habitat or Collector-gatherer (CG)

Functional Feeding Collector-filterer (CF) Group (FFG) Herbivore (H)

Detritivore (D)

124

Predator (P)

Thermal Preference Cold stenothermal/Cool

eurythermal, Cool/Warm

eurythermal, Warm

eurythermal

Habitat Burrower, Climber,

Sprawler, Clinger,

Swimmer, Skater

125

Table 3.2: Community functional richness for both stream reaches and between stream reaches within each sampling season. Percent functional richness represents the percent of functional space utilized out of the total functional space (convex hull) occupied by the community. Functional richness is represented by the volume of the convex hull calculated by NMDS results seen in Fig. 4 and ESM Fig. 3.

Honeysuckle Removal

Factor Reach (%) Reach (%) Total

Stream Reach 0.934 86.08 0.533 49.12 1.085

Spring 0.037 24.03 0.064 41.56 0.154

Summer 0.04 15.09 0.062 23.40 0.265

Autumn 0.487 77.06 0.332 52.53 0.632

Winter 0.216 67.08 0.077 23.91 0.322

126

Table 3.3: Functionally relevant indicator macroinvertebrate taxa for L. maackii and removal stream reaches. An asterisk indicates taxa that were unique to that stream reach.

Honeysuckle Reach Removal Reach

Observed P- Observed P-

Order Taxa Value Value Order Trait Value Value

Diptera Empididae* 0.168 <0.001 Diptera Chironomidae* 0.141 0.010

Diptera Tipulidae* 0.214 0.010 Diptera Dasyhelea* 0.216 <0.001

Lumbriculida Lumbriculidae* 0.291 <0.001 Diptera Forcipomyia* 0.135 0.010

Haplotaxida Naididae* 0.273 0.031 Diptera Maruina* 0.180 <0.001

Cerithioidea Goniobasis* 0.18 0.010 Tricladida Planariidae* 0.264 0.010

Planorboidea Physella* 0.188 <0.001 Ephemeroptera Baetidae* 0.264 0.041

127

Arhynchobdellida Erpobdella punctata* 0.273 <0.001

Rhynchobdellida Helobdella fusca* 0.23 <0.001

Rhynchobdellida Helobdella stagnalis* 0.23 0.010

Coleoptera Ectopria* 0.177 0.010

Coleoptera Lampyridae* 0.189 0.020

Coleoptera Stenelmis* 0.225 <0.001

Copepoda Copepoda* 0.249 0.020

Hemiptera Microvelia* 0.258 0.010

Isopoda Caecidotea* 0.205 0.010

Amphipoda Gammarus* 0.224 <0.001

Amphipoda Hyalella* 0.217 0.031

Zygoptera Agria* 0.221 0.010

128

Zygoptera Calopteryx* 0.245 <0.001

129

Table 3.4: Functionally relevant indicator macroinvertebrate taxa for each stream reach within each season. An asterisk indicates taxa that were unique to that stream reach within a season.

Honeysuckle Reach Removal Reach

Observed Observed

Season Order Taxa Value P-Value Order Trait Value P-Value

Spring Diptera Atrichopogon* 0.221 <0.001 Diptera Culicoides 0.215 <0.001

Diptera Tipulidae* 0.196 0.041 Tricladida Planariidae* 0.245 0.021

Diptera Ceratopogon* 0.177 <0.001 Ephemeroptera Baetidae 0.249 <0.001

Diptera Culicoides 0.219 <0.001

Amphipoda Hyalella* 0.199 0.020

Hydrachnidia Hydrachnidia* 0.185 0.010

Coleoptera Lampyridae* 0.172 0.041

Ephemeroptera Baetidae 0.258 <0.001

130

Summer Diptera Empididae 0.128 <0.001 Diptera Empididae 0.136 <0.001

Trichoptera Ceratopsyche 0.104 <0.001 Diptera Psychoda* 0.175 0.020

Trichoptera Hydroptila 0.122 <0.001 Diptera Tipulidae* 0.2082 0.031

Coleoptera Ectopria 0.128 <0.001 Trichoptera Ceratopsyche 0.084 <0.001

Coleoptera Lampyridae 0.154 <0.001 Trichoptera Cheumatopsyche* 0.094 <0.001

Coleoptera Stenelmis 0.19 <0.001 Trichoptera Hydropsyche* 0.114 <0.001

Copepoda Copepoda* 0.226 0.031 Trichoptera Hydroptila 0.114 <0.001

Basommatophora Ferrissia 0.16 <0.001 Cerithioidea Goniobasis 0.156 <0.001

Planorboidea Menetus dilatatus 0.185 0.041 Planorboidea Physella 0.16 <0.001

Planorboidea Physella 0.156 <0.001 Planorboidea Menetus dilatatus 0.182 0.010

Rhynchobdellida Helobdella fusca 0.192 0.010 Basommatophora Ferrissia 0.165 <0.001

Rhynchobdellida Helobdella stagnalis 0.192 <0.001 Coleoptera Ectopria 0.129 <0.001

Zygoptera Agria 0.173 <0.001 Coleoptera Haliplidae* 0.143 0.031

Zygoptera Calopteryx 0.21 <0.001 Coleoptera Lampyridae 0.164 <0.001

Decapoda Orconectes rusticus 0.21 <0.001 Coleoptera Stenelmis 0.204 <0.001

131

Amphipoda Gammarus 0.189 <0.001 Rhynchobdellida Helobdella fusca 0.228 0.031

Arhynchobdellida Erpobdella punctata 0.223 <0.001 Rhynchobdellida Helobdella stagnalis 0.228 0.020

Cerithioidea Goniobasis 0.127 <0.001 Decapoda Orconectes rusticus 0.244 <0.001

Hemiptera Microvelia 0.248 0.031 Hemiptera Microvelia 0.249 0.01

Isopoda Caecidotea 0.171 <0.001 Isopoda Caecidotea 0.192 0.010

Arhynchobdellida Erpobdella punctata 0.25 <0.001

Amphipoda Gammarus 0.201 <0.001

Zygoptera Agria 0.178 <0.001

Zygoptera Calopteryx 0.224 <0.001

Autumn Trichoptera Ceratopsyche 0.121 <0.001 Diptera Psychoda* 0.179 <0.001

Trichoptera Cheumatopsyche* 0.122 0.010 Trichoptera Ceratopsyche 0.134 0.031

Trichoptera Hydropsyche* 0.142 <0.001 Planorboidea Menetus dilatatus* 0.191 0.031

Amphipoda Gammarus* 0.217 0.020 Planorboidea Physella* 0.188 <0.001

Coleoptera Ectopria* 0.181 0.041

132

Coleoptera Haliplidae* 0.149 0.010

Cerithioidea Goniobasis* 0.189 0.010

Winter Diptera Culicoides 0.226 0.041 Diptera Culicoides 0.218 0.031

Diptera Dasyhelea* 0.207 <0.001 Tricladida Planariidae 0.242 <0.001

Tricladida Planariidae 0.252 <0.001 Ephemeroptera Baetidae 0.252 0.031

Emphemeroptera Baetidae 0.251 <0.001

133

FIGURE LEGENDS:

Figure 3.1: Macroinvertebrate density for L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

Figure 3.2: Taxonomic community dynamics between stream reaches within each sampling season. Panels represent 4-D NMDS results on a 2-D axis with standard error

95% confidence ellipses for stream reach and season.

Figure 3.3: Functional community dynamics between stream reaches within each sampling season. Panels represent 4-D NMDS results on a 2-D axis with standard error

95% confidence ellipses for stream reach and season.

Figure 3.4: Mean macroinvertebrate functional feeding group (FFG) relative abundance within L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated. Legend: CG – collector-gatherer, CG – collector-filterer, H – herbivore, P – predator, D – detritivore/shredder.

134

FIGURES:

Figure 3.1:

135

Figure 3.2:

136

Figure 3.3:

137

Figure 3.4:

138

SUPPLEMENTAL MATERIALS:

Supplemental Table 3.1: Description of functional trait states assigned to each taxonomic identification. Taxonomic names in parenthesis indicates taxa that were used to substitute unknown traits states for the corresponding taxon. Trophic habitat letter represent collector-gatherer (CG), collector-filterer (CF), herbivore (H), predator (P), and detritivore/shredder (D).

Ecology Life History Morphology

Trophic Thermal Develop- Size at Order Taxa Habitat Preference Habitat Voltinism ment Respiration Maturity Citations

Chironomidae Cold/Cool Fast Diptera CG Sprawler Univoltine Gill Small 5, 7 (Diamesinae) Stenothermal Seasonal

Simulium Cool/Warm Bi- or Fast Diptera CF Clinger Tegument Small 5, 7, 10 (Simuliidae) Eurythermal Multivoltine Seasonal

Cool/Warm Slow Diptera Tipulidae D Burrower Univoltine Gill Medium 5, 7 Eurythermal Seasonal

Dasyhelea Cold/Cool Fast Diptera CG Sprawler Univoltine Gill Small 7, 10 (Ceratopogonidae) Stenothermal Seasonal

Cool/Warm Slow Diptera Empididae P Sprawler Univoltine Tegument Medium 5, 7 Eurythermal Seasonal

Ceratopogon Cool/Warm Fast Diptera P Sprawler Univoltine Tegument Small 5, 7 (Ceratopogonidae) Eurythermal Seasonal

Psychoda Cold/Cool Bi- or Fast Diptera CG Burrower Tegument Small 5, 7, 10 (Maruina) Stenothermal Multivoltine Seasonal

Forcipomyia Cold/Cool Fast Diptera CG Sprawler Univoltine Gill Small 7, 10 (Ceratopogonidae) Stenothermal Seasonal

139

Atrichopogon Cool/Warm Fast Diptera P Sprawler Semivoltine Tegument Small 7, 10 (Ceratopogonidae) Eurythermal Seasonal

Culicoides Cold/Cool Fast Diptera P Burrower Univoltine Gill Small 7, 10 (Ceratopogonidae) Stenothermal Seasonal

Cold/Cool Fast Diptera Maruina H Clinger Univoltine Tegument Small 7 Stenothermal Seasonal

Cool/Warm Bi- or Non- Haplotaxida Naididae D Burrower Tegument Small 9, 10 Eurythermal Multivoltine Seasonal

Goniobasis Cool/Warm Bi- or Non- Cerithioidea H Clinger Gill Large 9, 10 (Pleuroceridae) Eurythermal Multivoltine Seasonal

Cold/Cool Slow Trichoptera Cheumatopsyche CF Clinger Univoltine Gill Medium 5, 7 Stenothermal Seasonal

Cool/Warm Slow Trichoptera Hydropsyche CF Clinger Univoltine Gill Medium 5, 7 Eurythermal Seasonal

