The spatio -temporal dynamics of fish and invertebrate communities within coastal wetland systems.

Rose Esther Stuart

A thesis submitted for the degree of Doctor of Philosophy in Zoology at the University of Otago Dunedin, New Zealand 2021

For my daughter Alice, Let your curiosity lead the way.

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Acknowledgements

There are many people who have helped to make this thesis a reality, I have thoroughly enjoyed the process of research and writing but could not have done it without such wonderful support.

First of all, tremendous thanks to my supervisors Gerry Closs and Travis Ingram, your guidance and advice has been invaluable. I have enjoyed all our conversations about fish, even when they were a little off topic.

Many thanks to the Waiau Trust and in particular Jan Riddell, I have had such a great time working in the wonderful wetland you have spent so much time and effort into creating, thank you for the opportunity to explore and study such a beautiful place.

I have been fortunate to have many people help out with my fieldwork, your company and conversation made the long car trips much more enjoyable and your assistance in the field made some of the more tedious tasks much more fun.

I have also had some great people help me with lab work. Thanks to Nicky McHugh for sharing your advice and expertise, your guidance definitely helped avoid some time-consuming mistakes; and to Jerusha Bennet for your help with processing otoliths, I really appreciate it.

Lastly, I need to thank my family and friends who provided so much encouragement and feedback, and who listened to my endless interesting facts about fish and wetlands and invertebrates and…

I am so grateful for the support from my husband Mike, your belief in me made this feel much more achievable, it would have been much harder without you by my side.

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Abstract

This thesis investigates the spatial and temporal dynamics of fish and invertebrate populations within coastal wetlands in Southland, New Zealand. Each chapter focusses on a different aspect of the community and its interaction with the environment. Initially fish in four systems were surveyed over two years, where great variation in abundance and diversity of species was observed, particularly in relation to connectivity to other waterways (Chapter 2). During the survey period, significant dewatering events occurred; the duration of these low water events had a significant impact on how rapidly the community recovered. Following this was an in-depth investigation of the invertebrate community dynamics over a spring to autumn sampling period, focussed in one of the initial wetland areas, a constructed wetland system with many small interconnected ponds (Chapter 3). A progression of dominant microcrustaceans were observed as was a succession of larger more predatory taxa. Spatial variation was also observed, related to hydrological connectivity and nutrient levels. An investigation into īnanga population dynamics was also conducted, as they are a key species in this system and there is a notable gap in knowledge on their habits within small ponds (Chapter 4). This found spring-hatching fish with a faster growth rate than those that hatch later in the year. Diet preferences indicate both benthic and pelagic foraging, and the presence and consumption of microplastic filaments. The study culminates in an in-depth investigation of the recovery of both fish and invertebrate communities after a drought within one system, tying together observations from the other chapters (Chapter 5). This highlighted the importance of spatial heterogeneity and connectivity for rapid invertebrate recovery, as well as the value of deep channels that are more resistant to drying in supporting fish recolonization. Overall, this research has provided insights into communities of fish and invertebrates, generating new information relevant to the specific systems and species as well as the wider body of knowledge on wetlands.

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Table of Contents ACKNOWLEDGEMENTS ...... II ABSTRACT ...... III LIST OF FIGURES ...... VI LIST OF TABLES ...... VIII CHAPTER 1. INTRODUCTION ...... 1

1.1 WETLANDS ...... 1 1.2 NEW ZEALAND WETLANDS ...... 2 1.3 WETLAND CONSERVATION ...... 4 1.4 THESIS OUTLINE ...... 5 CHAPTER 2. SPATIAL AND TEMPORAL VARIABILITY OF FISH COMMUNITIES WITHIN SOUTHLAND COASTAL WETLAND SYSTEMS ...... 9

2.1 ABSTRACT ...... 9 2.2 INTRODUCTION ...... 9 2.3 METHODS ...... 11 2.3.1 Study sites ...... 11 2.3.2 Sampling ...... 14 2.3.3 Data processing and statistics ...... 14 2.4 RESULTS ...... 15 2.4.1 Community composition and population variability ...... 15 2.4.2 Effects of hydrology ...... 18 2.5 DISCUSSION ...... 20 2.5.1 Differences between systems ...... 20 2.5.2 Impacts of hydrology ...... 21 2.5.3 Management implications ...... 22 2.6 CONCLUSIONS ...... 23 CHAPTER 3. DRIVERS OF INVERTEBRATE ABUNDANCE AND COMMUNITY COMPOSITION IN A SMALL POND COMPLEX 24

3.1 ABSTRACT ...... 24 3.2 INTRODUCTION ...... 24 3.3 METHODS ...... 26 3.3.1 Study site ...... 26 3.3.2 Environmental parameters ...... 27 3.3.3 Invertebrate sampling ...... 29 3.3.4 Data processing and statistics ...... 29 3.4 RESULTS ...... 31 3.4.1 Environmental factors ...... 31 3.4.2 Invertebrate community ...... 34 3.4.3 Relationship between environment and invertebrates ...... 37 3.5 DISCUSSION ...... 39 3.5.1 Environmental factors ...... 39

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3.5.2 Invertebrate community ...... 40 3.6 CONCLUSIONS ...... 42 CHAPTER 4. GROWTH AND DIET OF ĪNANGA WITHIN A COASTAL POND SYSTEM ...... 43

4.1 ABSTRACT ...... 43 4.2 INTRODUCTION ...... 43 4.3 METHODS ...... 45 4.3.1 Study site ...... 45 4.3.2 Īnanga sampling ...... 47 4.3.3 Environmental parameters ...... 48 4.3.4 Invertebrate sampling ...... 49 4.3.5 Data processing and statistics ...... 50 4.4 RESULTS ...... 51 4.4.1 Environmental patterns ...... 51 4.4.2 As previously described in Chapter 3, significant differences in the environmental conditions between sites were observed (F(5,46)=2.58, p=0.004), this difference appears to be primarily driven by water nutrient levels (NO2/NO3, ammonia, TN and TP). The conditions at Site 1 fall differently to the majority of the other sites in the RDA plot, indicating its influence on the spatial variation (Fig. 4.3). Collectively the environmental conditions showed no detectable variation over time (F(7,46)=1.4, p=0.13), but individually both water level (F(7,39)=8.01, p<0.001) and average daytime temperature (F(7,39)=42.87, p<0.001, ) did show a significant change. 51 4.4.3 Population dynamics ...... 52 4.4.4 Diet ...... 56 4.5 DISCUSSION ...... 59 4.5.1 Age and growth ...... 59 4.5.2 Diet ...... 60 4.6 CONCLUSIONS ...... 62 CHAPTER 5. RECOLONISATION OF A FISH AND INVERTEBRATE COMMUNITY IN A WETLAND FOLLOWING A DROUGHT: THE IMPORTANCE OF DEEP WATER REFUGIA...... 63

5.1 ABSTRACT ...... 63 5.2 INTRODUCTION ...... 63 5.3 METHODS ...... 65 5.3.1 Study site and design ...... 65 5.3.2 Fish populations ...... 68 5.3.3 Macroinvertebrate samples ...... 68 5.3.4 Zooplankton samples ...... 68 5.3.5 Data processing and statistics ...... 69 5.4 RESULTS ...... 69 5.5 DISCUSSION ...... 76 5.6 CONCLUSIONS ...... 78 CHAPTER 6. SUMMARY ...... 80

6.1 SUMMARY OF KEY FINDINGS ...... 80 6.2 SYNTHESIS ACROSS CHAPTERS ...... 81 6.3 FUTURE RESEARCH QUESTIONS ...... 83 6.4 MANAGEMENT RECOMMENDATIONS ...... 84 6.5 CONCLUSION ...... 85 REFERENCES ...... 86 v

List of Figures

FIG 2.1: LOCATION OF STUDY SITES WITHIN THE FOUR SELECTED SYSTEMS ALONG THE SOUTHLAND COAST. WITHIN WAITUNA LAGOON AND THE WAIAU PONDS (DESIGNATED AS THE MORE COMPLEX SITES) FOUR SAMPLING SITES WERE SELECTED, IN LAKE GEORGE AND BIG LAGOON (THE SIMPLER SITES) TWO SITES WERE SELECTED...... 13 FIG. 2.2: TOTAL CATCH (X-AXIS) COMPOSITION FOR ALL SITES (Y-AXIS) FISHED IN THE WAITUNA AND WAIAU SYSTEMS ACROSS 16 SAMPLING DATES...... ERROR! BOOKMARK NOT DEFINED. FIG 2.3: RELATIONSHIP BETWEEN NUMBER OF FISH CAUGHT AND WATER LEVEL BEFORE (MAY 2017 TO APRIL 2018) AND AFTER THE OPENING OF THE LAGOON (JUNE 2018 TO APRIL 2019). BEFORE OPENING P=0.09, R=-0.30, N=32, AFTER OPENING P=0.005, R=0.48, N=32...... 20 FIG. 3.1: LOCATION OF STUDY SITES WITHIN THE WAIAU POND SYSTEM ON THE SOUTHLAND COAST. LOCATION OF THE DEEP CHANNELS THAT RUN THROUGH THE SYSTEM ARE HIGHLIGHTED...... 27 FIG. 3.2: REDUNDANCY ANALYSIS PLOTS DISPLAYING THE ENVIRONMENTAL CONDITION POLYGONS FOR EACH OF THE SITES AND THE DIRECTION OF INFLUENCE OF THE INDIVIDUAL ENVIRONMENTAL PARAMETERS. P=0.001, N=9 PER SITE ...... 31 FIG. 3.3: (A) WATER LEVEL MEASURED ACROSS THE EIGHT DATES SAMPLED, P<0.001, N=6 SITES PER DATE. (B) AVERAGE WATER TEMPERATURE DURING DAYLIGHT HOURS FOR EACH OF THE EIGHT DATES SAMPLED, P<0.001, N=6 SITES PER DATE ...... 32 FIG. 3.4: PATTERN IN PC1 VALUE (CALCULATED FROM EIGHT ENVIRONMENTAL PARAMETERS) ACROSS THE SIX STUDY SITES (P=0.001, N=9 PER SITE) ...... 33 FIG. 3.5: ABUNDANCES OF FOUR INVERTEBRATE TAXA FOR EACH OF THE 6 SITES: DAPHNIA (P=0.022), CYCLOPOID COPEPODS (P=0.14), BOATMAN (P=0.047), AND HELICAL SNAILS (P=0.014). N=9 PER SITE...... ERROR! BOOKMARK NOT DEFINED. FIG. 3.6: REDUNDANCY ANALYSIS PLOT DISPLAYING INVERTEBRATE COMMUNITY CENTROIDS FOR EACH OF THE DATES SAMPLED, WITH ARROWS ILLUSTRATING THE DIRECTION OF CHANGE OVER TIME AND NAMES INDICATING THE DIRECTION OF INFLUENCE OF THE INDIVIDUAL TAXA. P=0.001, N=6 PER DATE ...... 36 FIG. 3.7: ABUNDANCES FOR THREE INVERTEBRATE TAXA OVER THE 9 DATES SAMPLED, THE ORDER OF GRAPHS ILLUSTRATES THE ORDER OF SUCCESSION OF THESE TAXA, AS ONE PEAK IN ABUNDANCE FOLLOWS THE NEXT. (A) CHIRONOMID (P=0.002), (B) BACKSWIMMER (P=0.03) AND (C) WATER BOATMAN (P=0.03). N=6 PER DATE...... 37 FIG. 3.8: (A) INVERTEBRATE ABUNDANCES EXCLUDING DAPHNIA FOR EACH OF THE 6 SITES (P=0.004). (B) NITRATE/NITRITE LEVELS ACROSS THE 6 SITES (P<0.0001). (C) THE RELATIONSHIP BETWEEN INVERTEBRATE ABUNDANCE AND NITRATE/NITRITE LEVELS (P=0.002, R=-0.43)...... 38 FIG. 3.9: RELATIONSHIP BETWEEN DAPHNIA ABUNDANCE AND PC1 (FIRST AXIS OF VARIATION IN A PCA RUN ON ENVIRONMENTAL PARAMETERS WITHIN THE POND SYSTEM)...... 38 FIG. 3.10: RELATIONSHIP BETWEEN DAPHNIA AND AVERAGE DAYLIGHT TEMPERATURE ...... 39 FIG. 4.1: LOCATION OF STUDY SITES WITHIN THE WAIAU POND SYSTEM ON THE SOUTHLAND COAST, HIGHLIGHTING THE LOCATION OF THE DEEP CHANNELS THAT RUN THROUGH THE SYSTEM ...... 46 FIG 4.2: (A) IMAGE OF PREPARED OTOLITH INDICATING THE INNER (WHITE ARROW) AND OUTER (BLACK ARROW) INCREMENTS IDENTIFIED FOR COUNTING, FROM FISH 8-1E. (B) IMAGE OF CLEAR INNER INCREMENTS INDICATING THE HATCH POINT (RED ARROW) FROM FISH 23-1A ...... 48 FIG. 4.3: OTOLITH GROWTH TRAJECTORIES (OTOLITH WIDTH OVER TIME) FOR EACH OF THE 17 ĪNANGA, AND FOR ALL FISH TOGETHER, THE RED LINE INDICATES THE FITTED SLOPE FOR EACH INDIVIDUAL AND IN THE FINAL PANEL THE OVERALL MODEL SLOPE FOR ALL FISH. ... 54 FIG. 4.4: OTOLITH GROWTH COEFFICIENT PLOTTED AGAINST DATE AT THE OTOLITH CORE (APPROXIMATION OF METAMORPHIC DATE). N=17, P=0.019, R=0.56 ...... 55 FIG. 4.5: AVERAGE DAYTIME WATER TEMPERATURE PLOTTED AGAINST DATE. N=188 DAYS ...... 55 FIG. 4.6: REDUNDANCY ANALYSIS PLOTS DISPLAYING THE ENVIRONMENTAL CONDITION POLYGONS FOR EACH OF THE SITES AND THE DIRECTION OF INFLUENCE OF THE INDIVIDUAL ENVIRONMENTAL PARAMETERS, HIGHLIGHTING SITE 1 WHERE ALL BUT ONE ĪNANGA WERE CAUGHT. P=0.001, N=9 PER SITE ...... ERROR! BOOKMARK NOT DEFINED.

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FIG. 4.7: COMPONENTS OF THE POND INVERTEBRATE COMMUNITY COMPARED TO THE GUT CONTENT FROM ĪNANGA CAUGHT IN THE POND OVER THE FOUR SAMPLING DATES WHERE ĪNANGA WERE RETAINED. (N=9 POND SAMPLES, N=17 GUT SAMPLES) ERROR! BOOKMARK NOT DEFINED. FIGURE 4.8: DIET SELECTIVITY RATIOS FOR TAXA FROM THE POND INVERTEBRATE COMMUNITY. POPULATION AVERAGE AND INDIVIDUAL BEHAVIOURS ARE DISPLAYED FOR EACH SAMPLING DATE. N=17 FISH ACROSS THE 4 DATES (JAN 8=5, JAN 22=7, FEB 2=3, MAR 10=2). RATIOS ARE CALCULATED BY COMPARING THE PROPORTIONAL ABUNDANCE OF INVERTEBRATE TAXA OBSERVED AT SITE 1 WHEN FISH WERE COLLECTED WITH THE PROPORTIONS OF THOSE SAME TAXA IN THE GUT CONTENT...... ERROR! BOOKMARK NOT DEFINED. FIG. 5.1: LOCATION OF STUDY SITES WITHIN THE WAIAU POND SYSTEM ON THE SOUTHLAND COAST. HIGHLIGHTING THE LOCATION OF KEY HYDROLOGICAL FEATURES AND STRUCTURES. SIX SITES WERE SELECTED FOR SAMPLING, SPREAD ACROSS THE KEY PONDS IN THE SYSTEM, SITES 1 AND 2 ARE IN THE WHITEHEAD POND SUITE, SITES 3 AND 5 ARE IN THE MAIN INDER POND WITH SITE 4 IN THE DIVERSION CHANNEL, AND SITE 6 IS IN THE MAIN MCCULLOCH POND...... 66 FIG. 5.2: TRENDS IN FISH CATCH NUMBERS BY MONTH POST-REFILL IN RELATION TO DISTANCE FROM THE REFUGE CHANNEL, SEPARATED INTO EELS AND NON-EEL FISH GROUPS. THERE WERE THREE SAMPLING OCCURRENCES IN MONTHS 1 AND 2 AND TWO IN MONTH 3 71 FIG. 5.3: TRENDS OBSERVED IN ZOOPLANKTON ABUNDANCES DURING THE INITIAL SAMPLING PERIOD, ILLUSTRATING THE TWO BLOOM EVENTS THAT OCCURRED, WITH A BREAKDOWN INTO (A) TOTAL ABUNDANCE AT EACH SITE AND (B) KEY TAXA ACROSS ALL SITES. ... 73 FIG. 5.4: BRAY-CURTIS BASED NON-METRIC MULTIDIMENSIONAL SCALING (NMDS) PLOT OF THE COMBINED MACROINVERTEBRATE AND ZOOPLANKTON SAMPLES (STRESS= 0.13). ELLIPSOIDS REPRESENT A 95% CONFIDENCE INTERVAL FOR THE COMMUNITY OF EACH SITE, WITH TAXON NAMES INDICATING THE DIRECTION OF THEIR INFLUENCE ON THE COMMUNITY...... 75 FIG. 5.5: COMPARISON OF INVERTEBRATE PROPORTIONS IN SAMPLES OVER THE FIRST THREE MONTHS AND A YEAR AFTER THE SYSTEM REFILLED. WHILE FEW OF THESE PATTERNS ARE SIGNIFICANT, THE OVERALL TREND IS A GENERAL DECLINE OVER 12 MONTHS FOR THE SMALL, GENERALIST PLANKTONIC TAXA (A), AND AN INCREASE IN LARGE, MORE SPECIALISED TAXA THAT ARE PRIMARILY EPIBENTHIC (B). SIGNIFICANT TRENDS ARE SEEN IN GYRAULUS AND PHYSA SNAILS (P=0.033, F=2.83, D.F=4/49 AND P=0.016, F=3.39,DF=4/49 RESPECTIVELY); CHIRONOMID LARVAE (P=0.074)AND ANISOPS (P=0.056) WERE NEAR SIGNIFICANT. SAMPLES WERE COLLECTED FROM 6 SITES ON 3 OCCURRENCES IN MONTHS 1 AND 2, TWICE IN MONTH 3 AND ONCE ON MONTH 12...... 75

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List of Tables

TABLE 3.1: PCA LOADINGS FOR THE FIRST THREE AXES OF VARIATION ...... 33 TABLE 4.1: DEMOGRAPHICS FOR THE 17 ĪNANGA CAPTURED FOR INDIVIDUAL ANALYSIS ...... 53 TABLE 4.2: OUTPUT FROM WIDESII SELECTIVITY ANALYSIS, INDICATING DIET PREFERENCE OF ĪNANGA. ONLY TAXA WHICH WERE FOUND IN GUT CONTENTS ARE INCLUDED IN TABLE. AVAILABLE VALUES INDICATE THE PROPORTION OF THE ASSEMBLAGE AS SAMPLED AT SITE 1, AND USED VALUES INDICATE THE PROPORTION OF THE ASSEMBLAGE AS SAMPLED AT SITE 1, AND USED VALUES INDICATE THE PROPORTION OF GUT CONTENT MADE UP BY EACH TAXA. RED TEXT INDICATES CONFIDENCE INTERVALS THAT DO NOT INCLUDE ZERO, MEANING THEY ARE CONSIDERED SIGNIFICANT...... 57 TABLE 5.1: DISTANCE FROM THE DEEP WATER REFUGE FOR EACH OF THE SIX STUDY SITES...... 67 TABLE 5.2: BRAUN-BLANQUET SCORE FOR MACROPHYTE COVER AT EACH SITE IMMEDIATELY AFTER REFILL AND AT THE END OF THE INITIAL SAMPLING PERIOD. INITIAL SCORE INCLUDES VEGETATION (E.G. GRASSES) THAT BECAME SUBMERGED DURING INUNDATION. BRAUN-BLANQUET SCORES ARE: R-<5% (FEW INDIVIDUALS), 1-<5% (MANY INDIVIDUALS), 2-5-25%, 3- 25-50%, 4-50-75%, 5-75-100% COVERAGE ...... 69 TABLE 5.3: AVERAGE FISH CATCH NUMBERS OVER THE FIRST THREE MONTHS OF SAMPLING POST REFILL OF THE SYSTEM (SAMPLING OCCURRED 3 TIMES IN MONTHS 1 AND 2, AND 2 TIMES IN MONTH 3) ...... 70 TABLE 5.4: ABUNDANCE AND PREVALENCE STATISTICS FOR THE MOST COMMON MACROINVERTEBRATE TAXA FOUND IN THE SWEEP NET SAMPLES ACROSS ALL SITES ...... 72 TABLE 5.5: ABUNDANCE AND PREVALENCE STATISTICS FOR THE MOST COMMON TAXA CAUGHT IN ZOOPLANKTON SAMPLES ACROSS ALL SITES...... 72 TABLE 5.6: FUNCTIONAL GROUPINGS FOR THE MOST ABUNDANT MACROINVERTEBRATE AND ZOOPLANKTON SAMPLES, INCLUDING HABITAT PREFERENCE, DISPERSAL POTENTIAL (LOW:10 M, MEDIUM: 1KM, HIGH:>1KM), FEEDING HABITS, DIETARY PREFERENCE AND REPRODUCTIVE POTENTIAL (R1:<100 OFFSPRING PER REPRODUCTIVE CYCLE, R2:100-1000 OFFSPRING, UNI: UNIVOLTINE, PLURI: PLURIVOLTINE). CATEGORISATIONS FROM NIWA MACROINVERTEBRATE TRAIT DATABASE (PHILIPS & SMITH 2018)...... 74

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Chapter 1. Introduction

1.1 Wetlands

Wetlands are defined as ecosystems that arise from inundation by water, where soils are dominated by anaerobic processes and the biota is adapted to flooding (Keddy 2010). In total wetlands make up around 8% of the Earth's land area (Zedler & Kercher 2005; Davidson, Fluet- Chouinard & Finlayson 2018), and are distributed from tropical to sub-polar climates. These systems are typically composed of areas with saturated soils, often interspersed with open water ponds. However, wetlands can vary widely in their appearance and permanence. In areas where prolonged dry periods are common, wetlands may dry out entirely, refilling when rains return; in other areas with more regular precipitation, wetlands are likely to be a permanent feature of the landscape (Boix et al. 2004; Culioli et al. 2006; Peterson et al. 2017; Pazin et al. 2006). Typically, wetland systems support communities of species adapted to semi-aquatic conditions; in particular, the vegetation found in a wetland system can be unique and specialised to survive in waterlogged soils. The specialisation of wetland flora is so extensive that its presence can be used as a descriptor in defining a wetland system (Johnson & Rogers 2003; Cronk & Fennessy 2016; McGlone 2009).

