Arachnid Ecology in New Zealand, Exploring

1 Arachnid ecology in New Zealand, exploring

2 unknown and poorly understood factors.

3 James Crofts-Bennett.

20 “A thesis submitted in fulfilment of the degree of Master of Science [1] in Botany [2] at the 21 University of Otago, Dunedin, New Zealand”

22 2020

24 Index

26 Abstract………………………………………………………………………………………5.

27 Chapter 1. Introduction……………………………………………………………………...7.

28 1.1 The importance of spiders………………………………………………………...7. 29 1.2 The influence of habitat structural complexity on spider distribution and 30 abundance…………………………………………………………………………...... 8. 31 1.3 Invasive rodents in the context of New Zealand Araneae………………………...9. 32 1.4 Thesis structure and aims………………………………………………………..14. 33 Chapter 2. The effect of habitat structural complexity on spider abundance and diversity..15.

34 2.1 Introduction ……………………………………………………………………..15. 35 Figure 2.1: Seasonal deciduous vegetation cover…………………………...16. 36 Figure 2.2: Seasonal deciduous vegetation cover with mistletoe parasites…16. 37 2.2 Methods…………………………………………………………………………17. 38 Figure 2.3: Examples of foliage samples……………………………………18. 39 Table 2.1: Sampling locations, dates and host data…………………………19. 40 2.2.1 Statistical Analyses……………………………………………………………20. 41 2.3 Results…………………………………………………………………………...20. 42 Figure 2.4: Total invertebrates sampled in summer, plotted………………..22. 43 Figure 2.5: Total invertebrates sampled in winter, plotted………………….23. 44 Table 2.2: Paired t-tests of host plant invertebrate populations……………..25. 45 2.4 Discussion……………………………………………………………………….26.

46 Chapter 3. A novel non-kill Araneae trap: test with regards to vegetation type versus 47 location 48 effects………………………………………………………………………………………..28.

49 3.1 Introduction……………………………………………………………………...28. 50 3.2 Materials and methods…………………………………………………………..30. 51 3.2.1 Trap Design…………………………………………………………………....30. 52 Figure 3.1: Internal structure of Sanctum Aranearum……………………….31. 53 Figure 3.2: Top-down view of Sanctum Aranearum………………………...32.

54 55 3.2.2 Field Testing…………………………………………………………………...33. 56 Figure 3.3: In field trap configuration……………………………………….34. 57 3.2.3 Statistical Analyses…………………………………………………………....35. 58 3.3 Results…………………………………………………………………………...35. 59 Table 3.1: Mean number of invertebrates by method……………………….35. 60 Table 3.2: Best fit models for total invertebrate abundance…………………36. 61 Figure 3.4: Total number of invertebrates by method, plotted………………37. 62 Figure 3.5: Total diversity of invertebrates by method, plotted……………..38. 63 Table 3.3: Best fit models for total spider abundance……………………….39. 64 Table 3.4: Best fit models for total invertebrate diversity…………………...40. 65 Table 3.5: Best fit models for total spider diversity…………………………41. 66 3.4 Discussion……………………………………………………………………….42. 67 Figure 3.6: Neoramia otagoa cohabitating Sancta tube……………………..44.

68 Chapter 4. The impact of vegetation type and predatory regime on Araneae abundance and 69 diversity……………………………………………………………………………………...46.

70 4.1 Introduction……………………………………………………………………...46. 71 4.2 Methods………………………………………………………………………….49. 72 4.2.1 Orokonui mouse data collection for 2019……………………………………..49. 73 4.2.2 Sampling regime………………………………………………………………49. 74 Figure 4.1: Sancta placement map for Orokonui……………………………50. 75 Figure 4.2: Sancta placement map for control site…………………………..51. 76 Figure 4.3: Distance between sampling sites………………………………..52. 77 4.2.3 Spiders as a component of mouse diet………………………………………...54. 78 4.2.4 Statistical Analyses……………………………………………………………54. 79 4.3 Results…………………………………………………………………………...54. 80 Figure 4.4: Total number of invertebrates sampled, plotted………………...55. 81 Table 4.1: Best fit models for spider abundance…………………………….56. 82 4.4 Discussion……………………………………………………………………….57. 83 Chapter 5. General Discussion and Conclusions…………………………………………...59.

84 5.1 Impacts of vegetation structural complexity and mammalian predators………..59. 85 5.2 Significance of this thesis……………………………………………………….60. 86 5.3 Next endeavours………………………………………………………………..61.

89 Acknowledgements…………………………………………………………………………63.

90 References. …………………………………………………………………………………64.

91 Appendices………………………………………………………………………………….77.

92 Appendix 1: Total invertebrate numbers sampled from mistletoes and host plants...77.

93 Appendix 2: Total spider numbers sampled by method…………………………….78.

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113 Abstract.

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115 Spiders (Araneae) are an important ecological component of most terrestrial ecosystems. 116 Comparatively, New Zealand spider ecology is poorly understood, particularly with regards 117 to how spiders have been affected by the influx of new organisms entering an otherwise 118 isolated habitat. This thesis aimed to observe the impact of vegetation structural complexity 119 and mammalian predation on coastal Otago spider abundance and diversity, focusing in 120 particular on the seasonal loss of vegetation complexity on introduced deciduous trees and 121 the impact of introduced rodents. In addition, in order to non-destructively sample spiders in 122 a predator-free ecosanctuary, a novel no-kill trap was designed and tested.

123 The native New Zealand green mistletoe (Ileostylus micranthus) provides a continuous area 124 of evergreen foliage on introduced deciduous trees throughout the winter. Recording 125 invertebrate numbers from mistletoe samples and comparing them to host plant samples 126 provided insight into how invertebrate populations react to loss of structural complexity on 127 introduced deciduous trees during leaf fall. There was no significant difference between 128 mistletoe and host plant invertebrate populations during the summer, but during the winter, 129 invertebrate numbers on deciduous hosts were significantly lower in comparison to evergreen 130 hosts and to mistletoes on either plant host type.

131 A novel, non-kill spider trap was developed to assist with sampling in environmentally 132 sensitive areas, such as ecosanctuaries. Dubbed the Sanctum Aranearum, the Sanctum utilizes 133 structural complexity to attract spiders and other invertebrates to settle within the trap, which 134 can then be readily collected without killing by-catch. The Sanctum was constructed from a 135 white cuboid bucket with windows cut near the surface to allow invertebrates to enter. The 136 internal structure included large and small plastic tubes and wire imbedded in sand. The 137 Sanctum was tested against pitfall traps as a close analogue to compare capture results and 138 found that the Sanctum was able to capture a higher abundance of invertebrates and a greater 139 diversity of arachnids.

140 The impact of predation by introduced rodents was assessed by sampling spiders at the 141 Orokonui Ecosanctuary, near Dunedin, Otago, while it underwent a rodent elimination 142 program. This elimination programme provided an optimal chance to observe how spider 143 populations would be affected by the removal of rodents from a habitat. Sancta were utilized

144 due to the sensitive nature of the ecosanctuary. A nearby Queen Elizabeth II National Trust 145 conservation area with comparable vegetation was used as a control site to compare spider 146 populations, as rodents were known to be present. Analyses were unable to separate the 147 impact of rodents from potential seasonal effects on spider populations. It was noted that 148 larger native spiders were more abundant in Orokonui than at the control site, but lack of 149 historical data made any conclusions purely speculative.