Cool/Warm Slow Trichoptera Ceratopsyche CF Clinger Univoltine Gill Medium 5, 7 Eurythermal Seasonal

Cool/Warm Slow Trichoptera Hydroptila H Clinger Univoltine Gill Small 5, 7 Eurythermal Seasonal

Cool/Warm Bi- or Fast Planorboidea Physella H Clinger Aerial Medium 9, 10 Eurythermal Multivoltine Seasonal

Cool/Warm Bi- or Fast Planorboidea Menetus dilatatus H Clinger Aerial Small 9, 10 Eurythermal Multivoltine Seasonal

Ferrissia Cool/Warm Non- Basommatophora (Ferrissia H Clinger Univoltine Aerial Small 1, 4 Eurythermal Seasonal rivularis)

Cool/Warm Non- Veneroidea Pisidium CF Burrower Univoltine Gill Small 9, 10, 2 Eurythermal Seasonal

140

Stenelmis Cool/Warm Non- Coleoptera CG Clinger Semivoltine Tegument Small 7, 10 (Elmidae) Eurythermal Seasonal

Cool/Warm Slow Coleoptera Ectopria H Clinger Semivoltine Gill Medium 5, 7 Eurythermal Seasonal

Haliplidae Cold/Cool Bi- or Slow Coleoptera H Clinger Gill Small 7, 10 (Brychius) Stenothermal Multivoltine Seasonal

Lampyridae Cool/Warm Slow Coleoptera P Climber Univoltine Tegument Medium 10, 3 (Luciola ficta) Eurythermal Seasonal

Erpobdella Warm Slow Arhynchobdellida P Climber Semivoltine Tegument Large 10 punctata Eurythermal Seasonal

Helobdella fusca Warm Bi- or Non- Rhynchobdellida P Climber Tegument Medium 10 (H. stagnalis) Eurythermal Multivoltine Seasonal

Helobdella Warm Bi- or Non- Rhynchobdellida P Climber Tegument Medium 9, 10 stagnalis Eurythermal Multivoltine Seasonal

Cold/Cool Bi- or Non- Lumbriculida Lumbriculidae D Burrower Tegument Large 9, 10 Stenothermal Multivoltine Seasonal

Planariidae Cool/Warm Bi- or Fast Tricladida P Climber Tegument Medium 9 (Turbellaria) Eurythermal Multivoltine Seasonal

Cool/Warm Bi- or Fast Ephemoropetera Baetidae (Baetis) CG Swimmer Gill Small 7 Eurythermal Multivoltine Seasonal

Microvelia Cool/Warm Bi- or Fast Hemiptera P Skater Aerial Small 5, 7 (Planorbidae) Eurythermal Multivoltine Seasonal

Cool/Warm Slow Zygoptera Agria P Clinger Semivoltine Tegument Medium 7 Eurythermal Seasonal

Cool/Warm Slow Zygoptera Calopteryx P Climber Semivoltine Gill Large 5, 7 Eurythermal Seasonal

Cold/Cool Non- Amphipoda Gammarus D Swimmer Univoltine Gill Medium 9, 10 Stenothermal Seasonal

141

Cold/Cool Bi- or Non- Amphipoda Hyalella D Climber Gill Small 9, 10 Stenothermal Multivoltine Seasonal

Cool/Warm Bi- or Non- Copepoda Copepoda D Swimmer Tegument Small 9 Eurythermal Multivoltine Seasonal

Orconectes Cool/Warm Non- Decapoda D Climber Semivoltine Gill Large 8, 9, 10 rusticus Eurythermal Seasonal

Cold/Cool Bi- or Non- Isopoda Caecidotea D Clinger Gill Medium 9, 10 Stenothermal Multivoltine Seasonal

142

SUPPLEMENTAL REFERENCES:

1. Burky, AJ. 1971. Biomass turnover, respiration, and interpopulation variation in

the stream limpet Ferrissia rivularis (Say). Ecol. Monographs. 41:235-251.

2. Burky, AJ, DJ Hornbach, and CM Way. 1981. Brief note: Growth of Pisidium

casertanum (Poli) in West Central Ohio. The Ohio J. Sci. 81:41-44.

3. Buschman, LL. 1984. Biology of the firefly Pyractomena lucifera (Coleoptera:

Lampyridae). The Florida Ent. 67:529-542.

4. Dillon, RT and JJ Herman. 2008. Genetics, shell morphology, and life history of

the freshwater pulmonate limpets Ferrissia rivularis and Ferrissia fragilis. J.

Freshwater Ecol. 24:261-271.

5. Merritt RW, Cummins KW, Berg MB (2008) An introduction to the aquatic

insects of North America, 4th edn. Kendall/Hunt, Iowa.

6. Pecharsky, BL, PR Fraissinet, MA Penton, DJ Conklin 1990. Freshwater

macroinvertebrates of Northeastern North America. Cornell. University Press,

Ithaca, NY.

143

7. Poff, N L, JD Olden, NKM Vieira, DS Finn, MP Simmons, and BC Kondratieff.

2006. Functional trait niches of North American lotic insects: traits-based

ecological applications in light of phylogenetic relationships. J. North Am.

Benthol. Soc. 25:730–755.

8. Prins, R. 1968. Comparative ecology of the crayfishes Orconectes rusticus

rusticus and Cambarus tenebrosus in Doe Run, Meade County, Kentucky. Int.

Rev. gesamten Hydrobiol. 53:667-714.

9. Thorp JW, Covich AP (2001) Ecology and classification of North American

freshwater invertebrates, 2nd edn. Academic Press, San Diego.

10. Vieiran NKM, NL Poff, DM Carlisle, SR Moulton, ML Koski, BC Kondratieff.

2006. A database of lotic invertebrate traits for North America. U.S. Geological

Survey Data Series 187. US Geological Survey, US Department of the Interior,

Reston, Virgina.

144

Supplemental Table 3.2: Macroinvertebrate taxonomic and abundance list that were sampled over a 31 month period in both the removal and honeysuckle stream reaches at BO park.

Honeysuckle Reach Removal Reach

Relative Total Relative Total

Order Taxa Abundance Abundance Abundance Abundance

Diptera Chironomidae 4032 47.98 9797 55.05

Diptera Simulium 104 01.24 448 02.52

Diptera Tipulidae 24 00.29 109 00.61

Diptera Dasyhelea 18 00.21 25 00.14

Diptera Empididae 13 00.15 46 00.26

Diptera Ceratopogon 3 00.04 2 00.01

Cyclorrhaphas- Diptera 2 00.02 1 00.01 Brachycera

145

Diptera Psychoda 2 00.02 6 00.03

Diptera Forcipomyia 1 00.01 2 00.01

Diptera Atrichopogon 0 00.00 2 00.01

Diptera Culicoides 0 00.00 4 00.02

Diptera Dolchopodidae 0 00.00 1 00.01

Diptera Ephydridae 0 00.00 1 00.01

Diptera Maruina 0 00.00 2 00.01

Diptera Muscidae 0 00.00 1 00.01

Haplotaxida Naididae 3066 36.49 5601 31.47

Cerithioidea Goniobasis 433 05.15 236 01.33

Trichoptera Cheumatopsyche 186 02.21 326 01.83

Trichoptera Hydroptila 93 01.11 495 02.78

146

Trichoptera Hydropsyche 41 00.49 132 00.74

Trichoptera Ceratopsyche 6 00.07 64 00.36

Planorboidea Physella 89 01.06 170 00.96

Planorboidea Menetus dilatatus 4 00.05 0 00.00

Basommatopho Ferrissia 1 00.01 1 00.01 ra

Veneroidea Pisidium 77 00.92 25 00.14

Coleoptera Stenelmis 65 00.77 79 00.44

Coleoptera Ectopria 3 00.04 3 00.02

Coleoptera Haliplidae 2 00.02 0 00.00

Coleoptera Lampyridae 1 00.01 0 00.00

Arhynchobdelli Erpobdella punctata 41 00.49 51 00.29 da

147

Lumbriculida Lumbriculidae 26 00.31 26 00.15

Tricladida Planariidae 25 00.30 34 00.19

Rhynchobdelli Helobdella stagnalis 17 00.20 10 00.06 da

Rhynchobdelli Helobdella fusca 1 00.01 2 00.01 da

Ephemeroptera Baetidae 12 00.14 21 00.12

Hemiptera Microvelia 3 00.04 0 00.00

Hemiptera Homoptera 2 00.02 3 00.02

Collembola Actaletidae 2 00.02 0 00.00

Collembola Bourletiella 1 00.01 0 00.00

Collembola Spinactaletes 1 00.01 2 00.01

Collembola Hypogastruridae 0 00.00 1 00.01

148

Zygoptera Argia 1 00.01 2 00.01

Zygoptera Calopteryx 1 00.01 4 00.02

Amphipoda Gammarus 1 00.01 4 00.02

Amphipoda Hyalella 0 00.00 1 00.01

Copepoda Copepoda 1 00.01 51 00.29

Decapoda Orconectes rusticus 1 00.01 3 00.02

Lepidoptera Noctuidae 1 00.01 0 00.00

Hydrachnidia Hydrachnidia 0 00.00 3 00.02

Isopoda Caecidotea 0 00.00 1 00.01

Total 8403 100.00 17798 100.00

Species Richness 40 43

149

Supplemental Table 3.3: Taxonomic and functional diversity metrics, FFG relative abundance, and ambient conditions statistical comparisons between L. maackii and removal stream reaches.

Parameter W-Value P-Value

Density 1807.5 < 0.0010

Diversity 2460.0 0.4434

Richness 0167.5 < 0.0001

Functional Richness 0188.0 0.5077

Functional Evenness 0200.0 0.9553

Functional Dispersion 0271.0 0.4400

Collector-Gatherer 2008.5 0.0760

Collector-Filterer 2190.0 0.6453

Herbivore 3241.0 0.0138

Predator 1380.5 0.3023

Detritivore 2014.5 0.9625

Canopy Cover 4986.0 < 0.0001

Light Availability 0063.0 < 0.0001

Phosphorous 0445.5 0.7444

150

Nitrite 0296.5 0.9932

Nitrate 0352.5 0.3145

Ammonia 0422.0 0.1642

TSS 0365.0 0.9419

151

Supplemental Table 3.4: ADONIS multiple permutation result for taxonomic and functional trait community dynamics between stream reaches, season, and their interaction and between stream reaches for each season.