Wetlands are a valuable feature of the environment. They provide habitat that many animals use for part, or all, of their lifecycle. For example, migratory birds use wetlands as waypoints for feeding during their migration, many fish species utilise wetland areas for spawning and rearing of offspring, and many mammals, particularly rodents, rely on wetland vegetation for shelter and foraging (Williams & Dodd 1978). In addition to creating habitat, wetlands also provide a wide range of ecosystem services that benefit surrounding human settlements. They can provide flood mitigation by storing and dispersing large volumes of water, reducing peak flood volume and velocity, protecting areas downstream, or preventing tidal surges from travelling inland (Johnston 1991; Mitsch & Gosselink 2000; Zedler & Kercher 2005; Engle 2011). Vegetation typical of wetlands can also improve the stability of shorelines, reducing erosion around lakes and coastal areas (Gedan et al. 2011). Wetland soils and vegetation can act as a biological filter, removing excess nutrients and heavy metal pollution as water flows through the system, and allowing sediment to settle, improving water quality downstream (Hansson et al. 2005). Wetlands also provide areas and opportunities for a wide range of recreational and cultural activities, making them valuable to many people.

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1.2 New Zealand wetlands

New Zealand has a wide range of wetland systems, as the relatively wet maritime climate provides ideal conditions for wetland formation. The majority of New Zealand wetland systems are permanent, with only a few semi-permanent or ephemeral systems, for example in inland Canterbury where precipitation is low due to rain shadow caused by the Southern Alps (Ausseil et al. 2008). In New Zealand, six wetland systems have been recognised and protected by the Ramsar Convention as sites of international significance, two in the South Island and four in the North Island (Hamilton & Hamilton 2006). Wetlands in New Zealand, as with wetlands globally, are declining, with a wide range of factors contributing to their loss. It is estimated that New Zealand has lost over 90% of its historical wetland area, with loss in some regions, particularly lowland areas, being much higher than in others (Ausseil et al. 2008). Wetland loss can be attributed primarily to drainage to allow for alternate land uses, especially agriculture (Myers et al. 2013). In some regions, wetland loss is continuing extensively on private land as farmable land value continues to increase, and drainage practices become more advanced (Robertson et al. 2018). In addition to drainage, factors such as the introduction of plant and animal pests, changes to upstream hydrology (such as dams), and pollution all contribute to the degradation of wetland systems.

A wide variety of animals utilise New Zealand wetland systems, many of which rely on the highly productive open water areas to support key stages of their lifecycle. The invertebrate fauna makes up a significant part of wetland ecosystems, providing the trophic link from the primary producers through to the larger vertebrate taxa. The slow flowing nature of wetlands allows planktonic species that are less common in flowing waters to develop large populations. These tend to occur in a boom-bust cycle when conditions allow rapid population growth until resources are depleted (Hebert 1978); and provide an abundant food source for higher trophic levels during the boom phase. While the primary groups of freshwater invertebrates are consistent with those found in systems around the globe, New Zealand invertebrate assemblages have some unique characteristics (Collier 2010; Batzer & Ruhí 2013). Unlike many other regions, New Zealand freshwater invertebrates do not have a subset of taxa that have evolved specifically to survive in non- permanent systems; instead, generalist species that have rapid colonisation abilities tend to be found in both permanent and non-permanent systems (Wissinger, Greig & McIntosh 2009).

Wetlands provide a range of habitats and resources used by a variety of birds, both transient and resident. Some wetlands, such as those on Farewell Spit are on the flight path for migratory birds,

2 such as the eastern bar-tailed godwit (Limosa lapponica) or the Pacific golden plover (Pluvialis fulva), that travel long distances across the Pacific (Williams et al. 2006). These wetlands provide a valuable area to rest and feed to replenish energy stores. Other bird species such as the brown teal (Anas chlorotis), and Australasian bittern (Botaurus poiciloptilus) spend the majority, if not all, of their lives in and around wetlands, meaning that the loss of wetland habitat has put both species under threat (O’Connor, Maloney & Pierce 2007; O’Donnell 2011).

Native fish species also extensively utilise wetland habitat. Mudfish (Neochanna sp.) are found exclusively in wetlands, and have the unusual ability to aestivate during periods where the waterbodies they inhabit dry entirely. All mudfish species are classified as at risk, with the Canterbury mudfish (Neochanna burrowsius) being identified as nationally critical due to its extensive loss of habitat, and the continuation of drainage and water abstraction from wetlands in its known range (Ling 2001). Other native fish species such as eels, bullies and galaxiids extensively use wetland habitat, but are not exclusively restricted to them. Eels, particularly shortfinned eels (Anguilla australis) are common throughout coastal wetlands. They will often utilise the system as a waypoint for elvers arriving from the sea, before continuing to move inland, or they may spend the entirety of their adult resident stage in wetlands before migrating to sea to spawn (Jellyman & Chisnall 1999). Galaxiids such as giant kōkopu (Galaxias argenteus) are typically stream-dwelling species, but have been found to utilise open water areas (for example Waituna Lagoon) for rearing of larvae and juveniles, indicating a flexibility in their diadromous life history pattern (David et al. 2004). In addition to the native species, a range of introduced species such as trout, catfish, and perch are also commonly found in wetlands, and have varying levels of impact on the ecosystem.

Galaxiids, specifically the five migratory species, are arguably the native fish most valued by the general public in New Zealand. This is because each spring thousands of people line the banks of key estuaries and river mouths around the country to catch the juveniles of these five species as they migrate upriver en masse to reach their adult habitat (Haggerty 2007; Yungnickel 2017). These small juvenile fish, known collectively as whitebait, are treated as a delicacy. They provide an important recreational activity, often undertaken by multiple generations. This makes whitebait a resource that provides both social, recreational and cultural significance, as well as having a high potential economic worth. Unfortunately, the galaxiid species that make up whitebait are under threat; three species, īnanga, giant kōkopu and kōaro, are listed as at risk-declining, and shortjaw kōkopu is classed as threatened (Dunn et al. 2018). The decline in their populations have not been attributed to any one factor, but habitat loss and overfishing of juveniles are both likely to contribute (Richardson & Taylor 2002; Haggerty 2007). 3

In addition to their value as habitat, wetlands also provide benefits to the people who live near these systems. The abundance of species found within wetlands means that many wetlands are considered mahinga kai sites by local Māori communities. These systems are valued for the bounty of food resources they support. Wetlands are used for a range of recreational activities, for example fishing or bush walks, both popular pastimes for many New Zealanders.

1.3 Wetland conservation

In response to the loss of wetland habitat, and the plight of many of the species that inhabit them, significant conservation efforts have been put towards protection of wetland systems throughout New Zealand. There have been several studies identifying the wetlands at greatest risk and those that are the most important, usually taking into account the biodiversity and key species that inhabit the system as well as the uniqueness of the system itself (Ausseil et al. 2008; Ausseil et al. 2011; Myers et al. 2013). Wetland conservation takes a number of forms and is overseen at a range of levels from government agencies through to individuals taking action on private property. The first step of wetland conservation, once an area of value is identified, is protection against any potential future degradation (Bay of Plenty Wetlands Forum 2007; Greater Wellington Regional Council 2009). This is typically done by preventing artificial drainage and creating buffers between the wetland and surrounding detrimental activities, either in the form of riparian planting or creating boundaries for land use. In many cases systems are already degraded and require restoration; this can involve a range of actions including replanting of vegetation, removing weeds and other pest species, restoring or improving water flows into the system, and improving water quality through management of upstream pollution sources. Once restoration has occurred the more sensitive or specialised native species may return. In some cases animals will be translocated as part of individual species conservation efforts or ecosystem restoration projects, rebuilding the diversity and functioning of the system as a whole (Craig et al. 2000; Department of Conservation 2012). In some cases, conservation efforts required are more extreme, and involve complete re- establishment of a system. This tends to be done in areas that were once wetland but have been drained and converted into farmland, typically with the goal of creating habitat for species that are facing declines due to lack of habitat. Re-establishment is a lengthy process involving the creation of a system through increased water flows and water retention, and may take an extended period of time to become stable. Wetland re-establishment can work to remediate some of the historic loss of wetlands while protection and restoration can reduce the current rate of loss.

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The Southland region has approximately 33,000 hectares of wetland remaining, second only to Westland among New Zealand’s regions. This is only 7.9% of the historical extent, but makes up over 13% of national wetland area (Ausseil et al. 2008). Included within this is Waituna Lagoon and the surrounding Awarua Wetland, which are identified and protected by the Ramsar Convention as wetlands of international significance. A considerable amount of effort from multiple agencies has been put towards conservation of this wetland. Eutrophication is becoming a major concern for the system, as runoff from farmland in the catchment is causing increased nutrient and sediment loading of the lagoon, and notable degradation of seagrass beds (Robertson & Funnell 2012; Scanes 2012). Other notable conservation efforts in Southland include the restoration of Big Lagoon from a swampy area of farmland back into a large open water lagoon, and several restoration and re- establishment projects along the Waiau River with the goal of providing and improving habitat for native species, particularly galaxiids (Riddell & Sutton 2018). There are also a wide range of wetland systems within Southland that, while not being actively managed, are monitored to ensure any changes in condition are observed and can be mitigated.

1.4 Thesis outline

This study investigates the coastal wetlands of Southland, with the overall aim of improving understanding of the fish and invertebrate communities that are found within this kind of wetland. Each data chapter is written as a standalone paper prepared for submission to peer-reviewed journals. This will include: (1) a broad survey of the fish species found within key coastal systems, and how their populations change over time; (2) an investigation into how invertebrate communities change over a summer season, and the influence of environmental factors on the community; (3) an analysis of īnanga populations in a wetland, focusing on their distribution, diet and growth within the system; and (4) observations of how a wetland community responds to drought, and how it recovers following rewetting;. The first chapter investigates four separate wetland systems throughout coastal Southland, and the following chapters all focus in on one of these wetlands – the Waiau pond system. This is because their structure provides an ideal study site to look at small scale spatial variability, as well as having a variety of species of interest. Overall, this thesis is hoped to provide insight into the fish and invertebrate communities of these specific wetlands; and to generate more widely applicable findings in relation to īnanga diet, age, and growth, and re-established wetland structure and functioning.

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The first data chapter (Chapter 2) focusses on the fish populations of four coastal Southland wetland systems. The key aim is to understand the spatial and temporal patterns in the fish community. The four systems are all areas with ecological and cultural significance, as well as being managed or monitored for conservation of the system and key species within. There has been research into some of the ecological properties of all but one of the systems of interest. Big Lagoon currently has very limited information on its fish community or any parameters that may influence the population and distribution of species within it. Lake George is known for the presence of giant kōkopu, but has little data on the broader fish community. Research in Lake George has been primarily focussed on the hydrology, water nutrients and macrophyte beds (Robertson & Stevens 2013; Schallenberg & Kelly 2012). The Waiau pond system has had several fish surveys to establish the restoration success of the system; however, each of these surveys have been conducted at the same time of year, meaning that there is no information on seasonal patterns and the dynamics that influence populations over a short time span (Riddell & Sutton 2014; Gross et al. 2013; Duggan & White 2010). Waituna Lagoon has the most extensive body of existing research, much of which investigates threats, considering conservation options and implications to protect and preserve the system (Thompson & Ryder 2003). Rupia beds are of particular concern as these are sensitive to increased nutrient and sediment levels that are an increasing threat as farming in the catchment intensifies, as well as being impacted by increased salinity from the artificial openings (Robertson & Funnell 2012, Sutherland, Taumoepeau & Wells 2016; Schallenberg & Tyrrell). Research into the fish of Waituna Lagoon has primarily focussed on giant kōkopu, revealing how this species utilises both the lagoon and its tributary streams to complete its typically diadromous lifecycle without the need to access the ocean (David et al. 2004). There is broad understanding of the species that are present within the lagoon (Thompson & Ryder 2003; Tait & Pearce 2019), but this data does not detail any temporal patterns, nor does it look at spatial patterns within the lagoon itself. Overall, this chapter should improve current knowledge on the fish species that exist within these four wetland systems, the differences that are present between the systems, and the way that fish communities vary across space and time.

Chapter Three looks at the macroinvertebrate dynamics within the Waiau pond system. Invertebrate diversity and abundance can be highly variable, with a wide range of potential factors driving the variation (Batzer & Ruhí 2013). The influence of factors such as water quality, conductivity, depth, substrate, riparian cover, and water temperature have been well studied in streams, but less so within pond systems (Collier 1995; Scarsbrook 2002; Smith, Vaala & Dingfelder 2003). Succession is common in systems that display variation in environmental parameters, as the

6 favourable conditions for one species change to that for another, allowing for a series of dominant taxa over time. This is particularly common for planktonic communities where the primary taxa, for example copepods or cladocerans, are able to reproduce rapidly under favourable conditions (Hebert 1978; Hairston 1996). Overall, this chapter aims to identify how the invertebrate community in the ponds changes over the course of a summer, with a particular focus on how environmental parameters influence spatial and temporal variation.

Chapter Four focusses on the īnanga population within the Waiau pond system. Īnanga have a broad natural distribution, being found throughout much of the temperate Southern Hemisphere in both lentic and lotic systems (Berra et al. 1996). They display a high level of plasticity in their life history with both diadromous and landlocked populations (McDowall 1972; Pollard 1971), and a wide range in the timing of their spawning period (Barbee et al. 2011; Taylor 2002; Egan et al. 2019). Īnanga diet also shows great variability; different habitats support different feeding behaviours and diet compositions. Stream populations have been found to drift feed, with a diet that is primarily Simulidae and Plecoptera (Tagliaferro et al. 2015). Īnanga in a small lake in Chile have a diet that is primarily planktonic, largely composed of copepods and Bosmina (Cervellini, Battini, and Cussac 1993; Modenutti, Balseiro, and Cervellini 1993), whereas in a larger lake in Australia, īnanga diet was composed of mainly amphipods and chironomids (Pollard 1973). There is currently little data on the diet of īnanga within small wetland pond systems, or even the diet of īnanga in New Zealand in general. This chapter aims to investigate the diet of īnanga within the Waiau pond system, identifying the overall composition and any selectivity preferences for particular taxa. Additionally, the population dynamics will be explored, looking at the spatial and temporal distribution, age, and growth rate, as well as the factors driving the patterns observed. Overall, this chapter should improve insight and understanding of īnanga dynamics within New Zealand and specifically within a small pond system.

Chapter Five has been published in the New Zealand Journal of Marine and Freshwater Research (Stuart et al. 2020). It investigates the recolonisation process for both fish and invertebrates in a wetland system affected by drought, with a particular focus on succession and spatial patterns that arise. Generally research on wetland recolonisation after drought has been conducted in areas with wet-dry seasonality, where pond drying is regular and predictable (Pazin et al. 2006; Florencio et al. 2009; Peterson et al. 2017). These studies typically observe a succession in the invertebrate community as new species establish; with the community being primarily composed of specialists adapted to temporary systems (Boix et al. 2004; Culioli et al. 2006). By comparison, New Zealand has no invertebrate taxa specifically adapted to temporary systems but rather a subset of the 7 general pond taxa that have the ability to withstand dry periods or rapidly colonise newly available ponds (Wissinger et al. 2009). Studies on the recolonisation of fish following drought typically focus on this process within a stream or river setting, rather than a pond (Griswold et al. 1982; Albanese et al. 2009; Cowx et al. 1984). Many of these studies note the importance of refugia for the survival of species within a system or to enable a rapid recolonisation back into the system from nearby (Magoulick & Kobza 2003; Vrdolijak & Hart 2007). The drying of the Waiau pond system provided an ideal opportunity to investigate the recolonisation process and use of refugia in a wetland that does not normally dry, which has potentially different dynamics to the more well- studied streams and seasonal wetlands.

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Chapter 2. Spatial and temporal variability of fish communities within Southland coastal wetland systems

2.1 Abstract Wetlands are highly valuable and complex systems that provide habitat to a wide range of taxa. They can be particularly important for fish species, providing breeding and rearing habitat, and areas of abundant food. Within New Zealand, the majority of the original wetland area has been drained, leading to increased conservation effort to protect remaining wetland areas. This study surveyed the fish populations of four coastal wetland systems in Southland with Fyke and Gee minnow traps over a two-year period, to identify patterns present in the fish community both spatially and temporally providing a longer term record of catch numbers not currently observed for many of the systems . Two of the systems – Lake George and Big Lagoon, had very low abundances of fish, attributed to Lake George’s structure and Big Lagoon’s weak connectivity to other waterways. The two more complex systems – Waituna Lagoon and Waiau pond system, had much higher abundances and greater diversity of species. Drought in the Waiau pond system, and artificial opening of Waituna Lagoon, caused low water levels that impacted the fish populations. While the Waiau pond system fish community rebounded from the event quickly, Waituna Lagoon’s recovery was much slower, indicating that changes in management practices may be a benefit to the Waituna Lagoon fish community in the future.

2.2 Introduction

Wetlands are important ecosystems that cover around 9% of the Earth's land area (Zedler & Kercher 2005). They provide a wide range of ecosystem services, as well as being able to support a wide diversity of plant and animal species (Clarkson, Ausseil & Gerbeaux 2013). Wetlands filter water, removing compounds such as nitrogen and phosphorous, and allowing sediment to settle out of the water column (Hansson et al. 2005). This can lead to an improvement in water quality and may be able to lessen the impact of pollution, such as farm runoff, or stormwater discharges, that occurs upstream (Johnston 1991). Wetlands are able to reduce the impacts of flooding on the surrounding land. In upper catchments and floodplains, wetlands can store water, reducing flood height and flow velocity, and coastal wetlands can provide protection from storm surges (Mitsch & Gosselink 2000; Zedler & Kercher 2005; Engle 2011). Coastal wetlands can also stabilise shorelines, reducing or preventing coastal erosion (Gedan et al. 2011).

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In New Zealand, a wide range of native species utilise wetlands, including many freshwater fish (Clarkson et al. 2013). The high level of productivity typical of many wetlands allows them to support large populations of fish, such as eels and bullies (Zedler & Kercher 2005; Joy & Death 2013). Wetlands also provide valuable spawning and nursery habitat for species such as common bullies and several galaxiids (Taylor 2002; David et al. 2004; Ellery & Hicks 2009). Coastal wetlands are particularly valuable for many species that have a marine phase of larval and/or juvenile development (e.g. īnanga, kōkopu, and eels); they provide good habitat for juveniles to recruit into once they complete their marine phase, and for larval rearing of species that are not obligately diadromous. The shoreline complexity and submerged vegetation creates ideal shelter from predators, while being productive enough to support the development and growth of the young fish (David et al. 2004).