150 This thesis has made a novel contribution in terms of understanding the role of mistletoes as 151 over-winter refugia for invertebrates. The findings in Chapter 2 have a multitude of 152 implications regarding how urban and agricultural invertebrate populations survive winter 153 with the introduction and proliferation of many deciduous plants to New Zealand. Both 154 beneficial and pest species could potentially utilize mistletoes and further research is 155 required.

156 The novel Sanctum Aranearum presented in Chapter 3 and successfully deployed in an 157 ecosanctuary in Chapter 4 is a valuable addition to arachnological methods and further work 158 to explore the potential field uses for it are already being planned.

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170 Chapter 1. Introduction

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172 1.1. The importance of spiders. 173 Araneae are a crucial order of chelicerates that provide a keystone function in virtually all 174 major terrestrial ecosystems on earth. They are both important predators of invertebrates 175 (Eggs & Sanders, 2013; Wimp, et al., 2013; Meehan, et al., 2009; Rypstra & Marshall, 2005; 176 Collier, et al., 2002; Moulder & Reichle, 1972) and a key prey item for other animals, 177 including a large number of vertebrates (Théry & Casas, 2002; Wise & Chen, 1999; Poinar Jr 178 & Early, 1990; Gunnarsson, 1983). Due to the sedentary nature of many spiders (Venner, et 179 al., 2000), combined with the effective nature of the web as a prey capturing tool (Rypstra & 180 Marshall, 2005: Venner, et al., 2000), spiders often form an important food/energy bank for 181 insectivorous animals to exploit during the winter, particularly for birds (Gunnarsson, 1983). 182 Therefore, it is crucial to maintain and protect spider populations and overall biodiversity, 183 both for crop pest management and to maintain food webs.

184 Practical uses for spiders as individuals, local populations and as a global community are well 185 known and have been applied to the fields of ecological restoration (Scott, Oxford & Selden, 186 2006), agricultural practices (Rajeswaran, Duraimurugan & Shanmugam, 2005) and even 187 medicine (Zachariah, et al., 2007). While it is true that there is an abundance of information 188 and data available regarding spiders, there are still a startling number of gaps in human 189 knowledge regarding the Araneae. This is further exacerbated by the overall abundance of 190 arachnologists, where arachnological research have only a limited number of experts working 191 towards furthering this knowledge. From 1940 to 1987 the majority of New Zealand’s 192 arachnological works were products of Ray and Lyn Forster (Forster & Forster, 1999). While 193 this work did bring to light and name quite a substantial proportion of the New Zealand 194 spider fauna, many of the specifics regarding ecology, biology and behaviour remain a 195 mystery. The most recent summary of New Zealand spider diversity is a decade old and 196 asserts 1126 described species of 57 families with high endemism (roughly 93%) and a 197 further 536 species to be described formally (Paquin, Vink, & Dupérré, 2010).

198 A critical limitation to research on spider ecology and biology is the ability to make 199 observations on live individuals. The primary method of field sampling spiders is the 200 destructive method of pitfall trapping (Hancock & Legg, 2012), which limits field

201 observations. Furthermore, as New Zealand’s spider fauna has yet to be fully described 202 taxonomically, let alone understood ecologically, widespread pitfall trapping could be putting 203 further pressure on endangered populations.

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205 1.2. The influence of habitat structural complexity on spider distribution and 206 abundance. 207 A reoccurring conclusion in the literature is that spider abundance and diversity is tied 208 directly to plant vegetation structural complexity (Vasconcellos-Neto, et al., 2017; Halaj, et 209 al., 2000; Marc, et al., 1999; Uetz, 1991; Greenstone, 1984). This factor has been extensively 210 reviewed and compared against other potentially important factors in spider success, 211 including prey availability (Greenstone, 1984), indicating that plant structural complexity is 212 the primary factor in spider abundance and diversity. The relationship between spiders and 213 plants is not only extensive but also mutualistic, with spider or spider-related paraphernalia 214 (such as silk) presence being enough to benefit plant success (Tahir, et al., 2019, 215 Vasconcellos-Neto, et al., 2017). The extent of the relationship is not ultimately known, but 216 observed interactions imply that both groups (spiders and plants) have intergenerational 217 impacts on each other. Spiders are not randomly distributed within plant vegetation and 218 various specialist groups have evolved that occupy different niches within a vegetated 219 environment (Vasconcellos-Neto, et al., 2017). Furthermore, plants can illicit ballooning 220 (passive flight via silk) behaviour in spiders, potentially playing a role in spider distribution 221 (Morley & Robert, 2018). In response, spiders provide plants with protection from herbivores 222 (Rypstra & Buddle, 2013; Rypstra & Marshall, 2005; Halaj, et al., 2000).

223 Spider seasonality varies between habitats. Gunnarsson (1983) found that winter is associated 224 with a high population decline for spiders through bird predation. Conversely, Dias et al., 225 (2006) found that seasonality in an urban forest fragment failed to show impacts on the local 226 spider population. There is some evidence in warmer climates that late spring to early 227 summer are optimal for spiders (Cardoso, et al., 2007), with higher temperatures negatively 228 impacting populations. There is evidence that New Zealand’s winter period has a negative 229 impact on spider populations, but the only available work specifically covers alpine species 230 (Sinclair, Lord & Thompson, 2001). Spider behavioural responses to vegetation complexity 231 and seasonality in other New Zealand ecosystems are poorly understood.

232 Spider habitat in New Zealand, like that of all native fauna, has undergone radical change 233 since the arrival of humans. New Zealand native forests are dominated by evergreen species, 234 but now the naturalized flora includes many deciduous trees, which can be dominant 235 components of both urban and rural landscapes (Wardle, 1978). As deciduous plants lose 236 structural complexity in winter it can be expected that spider abundance and diversity on 237 them subsequently drops in winter. An exception to this could be where evergreen habitat is 238 available as a refugium. New Zealand’s most widespread native mistletoe, Ileostylus 239 micranthus, is evergeen in nature and parasitises introduced deciduous trees as well as native 240 evergreen trees. If a deciduous tree is parasitised by a mistletoe and then sheds it leaves, 241 would that mistletoe act as a spider refuge? If so, the introduction of mistletoes could provide 242 an opportunity to improve spider habitat in highly modified environments. However, while 243 structural complexity is the primary driver in spider success, spiders are still impacted by 244 other environmental factors. Lövei, et al., (2019) notes that despite efforts to introduce urban 245 green spaces with vegetation structural complexity, spider diversity still dropped due to 246 unstable humidity levels.