Taxonomic Functional

Grouping Factor R2 F-Statistic P-Value R2 F-Statistic P-Value

Stream Reach 0.0135 3.8056 0.0004 0.0096 2.3969 0.0402

Season 0.0650 6.0897 0.0001 0.0429 3.5704 0.0001

Reach × Season 0.0166 1.5550 0.0357 0.0093 0.7769 0.6820

Spring 0.0459 2.6014 0.0190 0.0125 0.6859 0.5786

Summer 0.0451 1.6064 0.0979 0.0649 2.3606 0.0471

Autumn 0.0344 2.8554 0.0047 0.0089 0.7196 0.6032

Winter 0.1459 0.7551 0.6253 0.2294 1.1973 0.3092

152

Supplemental Table 3.5: Functionally relevant macroinvertebrate taxa for each season.

An asterisk indicates taxa that were unique to that season.

Grouping Observed P-

Factor Order Taxa Value Value

Spring Diptera Atrichopogon* 0.226 0.010

Diptera Ceratopogon* 0.182 <0.001

Diptera Culicoides 0.217 <0.001

Tricladidae Planariidae 0.258 <0.001

Ephemeroptera Baetidae 0.253 <0.001

Amphipoda Hyalella* 0.210 <0.001

Hydrachnidia Hydrachnidia* 0.200 0.020

Summer Diptera Empididae* 0.132 <0.001

Diptera Tipulidae* 0.201 <0.001

Coleoptera Stenelmis* 0.197 <0.001

Coleoptera Ectopria 0.129 <0.001

Coleoptera Haliplidae 0.143 <0.001

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Coleoptera Lampyridae* 0.160 <0.001

Cerithioidea Goniobasis 0.142 <0.001

Planorboidea Menetus dilatatus 0.184 <0.001

Planorboidea Physella 0.158 <0.001

Basommatophora Ferrissia* 0.163 <0.001

Arhynchobdellida Erpobdella punctata* 0.237 <0.001

Rhynchobdellida Helobdella fusca* 0.210 <0.001

Rhynchobdellida Helobdella stagnalis* 0.210 <0.001

Trichoptera Hydroptila 0.118 <0.001

Trichoptera Ceratopsyche 0.094 <0.001

Trichoptera Cheumatopsyche 0.106 <0.001

Trichoptera Hydropsyche 0.126 <0.001

Lumbriculida Lumbriculidae* 0.278 0.020

Zygoptera Agria 0.176 <0.001

Zygoptera Calopteryx* 0.217 <0.001

Amphipoda Gammarus 0.195 <0.001

Copepod Copepoda* 0.237 0.041

Decapoda Orconectes rusticus* 0.227 <0.001

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Hemiptera Microvelia* 0.249 <0.001

Isopoda Caecidotea* 0.181 <0.001

Autumn Diptera Psychoda* 0.183 <0.001

Coleoptera Ectopria 0.174 0.010

Coleoptera Haliplidae 0.149 0.010

Cerithioidea Goniobasis 0.180 0.041

Planorboidea Menetus dilatatus 0.191 0.020

Planorboidea Physella 0.185 <0.001

Trichoptera Hydropsyche 0.148 0.010

Trichoptera Ceratopsyche 0.128 0.011

Trichoptera Cheumatopsyche 0.127 <0.001

Trichoptera Hydroptila 0.140 0.010

Zygoptera Agria 0.224 0.041

Amphipoda Gammarus 0.225 0.020

Winter Diptera Chironomidae* 0.132 0.010

Diptera Culicoides 0.222 <0.001

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Diptera Dasyhelea 0.199 <0.001

Tricladida Planariidae 0.247 <0.001

Ephemeroptera Baetidae 0.252 <0.001

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SUPPLEMENTAL FIGURES:

Supplemental Figure 3.1: Taxonomic richness and diversity for L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

Supplemental Figure 3.2: Functional diversity metrics for L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

Supplemental Figure 3.3: Taxonomic and functional community dynamics between stream reach and sampling season. Panels represent 4-D NMDS results on a 2-D axis with standard error 95% confidence ellipses for stream reach and season.

Supplemental Figure 3.4: Mean macroinvertebrate functional feeding group (FFG) relative abundance between L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

Supplemental Figure 3.5: Mean above stream canopy cover between L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

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Supplemental Figure 3.6: Mean light at the surface of the stream between L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

Supplemental Figure 3.7: Ambient nutrient and total suspended solid dynamics L. maackii and removal stream reaches. Letters on the x-axis represent sampling months for years indicated.

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Supplemental Figure 3.1:

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Supplemental Figure 3.2:

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Supplemental Figure 3.3:

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Supplemental Figure 3.4:

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Supplemental Figure 3.5:

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Supplemental Figure 3.6:

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Supplemental Figure 3.7:

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CHAPTER 4: TERRESTRIAL-AQUATIC CONNECTIONS: RIPARIAN INVASION

OF LONICERA MAACKII INFLUENCES STREAM ALGAL GROWTH AND

THROUGHFALL CHEMISTRY

ABSTRACT:

Riparian forests are a critical interface between terrestrial and aquatic communities, allowing for the transfer of subsidies that support ecosystem processes and communities across the terrestrial-aquatic gradient. Invasion of Lonicera maackii throughout the Midwest has resulted in near-monocultures along headwater streams, substantially altering the riparian plant community and modifying riparian function along headwater streams. We investigated the effects of L. maackii on throughfall chemistry and algal communities in a L. maackii forest and headwater streams via two concurrent studies. During the growing season of 2015 throughfall collections were made in open conditions (Open), in a forest (Upper Canopy) and within the forest, under L. maackii shrubs (Honeysuckle) with n = 10 funnels in each habitat. In a second study, Lonicera maackii effects on stream algal growth was measured during spring 2015 and autumn

2014 in experimental L. maackii removal and non-removal reaches. Four leachate treatments were created: L. maackii leaves, flowers (spring only), and berries (autumn only) and Acer negundo leaves at 100% and 25% leachate concentrations. Algal communities were exposed to leachate solutions for 14 days in two headwater streams

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using nutrient diffusing substrates. Honeysuckle canopies substantially increased carbon species concentration and nitrogen deposition and reduced throughfall volume as compared to Upper Canopy and Open treatments (all P < 0.05). Lonicera maackii leachate had differential effects on algal growth that were primarily driven by light conditions created by riparian L. maackii (P < 0.01). Lonicera maackii leaves reduced algal growth regardless of light conditions during spring while flowers reduced algal growth primarily in low light conditions (P < 0.05). Lonicera maackii berries did not significantly influence algal growth. In summary, riparian L. maackii has the potential to alter nutrient subsidies during rain events that enter aquatic systems as throughfall, and suppress stream algal growth early in the growth season, impacting nutrient cross-system subsidies and one of the basal food resources in aquatic systems.

INTRODUCTION:

Anthropogenic nutrient loading to aquatic systems is a global phenomenon often leading to eutrophication of water systems with impacts on aquatic biota and broader ecosystem processes (Anderson et al. 2002, Smith 2003). For example, The Great Lakes,

USA and the coastal oceans of China have experienced massive algal blooms, resulting in public health concerns due to the threat of toxins present in these water bodies (Anderson et al. 2002, Hudnell 2010, Johnson et al. 2010, Zhang 1994). Point and non-point sources of pollution have been the primary contributors to increased nutrient levels in aquatic ecosystems (Carpenter et al. 1998, Dolan 1993, Heisler et al. 2008). Regulation of point sources (e.g., pipes) have reduced nutrient loading into aquatic waterbodies; however, progress with non-point sources (e.g., agricultural runoff) have been less successful

(Conroy et al. 2005). Riparian zones are natural buffers along aquatic habitats that help

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mediate non-point source pollution entering aquatic systems (Barling and Moore 1994,

Lowrance et al. 1984); therefore, it is imperative that we understand how these zones may also contribute to nutrient fluxes to aquatic systems.

Riparian zones act as an interface between terrestrial and aquatic ecosystems, supporting the transfer of subsidies (e.g. leaf litter) between systems, mediating storm water runoff, and reducing nutrient loading entering aquatic systems (Baxter et al. 2005,

Vannote et al. 1980). Riparian function is tightly linked with plant community composition in the riparian zone (Gregory et al. 1991); however, riparian zones have become degraded due to increased riparian forest fragmentation and the introduction of invasive plants (Chamier et al. 2012; Richardson et al. 2007). Riparian forests are highly susceptible to non-native invasive plants which can alter ecosystem function and potentially disrupt riparian zone function (Chamier et al. 2012, Greene 2014, Richardson et al. 2007). For example, invasive plants can suppressive the growth of potential competitors and riparian forests with low herbaceous ground cover tend to contribute a substantial amount of storm water runoff and nutrients to aquatic systems (Walsh et al.

2005). Invasion-related alterations of terrestrial-aquatic community dynamics may result in impacts on food-web interactions, nutrient cycling and ecosystem processing (e.g. decomposition) that occur between terrestrial and aquatic habitats (Baxter et al. 2005,

Davis et al. 2011, Greig et al. 2012).

Amur honeysuckle (Rupr.) Maxim. (Lonicera maackii) is one of the most successful invasive species in the North American Midwest. This species excels in edge habitats and is a significant invader of riparian and stream bank habitats (McNeish et al.

2012, 2015; Fig. 1 b & d). Amur honeysuckle possesses a suite of traits that make it a

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successful invader. It has a longer growing season than native species (McEwan et al.

2009), chemical resistance against insect herbivores (Dorning and Cipollini 2005;

McEwan et al. 2009), and also has the capacity to chemically suppress plants that it competes with (Cipollini et al. 2008, McEwan et al. 2010). This plant’s negative influences include impacting a variety of plant communities (Collier et al. 2002), terrestrial insects and pollinators ( McEwan et al. 2009, McKinney and Goodell 2010), and bird nesting success (Borgmann and Rodewald 2004, Rodewald 2009). Amur honeysuckle riparian forests also create dense leaf canopies overarching headwater streams that directly contribute a substantial amount of leaf litter to these streams

(McNeish et al. 2012, 2015; Fig. 1 b & d). Our previous work found L. maackii dominated riparian forests (1) contributed a significant amount of in-stream L. maackii leaf litter while reducing the overall availability of native leaf litter, and (2) L. maackii leaf breakdown was ~ 4 × faster than native leaf litter in headwater streams (McNeish et al. 2012, 2015). McEwan et al. (2012) and Arthur et al. (2012) demonstrated that L. maackii leaves are higher in nitrogen and leach significantly more cations during rainfall events compared to native plant species, potentially influencing soil chemistry and nutrient availability; therefore, it is imperative that we understand how L. maackii riparian forests impact nutrient availability between terrestrial and aquatic systems.