The Southland region of New Zealand has many wetlands that vary in size, complexity, connectivity to other water bodies, and surrounding land use. Unfortunately, due to the expansion and intensification of farming, particularly intensive dairy farming, many of these wetlands are threatened by nutrient run off and land reclamation (Chapman 1996; Brinson & Malvárez 2002; Johnson & Rogers 2003). In Southland the rate of wetland loss is dramatic, and higher than the national average, with only 7.9% of historic wetlands left, compared to 10.1% nationally (Ausseil, Gerbeaux & Chadderton 2008). In response to this rapid loss, wetland conservation is increasing, with projects involving the restoration of remaining wetlands as well as the re-establishment of previously drained wetlands.

The wide variability between wetlands has a significant impact on what species will be present in any given system, and how large their populations can become. Habitat preferences of wetland species range widely - some require very specific conditions (Jowett & Richardson 2003; Joy & Death 2013), whereas others, such as shortfinned eels, have far more generalised requirements (Jellyman & Chisnall 1999). Variability between wetlands is likely to cause significant differences in the fish communities and population sizes they support, with some species only being found in specific systems or sites within a system.

This study focusses on the fish populations of four coastal Southland wetland systems. It is expected that there will be significant variation in catch numbers across seasons, and between systems. Catch numbers are expected to be highest over summer, as fish activity, and therefore net encounter rate, tends to decline with colder temperatures over winter. The more complex systems (Waiau pond system and Waituna Lagoon) are expected to have larger catch numbers than the

10 other two systems (Lake George and Big Lagoon), as the more complex systems can potentially provide a greater variety of habitat for use by a wider range of species.

Discharge within the catchment, as an indicator of flood or drought conditions, is expected to have an influence on catch levels. Catch is expected to be greatest at annual median flow levels, and to decline during high or low flow events. Fish are likely to be less active or seeking refugia during low flows, and if movement occurs during a flood event there is a much greater depth and area of water available, reducing the net encounter rates during both events.

Community composition within systems is expected to remain consistent over time, given average hydrological conditions, with some variation in species occurrences between sites. Fish community composition is predicted to differ between systems, as not all species will have suitable habitat in all sites, in particular it is expected that more typically estuarine species will be found in Waituna Lagoon than the other systems, as it has greater connectivity to the sea (Riddell, Watson & Davis 1988, Thompson & Ryder 2003).

2.3 Methods

2.3.1 Study sites

Four systems within the Southland coastal area were selected as part of this study (Fig. 2.1, Table 2.1). All are major wetlands with ecological and cultural significance within the region, and have been subject to some level of conservation effort or management. These four wetland systems fall into two categories: the more complex systems partially connected to the sea – Waituna Lagoon and the Waiau pond system, and the less complex systems that are slightly further inland – Lake George and Big Lagoon. Within the complex systems four sites were selected for regular sampling every six weeks, these sites were selected based on ease of access while covering as much variability within the system as possible. Sites were selected in the same way at the less complex sites, however due to time and resource constraints, and indications that fish abundances are low in these systems, only two sites were selected, and these were sampled every 3 months.

Waituna Lagoon is one of the largest lagoons in New Zealand, with a surface area of 1350 hectares (Fig. 2.1). It is located within the Awarua wetland, a 20,000 hectare system that is internationally recognised by the Ramsar Convention. However it is threatened by intensification of land use in the small catchment, leading to high water abstraction rates and increased nutrient inputs (Thompson

11

& Ryder 2003; Ramsar 2020). The opening of the Waituna Lagoon is mechanically controlled, in order to manage the water level, primarily for flood mitigation and nutrient management. A digger is used to cut through the sand bank when the lagoon water level reaches 2.25m on the Waghorns Road gauge. Mechanical opening connects the lagoon to the sea until it naturally closes, in some cases over a year later (Thompson & Ryder 2003).

Big Lagoon has had extensive restoration works, as it was previously almost entirely drained to create farmland. It is now the second largest lagoon in Southland (after Waituna Lagoon), covering just over 31 hectares. Connectivity to the Waimatuku River and surrounding streams is limited, but a fish ladder has been installed to aid in the recolonization of native fish. However, the effectiveness of this in providing fish passage is yet to be investigated. To date there has been no research into the fish community or other system dynamics within the lagoon.

Lake George is a shallow dune lake, with an average depth of 0.8 m (maximum depth of 2 m), an area of 90-100 ha dependent on water level (Robertson & Stevens 2013), and limited connectivity to other water bodies. It is the least managed of the four systems, as much of the catchment is native forest and farming in the catchment is low intensity, hence the threat of degradation is comparatively low. It has a relatively short water residence time, meaning that phytoplankton blooms are rare (Schallenberg & Kelly 2012). However, it is regularly monitored to ensure any changes are recognised quickly and can be managed. This monitoring is primarily for water quality, and phytoplankton levels, with limited detail on the fish community, although indications are that fish abundances can be low (Robertson & Stevens 2013; Schallenberg & Kelly 2012).

The Waiau pond system is composed of a series of ponds that have been re-established on the Waiau river floodplain, in an area that had previously been drained to create farmland. The wetlands were created through a series of dams with connecting underground pipes, which are fed by two diversion channels that take water from the Waiau River. They then drain into Te Wai Wai Lagoon at the mouth of the river, and as a result, some of the lower ponds are subject to the intrusion of brackish lagoon water. The wetlands have been established in stages, with the construction of three suites of ponds (in 2009, 2012 and 2015), forming a total of 30ha of open water habitat (Riddell & Sutton 2017). Due to the way the ponds are set up they are highly dependent on the Waiau River flows - when the river is low, inflows to the wetland are significantly reduced. This makes the system highly variable, and prone to periodic low water levels and drying.

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Fig 2.1: Location of study sites within the four selected systems along the Southland coast. Within Waituna Lagoon and the Waiau ponds (designated as the more complex sites) four sampling sites were selected, in Lake George and Big Lagoon (the simpler sites) two sites were selected.

Table 2.1 Brief descriptions of the physical characteristics of each site sampled, including descriptions of substrate, submerged and emergent vegetation, and riparian plants System Site Site Description 1 Beach at far end from where opening occurs, cobbles and gravel, seagrass present, some reeds along the shoreline 2 Downstream of Waghorn’s Road bridge, deep channel with drop off along banks, Waituna mixed mud/silt and small stones, some riparian vegetation Lagoon 3 shallow side stream, dense reeds along the edges, established macrophyte bed, thick mud/silt 4 Near boat ramp, reeds along shoreline, seagrass present, mixed mud and gravel 1 Small pond, underground pipes connect to other ponds and lagoon, thick fine sediment, densely grassed riparian edge, established macrophytes with a variety of species Waiau 2 Main Whitehead pond, fed by siphon, submerged grasses and reeds, compact Ponds cobbled bed 3 Main McCulloch pond, thick fine sediment layer, some semi submerged reeds 4 Diversion channel, clay bed, patches of macrophyte, flowing water 1 Stream that feeds the lake, good riparian edge, grasses and overhanging trees, Lake muddy bottom George 2 Within the lake itself, very deep fine sediment layer, reeds and some riparian grasses 1 Within the Lagoon, thick anoxic mud, dense flax and reeds along shoreline Big 2 Near retaining wall, deep fine sediment, limited vegetation or macrophytes, Lagoon concrete rubble along shore

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2.3.2 Sampling

At each site one unbaited fyke net (1cm mesh, 0.5m high, 0.65m wide, 4.5m wing) and one G- minnow trap (3mm mesh, 20mm opening) were set overnight. The following day any captured fish were identified and counted, and the length and weight of the first 15 fish of each species at each site were measured. Sampling occurred every three months at Lake George and Big lagoon (four sampling dates a year), and every six weeks in the Waiau pond system and Waituna Lagoon (eight sampling dates in a year), and was planned to occur over two years (giving eight and sixteen sampling dates respectively).

It was concluded after the first year of sampling that catches were not sufficiently high in Big Lagoon and Lake George to justify the continuation of sampling for the following year, giving only four temporal replicates. Additionally, nets were unable to be set effectively in the Waiau pond system due to extremely low water levels as a result of drought (January 2018).

Daily average records of water level (Waituna Lagoon) and flow rate (Waiau River) were obtained from Environment Southland and NIWA respectively.

2.3.3 Data processing and statistics

All analyses were run in the statistical programme R (R Core Team 2019). Data was condensed to total catch numbers for each species at each site and date. Species richness (total number of taxa), Shannon diversity index (a measure of evenness within a community), and Simpson diversity index (a measure of the variety of species) values were calculated for each site and date. Timing and duration of the low water events at Waituna Lagoon and Waiau pond system were identified and included in the respective datasets.

To investigate patterns in the fish community, generalised linear models (GLMs) were run using Poisson distribution. Site and date were used as predictor variables, with number of eels, common bullies, īnanga and total catch number used as response variables; this was run for both the Waiau pond system and Waituna Lagoon, but not Big Lagoon or Lake George as too few fish were caught. Each of the three community indices (richness, Shannon, and Simpson) were used as response variables with site and date, or site and pre/post event used as the predictor variables. This was run for both Waituna Lagoon and the Waiau pond system, but not for the other two systems due to the low catch numbers.

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The relationships between water level and catch number both before and after the low water event in Waituna Lagoon were explored using Pearson’s correlation coefficients. This was not run for the Waiau pond system as the GLM analysis did not indicate any significant influence of the low water levels.

2.4 Results

2.4.1 Community composition and population variability

The number and composition of species that were caught across the four systems varied substantially; overall 6 species (117 fish in total) were captured in the Waiau pond system, dominated by, shortfin eel (30%), īnanga (28%), and common bully (26%), along with longfin eel (13%), brown trout (2%) and smelt (1%). Abundance and species richness in Waituna Lagoon was higher than what was observed in the Waiau pond system, with 589 individual fish and 8 species captured. Common bullies were the most abundant, making up 70% of fish caught, with longfin eel (11%), shortfin eel (10%) and īnanga (5%) making up majority of the remainder. Mullet, flounder, giant kōkopu and redfin bullies were also caught. Catch numbers were much lower at the other two sites; in Lake George ten fish were caught (7 shortfin eel, 2 giant kōkopu and 1 īnanga), and in Big Lagoon only two shortfin eels were caught. At the system level, across all sites sampled, Waituna Lagoon had significantly higher catch numbers (p<0.0001, R2=0.235)and species richness(p<0.0001, R2=0.095) than the Waiau pond system. There were not enough fish caught in Lake George or Big Lagoon to allow for a meaningful statistical comparison.

There was significant spatial variability observed within both the Waiau pond system and the Waituna Lagoon system (Fig. 2.2 & 2.3). The Waiau pond system showed between site variability in catch numbers (p=0.002, R2=0.33), and when broken down into individual species, there was significant variability in bullies p=0.008, R2=0.56) and īnanga (p=<0.0001, R2=0.93), and eels (p=0.019, R2=0.47). In comparison Waituna Lagoon had no significant difference between sites for the number of īnanga (p=0.29 R2=0.51) of eels (p=0.15, R2=0.64), but did show significant site differences for common bullies (p<0.0001, R2=0.48), and overall catch numbers (p=0.001, R2=0.43). īnanga

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Waiau ponds: Site 1 Waiau ponds: Site 2 18 16 16 14 14 12 12 10 10 8 8 6 6

4 4 Number fish of caught 2 Number fish of caught 2

0 0

03-12-18 30-05-17 17-07-17 02-09-17 18-10-17 02-12-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 03-12-18 22-01-19 10-03-19 19-04-19 30-05-17 17-07-17 02-09-17 18-10-17 02-12-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 22-01-19 10-03-19 19-04-19

Common bully Shortfin Longfin Inanga Brown trout Common bully Shortfin Longfin Inanga Brown trout

Waiau ponds: Site 3 Waiau ponds: Site 4 4 16 14 3 12 10 2 8 6

1 4 Number of fish fish of Number caught Number fish of caught 2

0 0

30-05-17 17-07-17 02-09-17 18-10-17 02-12-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 03-12-18 22-01-19 10-03-19 19-04-19 30-05-17 17-07-17 02-09-17 18-10-17 02-12-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 03-12-18 22-01-19 10-03-19 19-04-19

Common bully Shortfin Longfin Inanga Brown trout Common bully Shortfin Longfin Inanga Brown trout

Figure 2.2: Catch composition across the 16 sampling dates for the four Waiau pond sites

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Waituna Lagoon: Site 1 Waituna Lagoon: Site 2 45 35 40 30 35 25 30 25 20 20 15 15 10 10

5 5 Number fish of caught

0 Number fish of caught 0

02-09-17 30-05-17 17-07-17 18-10-17 02-12-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 03-12-18 22-01-19 10-03-19 19-04-19

03-03-18 30-05-17 17-07-17 02-09-17 18-10-17 02-12-17 16-01-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 03-12-18 22-01-19 10-03-19 19-04-19

common bully shortfin longfin common bully shortfin longfin inanga Giant kokopu mullet inanga Giant kokopu mullet flounder Brown trout redfin flounder Brown trout redfin

Waituna Lagoon: Site 3 Waituna Lagoon: Site 4 40 45 35 40 30 35 30 25 25 20 20 15 15 10 10

5 5 Number fish of caught

Number fish of caught 0 0

30-05-17 17-07-17 02-09-17 18-10-17 02-12-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 23-10-18 03-12-18 22-01-19 10-03-19 19-04-19

02-12-17 23-10-18 30-05-17 17-07-17 02-09-17 18-10-17 16-01-18 03-03-18 10-04-18 03-06-18 20-07-18 30-08-18 03-12-18 22-01-19 10-03-19 19-04-19

common bully shortfin longfin common bully shortfin longfin inanga Giant kokopu mullet inanga Giant kokopu mullet flounder Brown trout redfin flounder Brown trout redfin

Figure 2.3: Catch composition across the 16 sampling dates for the four Waituna Lagoon sites

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2.4.2 Effects of hydrology

Both Waituna Lagoon and the Waiau pond system experienced hydrological events that resulted in significant declines in water level over the study period. In the Waiau pond system, a period of drought lasted from 11 November 2017 to 1 February 2018. During this period Waiau River flows did not exceed 60 cumecs (m3/s), and were therefore not high enough to flow into the pond intake channel, resulting in the pond system becoming entirely dry. Discharge down the Waiau River varied substantially over the 2 years: prior to the drought event flows averaged 87.7 cumecs; during the drought period flow was reduced to 40.5 cumecs; and after the drought, averaged 110.8 cumecs (Fig 2.4). A flood with a peak of 299 cumecs refilled the ponds on 2 February 2018. Despite the severity of the drought, there was no notable difference in catch numbers or community indices (Shannon and Simpson diversity) before and after the event (catch numbers p=0.677 R2=0.051; Shannon diversity p=0.671, R2=0.01, Simpson diversity p=0.83, R2=0.017). In addition, there was no apparent relationship between catch numbers and flow rate recorded (p=0.663, R2=0.01, 95%CI= - 0.4, 0.6).

800

700

600

500

400

(cumecs) 300

200

100 Waiau river Waiau river flows atTuatapere

0

01-Jul-17 01-Jul-18

01-Jan-18 01-Jan-19

01-Jun-17 01-Jun-18

01-Oct-17 01-Oct-18

01-Apr-17 01-Apr-18 01-Apr-19

01-Sep-17 01-Feb-18 01-Sep-18 01-Feb-19

01-Dec-17 01-Dec-18

01-Aug-17 01-Aug-18

01-Nov-17 01-Nov-18

01-Mar-18 01-Mar-19 01-May-18 01-May-17

Fig. 2.4: Waiau River fows recorded at Tuatapere for the length of the study period, orange area highlights the period of low flows leading to dewatering of the wetland. Data provided by Meridian Energy

Waituna Lagoon also experienced a low water event; on May 30, 2018 the lagoon was artificially opened to the sea, due to high water levels triggering the lagoon opening protocol. Over the 2 years of data collection, the lagoon was at its highest level for the 5 days preceding the opening. Over the following 2 days, the water level dropped from 2278 mm to 541mm (Fig. 2.5). Lagoon opening had a lasting effect on water level for the following year, with an average water level of 859 mm

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compared to the previous year average of 1367 mm. The lagoon opening occurred in the middle of the two-year sampling period allowing for a full year to compare before and after opening. Unlike the Waiau drought, the lagoon opening event had a significant impact on the fish community: total

catch (F(1,62)=5.14, p=0.026 ), eel numbers (F(1,62)=5.46,p=0.022 ), and number of taxa (F(1,62)=10.9, p=0.0016) were all lower after the opening than they were prior to opening. The Simpson diversity

index (F(1,62)=13.48, p=0.0005) was higher after the opening event, most likely due to the reduced abundance of some taxa causing a more even community composition. Lagoon opening also changed the relationship between catch numbers and water level with a negative relationship prior to opening (Fig. 2.6, p=0.09, R2=0.09, 95% CI = -0.58, 0.055) whereas during the year after opening there was a positive relationship between water level and catch numbers (p=0.005, R2=0.23, 95% CI= 0.157, 0.709).

2500

2000

1500

1000

500

0

Water level Waterlevel atWaghorns Road (mm)

01-Jul-17 01-Jul-18

01-Jan-18 01-Jan-19

01-Jun-17 01-Jun-18

01-Oct-17 01-Oct-18

01-Apr-17 01-Apr-19 01-Apr-18

01-Sep-17 01-Feb-18 01-Sep-18 01-Feb-19

01-Dec-17 01-Dec-18

01-Aug-17 01-Aug-18

01-Nov-17 01-Nov-18

01-Mar-18 01-Mar-19

01-May-17 01-May-18

Fig. 2.5: Waituna Lagoon water level at Waghorns Road gauge for the length of the study. The lagoon was artificially opened on May 30 2018. Data provided by Environment Southland

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Fig 2.6: Relationship between number of fish caught and water level before (May 2017 to April 2018) and after the opening of the lagoon (June 2018 to April 2019). Before opening p=0.09, r=-0.30, n=32, after opening p=0.005, r=0.48, n=32.

2.5 Discussion

2.5.1 Differences between systems

Overall, there are a number of notable differences between the study systems, both in terms of number and variety of species. Significantly more fish were caught in the Waituna Lagoon than the Waiau pond system, a difference mostly explained by the high numbers of common bully. This was expected as Waituna Lagoon has large areas of suitable spawning and juvenile rearing habitat, such as cobble and gravel (Stephens 1982), whereas the Waiau pond system does not. The species found at both sites were similar, with the highest numbers of common bully, eels and īnanga. The key difference between the two sites was the presence in Waituna Lagoon of the more typically estuarine species such as mullet (Aldrichetta fosteri) and black flounder (Rhombosolea retiarii) that were common in the lagoon prior to opening. These fish likely entered during a previous open period and were able to persist even after the lagoon reclosed. 20

The very low number of fish caught in Big Lagoon was unexpected due to the substantial effort put into its restoration, however it does appear that providing habitat and access for fish was not the primary goal of the project. The lagoon requires improved connectivity to surrounding waterways for it to provide habitat for more than the occasional eel that makes its way into the system. The numbers of fish caught in Lake George were also lower than anticipated, this may be in part due to the shape and shallow gradient of the shoreline making it difficult to position the fyke nets effectively. It is also likely that fish are utilizing habitat closer to the middle of the lake, as this is where macrophytes are denser, and therefore refugia and foraging opportunities are greater (Robertson & Stevens 2013).

2.5.2 Impacts of hydrology

The Waiau pond system and Waituna Lagoon systems responded differently to the low water level events that occurred over the course of this study. The Waiau pond system, despite drying entirely, showed a rapid recovery with little evidence of a long-term impact on the fish community. By contrast, in Waituna Lagoon, markedly lowering the water level had an immediate reduction in numbers of fish captured, which did not rebound to pre-opening levels despite a full year of sampling post opening. Following lagoon opening there was also a reduction in the number of species present, which did not recover over the subsequent year of surveys.