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248 1.3. Invasive rodents in the context of New Zealand Araneae. 249 One of the major impacts of human dispersal around the globe has been the transmission of 250 non-native organisms to foreign lands (Burney, 1995). Species in the mammalian order 251 Rodentia have frequently been involved in such transmission, with the genera Mus and Rattus 252 (mice and rats) being noted as primary examples of invasive mammalian pest species 253 (Hardouin, et al., 2010; Burney, 1995). The impacts associated with this form of dispersal 254 include directly detrimental effects on humans, such as crop destruction (Stenseth, et al., 255 2003), to environmental devastation, such as the near extinction of many of New Zealand’s 256 native bird life (Clout & Russell, 2006). What is less obvious is the impact on native 257 invertebrates in areas where these animals have been introduced.

258 The pervasive traits of mice and rats include a high reproductive turn over (Caligioni, 2009), 259 strong incisors (von Koenigswald, 1985) and effective dispersal traits that allow rodents to 260 bypass small scale barriers, such as strong swimming capabilities and effective 261 climbing/jumping abilities (Pitt, et al., 2011; Sutherland & Dyck, 1984). Mus musculus offers 262 a primary example of the reproductive capabilities of invasive rodents, capable of incredibly 263 fast generational turn over for mammals (Caligioni, 2009). While generation time is short and

264 turnover is fast, mice do engage in parental investment and nurse young (Fuchs, 1982). This 265 optimises parental investment and reproductive output, which makes for a very pervasive and 266 resilient pest species that recovers from control methods quickly (Innes & Saunders, 2011).

267 The aforementioned traits associated with successful rodent dispersal have allowed mice and 268 rats to circumvent many barriers, both ecological and physical in nature. Both rats and mice 269 can directly bypass artificial barriers, given enough time, by gnawing through them (Lewis & 270 Hurst, 2004). This limits the capabilities of fencing such as “predator” free fences utilised in 271 ecosanctuaries, as regular maintenance limits fence length in relation to the number of 272 employees and volunteers an ecosanctuary can muster. As rats and mice have adaptations that 273 allow them to survive limited exposure to aquatic habitats, natural barriers such as rivers and 274 small oceanic gaps are ineffective at preventing rodent passage (Russell, et al., 2008; Russell, 275 et al., 2005).

276 Both rats and mice are noted omnivores, consuming plant matter and engaging in predatory 277 behaviour (Shiels & Drake, 2011; Bradley & Marzluff, 2003; Hobson, et al., 1999). It is well 278 recorded that vertebrates are part of the aforementioned rodent diets (Bradley & Marzluff, 279 2003; Hobson, et al., 1999) and, despite their naturally cautious nature, they are noted to 280 attack much larger animals. Mus musculus has been observed attacking the large skink, 281 Oligosoma otagense (Norbury, et al., 2014), and rodents can have detrimental impacts on 282 larger bird species by raiding nests for eggs and nestlings (Angel, et al., 2008; Bradley & 283 Marzluff, 2003).

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285 1.3.1 Rodent impacts on invertebrates

286 Current opinions on rodent impacts on invertebrates are mixed. While there appear to be clear 287 trends in the data that indicate rodents have a negative impact, it is also noted that studies are 288 clearly biased towards work done in New Zealand and only some tropical locations, despite 289 many tropical regions experiencing rodent invasions (St Clair, 2011).

290 New Zealand-based studies vary in observed impact of rodent control on invertebrate 291 populations. Several papers include or focus on land snails, showing strong treatment 292 responses for control of rodents in contemporary observations (Bennett, et al., 2002; 293 Craddock, 1997), while historically suspecting rodent introduction to be responsible for some 294 localised extinctions (Brook, 1999). Rodents exhibit preference for larger bodied

295 invertebrates (Craddock, 1997), but the size range is not consistent. Several studies set 4- 296 5mm as the lower preferred body length perimeter (Sinclair, et al., 2005; Bremner, et al., 297 1984), whereas another went as low as 3mm and failed to get a treatment response apart from 298 ant (Hymenoptera; Formicidae) abundance increasing (Rate, 2009).

299 Hawaii is noted to have been invaded by both rats and mice, with one study noting that more 300 than 90% of rodents trapped had evidence of invertebrate remains in stomach contents 301 (Shiels, et al., 2013). Shiels, et al., (2013) also noted that most of the invertebrates 302 successfully identified through stomach content analysis were invasive species. This brief 303 glimpse of rodent impacts in a tropical island setting already indicates a very different 304 environmental impact when compared to temperate New Zealand.

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306 1.3.2 Interactions between spiders and rodents

307 It is known, at least in some situations, that spiders are frequent prey of introduced rodents 308 (Green, 2002; Ruscoe, 2001), but the actual impact rodent predation has on spider 309 populations is poorly understood. Observational study has indicated that tropical regions have 310 higher population densities of rodents (Harper & Bunbury, 2015), but paradoxically, they 311 may not have that much impact on the ecosystem in these regions (Shiels, et al., 2013). 312 Currently it can only be speculated as to why invertebrates native to temperate regions, such 313 as New Zealand, appear to be more vulnerable to rodent predation.

314 In the context of New Zealand, spiders are primarily in a prey relationship with introduced 315 mice (Green, 2002; Ruscoe, 2001). Even in the wider context, predation of vertebrates by 316 spiders is only thought to be a significant factor in places that have sufficiently large spiders 317 (McCormick, et al., 1982). In terms of actual interactions between New Zealand spider 318 populations and invasive rodents, not enough work has been compiled to explore this. The 319 studies of spider-rodent interactions by Green (2002) and Ruscoe (2001) both only covered 320 the genera Uliodon and Meringa and various other assorted spiders over 5mm in length, 321 representing a fraction of New Zealand spider fauna. There are few viable spider habitats that 322 rodents cannot access. Ground-based spiders, such as Uliodon, are optimal prey items for 323 rodents as they have large body size and reasonably easy to access territories (Bowie & 324 Frampton, 2004). This likely extends to other large, ground dwelling spiders, such as 325 Cambridgea, although there is no current work exploring this. The climbing behaviour of 326 mice and rats places arboreal spiders also at risk of predation. Aquatic spiders are more

327 difficult for rodents to access, e.g. Pisauridae as represented by Dolomedes in New Zealand 328 (Williams, 1979). Dolomedes species are generally large spiders and capable of vertebrate 329 predation, typically targeting fish (Williams, 1979). It has been observed that Dolomedes is 330 capable of fending off attack from mice (Zimmermann & Spence, 1989), but it is unknown 331 how they interact with rats, though the sheer size difference would imply Dolomedes are 332 likely prey. Male Dolomedes are smaller than the females (Prenter, et al., 2006), which could 333 make them vulnerable to rodent attacks during their search for females. Female Pisauridae, 334 commonly known as nursery web spiders, spend a lot of time in the canopy of a shrub 335 guarding the eponymous nurseries. This may afford them some protection by limiting attack 336 direction to directly underneath the spider, although more observation is required to establish 337 this.