This study consisted of two experiments which aimed, in combination, to identify how L. maackii invasion of riparian forests (1) alters the chemistry of throughfall (rain water that passes through leaf canopies) that can enter aquatic systems directly or indirectly as runoff and (2) impacts stream algal growth – the dominant primary producers in stream ecosystems. In the first experiment we measured throughfall

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chemistry above and under L. maackii in a forest and, simultaneously, in an adjacent open area. In the second experiment, we assessed potential for in-stream effects of L. maackii riparian invasion using nutrient diffusing substrates. We predicted L. maackii forests will (H1) increase nitrogen, phosphorous, and carbon concentrations in throughfall, (H2) resulting in an increase in nutrient deposition. We also hypothesized that

(H3) stream algal growth would increase when exposed to L. maackii leachate due to high nutrient concentration in L. maackii leachate in comparison to leachate made from a native riparian plant. Finally, we hypothesized that (H4) L. maackii riparian forests will result in reduced in-stream light availability due to the dense canopy of L. maackii overarching headwater streams.

METHODS:

Study Sites:

The throughfall study took place at Taylorsville MetroPark located in southwestern OH, USA (39.52°N, 84.10°W). A forested site with a native upper canopy and dense shrub layer of L. maackii and an adjacent prairie were selected based on close proximity to each other within the park (Fig. 1a). The upper canopy was composed of a mix of deciduous forest trees including Platanus occidentalis L. (American Sycamore),

Ulmus spp. L. (Elm), and Acer spp. including Acer negundo L. (Box elder). The prairie flora consisted of Dipsacus L. (Teasel), Asclepias L. (Milkweed), and Schizachyrium scoparium (Michx.) Naxh (Small bluestem) and Andropogon gerardi Vitman (Big bluestem); however, the funnel installations were above the influence of these species.

Taylorsville MetroPark is part of the Five-Rivers Dayton MetroPark system and all aspects of the throughfall study took place in cooperation with land managers.

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Two head water streams, one located at Black Oak Park and the other located at

Fecher Park were used to study the effect of L. maackii leachate on algal growth (Fig. 1 b

& d). These streams are un-named tributaries (hereafter referred to as BO and FP) located in southwestern OH (BO: 36.63°N, 84.12°W; FP 36.67°N; 84.09°W). The streambeds were 1.5 – 5.0 m wide, formed on limestone geology, and L. maackii has been experimentally removed August 2010 from a portion of the study reach at BO

(McNeish et al. 2015, Schneider 1957). A L. maackii removal reach was created along a

160 m stream segment with a 5 m buffer on each stream side at BO. Lonicera maackii and all other invasive woody species were left intact upstream of the removal reach, creating a L. maackii and removal stream reaches. To prevent L. maackii regrowth

AquaNeat® Aquatic Herbicide, an Ohio EPA approved aquatic herbicide (EPA regulation number: 228-365; Nufarm Manufacturer; active ingredient Glyphosate N- glycine), was applied to cut stumps of invasive species within 48 h. Lonicera maackii maintenance removal took place twice post the 2010 removal event to further prevent regrowth. Lonicera maackii was also removed by a local land owner along a 20 m stream reach at FP summer 2014. Due to this opportunity, we used FP as a replicate stream since it had a removal and an upstream L. maackii site that was similar to BO. The

BO stream site was located within the Centerville-Washington Park District while FP site was located in the Bellbrook-Sugarcreek Park District and all aspects of the project were undertaken in cooperation with land managers.

Throughfall Experimental Design:

Throughfall rigs were set up within the forested and open prairie site and represented three conditions “Upper Canopy,” “Open” and “Honeysuckle.” All

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collections were made using a PVC apparatus that held 1 L amber Nalgene collection bottles, with 9.3 cm diameter funnels and these funnels were established just prior to a rain event and removed within 10 hours after the rain had stopped. Collection bottles were be capped during dry periods. Just prior to rain events bottle caps were replaced with funnels that had a 250 µm mesh insert to prevent debris from entering the bottles and contaminating throughfall and rain water samples. “Upper Canopy” and

“Honeysuckle” rigs were haphazardly installed along a transect in a L. maackii invaded forest. Rigs were at least 5 m apart and had a L. maackii crown of at least 2 m in width.

The “Upper Canopy” collection funnel was established at a height that placed the collection above the L. maackii shrub layer (>3 m in many instances). An additional bottle was attached approximately 70 cm above the ground to capture “Honeysuckle” throughfall that passed through both the forest canopy and the leafy canopy of the L. maackii shrub layer. Rigs were stabilized with 1 m rebar and a 122 cm fence post to prevent falling over and reduce swaying during wind events.

Samples were collected immediately after rain events on 7 June, 27 July, 31

August, and 5 November of the year 2015. Samples were processed immediately upon return to the lab for carbon, nitrogen and phosphorus content. Carbon analyses included total carbon (TC), total organic carbon (TOC), dissolved inorganic carbon (DIC), and dissolved organic carbon (DOC). All organic carbon measurements were acidified with

2N HCl to ~ 2.0pH. Dissolved OC was measured post acidification and filtered with a

0.45 µm filter. Dissolved IC was calculated as the difference between total dissolved carbon and DOC. All carbon measurements were conducted with a TOC-V analyzer.

Total nitrogen (TN) was determined by 10 mL samples that were ampulated with

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potassium persulfate added as an oxidizer and autoclaved at 121 ºC for four hours, which converted all forms of nitrogen to NO3-N. To determine dissolved nitrogen (DN) concentration 10 mL samples were first filtered through 10 mL a 0.45 µm filter and then processed as explained for TN. Nitrite (NO2-N), nitrate (NO3-N), and ammonia (NH3-N) were determined via standard colorimetric methods using the DREL 2800 water quality kit from Hach Company. Nitrate was identified with the cadmium reduction method, which created a pink colored reaction if NO3-N was present and colorimetrically determined with a spectrophotometer at 500 nm. The Nessler method was used to characterize NH3-N concentration, which created a yellow-orange colored reaction that was read at 524 nm. Nitrate concentrations were identified via the diazotization method and resulted in an amber color reaction that was measured at 507 nm. Total

-3 orthophosphate (PO4 ; hereafter referred to as P) was measured using the malachite green method (D’Angelo et al. 2001). A 1.75% (w/v) ammonium heptamolybdate tetrahydrate solution in 6.3N sulfuric acid was added to 5 mL samples and shaken with a benchtop shaker for 10 min. A 0.035% (w/v) solution of Malachite Green carbinol hydrochloride in 0.35% (w/v) aqueous polyvinyl alcohol solution was added to the sample and then shaken for an additional 20 min. to develop a light green-bluegreen color reaction. The solution was then read at 630 nm and used to calculate P concentrations.

An additional 5 mL was filtered through a 0.45 µm filter and then processed as explained for P for soluble reactive P (SRP: also known as dissolved reactive P). Deposition estimates were calculated for each nutrient as by correcting for the total volume of the sample and circular area of the collection funnel (concentration * volume/area based on radius; McEwan et al. 2012).

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Lonicera maackii Effect on Stream Algal Growth:

Seasonal L. maackii leachate effects on stream algal growth was assessed autumn

2014 and spring 2015. Fall leachate solutions were produced from L. maackii berries and

L. maackii and Acer negundo senesced leaves. Spring leachate solutions were created from L. maackii flowers and fresh L. maackii and A. negundo leaves. All plant materials were collected from the riparian forest at BO and FP. Leachate solutions were created in a 1:10 ratio of plant material to DI water similar to other plant leachate studies (e.g.

McEwan et al. 2010). Plant material leached for 24 hr at room temperature and then was strained out of the solution with a 250 µm sieve. A subset of each leachate was diluted to a 25% solution with DI water to create a concentration comparison to the 100% leachate concentration. The control treatment consisted of DI water that was also exposed to room temperature for 24 hr.

Nutrient diffusing substrates (NDS) were created from leachate treatments following methods outlined in Tank et al. (2007). Nutrient DS have been traditionally used to identify N and P nutrient limitation in aquatic systems. Algal communities respond to the presence of nutrients, and thus, are considered representative of nutrient limitation in aquatic systems (Tank et al. 2007). For this study NDS were used to identify the effect of L. maackii leachates on algal growth in comparison to the native riparian tree A. negundo and a control treatment. Substrates consisted of small plastic canisters filled with a 2% gelled agar solution that was made for all leachate treatments (n

= 7 treatments). The cap of each canister had a porous glass disc that allowed nutrients and chemicals from agar leachate to seep out of the canister and into the water column.

Autumn deployment occurred on 26 October 2014 at BO and FP and spring deployment

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occurred on 4 June 2015. Autumn 2014 data were used in lieu of autumn 2015 because of unexpected early leaf senesce and increased flooding events that occurred during the

2015 autumn season. NDS were attached the plastic L-bars which were then connected together three at a time to create an NDS block (Fig. 1 c). Blocks were anchored in the stream with 10 inch stakes within run habitats. Run habitats were chosen since the water was deeper compared to riffle habitats and would decrease the chance of exposure to the atmosphere and drying. Two replicate run habitats (n = 72 NDS/run) were chosen in both the removal and L. maackii stream reaches at BO while 1 run habitat within each reach was identified at FP due to habitat availability at FP. Algal growth on NDS were analyzed after 15 days following standard methods in the American Public Health

Association (1999) and Steinman and Lamberti (2007).

Algal growth is susceptible to variations in light, flow, and nutrient dynamics in aquatic systems (Besemer et al. 2007); therefore, stream ambient conditions were measured weekly at each NDS block (n = 3/reach). Above stream canopy cover was measured with a spherical densiometer and used to calculate percent cover. Ambient light conditions were measured at the surface of the water and each block (underwater) with a Milwaukee light meter.

Statistical Analyses:

Lonicera maackii impact on nutrient concentration and deposition during rain events were compared between Open, Upper Canopy, and Honeysuckle treatments throughout time. Outliers were identified for all parameters and removed if they were greater than 3 × the interquartile range for each treatment within each time point (Quinn and Keough 2004). Data were then screened for normality. Normal data were analyzed

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with repeated measures ANOVA (rmANOVA) while non-normal data were analyzed with friedmans test (Sokal and Rohlf 1981, Zar 1999). In some cases both parametric and non-parametric statistical results were presented due to mixed normality (i.e. at least one treatment failed normality tests within a time point) in order to balance statistical interpretation. Bonferroni pairwise post-tests were only conducted if main effects

(treatment, time) were significant with pairwise.t.test() with the R base package. The post.hoc.friedman.nemenyi.test() was used to conduct pairwise post-tests between treatment only when friedmans test was significant with the ‘PMCMR’ package in R

(Pohlert 2015).