There are a number of factors that may explain to the differing response in catch numbers in these systems, the biggest being the hydrology of the system and the nature of the event. Typically, the speed of recovery from drought or low flow conditions is related to the connectivity of the system and the length of time that water is depleted (Detenbeck et al. 1992; Matthews & Marsh‐Matthews 2003, Lennox et al. 2019). The Waiau pond system is connected to the Waiau River and fluctuates in level as the river does; this makes the ponds highly variable, but also means fish may be used to moving between the ponds and river as needed to cope with the changing water levels. The connectivity to the river also provides a nearby recruitment source for the system, and is likely the reason why the fish community was able to recover rapidly. Waituna Lagoon is quite different as it is such a large body of water that natural increases in water level are generally slow, except when extreme rainfall events occur. However, the artificial opening procedure results in a dramatic loss of water (almost 2 m over 2 days). Being a large water body fed by a few small streams means a potentially smaller pool of recruits being available to replenish the Waituna Lagoon population after a loss of water in comparison to the availability of recruits in the Waiau pond system. In addition, the lagoon itself acts as an important source of recruits back into to the connecting

21 streams (Thompson & Ryder 2003; David et al. 2004), meaning that loss or reduction of access to the lagoon could cause further detriment to all connected fish populations.

The nature of the low water level events is also very different between the two systems, and this likely influenced the ability of the system to recover. The development of drought in the Waiau pond system was a slow process, with the ponds drying gradually once river flows dropped below the level of the water intake to the wetlands system, with refilling occurring rapidly when the river rose above the critical level. This provided a sudden increase in suitable habitat which could be readily recolonized from the nearby river. In contrast the water level drop in Waituna Lagoon was rapid and occurred when the natural water level was at its highest. However, refilling was delayed as the opening takes time to be closed over by coastal sediment transport processes (Kirk & Lauder 2000). Unlike many other coastal lagoons, Waituna Lagoon is primarily a low salinity system. However, when lagoon opening does occur, salinity can increase rapidly and significantly, particularly in areas near the opening, although an increase is measurable across the lagoon in its entirety (Schallenberg et al. 2010; Sutherland et al. 2016). This salinity increases as a result of lagoon opening has potential to be highly detrimental to the existing population of typically freshwater species. This is a notable contrast to many more brackish lagoon systems dominated by marine species, where lagoon opening is beneficial as it increases the salinity and promotes the use of the habitat as a rearing habitat for many key species (Griffiths 1999; Saad & Beaumord 2002).

2.5.3 Management implications

Waituna Lagoon is considered a highly valuable wetland system, as it is internationally recognized by the Ramsar Convention, and has great significance for local iwi and the wider community. The lagoon provides habitat for many species that have been recognized as threatened or at risk of decline (Robertson et al. 2013; Dunn et al. 2018). This study found that opening of the lagoon caused an immediate reduction in both abundance and diversity of fish species, with a continued effect on the fish community for the year following the opening. This is consistent with findings from several other studies (Duggan & White 2010; Robertson & Funnell 2012). This impact on the fish community is counter to many conservation goals, and suggests that the opening practices of the lagoon require reevaluation. There are several reports investigating alternate management options, for example the use of a gated dyke to maintain a minimum water level but prevent excessive farmland flooding (DHI 2015; Thompson & Ryder 2003), although it has been suggested that much of the flooding of farmland results from heavy rainfall regardless of lagoon level (Jackson, Phillips & Ekanayake 2001). A modification to the current practice would likely prove

22 beneficial to the fish community, as well as the rest of the ecosystem, while still reducing the flooding risk for surrounding farmland.

The Waiau pond system has also been modified by human activities; the ponds themselves were created from farmland, that had previously replaced an existing wetland area. Further, the river that feeds them also has a highly modified flow regime. The control structure built at the confluence of the Waiau and Mararoa Rivers near Lake Manapouri has resulted in a significant reduction in river discharge, particularly over summer when only minimum flows of 16 m3/s-1 are required to be released, to maintain the level of Lake Manapouri for electricity generation (Jowett & Biggs 2006). The lack of any effect on the fish community as a result of the drought that occurred indicates that this system is resilient to fluctuations in water level. This is likely due to the strong connectivity to the Waiau River allowing for fish to move back into the system from the river as soon as it becomes habitable. Therefore, maintaining a strong connection to the river is vital for functioning of the system following any future drought events. In response to the severe dry period over the 2017-2018 summer, a land use consent was implemented to excavate areas around the intake of the pond system to improve connectivity to the river and reduce the risk of future dry periods draining the ponds entirely (Riddell & Sutton 2018). However, as the following summer was not dry for an extended period, the effectiveness of this has not yet been evaluated.

2.6 Conclusions

There is a wide variability in species and abundances of fish in Southland coastal wetland systems. It is evident that some systems can support far greater densities than others, as would be expected given the range in system size, connectivity and complexity. The effects of events that cause low water levels are variable, and would greatly benefit from long term study to investigate trends in population sizes over several years and multiple events. Overall, the abundance and diversity of fish species within these systems has the potential for increase given changes to management, for example, creating improved connectivity to Big Lagoon and the maintenance of water levels in Waituna Lagoon.

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Chapter 3. Drivers of invertebrate abundance and community composition in a small pond complex

3.1 Abstract Freshwater invertebrates are a key part of any pond ecosystem, with ecological roles and linkages through a range of trophic levels, and displaying great variability in abundance and diversity as the environment changes. This study investigated spatial and temporal variation in pond invertebrate communities over a spring to autumn period, and examined how environmental factors such as temperature, nutrients, and chlorophyll influenced abundance and diversity. Sampling was conducted every three weeks over a six-month timespan, with invertebrate community sampled via a semi buoyant plankton net tow, and environmental conditions measured through collected water samples and in situ data loggers. Invertebrate community composition varied significantly both spatially and temporally. Spatial variation appeared to be driven by hydrological connectivity and nutrient levels within the system. Temporal variation displayed a progression of dominant microcrustacean taxa, as well as a succession of larger more predatory taxa. Daphnia reached abundance levels greater than any other taxa but did not show a relationship with any measured environmental parameter. Abundance of invertebrates (excluding Daphnia) increased with distance from the main water inflow channel, and displayed a negative relationship with nitrate levels. This indicates that the filtering effect of the wetland improved water quality, and highlights the potential importance of a complex pond system for invertebrate communities.

3.2 Introduction

Aquatic invertebrates are a highly important part of any freshwater ecosystem. They provide the key linkage between primary productivity and higher trophic levels (Grubh & Mitsch 2004; Sorrell & Gerbeaux 2004). For example, planktonic crustaceans such as Daphnia spp. feed on phytoplankton and are in turn eaten by other invertebrates and small fish that feed in the water column. Larger macroinvertebrates also provide valuable food sources for larger fish species such as trout, as well as for birds (Hanson & Butler 1994; Sagar & Glova 1995).

The diversity of invertebrates in aquatic systems can vary hugely, potentially with high diversity in a single location (Batzer & Ruhí 2013). Abundances can also be variable, ranging across several orders of magnitude. This variation can occur both spatially and temporally within a system, typically as a result of environmental variation (Olerti et al. 2007; Marshall et al. 2006; Silvera et al. 24

2006). Invertebrate communities are influenced by a wide range of factors, although driving factors have been more commonly studied in streams and lakes than ponds. Studies on stream communities have found increases in populations of individual taxa, or community richness, driven by water velocity and quality, conductivity, depth, substrate, riparian cover, and water temperature (Collier 1995; Scarsbrook 2002; Smith, Vaala & Dingfelder 2003). Invertebrates in lakes have also been widely studied, particularly the planktonic community. Zooplankton populations show a strong response to changes in nutrient levels, particularly phosphorus (Jeppesen et al. 2003). Phosphorus limitation can cause declines in zooplankton through bottom up control, as the abundance of phytoplankton and algae that they feed on will be reduced at low nutrient levels (Abell et al. 2010). Temperature is also a key factor that influences invertebrate communities in lakes, with changes in abundance associated with seasonal variation (Burns & Mitchell 1980; Hart 1985). The same breadth of research is not found for ponds and wetlands, but it is expected that many of these drivers will still apply.

Small planktonic species are far more abundant in the slow flowing waters of lakes, ponds and wetlands than they are in flowing streams. They often undergo rapid population expansions (from here referred to as ‘blooms’) as a result of a high reproductive rate that is supported by an abundance of food (Hebert 1978; Mooij et al. 2003). Within New Zealand lakes these blooms are typically composed of small crustaceans (such as Daphnia, Bosmina, or copepods), and can develop rapidly, dominating a system (Burns & Mitchell 1980; Forsyth & McCallum 1980; Burns, Schallenberg & Verburg 2014). While studies of invertebrate communities in pond systems have seldom been done in New Zealand, there is some existing knowledge, primarily examining the difference between permanent and semi-permanent pond systems (Maxted, McCready & Scarsbrook 2005; Wissinger, Greig & McIntosh 2009; Greig, Wissinger & McIntosh 2013). These studies have shown that permanent ponds are dominated by snails, odonates and caddisflies, whereas semi-permanent ponds have greater abundances of beetles, microcrustaceans and chironomids. It has been shown that New Zealand invertebrate communities do not have a specific subset of species that are semi-permanent pond specialists, but that communities in semi- permanent systems are comprised of generalist taxa tolerant of drought (Wissinger, Greig & McIntosh 2009).

This study aims to investigate the abundance and diversity of aquatic invertebrates within a re- established wetland system, focusing on how small-scale spatial variability influences invertebrate community composition and how it changes over time. The pond system of interest is located at the mouth of the Waiau River in Southland New Zealand, and comprises of three suites of ponds built 25 over 6 years (in 2009, 2012 and 2015), forming a total of 30ha of open water habitat (Riddell & Sutton 2017). There are around ten large ponds and many smaller ones, which are interconnected, and fed via two diversion channels which take water from the Waiau River. Both spatial and temporal variability in the taxa present are investigated, as well as the abiotic and biotic factors that may be driving the changes observed

It is predicted that environmental parameters such as nutrients and chlorophyll levels will vary both spatially and temporally, with conditions in the diversion channel differing from the pond sites; and conditions differing over the peak of summer compared to spring and autumn. It is hypothesized that community composition will vary between sites due to the complex hydrological dynamics within the pond system; in particular it is expected that the diversion channel will support a different community as it has flowing water. Variation in community composition is also expected to occur over time, particularly through the peak of summer where temperatures are expected to be higher and water levels lower. It is expected that abundances will change in response to variation in temperature and other environmental factors; and that blooms of planktonic species (e.g. Daphnia) will coincide with peaks in nutrients and chlorophyll.

3.3 Methods

3.3.1 Study site

Six sites were chosen within the Waiau pond system (located at -46.195121, 167.630188, Fig. 3.1), a re-established wetland, created by a series of dams and interconnecting pipes, and fed by two diversion channels that take water from the Waiau River. The wetland system was constructed in three stages over a 6-year timespan (Whitehead ponds: 2009, McCulloch ponds: 2012, Inder ponds: 2015), and is now connected to allow water flow between all component ponds. The six sites selected are spread throughout the system at a range of distances from the diversion channel. Sites were selected to maximise the environmental variability samples within the system, while considering ease of access. Site 4 is located in the Inder diversion channel itself, at a point not far above the beginning of the main Inder pond. Sites 3 and 5 are both within the main Inder pond, with Site 3 being in a shallower section, and Site 5 near a deep channel remnant from an old river braid. Site 6 is within the McCulloch pond complex, in a pond that is not immediately connected to the inflow point of its diversion channel. Sites 1 and 2 are located in the Whitehead pond complex, a series of ponds that are fed via a siphon that takes water from the main Inder pond. Site 2 is located

26 within the pond where the siphon outflow is, and Site 1 is the furthest from the river intake, and is nearest to the lagoon.

Sampling trips occurred regularly covering the period from early spring through to the beginning of autumn. Sampling trips were approximately every 3 weeks, beginning on September 12th, 2018, with the last on March 10th, 2019, this resulted in a total of 9 sampling trips over the course of the study.

Fig. 3.1: Location of study sites within the Waiau pond system on the Southland coast. Location of the deep channels that run through the system are highlighted.

3.3.2 Environmental parameters

A variety of environmental parameters were measured over the course of the study. The parameters selected were those expected to have the greatest influence on the invertebrate community over the course of the summer. The progression of water through both the pipe network and the soil between sites is expected to cause spatial variation even at the reasonably small scale of the system, fed by the same water source. Water temperature, chlorophyll, and nutrients (both dissolved and total nitrogen and phosphorus) were parameters that were expected to show variance between sites, and have been shown to be important drivers of invertebrate

27 abundance in previous studies (Burns & Mitchell 1980; Hart 1985; Hassall, Hollinshead & Hall 2011). Water level within each pond has a less direct pathway for biological interaction; however, small changes can greatly change the surface area and dry margin of small ponds, altering the habitat available. It also provides an indication of the spatial and temporal variation in water movement within the system as well as the concentrating or diluting effects a change in water level may confer.

Prior to the beginning of sampling, stakes were positioned at each site to use as water level gauges. The stakes were marked with centimetre increments, positioned far enough into the ponds to ensure some stake would remain submerged during typical water level fluctuation, and hammered into the pond beds until secure. During each sampling trip water level at each site was recorded to the nearest centimetre. At the end of the sampling season an overall average water level, and above/below average levels for each date were calculated to allow for a comparison between ponds over the course of the season.

HOBO pendant temperature loggers were deployed at the beginning of the sampling season and logged temperature every 30 minutes for the duration of the study. Two loggers were set at each site, one attached to the water level stake, approximately 10cm above the bed of the pond, and a second connected to a float, with the logger hanging approximately 10cm below the surface. Unfortunately, due to weather conditions several of the surface floats were lost, all subsurface loggers remained functional throughout the study.

A series of water samples were collected from each site during each of the nine sampling trips. This included 500ml for chlorophyll analysis, 50ml for total nutrient analysis, and 50ml passed through a 0.3µm glass fibre filter for dissolved nutrient analysis. All samples were collected from an area of open water away from any plumes of disturbed sediment, and kept chilled in a dark chilly bin until they reached the lab for processing.

Immediately on return to the lab, and approximately 3 hours after collection, water samples for chlorophyll analysis were filtered through 0.3µm glass fibre filters using a vacuum filtration system. Filters were wrapped in foil and stored at -20°C until enough samples were collected to run the analysis. Chlorophyll was extracted overnight by ethanol and run on a plate reader (Fluostar Omega plate reader, BMG LabTech) at 665nm and 750nm. Chlorophyll concentration was then calculated following Biggs & Kilroy (2000).

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Water samples collected for nutrient analysis were stored at -20°C until enough samples were collected to run the analysis. Samples were then thawed and prepared for analysis in the autoanalyzer (Skalar SansPlus segmented flow analyser, De Breda, The Netherlands). Several analysis runs were completed to measure nitrate and nitrite (NO2/NO3), dissolved reactive phosphorus (DRP), ammonia, total phosphorus (TP), and total nitrogen (TN) (Valderrama 1981; Morris & Riley 1963; Murphy & Riley 1963; Solorzano 1969).

3.3.3 Invertebrate sampling

Invertebrate samples were collected using a buoyant 30 cm diameter plankton net with 250 um mesh which was thrown out into the pond and then drawn back through the water for approximately 5 meters. This sampled invertebrates in the water column in deeper water and in macrophyte beds in shallower water nearer the pond’s edge. The six sites were sampled in rotation, with 3 samples from each site collected over a two-hour period. This ensured that samples were more representative of the pond as a whole, minimising the risk of collecting from patches of higher or lower density being repeatedly sampling, and allowing time between samples to reduce the influence of net disturbance on subsequent samples. Unfortunately, due to equipment issues on November 13, the third rotation of sample collecting could not occur, giving only two replicates from each site. Invertebrate samples were emptied from the net and preserved in 90% ethanol. Samples were then processed, counting all invertebrates and identifying to the lowest practical taxonomic level. In some cases, the density of invertebrates in the samples was so high that counting the entire sample was not feasible. These samples were run through a plankton splitter and a subset of the sample was counted, before being multiplied back to estimate total number in the sample. Of the 156 samples that were collected and processed, 21 were split prior to counting, this includes 6 where half was counted, 9 where a quarter was counted, and 6 where an eighth was counted.

3.3.4 Data processing and statistics

Invertebrate and environmental data was collated, with some processing required to allow for analysis. Water level data was adjusted from above/below average measurements by adding 34 (the greatest measurement below zero) to all values; this maintains the relative differences between sites and dates while avoiding negative values. Data from the temperature loggers was

29 processed to provide daily daylight average, maximum and minimum temperatures. Since several surface loggers were lost, data was only used from the loggers positioned near the pond bed. The triplicate invertebrate samples collected were averaged to provide a single comparable community for each date at each site. Shannon and Simpson diversity indices, total number of taxa, total invertebrate abundance, and total abundance excluding Daphnia (by far the most common taxon) were calculated for each averaged sample. Where required for analysis, abundance data and environmental parameters (nutrients, temperature, chlorophyll, and water level) were transformed to improve normality. As abundance data contains many zero values a transformation of log(x+1) was used. All analysis was performed using the statistical programme R (R Core Team 2019).

To understand the environmental conditions within the ponds a series of analyses were run. Redundancy analysis (RDA), allowed visualisation of the spatial and temporal variation (between sites and between sampling dates). This was followed by PERMANOVA, run on Bray-Curtis dissimilarities, to test the significance of the observed differences, utilising the R package Vegan (Oksanen et al. 2019). One-way ANOVAs were run for each of the environmental parameters to test for differences between dates and between sites. Principal Component Analysis (PCA) was run to identify combinations of environmental factors that accounted for most of the variation observed. An ANOVA was used to test for variation in the primary axis of variation (PC1) between sites and between sampling dates.

The invertebrate community in the ponds was also investigated through a series of analyses. Spatial and temporal patterns in community indices for the overall invertebrate community were analysed through generalised linear models (GLMs). Community indices (Shannon diversity, Simpson diversity or taxon richness) were used as response variables and site and date used as the predictor variables. Poisson distribution was used for the number of taxa, and gaussian distribution for the two diversity indices.

Redundancy analysis (RDA), was used to visualise spatial and temporal variation (between sites and between sampling dates). This was followed by PERMANOVA, run on Bray-Curtis dissimilarities, to test the significance of the observed differences. One-way ANOVAs were run for each of the ten most abundant invertebrate taxa, and for overall abundances to identify individual and general spatial or temporal variation.

Analyses were also run to identify any relationships between the environmental factors and the invertebrate community. Pearson's correlation coefficients were calculated for relationships

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between environmental parameters and the most abundant invertebrate taxa, and between PC1 and key invertebrate taxa.

3.4 Results 3.4.1 Environmental factors

Spatial variation in the overall environmental conditions of the ponds was observed (F(5,46)=2.58, p=0.004). Site 1 was most dissimilar to the diversion channel (Site 4) with a progression of change through the rest of the sites. This difference appears to be driven primarily by the water nutrients

(NO2-NO3, ammonia, TN and TP, Fig. 3.2). When environmental factors were investigated individually, nitrate/nitrite level was the only parameter that showed significant variation among sites, with higher levels measured at site 4 (p<0.001, F=11.6, d.f=5/41), the closest site to the pond inflow.

Fig. 3.2: Redundancy analysis plots displaying the environmental condition polygons for each of the sites and the direction of influence of the individual environmental parameters. p=0.001, n=9 per site. NO2.N03 = dissolved nitrate and nitrite, NH3= ammonia, TP= total phosphorus, TN= total nitrogen. Based on the PERMANOVA run on the RDA, there was no detectable difference in the overall

environmental conditions of the ponds over time (F(7,46)=1.4, p=0.13). However when looking into

individual parameters, both water level (F(7,39)=8.01, p<0.001) and daylight water temperature

(F(7,39)=42.87, p<0.001) showed significant variation, with lower temperature seen during the first two dates (September 12 and October 3), and an increased water level on October 3 (Fig. 3.3).

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A B

Fig. 3.3: (A) System average water level measured across the eight dates sampled, p<0.001, n=6 sites per date. (B) Average water temperature during daylight hours for each of the eight dates sampled, p<0.001, n=6 sites per date, error bars are ± standard error.

Principal Component Analysis (PCA) of the environmental data showed the patterns relating the different variables. PC1 shows that samples with high water level and high nitrates tended to have low levels of total phosphorus, DRP and total phosphorus (Table 3.2). There is a significant

relationship between PC1 and site (F(5,41)=4.83, p=0.001), with a higher value for Site 4, and lower

for sites 1 and 6 (Fig. 3.4). There was not a relationship between PC1 and date (F(5,39)=1.43, p=0.22).