338 Mygalomorphae in New Zealand, such as Porrhothele antipodiana, utilise tunnel webs in 339 order to avoid predators such as hunting wasps (Laing, 1979). This protection may also 340 extend to mouse attacks and prevent rats reaching the spiders without digging. The spiders 341 are large and have been observed, but not recorded, taking mice as prey (Sirvid, n.d.). As 342 noted with Dolomedes and as a rule for most male spiders, the males will wander looking for 343 females at sexual maturity (Prenter, et al., 2006). Literature regarding many of New Zealand 344 spiders, behaviour during the males, dispersal when looking for potential mates is limited. 345 Porrhothele antipodiana males are observed to be aggressive and will strike at perceived 346 threats (Jackson & Pollard, 1990), but this represents a single species out of a speculated 347 1,662 species in the order Araneae present in New Zealand (Paquin, Vink, & Dupérré, 2010). 348 It is unknown if this is an effective defensive measure against rodent predators, but it is noted 349 in Bremner, et al., (1989) that surviving invertebrate populations that share overlapping 350 habitat with mammalian predators are shown to have heightened threat responses.

351 Green (2002) noted a decline in all spider species over 5mm in length in areas with no rodent 352 protection on off-shore islands. This would include the female native Katipō spider 353 (Latrodectus katipo) as well as several introduced Steatoda species, representing some of the 354 larger New Zealand Theridiidae. Katipō populations are already considered threatened due to 355 habitat loss (Lettink, et al., 2006), therefore the impact of rodents needs to be more closely 356 examined with this species. It is unknown how the much smaller male theridiids interact with 357 rodents.

358 Several of New Zealand’s threatened cave spider species are likely at risk of predation by 359 rodents but are not sufficiently studied to report on. Spelungula cavernicola is considered 360 range restricted and Maloides cavernicola is considered to be nationally threatened (Sirvid, et 361 al., 2012). S. cavernicola is found in cave systems and is the largest native spider in New 362 Zealand as measured by leg span (Sirvid, et al., 2012), making it a prime target for rat 363 predation. It is currently unknown how invasive rats react to cave environments in New 364 Zealand. M. cavernicola is known from a single female found in United Creek cave (Sirvid, 365 et al., 2012).

366 There has been insufficient study of introduced spiders in New Zealand to offer informed 367 commentary on how rodents may impact them in New Zealand. The largest introduced spider 368 is Delena cancerides, a colony-dwelling huntsman spider (Yip, et al., 2009). Though there is 369 no published work on how the New Zealand population handles rodents, previous evidence of 370 spider size in relation to spider defence would indicate that D. cancerides should be quite 371 capable of defending itself from mouse attack.

372 New Zealand offers some exclusive opportunities to explore arachnids in relation to other 373 ecological components. New Zealand was, prior to human colonization, devoid of mammals 374 (Gibbs, 2010) with some exceptions (O’Donnell, 2010). This has already provided 375 opportunities to research how invertebrates were adapting to the introduction of mammalian 376 pest species. For example, Bremner, et al., (1989) were able to discern that various 377 invertebrates, including spiders, had developed defensive behaviours that allowed for fast 378 responses against mice. Due to the opportunities to study the impact of mammalian invasion 379 New Zealand offers, research on the topic has developed an extensive body of work with a 380 distinct bias towards New Zealand (St Clair, 2011). One of the reoccurring conclusions is that 381 body size of invertebrates may affect susceptibility to rodent predation, but the critical size is 382 not consistent in the literature and there are ultimately not enough samples to get a solid 383 consensus (St Clair, 2011). With regards to spiders, the reoccurring pattern suggests that 384 spider populations decline with continued exposure to introduced rodents (St Clair, 2011). 385 Once again, there is simply not enough work to make solid conclusions.

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388 1.4. Thesis structure and aims: 389 This thesis aims to examine the effect of plant structural complexity and predation by 390 introduced rodents on the abundance and diversity of Araneae.

391 The importance of plant structural complexity to Araneae will be explored using the model 392 system of mistletoes and their hosts. Chapter 2 compares invertebrate populations of 393 evergreen mistletoes against invertebrate populations in both deciduous and evergreen hosts 394 in both summer and winter season settings. Specifically this chapter aims to determine 395 whether a mistletoe acts as a significant spider refuge when their deciduous host sheds it 396 leaves.

397 In Chapter 3, a novel non-kill trap designed to sample terrestrial Araneae in ecologically 398 sensitive areas is presented and tested.

399 Chapter 4 uses the trap design presented and tested in Chapter 3 to examine the impact of 400 vegetation type and mouse control on terrestrial Araneae abundance and diversity over 401 several months in and around Orokonui Ecosanctuary, Dunedin. Orokonui is surrounded by a 402 predator-proof fence that prevents the majority of introduced mammals from entering the 403 ecosanctuary. Despite this, Orokonui still has a mouse population, as excluding mice entirely 404 is thought to be impossible using current methods (pers comm. Elton Smith, Ecosanctuary 405 Manager, 2019). Mouse population control methods within the Ecosanctuary include poison 406 baiting and trapping programmes. This chapter compares spider populations before, during 407 and after one such control operation, with parallel sampling for comparison in nearby sites 408 with no rodent control program.

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418 Chapter 2. The effect of habitat structural complexity on spider

419 abundance and diversity.

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421 2.1 Introduction

422 New Zealand vegetation has undergone radical change since the arrival of humans 423 (McWethy, et al., 2010; McGlone, 1983). Native forests were dominated by evergreen 424 species, but now introduced deciduous trees are dominant components of both urban and 425 rural landscapes (Wardle, 1978). As deciduous plants lose structural complexity in winter, it 426 can be expected that spider abundance and diversity on them subsequently drops in winter. 427 An exception to this could be where evergreen habitat is available as a refugium. Specifically, 428 if a deciduous tree is parasitized by an evergreen mistletoe and then sheds it leaves, would 429 that mistletoe act as a spider refuge? If so, the presence or introduction of mistletoes could 430 provide an opportunity to improve spider habitat in highly modified environments.

431 Mistletoes are acknowledged as a keystone species in woodland and forest environments 432 (Watson & Herring, 2012; Watson, 2001). However, this is largely due to the impact 433 observed on vertebrate species (Watson, 2001). There is little information available regarding 434 the significance of mistletoes as a resource for invertebrates. Australian research regarding 435 the mistletoe, Decaisnina signata, suggests that invertebrate assemblages found within 436 mistletoes may be independent of the host plant (Anderson & Braby, 2009). This suggests 437 that more research is needed to explore the significance of mistletoes as both a resource and a 438 habitat for invertebrates. Anderson and Braby (2009) observed that Araneae were one of the 439 two dominant groups (the other being Lepidoptera) observed in mistletoe samples, being both 440 high in abundance and diversity.

441 As noted in the opening chapter of this thesis, spiders are positively correlated with plant 442 structural complexity (Vasconcellos-Neto, et al., 2017). Deciduous plants that seasonally 443 shed leaves, naturally reduce habitat with the loss of the leaves. However, mistletoes in the 444 family Loranthaceae are almost entirely evergreen and most retain leaves year-round (Kuijt & 445 Hansen, 2015). Waite (2012) observed that isolated trees in urban environments played a 446 crucial role in maintaining both arthropod and avian populations. Colour change in leaves of 447 some deciduous trees is known to impact invertebrate behaviour, such as host migration in 448 aphids on silver birch trees (Sinkkonen, et al., 2012). The relationship explored between

449 aphid and deciduous trees indicates an extensive evolutionary relationship that native New 450 Zealand fauna does not share with newly introduced deciduous plants.