Lonicera maackii leachate effects on stream algal growth were compared between leachate treatments within stream reaches. All treatments were screened for outliers and normality as explained previously for nutrient data. There was a significant light effect on algal response between the removal (high light) and L. maackii (low light) stream reaches; therefore, to control for the effect of light algal responses were analyzed within each stream reach (i.e. removal-high light and L. maackii-low light). Those data that were normal were analyzed with a two-way ANOVA with leachate type and concentration as main effects. Post-hoc tests were conducted for main effects that were significant with a Bonferroni correction using pairwise.t.test(). Non-normal data were analyzed with Kruskal-Wallis test and followed with post.hoc.kruskal.nemenyi.test() for significant models with the ‘PMCMR’ package in R (Pohlert 2015). All statistical analyses were conducted in R version 3.2.2.

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RESULTS:

Throughfall Chemistry:

The L. maackii canopy significantly influenced the chemistry of throughfall.

Throughfall was significantly intercepted by the L. maackii canopy resulting in a substantial reduction in throughfall in the Honeysuckle treatment as compared to that found in the Upper Canopy (Fig. 2; FTreatment = 38.68, P < 0.0001, df = 2; FTime = 153, P <

0.0001, df = 3). The Honeysuckle and the Upper Canopy measurements both exhibited significantly increased throughfall pH compared to the open treatment (Fig. 1; FTreatment =

180.5, P = 0.0022, df = 2; FTime = 5.197, P < 0.0001, df = 3; FInteraction = 13.47, P <

0.0001, df = 6). The Honeysuckle treatment substantially increased all carbon species

(except DOC) concentrations as compared to the Upper Canopy and Open treatments

(Suppl. Table 1; Figs. 3 & 4). Overall deposition of carbon was significantly greater for most carbon species as estimated from Honeysuckle treatment throughfall, with this result the strongest for DIC as compared to other treatments (Suppl. Table 2; Figs. 3 &

4). Lonicera maackii canopy subsidized throughfall from ~ 80 – 1300 % more with carbon relative to the Open treatment, whereas the Upper Canopy subsidized throughfall with carbon from ~ 3 – 55% as compared to the Open treatment (Table 1). Interestingly, time was an important factor in the availability of carbon in throughfall samples, especially for TOC concentration with concentrations increasing throughout the growing season (Suppl. Table 1).

Nitrogen concentration and deposition was differentially influenced by the presence of the L. maackii shrub canopy (Suppl. Tables 1 & 2). Total N and DN deposition were significantly reduced in throughfall collected in the Honeysuckle and

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Upper canopy collections throughout the growing season (Suppl. Table 2; Suppl. Fig. 1).

The Honeysuckle collections exhibited ~ 19 – 73% reduction of TN and DN deposition while ~ 3 – 43% reduction in deposition were recorded in the Upper Canopy collections

(Table 1). Interestingly, TN and DN concentration was statistically similar between treatments but were generally greater under L. maackii canopies (Suppl. Table 1; Suppl.

Fig. 1). Inorganic nitrogen species were generally not influenced by the Honeysuckle or

Upper Canopies, with the exception of nitrite (Suppl. Tables 1 & 2; Suppl. Fig.2).

Honeysuckle treatment throughfall was general greater in nitrite and ammonia concentrations while nitrate concentrations tended to be lower compared to the other treatments (Suppl. Fig. 2). Overall deposition of inorganic nitrogen species were reduced in Honeysuckle collections compared to the Upper Canopy and Open treatment (Suppl.

Fig. 2). Nitrite and nitrate concentrations significantly increased throughout the study period whereas ammonia decreased, with the same pattern true for deposition estimates

(Suppl. Table 1; Suppl. Fig. 2).

Honeysuckle and the Upper canopies tended to increase phosphorus and SRP concentration and deposition; although this effects was not significant, it was most apparent during the months of June and August (Suppl. Table 1 & 2; Suppl. Fig. 3).

Honeysuckle collections had greater P and SRP by ~ 49 – 4,800 % whereas the Upper

Canopy increased concentrations by nearly 4,000% during August (Table 1).

Phosphorous and SRP levels during July were nearly 4 × greater compared to all other sampling months (Suppl. Fig. 3).

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Lonicera maackii Effect on Stream Algal Growth:

Lonicera maackii leaf and flower leachate and riparian shading differentially influenced stream algal growth, with patterns generally similar across stream sites.

Riparian L. maackii resulted in a significant decrease in light within both stream reaches, creating a high light (removal) and low light environments (BO: df = 1, χ2 = 5.33, P =

0.0209l; FP: df =1, t = -9.33, P = 0.0112). Algal growth was significantly influenced by both plant leachate and concentration during the spring season (Suppl. Table 3, Fig. 5).

Lonicera maackii leaves and flowers reduced algal growth compared to A. negundo and control treatments within both removal and L. maackii stream reaches across stream sites

(Fig. 5). When leachate was diluted to a 25% solution L. maackii flowers tended to increase algal growth in both removal reaches (Fig. 5). Interestingly, L. maackii leaves substantially reduced algal growth in the L. maackii stream reach at FP (Fig. 5). Overall biomass of the community was influenced by leachate treatment and concentration across stream sites (Suppl. Table 3, Suppl. Fig. 4). Biomass significantly increased when exposed L. maackii flower leachate as compared to the L. maackii leaf leachate within the

100% concentration, with this pattern re-occurring across stream reaches and sites

(Suppl. Fig. 4). Lonicera maackii leaf leachate decreased biomass in most instances compared to control and A. negundo leachate treatments within the 25% concentration leachate treatment (Suppl. Fig. 4). The effect of L. maackii leaf and flower leachate was much more apparent when observed with the production to total biomass relationship

(P:B - chlorophyll a standing stock biomass : total biofilm community biomass). In low light environments (L. maackii reach) L. maackii leaves and flowers reduced P:B at both

100% and 25% concentrations (Suppl. Table 3, Suppl. Fig. 5). Interestingly L. maackii

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flowers increased the P:B ratio in the 25% concentration treatment under reduced light conditions (Suppl. Fig. 5). Production to biomass was zero for the 25% L. maackii leachate treatment at the reduced light site at FP (Fig. 5; Suppl. Fig 6). However, L. maackii leaves did support a substantial amount of total biomass for this treatment, suggesting under reduced light conditions L. maackii leachate can support non- photosynthetic microorganisms (Suppl. Fig. 4). In contrast to the reduced light treatment,

L. maackii leaves supported ~ 4 × greater P:B in the high light environment in FP (Suppl.

Fig. 5).

Autumnal leachate treatments tended not to influence stream algal growth but there were some interesting patterns between stream sites and light conditions (Suppl.

Table 4; Fig. 6). In high light conditions L. maackii berries at 100% concentration reduced algal growth compared to leaf leachate treatments with this pattern most prominent at BOP (Fig. 6). In reduced light conditions L. maackii leaves and berries increased algal growth at BOP regardless of concentration while L. maackii berries reduced algal growth at lower concentrations at FP (Fig. 6). Leachate effects were only significant at the low light conditions at FP (Suppl. Table 4). These differences highlight that leachate effects may be contingent upon stream conditions.

DISCUSSION:

Riparian zones are critical buffer habitats that act as an interface between aquatic and terrestrial ecosystems (Barling and Moore 1994, Baxter et al. 2005, Lowrance et al.

1984). These zones are susceptible to plant invasions which can alter terrestrial-aquatic connections and disrupt riparian zone function (Chamier et al. 2012, Greene 2014,

Richardson et al. 2007). Lonicera maackii is a dominant invasive shrub in the Midwest

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that is known to change plant composition and species richness in forests and riparian zones (Gould and Gorchov 2000, Musson and Mitsch 2003, White et al. 2014). Several studies have indicated that L. maackii leaf litter is high in nitrogen, breaks down rapidly, and influences the transformation of nitrogen (Arthur et al. 2012, McNeish et al. 2012,

Poulette and Arthur 2012, Trammell et al. 2012). Our previous work found that L. maackii riparian forests disrupts the availability of leaf litter and reduces macroinvertebrate density in stream systems (Fargen et al. 2015, McNeish et al. 2012,

2015). Furthermore, McEwan et al. (2012) found L. maackii substantially influenced the chemistry of throughfall compared to control treatments; therefore, L. maackii can potentially alter the availability of nutrient and habitat resources within aquatic systems.

This study supported the hypothesis that L. maackii would increase nutrient concentrations within throughfall (H1). Total carbon, TOC, and DIC significantly increased in the Honeysuckle collections compared to the Upper Canopy and Open collections; although, only DIC deposition was significantly influenced by the

Honeysuckle canopy. In this study the Upper Canopy collections had greater DOC concentration but DOC was reduced in Honeysuckle collections, suggesting L. maackii shrubs may intercept and absorb DOC directly from throughfall. Carbon fluxes during rain events can be driven in part by canopy exchange from leaves and stems (Hafner et al.

2005, Liu and Sheu 2003). Plant leaf traits have been linked to a plant’s ability to leach soluble compounds (Tukey 1966). Lonicera maackii does not have a thick waxy cuticle and tends to not be tough, which can make it susceptible to weathering, and thus, weaken the leaf resulting in increased nutrient concentrations in throughfall (Potter et al. 1991,

Tukey 1966). We observed a general increase in certain carbon and nitrogen species over

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the course of this study, which may be attributed to the phonological age of the leaves.

Young leaves are more hydrophobic, and thus, retain nutrients to prevent nutrient deficiency early during the growth season; however, leaves become less hydrophobic as they age and leach soluble compounds more readily (Tukey 1966).

Atmospheric deposition of nitrogen on plant canopies has been linked to an increase in nitrogen in forest and aquatic systems (Anderson et al. 2002, Lindberg et al.

1986, Lovett 1994). The estimated contribution of atmospheric deposition of nitrogen to coastal waters in eastern USA is 10% – 40% (Castro et al. 2000), suggesting atmospheric deposition of nitrogen can be an important factor in nitrogen fluxes. In this study TN and

DN concentrations were not influenced by L. maackii canopies; however, deposition was reduced under L. maackii canopies compared to the Upper Canopy and Open treatments

(H2). This finding suggests that atmospheric deposition of nitrogen may not have been important in this study, but deposition results were influenced by reduced throughfall volume under L. maackii canopies. Differences in L. maackii effects on concentration and deposition can be attributed to the fact that there was substantially less throughfall volume under L. maackii shrubs compared to the other treatments, indicating L. maackii canopies intercepted a significant portion of throughfall from the Upper canopy. In some cases this led to an increase in overall concentration, but when corrected for the volume of throughfall that actually accumulated within each treatment, there rarely was an effect on nutrient deposition. In comparison to a L. maackii throughfall study conducted by

McEwan et al. (2012) in KY, we found an overall increase in NH3-N concentration under

Honeysuckle canopies whereas they found a significant decrease in NH4-N; although we did find there was a reduction in NH3-N deposition under Honeysuckle that was a similar

182

pattern to McEwan et al. (2012). The availability of nitrogen has been linked to the external and internal concentration of the leaf and microbial activities (Jetten 2008,

Schjoerring et al. 2000). It is possible that the microbial community in conjunction with atmospheric availability may have influenced nitrogen concentration and deposition patterns; however, more research on direct links with L. maackii is needed. Location within the forest can also influence atmospheric deposition with forest edge habitats known to have higher atmospheric deposition of nitrogen compared to interior locations

(Weathers et al. 2001). Our forest throughfall rigs were set up in the interior forest, which may have impacted atmospheric deposition even though L. maackii shrubs have a large leaf crown.