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Table 3.1: PCA loadings for the first three axes of variation

PC1 PC2 PC3 Proportion of variance explained 0.43 0.15 0.13

Nitrates 0.36 -0.22 0.04 Dissolved reactive phosphorus -0.42 0.21 -0.26 Ammonia -0.31 0.56 0.00 Chlorophyll α -0.37 -0.28 -0.28

Water level 0.26 0.50 -0.53

Water temperature -0.27 0.18 0.71 Total nitrogen -0.37 -0.48 -0.26

Total phosphorus -0.42 0.06 -0.01

Fig. 3.4: Pattern in PC1 value (calculated from eight environmental parameters) across the six study sites (p=0.001, n=9 per site)

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3.4.2 Invertebrate community

Overall 28 different taxa were identified, with a maximum diversity of 18 taxa observed near the deep channel in the Inder pond (site 5) on March 10, and a minimum diversity of 5 taxa in the diversion channel (site 4) on November 13, and at site 5 on September 12. Daphnia dominated the community, comprising 95% of all invertebrates counted; the next most abundant taxa were cyclopoid copepods, making up 1%.

3.4.2.1 Spatial variability Significant differences between sites were observed for all community composition indices- Shannon diversity (p<0.0001, d.f=5/53, residual deviance=345.19) Simpson diversity (p=0.0005, d.f=5/53, residual deviance= 160.35), and total number of taxa (p=0.04, d.f=5/53, residual deviance=11.64). When looking at the invertebrate community in its entirety, there was no significant difference in the overall community composition between sites (F(1,53)=1.55, p=0.126). Of the ten most abundant taxa, four showed significant spatial variation when examined individually:

Daphnia (F(5,48)=2.91, p=0.022) and water boatmen (Sigara) (F(5,48)=2.44, p=0.047), both displayed increased abundances at sites 2 and 3, cyclopoid copepods (F(5,48)=3.16, p=0.014) had the greatest abundance at site 6, and helical snails (Gyraulis) (F(5,48)=3.17, p=0.014) had their greatest abundance at site 1 (Fig. 3.5). The total abundance of invertebrates sampled varied significantly between sites (F(5,48)=2.95, p=0.021), with a higher abundance at site 3 which is attributed to the high abundance of Daphnia. When Daphnia was removed from the dataset, significant variation was still observed; however, the pattern was different with a notably lower invertebrate abundance at

Site 4 (in the diversion channel) (F(5,48)=3.97, p=0.004).

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Fig. 3.5: Abundances of four invertebrate taxa for each of the 6 sites: Daphnia (p=0.022), cyclopoid copepods (p=0.14), boatman (p=0.047), and helical snails (p=0.014). n=9 per site.

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3.4.2.2 Temporal patterns When looking at the invertebrate community in its entirely there was a significant change in

composition over time (F(8,53)=2.06, p<0.001). During spring the community was dominated by Bosmina and chironomid larvae, progressing into summer with a shift to a community dominated

by Daphnia, and in autumn shifted towards ostracods and Potamopyrgus snails (F(8,45)=4.57, p<0.001 Fig. 3.6). This is supported by significant variation observed in all community indices over

time (Shannon diversity F(8,45)=6.2, p<0.0001, Simpson diversity F(8,45)=5.88, p=0.02, number of

taxa F(8,45)=3.83, p=0.001). Of the ten most abundant taxa, three showed a significant temporal

trend: chironomids (F(8,45)=3.62, p=0.002), backswimmers (Anisops) (F(8,45)=2.3, p=0.04), and water

boatmen (F(8,45)=2.33, p=0.04). The peaks in abundance of these three taxa are successive, with chironomids declining notably by November 13, at which time backswimmers peaked, and water boatmen began to increase in numbers (Fig. 3.7)

Fig. 3.6: Redundancy analysis plot displaying invertebrate community centroids for each of the dates sampled, with arrows illustrating the direction of change over time and names indicating the direction of influence of the individual taxa. p=0.001, n=6 per date

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C A B

Fig. 3.7: Abundances for three invertebrate taxa over the 9 dates sampled, the order of graphs illustrates the order of succession of these taxa, as one peak in abundance follows the next. (A) chironomid (p=0.002), (B) backswimmer (p=0.03) and (C) water boatman (p=0.03). n=6 per date, error bars are ± standard error.

3.4.3 Relationship between environment and invertebrates A significant negative relationship was observed between nitrate/nitrite levels and the total abundance of invertebrates excluding Daphnia, the highest observed nutrient levels coincide with the lowest abundances (p<0.002, R2=0.18, Fig. 3.8), although these sites are also those with the strongest connectivity to the river (the diversion channel and the deep area of the main Inder pond). In comparison there was no relationship between abundance excluding Daphnia and any of the other nutrients that were measured.

Comparison between PCA primary axis of variation (PC1) and Daphnia abundances appears to show a relationship between the two, where abundance of Daphnia is highest for intermediate values of PC1 and declines in abundance for higher or lower values (Fig 3.9). There also appears to be a relationship between Daphnia and water temperature, showing what is likely the optimum temperature range, with the greatest abundances observed when the daytime average temperature was between 17℃ and 19℃ (Fig. 3.10).

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A B C

Fig. 3.8: (A) Invertebrate abundances excluding Daphnia for each of the 6 sites (p=0.004). (B) Nitrate/nitrite levels across the 6 sites (p<0.0001). (C) The relationship between invertebrate abundance and Nitrate/nitrite levels (p=0.002, r=-0.43). Sites are ordered based on distance form the water diversion channel (Site 4), error bars are ± standard error.

Fig. 3.9: Relationship between Daphnia abundance and PC1 (first axis of variation in a PCA run on environmental parameters within the pond system)

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Fig. 3.10: Relationship between Daphnia and average daylight temperature

3.5 Discussion

3.5.1 Environmental factors

The differences observed between the sites in terms of environmental conditions appear to show patterns based on connectivity through the system, in particular the strength of the linkages to the Waiau River via the diversion channels. As hypothesised, the diversion channel (site 4) was consistently the most different from other sites, particularly when looking at the nitrate/nitrite levels or the PCA outputs. This is the newer of the two diversion channels, and has not yet developed a vegetated riparian strip or stable macrophyte cover over the clay bed. It follows that the conditions observed at site 4 will be most like the conditions in the Waiau River itself.

The difference in the overall environmental conditions observed was consistent with distance from the diversions. The two sites in the main Inder pond (sites 3 and 5) have the greatest similarity to the diversion channel that feeds it, whereas the site furthest from the diversion channel is consistently the most different. The two sites with an intermediate distance through the system (Sites 2 and 6) show similarity despite being fed by separate diversions. Wetland systems are known to act as a filter and a sink for nutrients, with a key ecosystem service they provide being an

39 improvement to water quality downstream (Hansson et al. 2005; Clarkson, Ausseil & Gerbeaux 2013; Craft, Vymazal & Kröpfelová 2018; Davidson et al. 2019). The progression of change in environmental condition moving away from the diversion channel illustrates this process in action. The further water has travelled through the system the longer the nutrients in the water are available to be utilised by algae and macrophytes or sequestered into the sediment (Jansson et al. 1994; Fisher & Acreman 2004; Li et al. 2019; Wang et al. 2021).

The ponds remained relatively stable over time, with the only variation observed being in temperature and water level early in the sampling season. This can be attributed to cooler temperatures and frosts typical of early spring, as well as a heavy rainfall event. The consistency over time of the other parameters measured, particularly nutrient levels, indicates a consistency of water quality in the Waiau River itself over the course of the study. This is understandable as the key sources of nutrient input into the river will be from agricultural runoff, which is typically more stable, and at lower levels over the summer study period (Monaghan et al. 2007; Monaghan et al. 2010). Higher rainfall levels during winter, combined with increased stock density for winter grazing leads to increased runoff with greater sediment and nutrient inputs. A year-long study of environmental conditions could prove valuable as it would be able to better identify seasonal variation, particularly over winter where farming practices such as break feeding can create more variable water quality (Monaghan et al. 2007; Southland Intensive Winter Grazing NES Advisory Group 2020).

3.5.2 Invertebrate community

The spatial variability observed in the composition of the invertebrate community was largely related to variation in the measured hydrological and environmental conditions. As predicted, the diversion channel (Site 4) had notably fewer taxa and lower abundances than any of the other sites; in addition to being the only site with flowing water, it also had the highest nitrate levels. While the effects of these two factors cannot be separated entirely, it is likely that they independently influence the pattern observed. Flowing waters support a different array of species than ponds, including those more suited to higher velocity conditions (Williams et al. 2004; Maxted et al. 2005; Datry et al. 2017); and elevated nitrogen concentrations are known to have impacts on invertebrate

40 abundances, particularly for taxa that are more sensitive to poor water quality (Quinn & Hickey 1990; Quinn et al. 1997; Mor et al. 2019).

The significant change in community composition over time showed a progression from one dominant taxon to another, although the link between this succession and the environmental conditions are not as clear as had been hypothesised. The dominant taxa driving the changes in community were typically smaller, planktonic taxa such as Bosmina, Daphnia, or ostracods, which are capable of rapid population growth. It is understandable that a population explosion, due to a shift in conditions favouring one taxon over another, can shift the community composition. In addition to this change in community composition, a succession of larger invertebrate taxa was also identified. Succession has often been studied in temporary ponds where the drying and flooding process is regular and predictable (Moorhead, Hall & Willig 1998; Boix et al. 2004; Nicolet et al. 2004; Culioli et al. 2006). These studies tend to find a shift over time from smaller taxa to larger ones, particularly for the predatory feeding group (Moorhead et al. 1998; Culioli et al. 2006; Cunillera-Montcusí et al. 2020). The succession observed in this study from smaller chironomids to larger backswimmers and then water boatmen is consistent with these observations derived from temporary ponds, but in a more permanent pond system.

Daphnia in the ponds were very abundant at times, making up 99% of the invertebrates in some samples, with notable increases in abundance during late spring and mid-summer (mid-November and January-February). The timing of these peaks in abundance was different to the typical spring bloom (known as the clearwater phase) observed in many zooplankton studies (Lampert et al 1986; Ferrara, Vagaggini & Margaritora 2002; Müller‐Navarra, Güss & von Storch 1997); however a number of New Zealand based studies have observed a wide variation in the timing of population peaks for a range of zooplankton taxa (Hall & Burns 2003, Burns & Mitchell 1980; Greenwood et al. 1999; Balvert, Duggan & Hogg 2009). Blooms in Daphnia populations are typically attributed to increases in food availability, with their primary food source being algae, bacteria, or other small planktonic organisms (Burns & Schallenberg 1996; Burns & Schallenberg 2001). This increase in food allows for rapid reproduction through one of two reproductive modes, parthenogenesis or sexually produced eggs (Schwartz & Hebert 1987; Lynch 1989; Hairston 1996; Green 1956; Lynch 1989; Kleiven, Larsson & Hobæk 1992). During processing a large proportion of Daphnia were observed to contain offspring, indicating that primarily the parthenogenic pathway was being utilized, although egg capsules were also present. Overall, there was no single factor that appeared as a clear driver of Daphnia abundance. Some sites displayed a pattern where a peak in Daphnia followed a peak in chlorophyll, however this was not consistent across all sites. In particular site 3 41 which had the largest numbers of Daphnia did not experience any peaks in chlorophyll; This may be because the Daphnia population was consuming the photosynthetic organisms as they increased, preventing any measurable changes. This idea is supported by the fact that within the pond system the highest levels of chlorophyll were observed when there were little to no Daphnia present. Additionally, the patterns observed with PC1 and with temperature indicate that there may be some relationship, but it is likely to do with an optimum range of conditions rather than a single factor trigger (Green 1956; Heugens et al. 2006).

As expected, there was a relationship between nutrients and abundances in the pond system, although the pattern of this was clearer than anticipated. The clear reduction in nitrates with distance from the river intake, and the corresponding increase in invertebrate abundance, highlights the benefit of creating a wetland that filters water effectively. Elevated nitrate levels can be common in rivers with agricultural inputs in their catchments (Julian et al. 2017; Harding et al. 1999), these rivers may then be used to feed constructed wetland systems such as this one. By designing wetlands to contain a series of ponds and connections, nitrates, and other pollutants, may be filtered more effectively, than a simple single pond system. Considering this when designing wetland systems may counteract the impacts of upstream pollution, and potentially allowing for greater invertebrate abundances, beneficial for the wetland ecosystem as a whole.

3.6 Conclusions

This study has shown that succession of small to large macroinvertebrate taxa occurs in this permanent pond system over the course of spring to autumn, displaying a similar pattern to what has been observed in non-permanent pond systems in other locations. The change in nutrient levels and the interaction between these levels and invertebrate abundances within this system highlights the importance of complexity in a wetland system. This complex connectivity allows for longer residence time of water; and may be a valuable characteristic to incorporate into the design of future wetland construction projects. Further investigation into the fine scale variability of nutrient levels across this system would be valuable to create a greater understanding of the specific design features that allow for the nutrient filtration seen in this study. A longer time scale covering several winters would also be valuable to observe fluctuations in nutrients and invertebrate community, particularly to understand if the system is capable of buffering the increased nutrient loads typical of winter agricultural runoff.

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Chapter 4. Growth and diet of īnanga within a coastal pond system

4.1 Abstract Īnanga (Galaxias maculatus) are a small freshwater fish species found in a range of freshwater habitats around much of the Southern Hemisphere. Although they are reasonably well studied within New Zealand, little is known about the populations found in pond systems. This study focussed on īnanga diet and population dynamics within a shallow pond complex in coastal Southland, investigating age, growth, distribution, and diet of juvenile īnanga over a spring to autumn period. Gee-minnow traps were set at six sites within the pond system over nine separate dates, all fish were measured with a subset retained for otolith and gut content analysis. Otolith analysis indicated a three-month period (September to November) where fish are metamorphosing and entering freshwater habitat, with earlier fish growing more slowly than later fish. Īnanga were found in only one location within the pond complex - a site with low water nutrient levels and more difficult to access by trout. Diet analysis showed high selectivity for chydorid cladocerans and chironomid larvae, indicating both benthic and pelagic foraging. Additionally, the majority of the guts dissected contained microplastic filaments, although the source of this material was not clear. Overall, this study has provided insight into pond īnanga population dynamics, which have not previously been observed, as well as highlighting the worrying pervasiveness of plastic pollution.

4.2 Introduction

New Zealand’s native fish display a variety of life history characteristics, often including diadromy, exemplified by the five migratory galaxiid species. Juveniles of these species rear at sea and return en masse to river mouths to reach their adult freshwater habitat. During this migration vast numbers of juvenile fish, known as whitebait, are caught to be eaten or sold. Whitebaiting is regarded as an important activity with high cultural value, although it is also considered to be one of the factors responsible for the decline in the adult fish populations (Department of Conservation 2019). This decline in several of the migratory galaxiid species has led to increased conservation efforts around adult habitat and spawning sites as well as an investigation into the options for improving management of the fishery (McDowall 1996; Franklin and Bartels 2012; Department of Conservation 2019).

Īnanga (Galaxias maculatus) make up a significant proportion of whitebaiting catches, typically comprising 70-100% of a catch (Yungnickel 2017). Of the five species they have the best

43 understood spawning behaviour (Taylor 2002), although there is much of their behaviour that is still not well understood. Īnanga are a lowland species, as they have limited ability to climb and navigate past barriers such as waterfalls or manmade structures like culverts (Franklin and Bartels 2012). They have the shortest lifespan of the migratory galaxiids, typically only living for a year (Stevens, Hickford, and Schiel 2016; Rojo, Figueroa, and Boy 2018). Īnanga spawn during spring tides in autumn, laying their eggs among grass stems at the upper limit of the salt wedge in an estuary, and on the following spring tide the larvae hatch and are swept out to sea. The adults typically die after spawning, although some survive and can spawn again the following year (Stevens, Hickford, and Schiel 2016; Rojo, Figueroa, and Boy 2018). After approximately six months at sea, the larvae then return to freshwater as whitebait in the following spring, where they will settle into their adult habitat (McDowall, Mitchell, and Brothers 1994).

Īnanga have one of the widest natural distributions for a freshwater fish species, being found in southern South America, southern Australia and New Zealand (Berra et al. 1996). They inhabit a variety of freshwater habitats, including both lentic and lotic waters. Īnanga diet and feeding behaviour has been studied in a range of system types, with significant variation apparent between studies, highlighting the plasticity of this fish species. In some small lakes their diet is entirely comprised of planktonic species such as copepods (Cervellini, Battini, and Cussac 1993; Modenutti, Balseiro, and Cervellini 1993), whereas in other larger lakes their diet also contains taxa such as chironomid larvae (Pollard 1973). In comparison diet of īnanga in some streams is largely made up of Simulidae and Plecoptera (Tagliaferro et al. 2015). Within New Zealand, studies have observed the drift feeding behaviour of īnanga (Jowett 2002), with a limited body of research into their diet (Caitlin, Collier & Duggan 2019).

In addition to the significant variation observed in the diet of īnanga there is also great variability observed in other features of their life history, and in particular their spawning and recruitment dynamics. Īnanga have a long potential spawning period, with spawning being observed in New Zealand fish across all but four months (June-September) (Taylor 2002). The peak spawning period is typically considered to be during autumn; however, this does appear to vary between populations, with earlier spawning occurring further south, indicating an effect of latitude on spawning timing (Barbee et al. 2011; Taylor 2002). This variation in timing enables them to capitalise on the conditions leading to the greatest larval survival and recruitment success, which will occur as different times in different areas. There is also significant plasticity in the requirements for completing their lifecycle, as although they are considered a diadromous species,

44 they can exist as landlocked populations using lakes for larval rearing rather than the ocean (McDowall 1972; Pollard 1971, David et al. 2019).

This study focusses on the īnanga population in a coastal pond system, with the aim of identifying and understanding patterns in their distribution, feeding and growth. Stream based populations have been the focus of New Zealand research, so there is a gap in knowledge based in ponds (Richardson and Taylor 2002; Bonnett and McIntosh 2004; Jowett and Richardson 2008). It is predicted that there will be both spatial and temporal variation observed in īnanga abundance, reflective of īnanga preference of particular areas as conditions vary over time and space. Differences in growth and age of fish caught over the course of the season will be investigated with the hope of revealing the effect of early or late hatching on future success of fish. This will be done through analysis of sagittal otolith banding; these ear bones form daily bands as the fish grow, and can then be counted and measured to give an estimate of age and rate of growth (Elsdon et al. 2008). It is expected that there will be a variety of ages of juvenile fish at any one point in time and that the fish that are younger later in the season will grow faster than those that hatched and metamorphosed earlier. This is expected as conditions, in particular temperature, are more likely to be advantageous further into the summer. It is hypothesised that the diet of īnanga will reflect that of the invertebrate community in the ponds, indicating non-selective feeding behaviour as has been observed in New Zealand streams (Jowett 2002)..

4.3 Methods

4.3.1 Study site

Six sites were chosen within the Waiau pond system (Located at -46.195121, 167.630188, Fig. 3.1), a re-established wetland, created by a series of dams and interconnecting pipes, and fed by two diversion channels that take water from the Waiau River. The wetland system was constructed in three stages over a 6-year timespan (Whitehead ponds: 2009, McCulloch ponds: 2012, Inder ponds: 2015), and is now connected to allow water flow between all component ponds. The six sites selected are spread throughout the system at a range of distances from the diversion channel. Sites were selected to maximise the environmental variability samples within the system, while considering ease of access. Site 4 is located in the Inder diversion channel itself, at a point not far above the beginning of the main Inder pond. Sites 3 and 5 are both within the main Inder pond, with Site 3 being in a shallower section, and Site 5 near a deep channel remnant from an old river

45 braid. Site 6 is within the McCulloch pond complex, in a pond that is not immediately connected to the inflow point of its diversion channel. Sites 1 and 2 are located in the Whitehead pond complex, a series of ponds that are fed via a siphon that takes water from the main Inder pond. Site 2 is located within the pond where the siphon outflow is, and Site 1 is the furthest from the river intake, and is nearest to the lagoon.

Sampling trips occurred regularly covering the period from early spring through to the beginning of autumn; these trips allowed for sample collection for both Chapter 3 and Chapter 4. Sampling trips were approximately every 3 weeks, beginning on September 12th, 2018, with the last on March 10th, 2019, this resulted in a total of 9 sampling trips over the course of the study.