451 With the introduction and cultivation of more deciduous plants in New Zealand, available 452 habitat for invertebrates may lower as seasonal leaf shedding occurs (Fig. 2.1). However, 453 evergreen mistletoes parasitizing deciduous host trees could potentially act as seasonal 454 refuges for invertebrates in deciduous tree plantings by offering otherwise unavailable 455 structurally complex habitat (Fig. 2.2).

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459 Figure 2.1. Illustration of vegetation leaf cover of a deciduous tree as seasonal progression 460 occurs.

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462 Figure 2.2. Illustration of vegetation leaf cover of a deciduous tree with several mistletoe 463 parasites as seasonal progression occurs.

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465 This chapter aims to determine if mistletoes are affecting invertebrate populations in general, 466 and spiders in particular, by surveying both mistletoes and hosts of evergreen and deciduous 467 nature in the Dunedin area. Non-random fluctuation in invertebrate numbers and diversity 468 noted between seasons and hosts could indicate if invertebrates rely on mistletoes to shelter 469 during the winter period when deciduous plants shed leaves.

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471 2.2 Methods

472 The common green New Zealand mistletoe, Ileostylus micranthus, was targeted for this 473 project due to its plentiful nature and ease of identification in field (Simpson, 1995). New 474 Zealand’s most widespread native mistletoe, Ileostylus micranthus, is evergreen in nature and 475 parasitises introduced deciduous trees as well as native evergreen trees (de Lange, Norton & 476 Molloy, 1995). Deciduous and evergreen host trees infected with Ileostylus micranthus were 477 selected for study if at least one mistletoe per host was large enough to sample without 478 compromising plant survival and was located between breast height to a maximum of 2.3m 479 off the ground. Extended handle loppers were not used to sample at greater heights as they 480 were found to be imprecise and clumsy when attempting to preserve invertebrates on 481 samples. Hosts were selected purely by deciduous or evergreen nature, and no distinction 482 between native and introduced was made. Approximately 500ml of leaf and small twig 483 material from each mistletoe and its host was harvested into plastic bags during two periods, 484 once during the summer of 2018-2019 and once during the winter of 2019. Care was taken to 485 collect host material at the same height and with the same aspect and light conditions as the 486 mistletoe material. Summer sampling occurred from late November to late-January. Winter 487 sampling occurred throughout August.

488 Mistletoe and host plants were sampled throughout the Dunedin area. Locations sampled 489 included a mix of rural farming areas such as Waitati and Taieri as well as urban areas such 490 as Queen’s Drive and the Dunedin Botanical Gardens (Table 2.1). Samples consisted of 491 freshly pruned branches, filling a 500ml ziplock plastic bag (or close analogue) (Fig. 2.3). 492 Host trees were divided into quadrants based on sampled mistletoe placement and host 493 samples were only taken from quadrants that did not contain mistletoes. The same hosts and 494 mistletoes were sampled between seasons when possible. Some mistletoes disappeared 495 between seasons.

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497 A total of 38 mistletoe samples were collected alongside 38 samples of the host of those 498 mistletoes. This number was evenly divided between seasons with 19 host-mistletoe sample 499 pairs collected in between late November 2018 and January 2019 (summer), and 19 pairs 500 collected in August 2019 (winter).

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504 Figure 2.3. Illustrated examples of evergreen and bare deciduous samples harvested to fill 505 500ml ziplock bags. More bare branches are required to fill a bag than branches bearing 506 vegetation. Overstuffing was avoided to prevent bag rupturing.

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515 Table 2.1. Sampling locations, date and host data from sample collection of Ileostylus 516 micranthus over a summer and winter period. Asterisk indicates host plant sample was 517 heavily covered in lichen when collected.

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Mistletoe sample Host species Location Date of collection 1 Pittosporum sp. Central Dunedin 45°51'52"S 170°30'40"E 11/20/2018 2 Magnolia sp. Opoho 45°51'22"S 170°31'46"E 11/29/2018 3 Hamamelis sp. Waitati 45°44'50"S 170°34'04"E 12/5/2018 4 Hamamelis sp. Waitati 45°44'50"S 170°34'04"E 12/5/2018 5 Coprosma crassifolia Taieri 46°01'25"S 170°13'47"E 1/8/2019 6 Crataegus monogyna Taieri 46°01'25"S 170°13'47"E 1/8/2019 7 C. crassifolia Taieri 46°01'25"S 170°13'47"E 1/8/2019 8 C. monogyna Queen's Drive 45°52'40"S 170°29'03"E 1/8/2019 9 Pittosporum sp. Queen's Drive 45°52'40"S 170°29'03"E 1/8/2019 10* C. monogyna Seacliff 45°40'44"S 170°37'34"E 1/5/2019 11 C. monogyna Seacliff 45°40'44"S 170°37'34"E 1/5/2019 12* C. monogyna Seacliff 45°40'44"S 170°37'34"E 1/5/2019 13 C. monogyna Queen's Drive 45°52'40"S 170°29'03"E 1/14/2019 14 Pittosporum sp. Queen's Drive 45°52'40"S 170°29'03"E 1/14/2019 15 Melicytus sp. Dunedin Botanical Gardens 45°51'17"S 170°31'03"E 1/25/2019 16 Coprosma intertexta Dunedin Botanical Gardens 45°51'17"S 170°31'02"E 1/25/2019 17 Oleria sp. Dunedin Botanical Gardens 45°51'17"S 170°31'02"E 1/25/2019 18 Rubus sp. Dunedin Botanical Gardens 45°51'16"S 170°31'02"E 1/25/2019 19 Coprosma virescens Dunedin Botanical Gardens 45°51'17"S 170°31'02"E 1/25/2019

20 Magnolia sp. Opoho 45°51'22"S 170°31'46"E 8/18/2019 21 C. monogyna Seacliff 45°40'44"S 170°37'34"E 8/22/2019 22* C. monogyna Seacliff 45°40'44"S 170°37'34"E 8/22/2019 23* C. monogyna Seacliff 45°40'44"S 170°37'34"E 8/22/2019 24 C. monogyna Seacliff 45°40'44"S 170°37'34"E 8/22/2019 25 C. monogyna Seacliff 45°40'44"S 170°37'34"E 8/22/2019 34 C. intertexta Dunedin Botanical Gardens 45°51'17"S 170°31'02"E 8/24/2019 35 Melicytus sp. Dunedin Botanical Gardens 45°51'17"S 170°31'03"E 8/24/2019 36 Oleria sp. Dunedin Botanical Gardens 45°51'17"S 170°31'02"E 8/24/2019 37 C. virescens Dunedin Botanical Gardens 45°51'16"S 170°31'02"E 8/24/2019 38 Coprosma robusta Dunedin Botanical Gardens 45°51'17"S 170°31'02"E 8/24/2019 26 C. monogyna Taieri 46°01'25"S 170°13'47"E 8/30/2019 27* C. monogyna Taieri 46°01'25"S 170°13'47"E 8/30/2019 28 C. monogyna Taieri 46°01'25"S 170°13'47"E 8/30/2019 19 C. monogyna Taieri 46°01'25"S 170°13'47"E 8/30/2019 30* C. monogyna Taieri 46°01'25"S 170°13'47"E 8/30/2019 31 C. crassifolia Taieri 46°01'25"S 170°13'47"E 8/30/2019 32 C. crassifolia Taieri 46°01'25"S 170°13'47"E 8/30/2019 519 33 C. crassifolia Taieri 46°01'25"S 170°13'47"E 8/30/2019