Stream algal growth is linked to the availability of light and nutrients in aquatic systems with an increase in light and nutrients typically resulting in increased in algal growth (Besemer et al. 2007, Tank and Dodds 2003). Previous work has demonstrated leachate from landfills and old field succession plants have negative effects on algal communities (Cheung et al. 1993, Parks and Rice 1969). This study demonstrated that stream algal growth was differentially influenced by plant leachate, concentration, and light conditions. Algal growth was greater in the high light (L. maackii removal reach) compared to the low light (L. maackii reach) conditions across stream sites (H4).

Lonicera maackii leachate generally did not influence algal growth; however, under low light conditions L. maackii leaf and flower leachate suppressed algal growth compared to control and native leaf leachate treatments (refuting H3). Several studies on L. maackii effect on plant growth have suggested that this invasive shrub has allelopathic capabilities. Since algal growth tended to be inhibited in reduced light conditions, we

183

hypothesize that algal communities may be susceptible to allelopathic interactions from

L. maackii, but these effects can be ameliorated in high light environments. It was not within the scope of our study to identify allelopathic links between L. maackii and algal growth; therefore, further research into this avenue will be needed in the future.

In summary, our data provide compelling evidence that L. maackii shrubs alter the chemistry of throughfall that serves as a resource for aquatic systems. Lonicera maackii shrubs reduce native plant survivorship and growth, resulting in decreased plant ground cover and potentially increasing storm water runoff into streams (Collier et al. 2002,

Gorchov and Trisel 2003, Hartman and McCarthy 2004, McEwan et al. 2012). This study further provides evidence that L. maackii not only impacts the detrital basal resource in streams (McNeish et al. 2012, 2015), but the primary producer resource as well – potentially influencing both basal resource pathways that support aquatic systems.

ACKNOWLEDGEMENTS:

We would like to thank the Centerville-Washington Park District, OH and Five-

Rivers MetroParks, OH for use of Black Oak Park stream site and Taylorsville MetroPark field site. Millie Hamilton, Mary Arthur, and Jim Crutchfield from the University of

Kentucky for guidance and processing of nutrient samples. Special thanks to Hannah

O’Callaghan, Shante Eisle, Caitlin Buccheim, and Claudia Garner and all the undergraduate students at the University of Dayton that assisted with field and lab work.

Casey Hanley for use of laboratory space and equipment. This work was supported by the National Science Foundation (NSF: DEB 1352995), Sigma Xi Grants-in-Aid

Research, and in part by the University of Dayton Office for Graduate Academic Affairs through the Graduate Student Summer Fellowship Program. Any opinions, findings, and

184

conclusions or recommendations expressed are those of the author and do not necessarily reflect the views of the National Science Foundation.

185

LITERATURE CITED:

ANDERSON, D., P. GLIBERT, AND J. BURKHOLDER. 2002. Harmful algal blooms and

eutrophication: Nutrient sources, composition, and consequences. Estuaries and

Coasts 25:704–726. Springer New York. Retrieved from Estuaries and Coasts.

ARTHUR, M. A., S. R. BRAY, C. R. KUCHLE, AND R. W. MCEWAN. 2012. The influence of

the invasive shrub, Lonicera maackii, on leaf decomposition and microbial

community dynamics. Plant Ecol. 213:1571–1582. Retrieved June 13, 2013, from

Plant Ecology. .

ASSOCIATION, A. P. H. 1999. Standard methods for the examination of water and

wastewater, 20th edition. American Public Health Association, Washington, DC,

USA.

BARLING, R. D., AND I. D. MOORE. 1994. Role of buffer strips in management of

waterway pollution: A review. Environ. Manage. 18:543–558.

BAXTER, C. V., K. D. FAUSCH, AND W. CARL SAUNDERS. 2005. Tangled webs: reciprocal

flows of invertebrate prey link streams and riparian zones. Freshw. Biol. 50:201–

220. Retrieved May 25, 2013, from Freshwater Biology.

BESEMER, K., G. SINGER, R. LIMBERGER, A.-K. CHLUP, G. HOCHEDLINGER, I. HÖDL, C.

BARANYI, AND T. J. BATTIN. 2007. Biophysical controls on community succession in

stream biofilms. Appl. Environ. Microbiol. 73:4966–74. Retrieved June 12, 2013,

from Applied and environmental microbiology.

186

rez&rendertype=abstract>.

BORGMANN, K. L., AND A. D. RODEWALD. 2004. NEST PREDATION IN AN

URBANIZING LANDSCAPE : THE ROLE OF EXOTIC SHRUBS. Ecol. Appl.

14:1757–1765.

CARPENTER, S. R., N. F. CARACO, D. L. CORRELL, R. W. HOWARTH, A. N. SHARPLEY, AND

V. H. SMITH. 1998. NONPOINT POLLUTION OF SURFACE WATERS WITH

PHOSPHORUS AND NITROGEN. Ecol. Appl. 8:559–568.

CASTRO, M., C. DRISCOLL, T. JORGAN, W. REAY, W. BOYNTON, S. SEITZINGER, R.

STYLES, AND J. CABLE. 2000. Contribution of atmospheric deposition to the total

nitrogen loads of thrity-four estuaries on the Atlantic and Gulf Coast of the United

States. Pages 77–106 In R. Valigura [ed.], Atmospheric Nitrogen Deposition in

Coastal Waters, Coaster Estuarine Science Series. American Geophysical Union

Press, Washington, DC, USA.

CHAMIER, J., K. SCHACHTSCHNEIDER, D. C. LE MAITRE, P. J. ASHTON, AND B. W. VAN

WILGEN. 2012. Impacts of invasive alien plants on water quality, with particular

emphasis on South Africa. Water Sa 38:345–356. Retrieved from Water Sa. <

ISI>://WOS:000306368200019>.

CHEUNG, K. C., L. M. CHU, AND M. H. WONG. 1993. Toxic effect of landfill leachate on

microalgae. Water, Air, Soil Pollut. 69:337–349.

CIPOLLINI, D., R. STEVENSON, AND K. CIPOLLINI. 2008. Contrasting effects of

allelochemicals from two invasive plants on the performance of a nonmycorrhizal

187

plant. Int. J. Plant Sci. 169:371–375. Retrieved June 13, 2013, from International

Journal of Plant Sciences. .

COLLIER, M. H., J. L. VANKAT, AND M. R. HUGHES. 2002. Diminished plant richness and

abundance below Lonicera maackii, an invasive shrub. Am. Midl. Nat. 147:60–71.

CONROY, J. D., D. D. KANE, D. M. DOLAN, W. J. EDWARDS, M. N. CHARLTON, AND D. A.

CULVER. 2005. Temporal Trends in Lake Erie Plankton Biomass: Roles of External

Phosphorus Loading and Dreissenid Mussels. J. Great Lakes Res. 31:89–110.

Retrieved from Journal of Great Lakes Research.

D’ANGELO, E., J. CRUTCHFIELD, AND M. VANDIVIERE. 2001. Rapid, sensitive, microscale

determination of phosphate in water and soil. J. Environ. Qual. 30:2206–9.

Retrieved from Journal of environmental quality.

DAVIS, J. M., A. D. ROSEMOND, AND G. E. SMALL. 2011. Increasing donor ecosystem

productivity decreases terrestrial consumer reliance on a stream resource subsidy.

Oecologia 167:821–834. Retrieved from Oecologia. <

ISI>://WOS:000295984800021>.

DOLAN, D. M. 1993. Point Source Loadings of Phosphorus to Lake Erie: 1986–1990. J.

Great Lakes Res. 19:212–223. Elsevier. Retrieved June 13, 2013, from Journal of

Great Lakes Research.

DORNING, M., AND D. CIPOLLINI. 2005. Leaf and root extracts of the invasive shrub,

188

Lonicera maackii, inhibit seed germination of three herbs with no autotoxic effects.

Plant Ecol. 184:287–296. Retrieved June 13, 2013, from Plant Ecology.

FARGEN, C., S. M. EMERY, AND M. M. CARREIRO. 2015. Influence of Lonicera maackii

Invasion on Leaf Litter Decomposition and Macroinvertebrate Communities in an

Urban Stream. Nat. Areas J. 35:392–403.

GORCHOV, D. L., AND D. E. TRISEL. 2003. Competitive effects of the invasive shrub,

Lonicera maackii (Rupr.) Herder (Caprifoliaceae), on the growth and survival of

native tree seedlings. Plant Ecol. 166:13–24.

GOULD, A. M. A., AND D. L. GORCHOV. 2000. Effects of the Exotic Invasive Shrub

Lonicera maackii on the Survival and Fecundity of Three Species of Native

Annuals. Am. Midl. Nat. 144:36–50.

GREENE, S. L. 2014. A roadmap for riparian invasion research. River Res. Appl. 30:663–

669.

GREGORY, S. V, F. J. SWANSON, W. A. MCKEE, AND K. W. CUMMINS. 1991. An ecosystem

perspective of riparian zones. Bioscience 41:540–551.

GREIG, H. S., P. KRATINA, P. L. THOMPSON, W. J. PALEN, J. S. RICHARDSON, AND J. B.

SHURIN. 2012. Warming, eutrophication, and predator loss amplify subsidies

between aquatic and terrestrial ecosystems. Glob. Chang. Biol. 18:504–514.

Retrieved from Global Change Biology. <://WOS:000299042500010>.

HAFNER, S. D., P. M. GROFFMAN, AND M. J. MITCHELL. 2005. Leaching of dissolved

189

organic carbon, dissolved organic nitrogen, and other solutes from coarse woody

debris and litter in a mixed forest in New York State. Biogeochemistry 74:257–282.

HARTMAN, K. M., AND B. C. MCCARTHY. 2004. Restoration of a forest understory after

the removal of an invasive shrub, Amur Honeysuckle (Lonicera maackii). Restor.

Ecol. 12:154–165. Retrieved from Restoration Ecology.

synergy.com/links/doi/10.1111%2Fj.1061-2971.2004.00368.x>.

HEISLER, J., P. M. GLIBERT, J. M. BURKHOLDER, D. M. ANDERSON, W. COCHLAN, W. C.