Fig. 4.1: Location of study sites within the Waiau pond system on the Southland coast, highlighting the location of the deep channels that run through the system

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4.3.2 Īnanga sampling

Nets were set overnight at the six study sites during each of the sampling trips. The trapping regime included a metal Gee minnow trap (3mm mesh, 20mm opening), and a soft mesh minnow trap (1mm mesh, 70mm opening). Traps were retrieved early the following morning and all fish caught were identified, counted and measured. Some juvenile galaxiids (whitebait) were caught that were too small to identify in the field but probability is high that these were īnanga, as they make up majority of the whitebait population in this area (Yungnickel 2017). A total of 17 īnanga were euthanised in water with a high concentration of Aqui-S for further analysis (isoeugenol solution produced by Aqui-S New Zealand Ltd.); the number of fish collected on each date varied depending on the overall number caught. Larger īnanga (those greater than 85 mm) were excluded from collection as they are more likely to be adults from the previous season than juveniles recruiting into the ponds. It is recognised that this is a small sample size for this type of study, however as the study site is a conservation area, caution was taken to ensure that the number of fish removed was not too high.

In the lab, retained īnanga were measured for standard length, then dissected to remove the digestive tract and both otoliths. Guts were examined under microscope, identifying as much of the contents as possible, and counting both identified and unidentifiable items. Invertebrate content was identified to the lowest practical taxonomic level where possible or categorised by type, for example ‘unspecified insect leg’. Unexpectedly, microplastic was observed in the guts, this content was identified as either fibre or particulate and the colour noted.

One otolith from each fish was mounted on a microscope slide with a thermoplastic adhesive and polished to reveal the core (Fig 4.2). Daily otolith increments were independently counted three times for each otolith, by the same person. If variation between counts was more than 5%, the second otolith from that fish was prepared and counted, with variation rechecked. In all cases where a second otolith was prepared, the variation between counts was acceptable, and it was therefore used. This was done to ensure that the best quality otolith was used. For each otolith, the clearest line from edge to core was identified within the same sector. Measurements of daily increments and total radius were made along this line using ImageJ software (Wayne Rasband, NIH, USA) from photographs taken at low-power (x40 or lower) magnification on an Olympus BX51 compound microscope mounted with BP25 camera attachment. Counting began at the first clear

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increment from the centre at the magnification used, which is representative of the hatch point of the fish (McDowall, Mitchell & Brothers 1994; McClintock 2018; Egan Hickford & Schiel 2019).

A B

Fig 4.2: (A) Image of prepared otolith indicating the inner (white arrow) and outer (black arrow) increments identified for counting, from fish 8-1E. (B) Image of clear inner increments indicating the hatch point (red arrow) from fish 23-1A

4.3.3 Environmental parameters

A variety of environmental parameters were measured over the course of the study, this data is used in both the Chapter 3 and Chapter 4 analyses. This information was collected to observe the general condition of the ponds and may help to explain any variation observed in īnanga distribution or population sizes within the system. Prior to the beginning of sampling, stakes were positioned at each site to use as water level gauges. The stakes were marked with centimetre increments, positioned far enough into the ponds to ensure some stake would remain submerged during typical water level fluctuation, and hammered into the pond beds until secure. During each sampling trip water level at each site was recorded to the nearest centimetre. At the end of the sampling season an overall average water level, and above/below average levels for each date were

48 calculated. As negative values are problematic for some analyses, water level data was adjusted from above/below average measurements, by adding 34 (the greatest measurement below average) to all values, this maintains the relative differences between sites and dates while enabling analysis.

Temperature loggers were deployed at the beginning of the sampling season and logged temperature every 30 minutes for the duration of the study. One logger was set at each site, attached to the water level stake, approximately 10cm above the bed of the pond. Data from the temperature loggers was subsequently processed to provide daily daylight average, maximum, and minimum temperatures.

A series of water samples were collected from each site during each sampling trip, including 500ml for chlorophyll analysis, 50ml for total nutrient analysis, and 50ml passed through a 0.3µm glass fibre filter for dissolved nutrient analysis. All samples were collected from an area of open water away from any plumes of disturbed sediment, and kept chilled in a dark chilly bin until they reached the lab for processing.

Water samples for chlorophyll analysis were filtered through 0.3µm glass fibre filters using a vacuum filtration system immediately on return to the lab. Filters were wrapped in foil and stored at -20°C until enough samples were collected to run the analysis. Chlorophyll was extracted by ethanol and run on a plate reader (Fluostar Omega plate reader, BMG LabTech) at 665nm and 750nm. Chlorophyll concentration was then calculated following Biggs & Kilroy (2000).

Water samples collected for nutrient analysis were stored at -20°C until enough samples were collected to run the analysis. Samples were then thawed and prepared for analysis in an autoanalyzer (Skalar SansPlus segmented flow analyser, De Breda, The Netherlands). Several analysis runs were completed to measure nitrate and nitrite (NO2/NO3), dissolved reactive phosphorus (DRP), ammonia, total phosphorus (TP), and total nitrogen (TN) (Valderrama 1981; Morris & Riley 1963; Murphy & Riley 1963; Solorzano 1969).

4.3.4 Invertebrate sampling

Invertebrate samples were collected to investigate the relationship between what īnanga consume and what prey is available in the system. Samples were taken using a buoyant 30 cm diameter plankton net with 250 um mesh which was thrown out into the pond and then drawn back through

49 the water for approximately 5 meters. This sampled invertebrates in the water column in deeper water and in macrophyte beds in shallower water nearer the ponds edge. The six sites were sampled in rotation, with 3 samples from each site collected over a two-hour period. This ensured that samples were more representative of the pond as a whole if patches of higher or lower density were sampled on a single tow, and allowing time between samples reduced the influence of net disturbance on subsequent samples. Invertebrate samples were emptied from the net and preserved in 90% ethanol. Samples were then processed, counting all invertebrates and identifying to the lowest practical taxonomic level. In some cases, samples were so dense that counting the entire sample was not feasible. In these cases, samples were run through a Folsom plankton splitter and a subset of the sample was counted instead. Invertebrates were collected throughout the sampling season, but only data from dates when īnanga were sampled was used in analysis.

4.3.5 Data processing and statistics

Otolith data was processed to convert daily increment measures into a cumulative growth measure. The number of daily increments was used to back-calculate approximate hatch date (date at core) from the date of capture. Otolith width was plotted against time in a linear model, to calculate the slope coefficient for each fish, which is then used as a growth coefficient (of µm/day) for subsequent analysis. Invertebrate community data from dates where īnanga were collected was converted into proportions, after any groups deemed too large for īnanga to be able to consume (for example adult damselflies) were excluded from the dataset. Gut content data was also converted into numeric proportions to allow for diet analysis.

ANOVAs were run to analyse variation in numbers of post-metamorphic īnanga, separately investigating spatial trends (between sites) and temporal trends (between dates). This pair of ANOVAS were also run for the total number of galaxiids caught (īnanga and whitebait combined).

Population dynamics were analysed using Pearsons correlations, examining relationships between age, length, otolith radius, gut content parameters (total number of food particles and number of microplastics), otolith core date, and otolith growth coefficient being investigated.

A mixed effects model was used to identify trends in otolith growth, and the influence of water temperature. The random intercepts model was constructed with otolith width (as a sum of the daily increments to the core for each date) as the response variable, date and water temperature for each date as explanatory variables, and fish ID as a random effect factor.

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Redundancy analysis (RDA) was used to analyse the environmental conditions, allowing visualisation of the spatial and temporal variation (between sites and between sampling dates). This was followed by PERMANOVA to test the significance of the observed differences. A particular focus was placed on the sites and dates when īnanga were caught to identify any differences with those where they were not caught.

Diet selectivity analysis was run using the WidesII function from the R package adehabitatHS. This computes resource selectivity scores from the resources used by individuals (invertebrates consumed) and the resources available to those individuals (invertebrates in the pond), and tests the hypothesis that resources are used in proportion to their availability.

4.4 Results

4.4.1 Environmental patterns

4.4.2 As previously described in Chapter 3, significant differences in the environmental conditions

between sites were observed (F(5,46)=2.58, p=0.004), this difference appears to be primarily

driven by water nutrient levels (NO2/NO3, ammonia, TN and TP). The conditions at Site 1 fall differently to the majority of the other sites in the RDA plot, indicating its influence on the spatial variation (Fig. 4.3). Collectively the environmental conditions showed no detectable

variation over time (F(7,46)=1.4, p=0.13), but individually both water level (F(7,39)=8.01, p<0.001)

and average daytime temperature (F(7,39)=42.87, p<0.001, ) did show a significant change.

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Fig. 4.3: Redundancy analysis plots displaying the environmental condition polygons for each of the sites and the direction of influence of the individual environmental parameters, highlighting site 1 where all but one inanga were caught. p=0.001, n=9 per site

4.4.3 Population dynamics

In total 84 galaxiids were caught, of these 49 were identifiable as īnanga and the rest were classified as whitebait as they were too small to be definitively identified in the field. Length of īnanga caught in the field ranged between 47mm and 135mm, with an average of 76mm; the subset of fish collected for analysis ranged from 47mm to 83mm with an average of 60mm.

There was significant variation in both number of īnanga and total galaxiids between sites

(F(5,48)=9.26, p<0.0001and F(5,48)=3.95, p=0.004 respectively) but no variation observed over time

(F(8,45)=0.5, p=0.8, and F(8,45)=0.74, p=0.6). All fish captured were caught at site 1 with the exception of a single large īnanga caught at Site 5 on March 10th, maximum catches for both īnanga (9 fish) and whitebait (25 fish) occurred on October 24th.

Otolith analysis showed fish were aged between 73 and 141 days when caught. By back calculating, the date at age 0, ranged from 4 September to 11 November (Table 4.1). Length and age were significantly correlated (p=0.02, R2=0.3).

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Table 4.1: Demographics for the 17 īnanga captured for individual analysis Fish ID Length Date Age (days) Date at core Otolith growth (mm) caught coefficient (µm/day) 8-1A 54 08-01-19 83 17-Oct-18 4.577 8-1B 60 08-01-19 120 10-Sep-18 3.65 8-1C 47 08-01-19 73 27-Oct-18 4.84 8-1D 50 08-01-19 89 11-Oct-18 3.485 8-1E 54 08-01-19 120 10-Sep-18 3.494

23-1A 69 23-01-19 127 18-Sep-18 4.205 23-1B 68 23-01-19 104 11-Oct-18 3.918

23-1C 57 23-01-19 79 05-Nov-18 5.331 52 23-01-19 79 05-Nov-18 4.089 23-1D 23-1E 79 23-01-19 118 27-Sep-18 4.258 52 23-01-19 141 04-Sep-18 3.428 23-1F 23-1G 65 23-01-19 113 02-Oct-18 4.332 3-2A 55 03-02-19 105 21-Oct-18 3.457 3-2B 58 03-02-19 100 26-Oct-18 3.476 3-2C 56 03-02-19 87 08-Nov-18 4.796 11-3A 66 11-03-19 120 11-Nov-18 4.394 11-3B 83 11-03-19 129 02-Nov-18 4.481

Fish showed individual variability in their otolith growth pattern, with notable differences in growth rate (Fig. 4.4, Table 4.1). Otolith growth coefficient was significantly correlated with otolith core date, fish with an earlier otolith core date have a lower growth rate than fish with a later otolith core date (p=0.019, R2=0.31, Fig. 4.5). Daytime average water temperature at Site 1 ranged from 7.6℃ to 24.5℃ with an average of 16.4℃, and showed an increasing trend towards summer before beginning to cool in March (Fig. 4.6). Average daytime water temperature showed a significant positive relationship with otolith growth (F(1,1764)=824, p<0.001, R2=0.318), as both are higher later in the season.

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Fig. 4.4: Otolith growth trajectories (otolith width over time) for each of the 17 inanga, and for all fish together, the red line indicates the fitted slope for each individual and in the final panel the overall model slope for all fish.

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Fig. 4.5: Otolith growth coefficient plotted against date at the otolith core (approximation of metamorphic date). n=17, p=0.019, r=0.56

Fig. 4.6: Average daytime water temperature plotted against date. n=188 days

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4.4.4 Diet

All guts contained food particles, with the total number per individual ranging from 8 to 129 pieces. A large proportion of gut content items were too digested to be identified. What was identifiable was dominated by chironomid larvae, but also with notable amounts of Chydoridae (Fig. 4.7). In comparison, the invertebrate community at Site 1 was dominated by Daphnia, with high numbers of cyclopoid copepods, chironomid larvae, and Bosmina also caught.

Analysis of diet showed no evidence that īnanga were utilising the food resources in proportion to their availability (p=0.47, 0.52, 0.38, 0,96 for each of the dates īnanga were collected), in other words, they are feeding selectively. Of the invertebrates found in the pond system during the time īnanga were sampled, selectivity preference was shown for Chydoridae and chironomid larvae, as well as for ostracods, amphipods and microplastics (Table 4.2), although selectivity varied notably between individuals .

Of the 17 fish sampled, 12 contained at least one piece of microplastic, with one fish containing ten pieces. Microplastic consumption is not correlated with total number of food particles consumed (p=0.98, df=15, t=-0.018). It also does not appear to be related to otolith growth coefficient (p=0.5, df=15, t=0.68) or individual length (p=0.45, df=15, t=0.77)

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Table 4.2: Output from WidesII selectivity analysis, indicating diet preference of īnanga. Only taxa which were found in gut contents are included in table. Available values indicate the proportion of the assemblage as sampled at Site 1, , and used values indicate the proportion of gut content made up by each taxa. Red text indicates confidence intervals that do not include zero, meaning they are considered significant.

Taxa Available Used Selectivity score Confidence interval January 8 Amphipod 0.0001 0.0334 333.747 -649.57 1317.07 5 guts sampled Chydoridae 0.0001 0.0121 121.362 -244.19 486.92 Chironimid 0.0069 0.3255 47.033 -26.79 120.85 Austrolestes 0.0017 0.0121 7.012 -14.11 28.13 Microplastic 0.0602 0.3118 5.177 -3.39 13.75 Daphnia 0.6402 0.3051 0.477 -0.60 1.55 January 22 Chydoridae 0.0001 0.0381 380.925 -749.49 1511.34 7 guts sampled Chironimid 0.0152 0.3023 19.955 -2.77 42.68 Microplastic 0.0321 0.5368 16.733 1.70 31.77 Austrolestes 0.0009 0.0132 14.765 -29.05 58.58 Ostracod 0.0009 0.0038 4.274 -8.41 16.96 Physa 0.0356 0.0132 0.369 -0.73 1.46 Daphnia 0.7637 0.0926 0.121 -0.10 0.35 February 2 Ostracod 0.0001 0.0359 359.179 -600.20 1318.56 3 guts sampled Chydoridae 0.0001 0.0359 359.179 -600.20 1318.56 Chironimid 0.0518 0.8057 15.562 12.81 18.32 Potamopyrgus 0.0035 0.0320 9.279 -17.80 36.36 Microplastic 0.0414 0.0359 0.867 -1.45 3.18 Physa 0.0207 0.0107 0.516 -0.99 2.02 Daphnia 0.7076 0.0439 0.062 -0.14 0.26 March 10 Chironimid 0.0084 0.5294 62.982 -48.94 174.90 2 guts sampled Chydoridae 0.0056 0.1176 20.994 -52.00 93.98 Ostracod 0.0196 0.0588 2.999 -7.43 13.43 Hydroptilidae 0.0252 0.0588 2.333 -3.34 8.01 Microplastic 0.0614 0.1176 1.917 -4.75 8.58 Daphnia 0.3783 0.1176 0.311 0.15 0.47

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1.2

1

0.8

0.6 Proportion of sample

0.4

0.2

0 Pond Gut

Daphnia Cyclopoid Gyralis Chironomid Bosmina Amphipod Chydoridae insect leg Microplastic Other

Fig. 4.7: Components of the pond invertebrate community compared to the gut content from inanga caught in the pond over the four sampling dates where inanga were retained. (n=9 pond samples, n=17 gut samples)

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4.5 Discussion

The presence of young īnanga in only one pond provides an interesting insight into habitat usage within this system, and shows a clear spatial pattern as was hypothesised; although it is acknowledged that this limited distribution does reduce the wider application of findings. The pond where īnanga were found has the least connectivity to the main river, and the greatest distance from the intake. This results in Site 1 having a different water quality to much of the rest of the system, with lower levels of NO2/NO3, likely due to the filtering effect of the wetland (Chapter 3). In addition, over the duration of the study no trout were caught at Site 1. Trout are known to pose significant predation risk to galaxiids, as well as creating competition for food resources, with their presence often being found to have a negative influence on abundance of stream dwelling galaxiids (McIntosh, Townsend, and Crowl 1992; McDowall 2003; McIntosh et al. 2010). The complete drying of this pond during the previous summer, with no subsequent flood events significant enough to overtop between ponds, means that it is unlikely for any large īnanga predators to have successfully accessed the pond. This pond is also the nearest to the lagoon itself, with the least complex connectivity to the lagoon, where incoming whitebait may be recruiting from. The combination of connectivity, environmental conditions and lack of competition and predation risk may explain the presence primarily in this pond compared to others in the system. The lack of a temporal pattern in abundance as was predicted is likely due to the small number of fish caught and their limited distribution. More intensive sampling would likely improve the clarity of any patterns both temporal and spatial within this system.

4.5.1 Age and growth The range of dates at otolith core were unexpected as they are indicative of spring hatching rather than the classical dates that would be expected following an autumn spawning period of March to May (Taylor 2002; Smith 2014). While spawning has not been observed within the wider Waiau River area, a survey of other potential spawning sites around Southland has observed spawning beginning in February and continuing through June (Hicks, Leigh, and Dare 2013). It is not known the origin location of the fish from this study, but the age range observed indicates that spawning occurs over an even broader time scale than has currently been observed in Southland rivers. īnanga. This spring hatch pattern has also been observed by Egan et al. (2019) who found that there was notable variation in growth rate patterns between īnanga hatched over different seasons, although not in īnanga populations as far south as this one. Surveys for spawning activity at

59 potential spawning sites along the Waiau River mouth would be valuable to corroborate these findings.

The age of fish sampled in this study are notably younger than a previous study in this area (McClintock 2018). The previous study caught whitebait upon entering the Waiau River mouth, and found that ages of fish caught were on average 160-170 days, compared to the fish in this study caught within the Waiau ponds themselves which did not exceed 130 days. This significant difference creates some interesting questions about the origin of the fish within the ponds, in particular, could the pond system contain a non-diadromous population? Diadromy is not considered an obligate trait for īnanga, with non-didromous populations typically using lakes as rearing habitat (Gorski et al. 2018). These non-diadromous populations can display differences in phenotypes that distinguish them from migratory populations (Rojo et al. 2020; Cussac et al. 2020). To explore this hypothesis would require analysis of isotopes within the otoliths of īnanga from this pond system to determine if they contain evidence of a marine signature or lack thereof.

4.5.2 Diet

Feeding behaviour of pond dwelling īnanga, as observed in this study, has not been studied previously in New Zealand, and the consumption of microplastics has also not been observed, although I acknowledge that only a small sample size was analysed. Īnanga in streams show different feeding behaviour, relying on drift feeding as their foraging method (Jowett 2002). Drift feeding is common for many galaxiid species (Cadwallader 1975; Hayes 1996; David and Closs 2003), but it is not a suitable feeding method for ponds as there is not sufficient flow. In this system, īnanga show selective preference for Chydoridae and chironomid larvae, which make up only a small proportion of the invertebrate community. Chydorids and chironomids are primarily benthic, indicating that īnanga are likely actively foraging for desirable prey species within the benthos. By contrast Daphnia, a pelagic taxon which was highly abundant in the invertebrate community was in much lower numbers in the gut content, indicating that they are likely not a preferred food source. This contrasts with the original hypothesis, expecting to see non-selective feeding within the appropriate size class as is observed in stream feeding populations.

A study by Pollard (1973) on the diet of G. maculatus in a lake in south-west Victoria (Australia) found that amphipods made up majority of the diet followed by chironomid larvae and copepods. That study also found that both chironomid larvae and amphipods, as well as large ostracods were

60 selected for in relation to the lake community proportions. Studies based in a small lake system in Chile found that diet of juvenile G. maculatus was predominantly planktonic, dominated by Bosmina, harpaticoid copepods and cyclopoid copepods (Cervellini, Battini, and Cussac 1993; Modenutti, Balseiro, and Cervellini 1993). The diet observed in a stream dwelling population showed selectivity for Simuliidae and Plecoptera, although chironomids were also a major component (Tagliaferro et al. 2015). The diet observed in the Waiau pond system is not entirely consistent with observations from any of these studies, but does show some similarities with both the small and larger lake populations.