520

521 Samples were stored at five degrees Celsius for two days to slow the movement of 522 invertebrates (MacMillan & Sinclair, 2011; Mellanby, 1939) following collection. Samples 523 were hand checked under a dissection microscope on an individual branch basis. All 524 invertebrates collected were transferred directly to 70% ethanol to preserve samples for 525 identification, with the exception of one Phasmatodea that was too large for the storage tubes 526 being used. As this Phasmatodea was collected during the final sampling, was easily 527 identified without preservation, and there was no need for further collection, the individual 528 was returned to the corresponding mistletoe the insect was collected from. Non-arachnid 529 invertebrates were identified to Family level, arachnids were identified to species level or 530 lowest taxonomic level possible.

531 2.2.1 Statistical Analyses

532 Paired T-tests were used to test for statistically significant differences in total invertebrate 533 abundance, total spider abundance, and invertebrate and spider taxonomic richness between 534 individual mistletoe samples and the paired foliage samples from their host plant, first for all 535 hosts combined then the subset of deciduous hosts only. These analyses were conducted 536 using Statistix v.9 (Analytical Software).

537 2.3 Results

538 Total invertebrate numbers came to 1374 of which the majority consisted of mites from two 539 host plants (Rubus sp. and Coprosma virescens). After removing domatia dwelling mites as 540 an outlying group, total invertebrates consisted of 211 individuals. After adjustments, mites 541 were still the dominant group, with 73 individuals identified. Barklice were also abundant, 542 with 68 individuals identified. Total diversity amounted to 28 taxa, of which 24 were found 543 on mistletoes and 17 were found on host plants and a total of 12 taxa shared between both 544 mistletoe and host (Appendix 1). Arachnid diversity totaled 11 species with five Araneae and 545 six Acari from a total of 9 families. Acari could only be identified to family level. Seasonal 546 variation saw the majority of invertebrates counted during the summer period with 150 547 individuals compared to 61 sampled during winter. Total diversity sampled in summer was 548 higher, with 20 taxa counted compared to 14 noted in winter.

549 Invertebrate abundance and diversity of species collected in the summer did not significantly 550 vary between mistletoes and their hosts (Fig. 2.4). Spiders were not commonly found during 551 the summer period, with only eight sampled at four sites (Fig. 2.4). Invertebrate numbers

552 collected from evergreen hosts and mistletoes attached to evergreens were statistically 553 indistinguishable regardless of seasonal effect.

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555

556

557 Figure 2.4. Invertebrate numbers sampled during the summer period from host plants (a) and 558 mistletoe parasites (b). Samples 15 and 16 have been omitted due to domatia mite numbers 559 vastly outnumbering invertebrates found on all other plants sampled.

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562

563 Figure 2.5. Invertebrate numbers sampled during the winter period from host plants (a) and 564 mistletoe parasites (b). Samples 1, 13 and 14 have been omitted due to no invertebrates being 565 found on either host or mistletoe sample.

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570 When all hosts were considered, winter sampling revealed invertebrates were more 571 consistently found in mistletoes than host plants (Fig. 2.5) but failed to find a significant 572 difference between hosts and mistletoes. However, focusing on deciduous hosts and their 573 mistletoes sampled during the winter period found that total spider population, total non- 574 spider invertebrate population, total spider diversity and total non-spider diversity (Table 2.2) 575 in mistletoes was significantly higher than on their deciduous host plants.

576

577 Table 2.2. Paired t-test analysis of hosts and invertebrate parameters compared against 578 associated mistletoe samples. All statistically significant p-values are highlighted in bold text. 579 Negative mean indicates mistletoe samples were greater than the host samples via direction 580 of t-test.

Deciduous hosts versus total inverterbrates Parameters Mean DF t p Both seasons -1.95 21 -1.35 0.1915 Summer -1.67 11 -0.63 0.543 Winter -2.3 9 -3.63 0.0055 Deciduous hosts versus total spiders Parameters Mean DF t p Both seasons -0.36 21 -1.89 0.0726 Summer 0 11 0 1 Winter -0.8 9 -2.75 0.0224 Deciduous hosts versus total invertebrate diversity Parameters Mean DF t p Both seasons -1.05 21 -3.15 0.0049 Summer -0.58 11 -1.13 0.2808 Winter -1.6 9 -4.17 0.0011 Deciduous hosts versus total spider diversity Parameters Mean DF t p Both seasons -0.18 21 -1.45 0.1621 Summer 0.0833 11 0.56 0.5863 581 Winter -0.5 9 -3 0.015

Evergreen hosts versus total inverterbrates Parameters Mean DF t p Both seasons 69.56 15 1.31 0.2099 Summer 158.29 6 1.35 0.2261 Winter 0.56 8 0.63 0.5471 Evergreen hosts versus total spiders Parameters Mean DF t p Both seasons 0.38 15 1.57 0.138 Summer 0.29 6 0.6 0.5686 Winter 0.44 8 1.84 0.1038 Evergreen hosts versus total invertebrate diversity Parameters Mean DF t p Both seasons -0.06 15 -0.16 0.8755 Summer -0.57 6 -0.88 0.4128 Winter 0.33 8 0.71 0.4996 Evergreen hosts versus total spider diversity Parameters Mean DF t p Both seasons 0.25 15 1.46 0.1639 Summer 0 6 0 1 582 Winter 0.44 8 1.84 0.1038

583

584 2.4 Discussion

585 The majority of papers reviewing the ecological role of Loranthaceae, or mistletoes in 586 general, focus on vertebrate interactions with limited focus on invertebrate ecology 587 (Anderson & Braby, 2009). Anderson and Braby (2009) studied Decaisnina signata and 588 found evidence suggesting higher invertebrate taxonomic richness within the mistletoe that 589 led to the hypothesis that invertebrate assemblages in mistletoes are independent from that of 590 the host plant. In contrast, this study found that in summer and on evergreen hosts, 591 invertebrate abundance and taxonomic diversity on Ileostylus micranthus did not differ from 592 that of its host.

593 The lack of variation between evergreen hosts and parasitic I. micranthus implies that it is the 594 deciduous component, rather than the mistletoe itself, that is impacting invertebrate 595 populations during the winter period. This favours the hypothesis that deciduous leaf 596 shedding and the resulting drop in available vegetation habitat is resulting in lower 597 invertebrate numbers. This is backed by the literature regarding both invertebrate and 598 specifically spider reactions to available environmental structural complexity (Vasconcellos- 599 Neto, et al., 2017; Halaj, et al., 2000; Marc, et al., 1999; Greenstone, 1984).