DENNISON, Q. DORTCH, C. J. GOBLER, C. A. HEIL, E. HUMPHRIES, A. LEWITUS, R.

MAGNIEN, H. G. MARSHALL, K. SELLNER, D. A. STOCKWELL, D. K. STOECKER, AND

M. SUDDLESON. 2008. Eutrophication and harmful algal blooms: A scientific

consensus. Harmful Algae 8:3–13.

HUDNELL, H. K. 2010. The state of US freshwater harmful algal blooms assessments,

policy and legislation. Toxicon 55:1024–1034. Retrieved from Toxicon. <

ISI>://WOS:000276436400015>.

JETTEN, M. S. M. 2008. The microbial nitrogen cycle. Environ. Microbiol. 10:2903–2909.

JOHNSON, P. T. J., A. R. TOWNSEND, C. C. CLEVELAND, P. M. GLIBERT, R. W. HOWARTH,

V. J. MCKENZIE, E. REJMANKOVA, AND M. H. WARD. 2010. Linking environmental

nutrient enrichment and disease emergence in humans and wildlife. Ecol. Appl.

20:16–29. Ecological Society of America. Retrieved from Ecological Applications.

LINDBERG, S., G. LOVETT, D. RICHTER, AND J. DW. 1986. Atmospheric deposition and

canopy interactions of major ions in a forest. Science (80-. ). 231:141–145.

190

LIU, C. P., AND B. H. SHEU. 2003. Dissolved organic carbon in precipitation, throughfall,

stemflow, soil solution, and stream water at the Guandaushi subtropical forest in

Taiwan. For. Ecol. Manage. 172:315–325.

LOVETT, G. 1994. Atmospheric deposition of nutrients and pollutants in North America:

an ecological perspective. Ecol. Appl. 4:629–650.

LOWRANCE, R., R. TODD, J. FAIL, O. HENDRICKSON, L. ASMUSSEN, R. LOWRANCE, R.

TODD, J. FAIL, O. HENDRICKSON, R. LEONARD, AND L. ASMUSSEN. 1984. Riparian

filters in agricultural watersheds. Bioscience 34:374–377.

MCEWAN, R. W., M. A. ARTHUR, AND S. E. ALVERSON. 2012. Throughfall chemistry and

soil nutrient effects of the invasive shrub Lonicera maackii in deciduous forests.

Am. Midl. Nat. 168:43–55.

MCEWAN, R. W., L. G. ARTHUR-PARATLEY, L. K. RIESKE, AND M. A. ARTHUR. 2010. A

multi-assay comparison of seed germination inhibition by Lonicera maackii and co-

occurring native shrubs. Flora - Morphol. Distrib. Funct. Ecol. Plants 205:475–483.

Elsevier. Retrieved June 13, 2013, from Flora - Morphology, Distribution,

Functional Ecology of Plants.

MCEWAN, R. W., M. K. BIRCHFIELD, A. SCHOERGENDORFER, AND M. A. ARTHUR. 2009.

Leaf phenology and freeze tolerance of the invasive shrub Amur honeysuckle and

potential native competitors. J. Torrey Bot. Soc. 136:212–220.

MCEWAN, R. W., L. K. RIESKE, AND M. A. ARTHUR. 2009. Potential interactions between

invasive woody shrubs and the gypsy moth (Lymantria dispar), an invasive insect

191

herbivore. Biol. Invasions 11:1053–1058. Retrieved May 23, 2013, from Biological

Invasions. .

MCKINNEY, A. M., AND K. GOODELL. 2010. Shading by invasive shrub reduces seed

production and pollinator services in a native herb. Biol. Invasions 12:2751–2763.

Retrieved June 13, 2013, from Biological Invasions.

MCNEISH, R. E., M. E. BENBOW, AND R. W. MCEWAN. 2012. Riparian forest invasion by

a terrestrial shrub (Lonicera maackii) impacts aquatic biota and organic matter

processing in headwater streams. Biol. Invasions 14:1881–1893. Retrieved June 13,

2013, from Biological Invasions.

0199-8>.

MCNEISH, R. E., E. M. MOORE, M. E. BENBOW, AND R. W. MCEWAN. 2015. Removal of

the invasive shrub, Lonicera maackii, from riparian forests influences headwater

stream biota and ecosystem function. River Res. Appl. 31:1131–1139.

MUSSON, J., AND W. J. MITSCH. 2003. The effects of the invasive shrub Lonicera maacki

on species richness and soil moisture in the bottomland hardwood forest at the

ORWRP.

PARKS, J. M., AND E. L. RICE. 1969. Effects of certain plants of old-field succession on

the growth of Blue- Green Algae. Bu 96:345–360.

POHLERT, A. 2015. PMCMR: Calculate multiple comparisons of mean rank sums. 4.0.

Retrieved from 4.0.

project.org/web/packages/PMCMR/PMCMR.pdf>.

192

POTTER, C. S., H. RAGSDALE, AND W. SWANK. 1991. Atmospheric deposition and foliar

leaching in a regenerating southern Appalachian forest canopy. J. Ecol. 79:97–115.

POULETTE, M. M., AND M. A. ARTHUR. 2012. The impact of the invasive shrub Lonicera

maackii on the decomposition dynamics of a native plant community. Ecol. Appl.

22:412–424.

QUINN, G., AND M. KEOUGH. 2004. Experimental design and data analysis for biologists.

Cambridge University Press, Cambridge, UK.

RICHARDSON, D. M., P. M. HOLMES, K. J. ESLER, S. M. GALATOWITSCH, J. C.

STROMBERG, S. P. KIRKMAN, P. PYŠEK, AND R. J. HOBBS. 2007. Riparian vegetation:

degradation, alien plant invasions, and restoration prospects. Divers. Distrib.

13:126–139. Blackwell Publishing Ltd. Retrieved from Diversity and Distributions.

RODEWALD, A. D. 2009. Urban-associated habitat alteration promotes brood parasitism of

Acadian Flycatchers. J. F. Ornithol. 80:234–241. Retrieved June 13, 2013, from

Journal of Field Ornithology.

9263.2009.00226.x>.

SCHJOERRING, J. K., S. HUSTED, G. MÄCK, K. H. NIELSEN, J. FINNEMANN, AND M.

MATTSSON. 2000. Physiological regulation of plant-atmosphere ammonia exchange.

Plant Soil 221:95–102.

SCHNEIDER, W. J. 1957. Relation of geology to stream flow in the Upper Little Miami

Basin. Ohio J. Sci. 57:11–14. Retrieved from The Ohio Journal of Science.

193

SMITH, V. 2003. Eutrophication of freshwater and coastal marine ecosystems a global

problem. Environ. Sci. Pollut. Res. 10:126–139.

SOKAL, R., AND F. ROHLF. 1981. Biometery: the Principles and Practice of Statistics in

Biological Research, 2nd edition. W.H. Freeman and Company, NewYork.

STEINMAN, A. D., AND G. A. LAMBERTI. 2007. Biomass and pigments of benthic algae.

Pages 357–379 In R. F. Hauer and G. A. Lamberti [eds.], Methods in stream

ecology, 2nd edition. Academic Press, San Diego.

TANK, J. L., M. J. BERNOT, AND E. J. ROSI-MARSHALL. 2007. Nitrogen limitation and

uptake. Pages 213–238 In R. F. Hauer and G. A. Lamberti [eds.], Methods in Stream

Ecology, 2nd edition. Academic Press, New York, NY.

TANK, J. L., AND W. K. DODDS. 2003. Nutrient limitation of epilithic and epixylic

biofilms in ten North American streams. Freshw. Biol. 48:1031–1049. Retrieved

from Freshwater Biology.

2427.2003.01067.x>.

TANK, J. L., E. J. ROSI-MARSHALL, N. A. GRIFFITHS, S. A. ENTREKIN, AND M. L. STEPHEN.

2010. A review of allochthonous organic matter dynamics and metabolism in

streams. J. North Am. Benthol. Soc. 29:118–146. Retrieved May 27, 2013, from

Journal of the North American Benthological Society.

TRAMMELL, T. L. E., H. A. RALSTON, S. A. SCROGGINS, AND M. M. CARREIRO. 2012.

Foliar production and decomposition rates in urban forests invaded by the exotic

invasive shrub, Lonicera maackii. Biol. Invasions 14:529–545.

194

TUKEY, H. 1966. Leaching of metabolites from above-ground plant parts and its

implications. Bull. Torrey Bot. Club 93:385–401.

VANNOTE, R. L., G. W. MINSHALL, K. W. CUMMINS, J. R. SEDELL, AND C. E. CUSHING.

1980. The River Continuum Concept. Can. J. Fish. Aquat. Sci. 37:130–137.

Retrieved from Canadian Journal of Fisheries and Aquatic Sciences.

WALSH, C. J., A. H. ROY, J. W. FEMINELLA, P. D. COTTINGHAM, P. M. GROFFMAN, AND R.

P. MORGAN. 2005. The urban stream syndrome: current knowledge and the search

for a cure. J. North Am. Benthol. Soc. 24:706–723.

WEATHERS, K. C., M. L. CADENASSO, AND S. T. A. PICKETT. 2001. Forest edges and

pollutant concentrations: potential synergisms between fragmentation, forest

canopies, and the atmosphere. Conserv. Biol. 15:1506–1514.

WHITE, R. J., M. M. CARREIRO, AND W. C. ZIPPERER. 2014. Woody plant communities

along urban, suburban, and rural streams in Louisville, Kentucky, USA. Urban

Ecosyst. 17:1061–1094. Retrieved from Urban Ecosystems.

ZAR, J. 1999. Biostatistical Analysis, 4th edition. Prentice-Hall, New Jersey.

ZHANG, J. 1994. Atmospheric wet deposition of nutrient elements - Correlation with

harmful biological blooms in Northwest Pacific Coastal Zones. Ambio 23:464–468.

Retrieved from Ambio.

BD7C-4444-840B-58877DB4DF73>.

195

TABLES:

Table 4.1: Percent difference of nutrient concentration and deposition for honeysuckle and upper canopy throughfall relative to the open treatment.