The presence of microplastic fibres within the pond system and particularly in the gut of these fish was unexpected. This pond system is reasonably remote, with no large urban areas within the catchment of the Waiau River, highlighting the pervasiveness of microplastics throughout our waterways. Microplastics are an increasingly recognised threat to many freshwater systems globally (Horton et al. 2017; Wagner et al. 2014), as well as to marine and terrestrial systems (Cauwenberghe et al. 2013; Ivar du Sul and Costa 2014; Wagner et al. 2014; Horton et al. 2017). These small plastic particles can be consumed by a range of species allowing them to enter the food chain, and have a range of impacts on growth, reproduction, and survival for individuals that consume them (Eerkes-Medrano, Thompson & Aldridge 2015; Jovanovic 2017; Foley et al. 2018; Anbumani and Kakkar 2018). The source of the microplastics observed in these īnanga guts was not determined; however, the colours of the fibres were not consistent with any of the sampling gear used in the ponds, and therefore cross contamination can be ruled out with some certainty. Microplastic fibres are abundant in outflows from wastewater treatment plants, and have also been observed as airborne particles (Prata et al. 2020; Gaylarde, Baptista-Neto & da Fonseca 2021). Within the Waiau River there are several wastewater management schemes, Tuatapere (population 550) has a wastewater discharge that is a potential nearby source of microplastic fibres, the towns of Te Anau and Manapouri (combined population 3000) are also upstream, although at a far greater distance.

The significant proportion of juvenile īnanga that consumed microplastics provides even greater cause for concern, particularly since the fish indicate a selective consumption for the fibres. Unfortunately with this style of direct observation, it cannot be determined if the microplastics are at a higher level in the gut because they are being actively selected by the fish, or because they are accumulating within the gut; although studies suggest that microplastics can pass relatively easily through the digestive tract of some fish species (Critchell and Hoogenboom 2018; Foley et al. 2018). A future laboratory-based experiment would be valuable to understand the feeding 61 behaviour and digestive ability of īnanga in respect to these microplastic fibres, and to identify what impacts plastic consumption has on energy levels, body condition, or survival rate.

4.6 Conclusions

The age of fish found indicates a spawning and hatch period outside of the typical time period for īnanga, this finding along with the relationship between growth rate and otolith core date indicates the importance of spawning timing for juvenile īnanga. The diet and feeding habits of īnanga within this small New Zealand pond system show similarities to diet studies from larger lake systems elsewhere in the Southern Hemisphere, and as expected are notably different from stream-based studies. The presence of microplastics both in the pond system and within the gut of the fish provides further evidence of the pervasiveness of this environmental pollutant. This leads to a range of further questions on the origins and impacts of these microplastic fibres within this system.

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Chapter 5. Recolonisation of a fish and invertebrate community in a wetland following a drought: the importance of deep water refugia.

5.1 Abstract

Wetland restoration or re-establishment is a common conservation goal, and the ability of a wetland community to recover from events such as drought is key to its long-term success. At the mouth of the Waiau River in Southland, New Zealand, a series of wetland ponds have been created by diverting water from the river. A period of particularly low river flows resulted in the wetland drying out entirely, except for a few deep channels. A subsequent flood inundated the ponds, creating an opportunity to study recolonization of the system. The deep channels provided an important refuge habitat for eels (Anguilla spp.) during dry periods, with significantly higher numbers captured close to the channels immediately after refilling than at sites further away. Smaller fish (Gobiomorphus and Galaxias spp.) took much longer to reappear in the ponds. The overall macroinvertebrate community showed no significant variation over the short term or up to a year post-refill, although individual snail and water bug taxa increased and microcrustaceans decreased in abundance. Daphnia and ostracods were observed soon after refilling and dominated the community. This study highlights the importance of spatial heterogeneity in constructed wetland systems, as this habitat variation likely increases the recovery rate after drying.

5.2 Introduction Wetlands are an important ecosystem found throughout temperate, tropical and even polar regions of the world. They are defined as an ecosystem of water-inundated soils dominated by anaerobic processes, which supports biota adapted to flooding or submergence (Keddy 2010). Wetlands can exhibit varying patterns of permanence, and can be permanently inundated, rhythmic (displaying seasonal changes between wet and dry), or arrhythmic (intermittently or rarely inundated with water, often as a result of high levels of precipitation) (Deil 2005). Non-permanent wetlands are common throughout the world, particularly in areas that have a wet and dry seasonality. Many non- permanent wetlands, particularly those that are only periodically wet, have aquatic communities dominated by species that are adapted to survive dry periods. These adaptations include traits such as aestivation within the drying substrate, or an ability for individuals or eggs to withstand desiccation (Brock et al. 2003; Strachan et al. 2014). Other species that have not specifically evolved 63 to withstand drying can also be successful in these systems through a high level of mobility and an ability to rapidly colonise newly inundated areas and reproduce before future drying occurs (Galatowitsch 2014, Frouz, Matena & Ali 2003).

The recolonisation of previously dry wetlands has been studied in several locations, particularly in areas that have predictable dry and wet periods (Boix et al. 2004; Culioli et al. 2006; Pazin et al. 2006; Florencio et al. 2009; Peterson et al. 2017). Typically, community succession occurs as new species establish over time (Boix et al. 2004; Culioli et al. 2006; Peterson et al. 2017). In areas where wetlands are more stable and drying infrequent, the recolonisation process has been less studied, largely due to the unpredictability of drying events. In most cases data comes from the colonisation of artificially created ponds rather than the recolonisation of existing ponds that have become uninhabitable for a period of time (Williams et al. 2007; Jeffries 2011).

Non-permanent or ephemeral waterways are uncommon throughout much of New Zealand due to the relatively wet maritime climate, and the drainage of land to allow development of many areas for agriculture (McGlone 2009). Extended periods of unusually low rainfall, or increased water abstraction, can lead to the drying of systems that do not typically dry (Brinson and Malvárez 2002). Compared to many other regions with ephemeral wetland systems, New Zealand does not have a large assemblage of specialist fauna adapted to regular drying events, but rather opportunist species able to utilise both permanent and non-permanent waterways (Wissinger et al. 2009, Galatowitsch & McIntosh 2016). While studies of invertebrate communities in pond systems have seldom been done in New Zealand, there is some existing knowledge primarily from experimental studies, examining the effects of drying and other environmental stressors (Maxted, McCready & Scarsbrook 2005; Wissinger, Greig & McIntosh 2009; Greig, Wissinger & McIntosh 2013)

Representatives from the majority of New Zealand’s aquatic invertebrate and fish groups can be found in both permanent and ephemeral systems, with those found in ephemeral systems typically being species that are able to rapidly recolonise from dormant eggs or nearby permanent water bodies (Barclay 1966). Invertebrate species that are most successful have a variety of different strategies. For example, Ostracoda and Daphnia produce desiccation tolerant eggs which remain within a drying pond while Coleoptera and Diptera have highly mobile adults that disperse and lay eggs once habitat is available (Barclay 1966; Crumpton 1979; Hairston 1996; Strachan et al. 2015). Some fish species such as eels (Anguilla spp.) have a relatively high tolerance for poor conditions including low oxygen levels (Dean and Richardson 1999). This ability may enable them to quickly

64 recolonise a dried waterway before it is suitable for other species, if there is some connection to a permanent water body. Eels are also highly mobile so this connection could simply be an area of damp field that they can travel across during the night (Sayer 2005).

Many organisations within New Zealand have developed restoration or re-establishment projects for wetland areas. In part, this is to counteract the rapid nationwide loss of wetlands (Ausseil et al. 2008), but is also because of the wide range of ecosystem services and benefits provided by an established wetland system (Johnston 1991; Zedler & Kercher 2005; Gedan et al. 2011; Hansson et al. 2005). For these projects to succeed they must be capable of withstanding drought conditions or be designed to facilitate recolonisation after a drought has occurred. This is particularly important in the early years of a wetland project, as it takes time for vegetation to establish and for pond beds to accumulate the clay layer that improves water retention (Auckland Regional Council 2003). One feature that has potential to aid in both drought resistance and recolonisation is the presence of deep channels, potentially with connections to ground water or to permanent water bodies. This could decrease the probability of a wetland drying completely and provide a refuge from which recolonisation can progress.

During the summer of 2017-2018, a period of drought led to the near-complete drying of a re- established wetland complex at the mouth of the Waiau River in Southland. The drought was followed by a flood that inundated the dry pond beds. These events provided a unique opportunity to study how fish, macroinvertebrates, and zooplankton recolonise a complex pond system. We predicted that there would be an increase in abundance over time, as taxa utilised the newly available space and resources, as well as an increase in taxonomic diversity as community structure re-established. We also expected that fish and other taxa with limited dispersal ability would first appear in areas close to the deep channels in the pond system, which were predicted to act as refugia.

5.3 Methods

5.3.1 Study site and design

At the mouth of the Waiau River in Southland, a series of wetland ponds have been re-established in an area previously used for agriculture. (Fig. 5.1). The wetland system was constructed in three stages over a 6-year timespan (Whitehead ponds: 2009, McCulloch ponds: 2012, Inder ponds:

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2015), and is now connected to allow water flow between all component ponds. The Waiau River has been highly modified, with a control structure built across it in 1977 to divert river flows into Lake Manapouri to maximise hydroelectric generation potential. The control structure maintains the lake level within a consented range and requires a minimum outflow of 16 m3/s. This has resulted in a significant decrease in Waiau River flow from the historic average of 400-550 m3/s (Ellis & Palliser 2019, Riddell & Sutton 2014). The Waiau pond system has been constructed as a means of mitigating the environmental impacts of the hydro-generation scheme; the reduction in flow causes restriction of the river, decreasing connectivity across the flood plain. The pond system is formed by a series of inter-connected underground pipes and fed via two diversion channels that take water from a side braid of the river. Water is retained within the system by several retaining walls, and then drains into Te Waewae lagoon at the mouth of the Waiau River. Due to the position and connectivity of the diversion channel intakes, water flow into the wetlands reduces significantly when Waiau River discharge drops below 50-60 m3/s.

Fig. 5.1: Location of study sites within the Waiau pond system on the Southland coast. Highlighting the location of key hydrological features and structures. Six sites were selected for sampling, spread across the key ponds in the system, sites 1 and 2 are in the Whitehead pond suite, sites 3 and 5 are in the main Inder pond with site 4 in the diversion channel, and site 6 is in the main McCulloch pond.

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An unusually dry year in 2017 culminated in an 80-day period between November and January where discharge did not exceed 60 m3/s. As a result, the majority of the pond system dried out, except for a deep, well-vegetated channel located along an old river braid. Following this extended period of drought, a significant rainfall event and subsequent flood of the Waiau River occurred. This resulted in a water level high enough to inundate the pond system.

A single deep channel (approximately 2 m deep) runs through the Inder pond system, and this channel continues into the McCulloch system but becomes much shallower (see Fig. 5.1). Water depth throughout the rest of the system is reasonably consistent with a maximum depth of approximately 1 to 1.2 m. A siphon is located within the main Inder pond which feeds water through into the Whitehead pond system on the other side of the Holly Burn.

We selected 6 sites within the Waiau pond system to represent the variation within the system, including proximity to the deeper refuge channel, while also taking into consideration ease of access (Fig. 5.1). Distance from the deepest point of the refuge to each sampling site ranged from 22 to 935 m (Table 5.1). We measured distance, on an aerial image in the programme IC measure (from: www.theimagingsource.com), as the most direct route through waterways (utilising the underground pipe network), rather than in a straight line, to best represent the connectivity for aquatic taxa.

Table 5.1: Distance from the deep water refuge for each of the six study sites.

Site Distance from refuge (m)

1 935 2 574 3 448 4 472 5 22 6 567

We sampled weekly for 1 month following the inundation of the ponds on 2 February 2017), then fortnightly until the system stabilised, completing sampling on 4 May 2017. This was determined by all ponds being full at a stable level and macrophytes becoming established. We also sampled 1 year after the ponds refilled on 3 February 2018, to provide a comparison to a summer pond community

67 most likely little impacted by drought. We collected a zooplankton sample, a macroinvertebrate sweep sample, and surveyed the fish population at each site. We also recorded the abundance of macrophytes immediately after refill on 2 February 2017, and again on 4 May 2017

5.3.2 Fish populations We set 1 unbaited fyke net (10mm mesh, 0.5m high, 0.65m wide, 4.5m wing) and 1 Gee minnow trap (3mm mesh, 20mm opening) overnight at each of the 6 sites. Fyke nets were positioned perpendicular to the shoreline with the leader as close as possible to the pond edge as the gradient allowed. Gee minnow traps were submerged near macrophytes if they were present. The following morning, we identified, counted, weighed and measured the length of fish before returning them to the water.

5.3.3 Macroinvertebrate samples We collected 1 macroinvertebrate sample at each site by sweeping a fine mesh net (660 µm) back and forth for 30 seconds, along the margin of the pond and near macrophytes if present. Any large pieces of macrophyte collected were removed and rinsed over the net, and the remaining contents then preserved in 80% ethanol. Invertebrates in each sample were later counted and identified to the lowest practical taxonomic level using Chapman et al. (2011), Winterbourn et al. (1989) and NIWA invertebrate identification guides. In some cases, where samples were too large to count within a reasonable time frame, we used a plankton splitter to subsample which was counted and multiplied to estimate abundances in the whole sample. Additionally, we inspected the entire sample for larger rare taxa to ensure these were not missed by the fractioning of the sample. The fraction of sample counted varied between 1/2 and 1/8 and was determined by density in the counting tray – individuals needed to settle in a single layer to ensure accurate counts, this resulted in a maximum of 14,000 individuals counted for a sample.

5.3.4 Zooplankton samples We filtered 20 L of open water through a 250um plankton net at each site. The plankton net was then rinsed, and the sample emptied into a pottle and preserved in 10% formalin. Invertebrates were counted and identified to the lowest practical taxonomic level using Chapman et al. (2011), Winterbourn et al. (1989) and NIWA invertebrate identification guides.

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5.3.5 Data processing and statistics All data analysis was conducted using the statistical programme R (R Core Team 2019). Fish species were grouped into eels and non-eel taxa. We used two-way ANOVAs to test whether eels and non- eel abundances varied with distance from refuge and time since refill (in months). Fish data was averaged to give monthly catch levels (three sampling trips per month). We calculated summary statistics for abundant macroinvertebrate and zooplankton taxa. We then combined data from the macroinvertebrate and zooplankton samples, giving overall abundances, for analysis of invertebrate community composition because there was clear overlap in taxa. We used non-metric multidimensional scaling (NMDS) with Bray-Curtis dissimilarity to explore how invertebrate communities varied both spatially (between sites) and temporally (between sampling dates). We used a permutational ANOVA (using function adonis in R package vegan) on the matrix produced in the NMDS to test for differences between sites and over time (Oksanen et al. 2019).

5.4 Results During the initial 3-month study period (2 February 2017 – 4 May 2017), the Waiau River experienced seven instances where river flows exceeded 200 m3/s, but all were less than 450 m3/s. This means that the pond system was supplied with influxes of water on a regular basis without being so high as to inundate the system completely. Macrophyte cover at each site increased notably over time (Table 5.2).

Table 5.2: Braun-Blanquet score for macrophyte cover at each site immediately after refill and at the end of the initial sampling period. Initial score includes vegetation (e.g. grasses) that became submerged during inundation. Braun-Blanquet scores are: r-<5% (few individuals), 1-<5% (many individuals), 2-5-25%, 3-25-50%, 4-50-75%, 5-75-100% coverage

Site One week post refill 3 months post refill 1 3 5 2 1 3 3 r 2 4 2 2 5 r 2 6 2 5

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Fish numbers were generally low throughout the sampling period (Fig. 5.2, Table 5.3). The catch was dominated by shortfin eels (Anguilla australis), with longfin eels (A. dieffenbachia), common bullies (Gobiomorphus cotidianus), īnanga (Galaxias maculatus) and brown trout (Salmo trutta) also caught.

Eel numbers differed significantly by distance from a deep-water refuge (F(1,46)= 10.16, p=0.0026 ) but non-eel fish numbers did not (F(1,46)=0.029, p=0.866 ). A significant interactive effect between time and distance from the refuge was observed during the initial sampling period for both eels

(F(1,44)=11.53, p=0.001) and non-eel fish (F(1,44)=5.87, p=0.019), although the direction of this effect differed (Fig. 5.2, Table 5.3). Eels had highest abundance near the refuge in the month immediately after refilling, with a more even distribution across the system in the second month and but few were caught in the third month. Non-eel fish had very low abundance throughout the system in the first two months and a large increase in common bully and īnanga abundance far from the refuge by month three. In general, non-eel fish tended to have higher numbers at locations where eel numbers were lower. Fish numbers a year post refill reflected a similar pattern to the end of initial sampling with high numbers of non-eel fish present furthest from the refuge.

Table 5.3: Average fish catch numbers over the first three months of sampling post refill of the system (sampling occurred 3 times in months 1 and 2, and 2 times in month 3)

Distance from First month post refill Second month post refill Third month post refill refuge (site) Total Eels Non-eels Total Eels Non-eels Total Eels Non-eels 22 (5) 7 6.66 0.33 4.33 4 0.33 2 0.5 1.5 448 (3) 2.33 2.33 0 4.33 4.33 0 0 0 0 472 (2) 5 4 1 2 2 0 3.5 3 0.5 567 (6) 0.33 0.33 0 1.67 1.67 0 0.5 0.5 0 574 (2) 0 0 0 2.67 1.33 1.33 0 0 0 935 (1) 0.33 0.33 0 1.33 1.3 0 3.5 0.5 3

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Fig. 5.2: Trends in fish catch numbers by month post-refill in relation to distance from the refuge channel, separated into eels and non-eel fish groups. There were three sampling occurrences in months 1 and 2 and two in month 3

In total, 32 taxa were collected in the macroinvertebrate sweep samples. The species richness of individual samples varied from 2 taxa observed at Site 4 on 27th March and 10th April to 18 taxa observed at Site 1 on 14th March (40 days after refilling). The five most frequently observed taxa were Daphnia, ostracods (Cypridinidae), cyclopoid copepods, water boatmen (Sigara sp.), and Physa snails (Table 5.4). Daphnia and ostracods were also the most abundant taxa across all samples.

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Table 5.4: Abundance and prevalence statistics for the most common macroinvertebrate taxa found in the sweep net samples across all sites. Difference after one year symbols: += increase in proportion, -=decrease in proportion

Taxa Total Prevalence (out Maximum Date of Proportion of Proportion Difference abundance of 48 samples) abundance maximum individuals after 1 year after 1 year Daphnia 193236 45 113688 14-03-18 0.843 0.615 - Ostracod 19834 40 6872 10-04-18 0.087 0.002 - Physa 5532 34 2273 14-03-18 0.024 0.179 + Potamopyrgus 2121 16 626 14-03-18 0.009 0.009 + Sigara 1428 34 227 14-03-18 0.006 0.033 + Gyraulus 1230 21 339 10-04-18 0.005 0.083 + Cyclopoid copepod 897 35 192 04-03-18 0.004 0.001 - Anisops 715 30 96 04-03-18 0.003 0.066 + Hydroptilidae 397 20 87 07-02-18 0.002 0.002 - Chironomid larvae 339 21 96 03-05-18 0.001 0.003 +

In total, 22 taxa were observed in the zooplankton samples, with the maximum diversity of 12 taxa observed at Site 1 on 14th March, and the minimum diversity of 1 taxon observed at Site 4 on 27th March. The five most frequently observed taxa were Daphnia, ostracods (Cypridinidae sp.), Bosmina, harpaticoid copepods and cyclopoid copepods; these were also the most abundant across all samples (Table 5.5).

Table 5.5: Abundance and prevalence statistics for the most common taxa caught in zooplankton samples across all sites.

Taxa Total Prevalence Maximum Date of Proportion of Proportion Difference abundance (out of 48 abundance maximum individuals after 1 year after 1 year samples) Daphnia 4667 45 1721 14-03-18 0.358 0.939 + Ostracod 3878 27 2040 03-05-18 0.298 0.002 - Bosmina 2713 37 1448 03-05-18 0.208 0.002 - Harpaticoid 1166 34 372 03-05-18 0.089 0 - copepod Cyclopoid copepod 177 27 64 14-03-18 0.014 0 - Chironomid larvae 108 15 76 03-05-18 0.008 0.012 + Calanoid copepod 48 11 33 14-03-18 0.003 0.002 - Physa 47 12 15 14-03-18 0.003 0.006 +

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There were two peaks in zooplankton numbers that occurred during the initial study period (referred to from here onwards as ‘blooms’). These blooms were not statistically significant occurrences as they were not observed across the entire system, however they were notable events at specific sites (Fig. 5.3). The first bloom was comprised of only Daphnia, the second bloom was comprised of ostracods, Bosmina, and to a lesser extent Daphnia.

Fig. 5.3: Trends observed in zooplankton abundances during the initial sampling period, illustrating the two bloom events that occurred, with a breakdown into (A) total abundance at each site and (B) key taxa across all sites.