600 It can be concluded that New Zealand invertebrates in general, and spiders in particular, are 601 using evergreen mistletoes as winter refugia on deciduous host plants. This process is a 602 simple result of invertebrates favouring areas of increased structural complexity. The stark 603 drop in invertebrate numbers in the winter samples of deciduous hosts infers that deciduous 604 plants are not viable invertebrate habitats in New Zealand, being viable during the summer 605 but largely devoid of life during the winter. This effect should be considered in biodiversity- 606 focused projects, ranging from full ecological restoration projects to simple horticultural 607 ventures. Ecological restoration projects are unlikely to negatively impact New Zealand 608 invertebrates due to the abundance of evergreen plants native to New Zealand. However, any 609 planting projects revolving around ornamental, agricultural or even native monocultures of 610 deciduous plants could result in negative impacts to the local invertebrate population. In 611 particular, fruit orchards such as apple and kiwifruit and other deciduous fruit plants that are 612 commonly grown in New Zealand (Müller, et al., 2015; Wearing, et al., 1992) could be 613 negatively impacting related invertebrate populations during winter periods if no evergreen 614 refugia are available. Presence of evergreen refugia in these areas may impact numbers of 615 both beneficial and destructive pest invertebrates. Further research into this is advised. The

616 results of this study also raise questions regarding how helpful gardens are in providing 617 habitat for invertebrates (Barratt, et al., 2015). Garden composition is fickle depending on the 618 owner’s tastes and trends in social gardening behaviour (Goddard, et al., 2010), which has 619 both direct (removal of mistletoes as pests) and indirect (what hosts are available for 620 mistletoes) impacts on what habitats are available throughout the season.

621 Despite the implications, this study was simply one project over the passage of one year. 622 Mistletoe invertebrate ecology is still poorly understood and much more work will be needed 623 to further clarify how important the relationship between mistletoes and invertebrates are.

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641 Chapter 3. A novel non-kill Araneae trap, test with regards to

642 vegetation type versus location effects.

643

644 3.1 Introduction

645 Terrestrial invertebrate sampling in the field often relies on the use of pitfall traps (Hancock 646 & Legg, 2012). This method relies on mobile invertebrates accidentally falling into a 647 container where they are killed and preserved for later collection. The pitfall trapping method 648 is known to produce sampling biases, for example small-bodied and less mobile invertebrate 649 groups are likely to be under-sampled (Hancock & Legg, 2012, Spence & Niemelä, 1994, 650 Olson, 1991). However, methods have been developed to accommodate these biases in 651 quantitative estimates of abundance derived from pitfall trap data (Hancock & Legg, 2012). 652 In studies targeting specific taxa, by far the greatest disadvantage of the kill trap method of 653 pitfalling, is the high bycatch and unnecessary invertebrate death (Seldon & Beggs, 2010). 654 Considering the current “invertebrate extinction crisis” (Régnier, et al., 2015; Dunn, 2005), 655 these characteristics of pitfall traps can not only trigger negative public responses to field 656 entomology/ecology, but could also be contributing directly to local invertebrate decline 657 (Cardoso & Leather, 2019). It has been suggested that traps designed to target specific 658 invertebrate orders of interest (Seldon & Beggs, 2010) would lower bycatch fatality. 659 However, as Seldon & Beggs (2015) noted, the high diversity of invertebrates makes it 660 difficult to bait a pitfall trap in a manner that would appeal to a target invertebrate order, 661 without also attracting other orders.

662 With small-bodied invertebrates being underrepresented in pitfall traps (Hancock & Legg, 663 2012, Spence & Niemelä, 1994, Olson, 1991), combined with the overall effectiveness of the 664 pitfall trap for capturing spiders (Hancock & Legg, 2012), logic dictates a bias against small- 665 bodied spiders in pitfall trap-centric methods.

666 Arachnids are an important component of the global invertebrate community, the abundant 667 Araneae, i.e. spiders (Agnarsson, et al., 2013), being a notable ecological component strongly 668 related to plant community structure (Vasconcellos-Neto, et al., 2017). The relationship 669 between spiders and plants could be considered mutualistic, with spiders being an effective 670 control for herbivory (Knauer, et al., 2018; Del-Claro, et al., 2017) and plants providing a 671 diverse, structurally complex environment for spider habitation (Villanueva-Bonilla, et al.,

672 2017; Amaral, et al., 2016; Gómez, et al., 2016). The correlation between spider abundance 673 and vegetation structural complexity has been observed to be stronger than the relationship 674 between spider abundance and prey abundance (Lövei, et al., 2019; Halaj, et al., 2000; 675 Greenstone, 1984). As with invertebrates in general (Cardoso & Leather, 2019), arachnid 676 extinction risks are not well understood and require further study to better understand 677 causation of observed decline (Régnier, et al., 2015).

678 This chapter reports on a proof of concept study to test whether the provision of desirable 679 habitat with structural complexity that appeals to a specific order, namely Araneae, could 680 form the basis of a non-kill sampling method equal in effectiveness to the standard pitfall 681 trapping method. It also tests whether a habitat trap targeted at spiders would minimise 682 bycatch of non-target organisms. Furthermore, if bycatch is unavoidable, the trap design 683 could allow for efficient sampling of target invertebrates without killing said bycatch. Spiders 684 are not scattered amongst vegetation in an unpredictable manner; they specifically associate 685 with areas of high structural complexity (Vasconcellos-Neto, et al., 2017; Observations in 686 Chapter 2). Spiders have been observed to select for vegetation traits that denote greater 687 structural complexity over plant species specificity (Villanueva-Bonilla, et al., 2017). 688 Spiders do not exclusively benefit from plant structural complexity; many other invertebrate 689 orders benefit from three dimensional complexity of leaves and vegetation, for example, 690 mites and leaf domatia (Weber, et al., 2016), and the relationship observed with multiple 691 invertebrate orders exhibiting higher abundance in relation to higher densities of Douglas fir 692 needles per branch (Halaj, et al., 2000). Using structural complexity as bait will not be 693 exclusively attractive to the order Araneae, thus a non-fatal approached would be necessary 694 to prevent bycatch fatalities. Any new non-kill invertebrate trapping system would also need 695 to be viable after prolonged environmental exposure. Pitfall traps often require rain guards to 696 prevent flooding and protect the contents of the traps (Brown & Matthews, 2016). Physical 697 robustness is also required to prevent the trap from being destroyed by environmental factors 698 such as falling branches or being trampled by larger animals.

699 The expectation is that a habitat trap that emphasises structural complexity will have a higher 700 efficiency in capturing spiders, especially small-bodied taxa, than the standard pitfall. 701 Structural complexity will be provided by accessible components within the habitat trap that 702 maximise surface area within the limited volume. Variation in the types of available 703 structural complexity will also be important to the functioning of a habitat trap, as different 704 spiders will seek different forms of structural complexity within the trap.