Concentration

Month Treatment TC TOC DOC DIC TN DN NO2 NO3 NH4 P SRP

June Honeysuckle 135.77 240.15 80.52 0573.62 -05.46 -06.84 282.66 000.00 0029.39 0049.97 00050.46

Upper Canopy 008.69 003.54 49.19 0021.45 -05.35 -11.64 329.62 137.77 -0035.43 -0070.07 -00085.39

July Honeysuckle 212.87 134.25 -06.60 1288.94 38.36 00.00 069.76 036.84 0061.42 0006.86 -00000.93

Upper Canopy 009.65 020.67 55.32 -0008.31 -01.81 -29.17 013.95 -035.08 -0021.42 0012.07 00001.79

August Honeysuckle 212.83 134.26 -06.62 1289.21 35.48 13.26 014.03 -048.91 1756.25 4853.21 12543.40

Upper Canopy 003.26 015.26 63.54 -0008.27 06.09 00.47 029.82 -008.69 1257.14 3883.33 03256.39

November Honeysuckle ------025.49 037.50 -0087.75 0075.16 00031.47

Upper Canopy ------013.72 009.37 -0094.04 0023.96 -00003.43

Deposition

Month Treatment TC TOC DOC DIC TN DN NO2 NO3 NH4 P SRP

June Honeysuckle -26.99 -18.79 -57.33 077.49 -59.49 -73.06 -54.17 -046.47 -0033.31 -0022.06 -0026.61

196

Upper Canopy -31.34 -35.23 46.81 -051.19 -33.53 -11.85 12.74 124.93 -0031.49 -0080.54 -0075.97 July Honeysuckle 69.74 54.79 -04.38 275.16 -19.50 -37.91 00.52 -019.22 -0009.85 -0033.21 -0024.46

Upper Canopy -28.52 08.25 19.52 -063.53 -30.69 -43.73 -04.75 -042.61 -0036.79 0001.50 -0001.54

August Honeysuckle 39.59 66.03 -68.17 679.09 -37.54 -40.85 -36.69 -071.36 0888.48 2500.28 6461.35

Upper Canopy -24.12 -00.89 -07.57 -051.44 -15.14 -03.66 -01.48 -027.75 1076.89 1984.95 5206.91

November Honeysuckle ------12.63 -002.68 -0091.08 0002.27 0035.19

Upper Canopy ------28.25 -008.89 -0095.54 -0007.11 -0020.17

197

FIGURE LEGENDS:

Figure 4.1: Taylorsville MetroPark throughfall field site with invaded Lonicera maackii

(a), and Black Oak Park and Fecher Park stream sites with riparian invasion of Lonicera maackii (b, d), and plant leachate nutrient diffusing substrates anchored within Black Oak

Park stream site (c).

Figure 4.2: Mean (±SEM) throughfall and rain water volume and pH collected from open and under upper Lonicera maackii canopies.

Figure 4.3: Mean (±SEM) total carbon and organic carbon concentration and deposition from throughfall and rain water collected from open and under upper Lonicera maackii canopies.

Figure 4.4: Mean (±SEM) dissolved organic and dissolved inorganic carbon concentration and deposition from throughfall and rain water collected from open and under upper Lonicera maackii canopies.

198

Figure 4.5: Spring mean (±SEM) algal grown (standing stock chlorophyll a) grown on nutrient diffusing substrates made with Lonicera maackii flowers (LM Flowers) and leaves (LM Leaves), Acer negundo leaves (BE Leaves), and control treatments during

2015.

Figure 4.6: Autumn mean (±SEM) algal grown (standing stock chlorophyll a) grown on nutrient diffusing substrates made with Lonicera maackii berries (LM Berries) and leaves

(LM Leaves), Acer negundo leaves (BE Leaves), and control treatments during 2014.

199

FIGURES:

Figure 4.1:

200

Figure 4.2:

201

Figure 4.3:

202

Figure 4.4:

203

Figure 4.5:

204

Figure 4.6:

205

SUPPLEMENTAL MATERIALS:

Supplemental Table 4.1: Two-way ANOVA and Friedmans statistical results for throughfall nutrient concentration with treatment and time as main effects.

Nutrient Factor df F-Stat P Value Χ2 P Value

Total Carbon - 2 - - 6.00 0.0497

Total Organic Carbon Treatment 2 062.13 < 0.0001 6.00 0.0497

Time 2 027.52 < 0.0001 - -

Interaction 4 000.93 0.3384 - -

Dissolved Organic Carbon Treatment 2 001.43 0.2335 - -

Time 2 000.80 0.3730 - -

Interaction 4 004.07 0.0466 - -

206

Dissolved Inorganic Carbon Treatment 2 091.21 < 0.0001 4.67 0.0969

Time 2 000.22 0.6396 - -

Interaction 4 001.88 0.1741 - -

Total Nitrogen - 2 - - 0.67 0.7165

Dissolved Nitrogen - 2 - - 2.36 0.3067

Nitrite Treatment 2 007.94 0.0056 3.50 0.1738

Time 3 057.37 < 0.0001 - -

Interaction 6 000.45 0.5029 - -

Nitrate Treatment 2 000.23 0.6260 0.14 0.9355

Time 3 124.02 < 0.0001 - -

Interaction 6 000.11 0.7319 - -

Ammonia - 2 - - 4.50 0.1054

207

Phosphorous - 2 - - 4.50 0.1054

Soluble Reactive Phosphorous - 2 - - 1.50 0.4724

208

Supplemental Table 4.2: Two-way ANOVA and Friedmans statistical results for throughfall nutrient deposition with treatment and time as main effects.

Nutrient Factor df F-Stat P Value Χ2 P Value

Total Carbon Treatment 2 01.96 0.1733 4.67 0.0969

Time 2 12.50 0.0016 - -

Interaction 4 02.02 0.1675 - -

Total Organic Carbon Treatment 2 2.79 0.1065 2.67 0.2636

Time 2 17.77 0.0002 - -

Interaction 4 01.86 0.1843 - -

Dissolved Organic Carbon Treatment 2 06.66 0.0158 4.67 0.0969

Time 2 24.56 < 0.0001 - -

209

Interaction 4 05.24 0.0304 - -

Dissolved Inorganic Carbon Treatment 2 - - 6.00 0.0497

Total Nitrogen Treatment 2 10.94 0.0036 - -

Time 2 17.36 0.0005 - -

Interaction 4 03.92 0.0624 - -

Dissolved Nitrogen Treatment 2 14.09 0.0010 4.67 0.0969

Time 2 20.53 0.0001 - -

Interaction 4 05.58 0.0270 - -

Nitrite - 2 - - 1.50 0.4724

Nitrate - 2 - - 3.50 0.1738

Ammonia - 2 - - 1.50 0.4724

Phosphorous - 2 - - 0.50 0.7788

210

Soluble Reactive Phosphorous - 2 - - 1.50 0.4724

211

Supplemental Table 4.3: ANOVA and Kruskal-Wallis statistical results for stream algal growth response when exposed to control, honeysuckle leaves and flowers, and Acer negundo leaves at 100% and 25% concentrations. The full model represented a 3-way

ANOVA with leachate, concentration, and stream reach as main effects. Two-way ANOVA results are represented for removal and honeysuckle stream reach with leachate and concentration as main effects.

Black Oak Park Fecher Park

Reach Factor df F-Stat P Value Χ2 P Value F-Stat P Value Χ2 P Value

Removal Algal Leachate 3 16.731 0.0103 01.19 0.3219 - -

Growth 03.24 0.0261

Concentration 1 05.48 0.0216 - - 02.07 0.1559 - -

Leachate:Con 2 02.17 0.1206 - - 02.09 0.1345 - -

centration

Biomass Leachate 3 00.77 0.5129 10.54 0.1037 01.59 0.2029 14.68 0.0229

Concentration 1 02.30 0.1331 - - 13.63 0.0005 - -

212

Leachate:Con 2 02.23 0.1137 - - 00.67 0.5163 - -

centration

P:B Leachate 3 - - 07.66 0.2638 01.12 0.3487 11.69 0.0691

Concentration 1 - - - - 06.18 0.0163 - -

Leachate:Con 2 - - - - 01.23 0.3001 - -

centration

Honeysuckle Algal Leachate 3 13.64 0.0339 01.44 0.2405 01.12 0.0550

Growth 05.95 0.0008

Concentration 1 00.11 0.7391 - - 04.88 0.0317 - -

Leachate:Con 2 - - 01.85 0.1672 - -

centration 00.04 0.9595

Biomass Leachate 3 01.73 0.1638 21.12 0.0017 02.98 0.0396 14.92 0.0208

Concentration 1 00.01 0.9384 - - 05.69 0.0208 - -

Leachate:Con 2 - - 00.73 0.4843 - -

centration 00.34 0.7099

213

P:B Leachate 3 05.08 0.0024 12.50 0.0517 - - 11.42 0.0762

Concentration 1 00.01 0.9316 ------

Leachate:Con 2 00.21 0.8111 ------

centration

214

Supplemental Table 4.4: ANOVA and Kruskal-Wallis statistical results for stream algal growth response when exposed to control, honeysuckle leaves and berries, and Acer negundo leaves at 100% and 25% concentrations. The full model represented a 3-way

ANOVA with leachate, concentration, and stream reach as main effects. Two-way ANOVA results are represented for removal and honeysuckle stream reach with leachate and concentration as main effects.

Black Oak Park Fecher Park

Model Factor df F-Stat P Value F-Stat P Value Χ2 P Value

Removal Leachate 3 01.7394 0.1740 00.1996 0.8961 06.00 0.4232

Concentration 1 00.1369 0.7133 00.8701 0.3560 - -

Leachate:Concentration 2 00.8123 0.4508 00.0777 0.9254 - -

Honeysuckle Leachate 3 01.1577 0.3393 01.6076 0.2019 12.61 0.0495

Concentration 1 00.0165 0.8985 00.6777 0.4150 - -

215

Leachate:Concentration 2 00.5653 0.5732 01.9487 0.1551 - -

216

SUPPLEMENTAL FIGURES:

Supplemental Figure 4.1: Mean (±SEM) total and dissolved nitrogen concentration and deposition from throughfall and rain water collected from open and under upper Lonicera maackii canopies.

Supplemental Figure 4.2: Mean (±SEM) nitrite, nitrate, and ammonia concentration and deposition from throughfall and rain water collected from open and under upper Lonicera maackii canopies.

Supplemental Figure 4.3: Mean (±SEM) total orthophosphate (phosphorous) and soluble reactive phosphorous (SRP) from throughfall and rain water collected from open and under upper Lonicera maackii canopies.

Supplemental Figure 4.4: Spring mean (±SEM) algal biomass (ash-free-dry-mass) grown on nutrient diffusing substrates made with Lonicera maackii flowers (LM Flowers) and leaves (LM Leaves), Acer negundo leaves (BE Leaves), and control treatments during

2015.

217

Supplemental Figure 4.5: Spring mean (±SEM) algal P:B (standing stock chlorophyll a : biomass) response to nutrient diffusing substrates made with Lonicera maackii flowers

(LM Flowers) and leaves (LM Leaves), Acer negundo leaves (BE Leaves), and control treatments during 2015.

218

Supplemental Figure 4.1:

219

Supplemental Figure 4.2:

220

Supplemental Figure 4.3:

221

Supplemental Figure 4.4:

222

Supplemental Figure 4.5:

223