Combined analysis of the total abundance from the macroinvertebrate sweep and zooplankton

samples showed that the invertebrate community varied significantly between sites (F(1,53)=2.81, p=0.004) with Site 1 (the furthest from a deep water refuge) supporting the most distinctive community, driven by an abundance of snail taxa. Site 4 (the diversion channel) was also relatively distinct, while the remaining four sites clustered together (Fig. 5.4). There was no significant

temporal variation in the community composition (F(8,53)=0.79, p=0.92). Although individual taxa did exhibit marked variation in abundance and proportion of the community they comprised over time (Table 5.3 & 5.4, Fig. 5.5), although this change in abundance was only significant for Gyraulus

and Physa snails (F(4,49)=2.83, p=0.033 and F(4,49)=3.39, p=0.016 respectively), with chironomid larvae (p=0.074)and Anisops sp. being near significant (p=0.056). The direction of change in

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abundance differed depending on the functional groupings they fell into (Table 5.6), with the smaller, typically planktonic species, such as copepods, declining over time, and the larger, often epibenthic taxa, such as snails, increasing over time.

Table 5.6: Functional groupings for the most abundant macroinvertebrate and zooplankton samples, including habitat preference, dispersal potential (low:10 m, medium: 1km, high:>1km), feeding habits, dietary preference and reproductive potential (R1:<100 offspring per reproductive cycle, R2:100-1000 offspring, uni: univoltine, pluri: plurivoltine). Categorisations from NIWA macroinvertebrate trait database (Philips & Smith 2018).

Taxa Habitat Dispersal Feeding habits Dietary preference Reproductive potential potential Anisops Open water high Predator Specialist R1: pluri Bosmina Open water Low-medium Filter feeder Generalist R1: pluri Calanoid Open water/ low Filter feeder/ Generalist R1: pluri copepod epibenthic predator/scraper Chironomid Epibenthic/ Medium-low Scraper Generalist R1-R2: pluri larvae infaunal Cyclopoid Open water/ low Filter feeder/ Generalist R1: pluri copepod epibenthic predator/scraper Daphnia Open water Low-medium Filter feeder Generalist R1-R2: pluri Gyraulus Epibenthic Low-medium Scraper Generalist R1: pluri-uni Harpaticoid Open water/ low Filter feeder/ Generalist copepod epibenthic predator/scraper Hydroptilidae Epibenthic low Scraper/ algal piercer Strong- moderate R2: uni specialist Ostracod Open water/ low Deposit feeder/ Generalist R1: pluri epibenthic predator/scraper Physa Epibenthic Low-high Scraper Generalist R1: pluri Potamopyrgus Epibenthic Low-medium Scraper Generalist R1: pluri Sigara Open water/ high Scraper/predator Moderate specialist- R1: pluri epibenthic generalist

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Fig. 5.4: Bray-Curtis based non-metric multidimensional scaling (NMDS) plot of the combined macroinvertebrate and zooplankton samples (stress= 0.13). Ellipsoids represent a 95% confidence interval for the community of each site, with taxon names indicating the direction of their influence on the community.

Fig. 5.5: Comparison of invertebrate proportions in samples over the first three months and a year after the system refilled. While few of these patterns are significant, the overall trend is a general decline over 12 months for the small, generalist planktonic taxa (A), and an increase in large, more specialised taxa that are primarily epibenthic (B). Significant trends are seen in Gyraulus and Physa snails (p=0.033, F=2.83, d.f=4/49 and p=0.016, F=3.39,df=4/49 respectively); chironomid larvae (p=0.074)and Anisops (p=0.056) were near significant. Samples were collected from 6 sites on 3 occurrences in months 1 and 2, twice in month 3 and once on month 12.

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5.5 Discussion The findings of this study provide insights into the process of recolonisation by both fish and invertebrates in a wetland system inundated following drought. Additionally, several features of this reconstructed wetland appear to have aided the rate in which recolonization occurred, which may be valuable to consider for other wetland conservation projects.

Eel numbers remained relatively constant throughout the sampling period. This most likely indicates that the same eels are using the system, spreading out from the deep channel as more habitat becomes available, and retreating towards the deeper channels when water levels drop. Eels are known to emigrate from ponds when conditions no longer meet their requirements (Cucherousset et al. 2007), and their ability to travel over land allows them to seek out refugia if a pond becomes isolated (Sayer 2005). The lower catch numbers for the last two sample times can be attributed to the beginning of winter, as lower temperatures are associated with a decline in eel activity (Ryan 1984). Dispersal of eels from the deep channel over time indicates that this habitat is not necessarily preferred habitat but does allow persistence within the system. This is highlighted by the significant decline in eel numbers at the refuge between the first and second month. This suggests the deep channel is valuable to the system as it allows rapid recolonisation from within the system rather than relying on colonists from the main river.

Whilst eels were able to rapidly recolonise the pond system, this was not the case for other fish species. Bullies and galaxiids took much longer to return to pre-drought abundances (Chapter 2) with populations only beginning to increase three months after the refilling. This is likely partially because ponds where the highest levels of bullies and galaxiids were found pre-drought are ponds that dried early and entirely. These are also the ponds that had less direct connections to the river, meaning that it was more difficult to re-access; they are closer to the lagoon, however connectivity may be difficult via the underground pipe network. Temperature may also have affected recolonization; non-eel fish that did not recolonise early on are unlikely to recolonise during winter. There is also the possibility that the non-eel fish did not utilise the refuge channel to the same extent as eels due to predation risk from the increased eel density. Fish such as bullies and galaxiids are a key dietary component for larger eels (Cadwallader 1975; Jellyman 1989); this may also explain the lack of smaller eels and elvers, as cannibalism by larger eels is also common (Jellyman 1989).

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There was no significant change in the overall macroinvertebrate community composition during the initial three-month sampling period, and relatively high numbers of invertebrates were constantly observed. This indicates a rapid invertebrate recovery. It is potentially aided by the initial absence, and then slow subsequent recovery of fish populations, particularly, key predators such as īnanga and common bully that feed on the micro-invertebrates which dominated the community composition (Chapter 4, Wilhelm et al. 2007). Invertebrate recovery may also be aided by the rapid recovery of their own key food sources. The invertebrate community composition did show some changes after a year post refill with notable increases in the proportion of water bugs (Sigara and Anisops) and snail taxa, and declines in abundance of the smaller crustacean groups (notably Daphnia and ostracods). This indicates that over a longer time span the smaller, open water generalist taxa are displaced by larger taxa with more specialist diet or habitat requirements.

Overall, Daphnia was the most abundant taxa, most likely due to their reproductive strategies that allow rapid recovery and population growth. Daphnia can produce resting eggs, which lie dormant in dry conditions and hatch once water returns, allowing much faster return than other taxa (Schwartz and Hebert 1987; Hairston 1996). They also reproduce both sexually and asexually, maximising population growth when conditions are favourable (Schwartz and Hebert 1987). There was a significant bloom in cypridinid ostracods later in the sampling period, which was also likely aided by the ability of non-marine ostracods to reproduce sexually and asexually (Butlin et al. 1998). In addition, the timing of the bloom corresponds to increased presence and density of macrophytes in some ponds. This would provide ideal habitat as ostracods are not fully planktonic, requiring some substrate to inhabit and aid in feeding (Chapman 1963). The low numbers of bullies and galaxiids may have also aided in the large blooms observed. Small crustaceans such as Daphnia and ostracods comprise key components of the diets of both īnanga and common bully (Pollard 1973; Wilhelm et al. 2007) Their absence may therefore have left populations released from top- down control, particularly in ponds where high numbers of fish are common under normal conditions.

The results of this study are consistent with findings from previous investigations of invertebrates of non-permanent pond systems in New Zealand. Wissinger et al. (2009), Byars (1959) and Barclay (1966) found that the temporary ponds they studied tended to be dominated by taxa such as microcrustaceans, chironomids, water bugs, and beetles. The Waiau pond system also had high numbers of some of these groups, particularly the microcrustaceans, although beetles were uncommon. Interestingly the temporary ponds studied by Wissinger et al (2009) in inland

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Canterbury, and Barclay (1966) near Auckland, lacked Potamopyrgus, a widespread snail that was one of the more abundant taxa in the Waiau pond system, although they did find other snail taxa present. Those studies also found that odonates were very common in permanent systems but seldom found in temporary ponds, this was consistent with findings in the Waiau pond system where odonates were found in few samples, typically only as single individuals. This pattern can be attributed to the comparatively long larval stage for odonates before the mobile, but short-lived, adult stage can occur and allow dispersal (Conrad et al. 1999).

A study of invertebrates in a non-permanent pond system near Auckland by Barclay (1966) found high abundances of cladocerans and ostracods; and an approximately 5-week delay for some cladoceran taxa to appear in samples. This delay was attributed to time taken for the small number of individuals hatched from resting eggs to mature and begin reproducing. Ostracods and cladocerans were also found in high numbers in the Waiau pond system, and although Daphnia were found in samples immediately, it took approximately five weeks for numbers to bloom. This indicates that a similar pattern of delay occurred in both systems.

Studies of temporary pond systems overseas have shown much stronger successional patterns, particularly for macroinvertebrate taxa (Moorhead et al. 1998; Boix et al. 2004; Nicolet et al. 2004; Culioli et al. 2006). This may be due to regular, predictable wet-dry seasonality of the studied systems. In contrast, temporary pond systems in New Zealand are dominated by generalist taxa that can survive the pressures of rare and unpredictable drying. Additionally, the diversity of New Zealand ponds is relatively low when compared to that of other countries, meaning that a clear succession is less likely to be observed.

5.6 Conclusions This study has allowed several conclusions to be drawn about what makes a reconstructed wetland successful. These insights can hopefully be used to guide future conservation projects to ensure they are as effective as possible. Having deep channels that do not dry easily (potentially connecting to groundwater) provides refuge, allowing for rapid recolonisation following a period of drought. This is particularly true for eels which are more mobile and able to cope with poorer water conditions (lower oxygen and higher nutrient loads) than other fish species. Variability in connectivity and water flow is also valuable as it enables different communities to form within one larger system, creating greater resilience. While drought is never ideal for a typically permanent

78 wetland there are features that can make the impact of drying less severe, if considered during construction or restoration. This is particularly important to consider given the potential for more frequent dry periods and droughts in the future (Mullan et al. 2005).

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Chapter 6. Summary

6.1 Summary of key findings

Overall, this thesis has investigated four separate but related components of the macrofaunal community in coastal wetland systems. Some of the findings are specific to these sites within Southland, for example the distribution of fish species across the four systems surveyed, but much of the information gained has relevance to wetland systems in general.

The survey of Southland wetlands showed wide variability in the number of species and population sizes of fish in the four wetland systems sampled. Variation was expected between systems, but the low catch rates in Lake George and Big Lagoon were not. Within the Waiau pond system and Waituna Lagoon systems the same species – short and longfin eels, common bullys and īnanga dominated, although there were significant differences in both abundance and distribution of these taxa. The occurrence of low water events in both systems during the survey period allowed for investigation into how the systems respond to this type of event, as well as the differences in response between systems. Fish community in the Waiau pond system showed little long-term influence from the drying of the pond system, whereas the fish community in Waituna Lagoon showed significant declines in abundance and total number of taxa that did not recover during the year following the artificial lagoon opening.

Succession of the invertebrate community was also observed during the following summer, with a clear progression from Bosmina and chironomid larvae in spring through Daphnia over summer and on to ostracods and Potamopyrgus snails into autumn. There were also differences observed in the invertebrate community between sites, with some taxa displaying clear spatial patterns. There were significant differences in the overall environmental conditions between sites, but not over time. The overall invertebrate abundance (excluding Daphnia) showed a significant negative relationship with NO2 and NO3 concentration, which showed a pattern of increasing abundance and decreasing NO2 and NO3 as distance from the diversion channel increased.

Īnanga populations within the Waiau pond system were also studied in detail, though only one of the sites surveyed had īnanga present. This site had a notably different overall environmental condition when compared to other sites, particularly in terms of nutrient levels, although this is unlikely to be the sole reason for the distribution pattern. Otolith analysis revealed a spring spawning population, with a hatch date of September to November, a pattern that has not previously been observed in Southland, but has been seen in populations further North. Growth

80 patterns indicate that fish which hatched later grow faster than those that were earlier. Diet of fish showed strong selection for Chydoridae and chironomids when compared to abundance of these taxa within the pond. Unexpectedly, microplastic fibres were observed in both the guts and pond samples, and showed significant selection for consumption.

The recolonisation process within the Waiau pond system following the near complete drying of the ponds showed a range of spatial and temporal patterns in both the fish and invertebrate communities. The eel population maintained a stable size over the first two months following rewetting, and indicated the use of deep channels as refugia during the drought from which eels subsequently dispersed into the newly available habitats. The smaller fish taxa (bullies and galaxiids) were slower to recover than eels, and did not display the same use of refugia, possibly due to the abundance of eels in the deep refuge channel. The macroinvertebrate community showed no significant temporal variation; it did display variability between sites, although this spatial variability was not linked to the deep channels as it was for eels. Temporal variation was observed in the zooplankton community as the dominant taxa changed over time. Two blooms occurred, with significant population spikes for Daphnia and then for Bosmina and ostracods. Overall highlighting the importance of spatial heterogeneity and deep-water areas within wetlands, particularly constructed systems that may be more prone to dewatering.

6.2 Synthesis across chapters

Water level variation shows an effect on the community of a wetland system, but only at an extreme level. The small-scale variations observed in the Waiau pond system over the 2018-19 summer season had little effect on the invertebrate or fish communities (Chapters 3 & 4), particularly when looked at in comparison to the changes that occurred after the drought the previous summer (Chapter 5). In particular, the fish population displayed a much more stable community dynamic. The natural fluctuations of Waituna Lagoon (prior to lagoon opening) also appeared to have little consequence to the fish community overall (Chapter 2). However the opening of the lagoon during winter 2018 had a significant impact on both the abundance and diversity of fish that were caught, the lagoon opening also showed a more significant long term impact when compared to the Waiau drought. This demonstrates the importance of scale, both physical and temporal, when considering the impact of an event. Small water level fluctuations are unlikely to trigger changes in the community composition of a system whereas larger events may (Davey & Kelly 2007; Matthews & Marsh-Matthews 2003). The rapid return of the Waiau pond system to its pre-drought water level

81 created habitat that was similar to what was available prior to the drought, and that remained stable after refilling. In comparison Waituna Lagoon remained in an altered state of low water for an extended period of time, so although there was habitat available, it did not function as it had pre- opening. This is important to consider as the risk of droughts and flooding becomes more common (Mullan et al. 2005; Tait & Pearce 2019), requiring changes to current activities within the system or increased management interventions to maintain a balance between human activities and the natural functioning of the system.

Findings from chapters based in the Waiau pond system have demonstrated a clear progression through the system, both for nutrients and water conditions (Chapter 3), and for fish and invertebrate dynamics (Chapters 3, 4 & 5). This progression is shown through a measurable difference between ponds as the distance and complexity of connection increases. The transitions between ponds allow for filtration and improvement of water quality; they also restrict movement of taxa, both fish and invertebrates, without preventing it entirely. This allows for a diversity of communities to form rather than a single homogenous community throughout the system. For example, the abundance and success of īnanga within one of the ponds and not in the rest of the system may be related to the different water quality, invertebrate community, and absence of larger predatory fish species. All of these factors are driven by the flow of water through the system originating from the diversion channel. The pattern observed indicates that the ‘artificial’ origin of this wetland system has successfully mimicked what would be found in a natural wetland (Scholz & Lee 2005).

Invertebrate community composition (chapter 3), and īnanga growth rates (chapter 4) both vary over the sampling season, as would be expected with changes in temperature and other environmental conditions over the seasons. Interestingly, chydorids, which have the highest selectivity in īnanga diet, dominate the invertebrate community during early-mid October before declining and remaining at a low level. This high abundance occurs prior to when īnanga are likely to be present, and suggests that īnanga may be supressing population size. The alternative explanation is that increases in competitive taxa reduce the chydorid population, but that they are a preferred prey, meaning that īnanga then need to seek out the less abundant taxa. Although fish are known to exert pressure on populations of lower trophic levels (Detmer & Wahl 2019), in this pond the second hypothesis is more likely as īnanga numbers are relatively low to be effecting the invertebrate community.

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6.3 Future Research Questions

This study has made several observations which have created further questions and grounds for future research. Of particular interest are the īnanga within the Waiau area and the impacts of microplastics in this system. The spawning location and the timing of spawning around the Waiau Mouth is unknown, as is the origin of the fish found within the ponds themselves. The hydrology of the Waiau Mouth is highly dynamic and does not have an obvious salt wedge area where spawning is likely to occur. A survey investigating potential spawning locations as well as the timing of spawning would be valuable. This could inform areas and times to implement protection which may in turn allow for greater success and more whitebait recruiting back into the wider system. If non-diadromy is recognised within the Waiau pond system then changes may need to be made to the way connectivity and riparian edges of the ponds are managed.

The presence of microplastics in this system was unexpected and has created many questions. A longer-term study would be valuable to quantify the prevalence of plastic filaments over broader spatial and temporal scales. However, in the Waiau pond system there are no large urban areas upstream, which makes the origin of these fibres of interest. The primary source of microplastic fibres is considered to be from wastewater in urban areas as the treatment process does not filter them out (Horton et al. 2017). If they are originating from the few townships along the Waiau River it will highlight the abundance and mobility of these contaminants. The impact of the fibres on the taxa that consume them is also of interest. Do the plastics accumulate, or do they pass through the gut easily? Are there long-term effects on fish that consume more fibres? A lab-based experiment providing īnanga with food with and without microplastic and an observation of the digestibility and any negative effects would be highly valuable to begin understanding the consequence of these fibres within the environment. Additionally, it is unclear what impacts microplastics have on macroinvertebrates. If they are also consuming them, this may be a means of transmission through the food web to larger species and may lead to bioaccumulation of microplastics in species such as eels that feed at higher trophic levels.

Understanding the consequences of Waituna Lagoon’s water level changes on the absolute abundances of fish would be valuable information to aid in the argument for changing the current management practices. Knowing more accurately how many fish are present and their distributions both before, and after lagoon opening would create a much clearer picture of the impacts.

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6.4 Management Recommendations

This study has identified several management recommendations, some of which suggest changes to current protocols, particularly around the opening regime in Waituna Lagoon, and others which recognise successes of current methods. The significant reduction in fish abundance and diversity following the opening of Waituna Lagoon along with the observation of large areas of exposed seagrass where water levels had receded, provides grounds for reconsideration of current management protocols. The potential for increased rate of loss of the seagrass beds as well as negative impacts on the fish community are both impacts which work against conservation goals for the lagoon. There are several reports in existence which have in depth analysis of several options for lagoon management (DHI 2015; Thompson & Ryder 2003). They take into consideration maintaining lagoon level while still allowing nutrient flushing and reduction of flooding in the catchment. Many of these options would be preferable to the current opening practices.

Current climate predictions indicate that the Waiau area may experience an increase in both rainfall and frequency of dry days, making the hydrologic regime within the pond system less stable (Ministry for the Environment 2018). Minimising the likelihood of reduced inflow during dry periods would be beneficial to the system. Some level of mitigation has already been implemented following the dry summer in 2017-18 through earthworks at the intake to the diversion channel. However, as the following summer was much wetter, the success is unclear. The presence of deep areas within the system have been shown to be highly valuable to the macrofaunal community during a dry period. This feature would be beneficial to consider on any similar conservation projects, as it can provide refugia, particularly when the wetland is new, more unstable, and more likely to dry.

The functioning of the Waiau pond system itself shows the benefit of this constructed wetland design. The series of ponds connected through underground piping appears to encourage the system to function more similarly to a natural wetland than a single pond system would. There is clear filtering progression through the wetland as well as the creation of unique communities allowing for increased diversity. Both of these features are desirable in any wetland conservation project.

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6.5 Conclusion This thesis has provided an opportunity to explore the dynamics and interactions of fish and invertebrate populations within coastal wetland systems. It has illustrated the variability that exists both spatially and temporally at the large scale between waterbodies, and the smaller scale within a single system. Seasonal succession of invertebrate taxa has been observed in both drought and non- drought years, highlighting the importance of both spatial heterogeneity and connectivity for the functioning of a constructed system. Īnanga population dynamics have been identified which have not previously been seen in this region of Southland, while gut analysis has created a basis of understanding about their diet within shallow pond systems. Overall, this thesis has generated an increased depth of knowledge on the functioning of these systems, the traits of the species that are found within them, and what makes for a successful re-establishment of a wetland complex.

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