705 3.2 Materials and methods

706 3.2.1 Trap design

707 The Sanctum Aranearum was built to encourage residence rather than completely trap the 708 invertebrates, hence the use of the word Sanctum rather than trap. The outer casing consisted 709 of a white, cuboid, plastic pail (180mm length x 180mm width x 190mm height); the white 710 colouration allows exposed Sancta Aranearum to mitigate internal temperature built up in the 711 field through reflectance (Shiukhy, et al., 2015). The lid was retained and used on the pail to 712 shield the internal Sanctum Aranearum structure from weather and prevent birds easily 713 accessing any invertebrates inside. Two windows were cut near the top of the pail on each 714 chassis wall to allow animals to enter and exit at will. The size of these windows can be 715 customised depending on preferences; in our case the windows were 2cm high and 4cm wide 716 due to the presence of the large spider Dolomedes minor (L. Koch, 1876) at the study sites. 717 The casing was filled with sand level with the windows to create a near-seamless transition 718 from outside to inside habitat. The sand also controls humidity levels within the Sanctum 719 Aranearum, trapping moisture and slowly releasing it during drought periods to keep the 720 inside humid once it equilibrates with the surrounding environment (Vriend, et al., 2017). 721 This also prevented flooding from intermittent rain. Twelve 50ml centrifuge tubes and twenty 722 2ml microcentrifuge tubes were used to emulate crevices in rocks and bark within the 723 Sanctum Aranearum and were buried in the sand with only the apertures exposed to allow 724 animals access. To further increase structural complexity within the Sanctum Aranearum, 725 two 40cm lengths of wire were bent into varying arcing shapes and anchored into the sand at 726 both ends to create a zone of wiring above the tube apertures (Figs 3.1, 3.2).

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730 Figure 3.1. Simplified internal structure of the Sanctum Aranearum. Incomplete lines

731 indicate structures submerged in sand. Front right corner and portion of front face of Sanctum

732 Aranearum have been removed to show internal structure.

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736 Figure 3.2. Top-down view of Sanctum Aranearum with lid removed.

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743 The resulting complete Sanctum had combined entry apertures of 64cm2, flat internal surface 744 area of 924cm2 (excluding wires) and a total tube volume of 416.24cm 3 (centrifuge tubes = 745 377.04cm3 and microcentrifuge tubes = 39.2cm3). This total volume did not include potential 746 volume generated by animals that burrow into the interior sand of the Sanctum Aranearum in 747 situ.

748 3.2.2 Field testing

749 The study was conducted from October 7th (Southern Hemisphere spring) 2019 to February 750 2020 on two rural properties within 30km of Dunedin, South Island, New Zealand. On each 751 property, areas of grass-dominated pasture and adjacent areas of regenerating native forest 752 dominated by Kunzea robusta (Myrtaceae) were selected for trap placement. At the northern 753 site, Waitati (lat 45.7466°S, long 170.5655°E, 40 m a.s.l.), the available pasture was 100m 754 distant from the forest margin, however at the southern site, Taieri (45.0221°S, 170.2296°E, 755 20m a.s.l.), the available pasture was immediately adjacent to the forest margin. Field sites 756 were surveyed for plant diversity prior to sampling. Despite this difference in proximity, the 757 pasture communities were structurally and compositionally similar between sites, as were the 758 two forest communities. Four Sancta Aranearum and eight pitfall traps were placed in each 759 vegetation type at each site, giving a total sampling effort of 16 Sancta Aranearum and 32 760 pitfall traps. Each sampling station contained one Sanctum Aranearum at a distance of ten 761 metres from two adjacent pitfall traps within a vegetation type to offer equivalent active trap 762 surface area but minimise risks of interference (Digweed, 1995), without compromising the 763 similarity of vegetation being sampled (Fig. 3.3). Pitfall traps consisted of small plastic cups 764 with a 7cm aperture diameter and 10cm depth generating a total trap capacity of 384.85cm3 765 per pitfall with an aperture of 38.48cm2. Pitfalls were primed with 2cm of ethylene glycol, 766 resulting in a volume of 76.97cm3 per pitfall. A plastic rain guard was secured over each 767 pitfall to prevent rainfall flooding the traps. Sancta Aranearum were semi-buried so that the 768 ground was level with the sand within the traps. Pitfall traps were buried so that the rim of the 769 pitfall was at ground level. Sampling stations within each site were a minimum of ten metres 770 apart to avoid interference between replicates.

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777 Figure 3.3. Diagram of Sanctum Aranearum and pitfall trap pair in field.

778

779 All Sancta Aranearum and pitfalls were emptied monthly over the three-month sampling 780 period. The contents of paired pitfall traps were pooled and directly transferred to 70% 781 ethanol for storage. All invertebrates were identified to family level or as close as possible by 782 consulting the primary literature for keys or consulting specialists. Araneae were identified to 783 species or as close as possible via consulting the literature, primarily Spiders of New Zealand, 784 Annotated Family Key & Species List (2010). For the purposes of this proof-of-concept, the 785 contents of Sancta Aranearum were also harvested destructively following the same protocol 786 as with pitfall traps. In future projects involving the Sanctum Aranearum, invertebrates can 787 be collected, photographed and then released back into the field site without unnecessary 788 fatalities. For a few highly abundant and fast moving or small invertebrate taxa, numbers in 789 the traps were capped at fifty when more than fifty individuals were clearly observed, to 790 minimise bias. For example, one Sanctum Aranearum sampled contained more than 200 791 individual amphipods when opened; the animals quickly scattered and only a fraction could 792 be collected, so an estimation of fifty individuals was used. This also applied to pitfall traps 793 and Sancta Aranearum when sampling nematodes and springtails, which could reach numbers 794 well over 200 per trap, but were unlikely to have been fully sampled during collection.

795 Fragmentary remains of invertebrates, including exuviae and prey remains, were noted but 796 not added to abundance or diversity measures.

797 3.2.3 Statistical Analyses

798 Generalised Linear Models were used to determine the effect of site, sampling month, 799 vegetation type and trapping method on total invertebrates caught, total arachnids caught, 800 total invertebrate family diversity and total arachnid family diversity. All models were tested 801 against both a negative binomial and a Poisson distribution. Akaike’s Information Criterion 802 (AIC) values were used to compare alternative models (Symonds & Moussalli, 2011).

803

804 3.3 Results

805 A total of 4041 invertebrates, representing 51 invertebrate families were collected during the 806 study, of which a total of 1328 were arachnids, representing 14 families (Appendix 2). More 807 invertebrates were captured from grass areas (974 individuals) than from forest (354 808 individuals) (Table 3.1). Predatory behaviour and cannibalism were noted within Sancta 809 Aranearum, but this failed to reduce capture rates compared to pitfall traps. In general, more 810 arachnids were collected from Sancta Aranearum (938 whole individuals) than from pitfall 811 traps (390 individuals), confirming the attractiveness of the design.

812

813 Table 3.1. Mean (standard deviation) number of non-arachnid invertebrates and arachnids 814 and number of families represented in each method and vegetation type.

Trap type Vegetation type Pitfall Sanctum Aranearum Forest Grass Total non-Arachnid inverterbates 89 (74) 137 (83) 95 (75) 131 (85) Total Arachnids 9 (1) 20 (3) 9 (1) 20 (3) Non-Arachnid family diversity 7 (0.4) 7 (0.4) 6 (0.4) 8 (0.4) Arachnid family diversity 2 (0.2) 3 (0.3) 2 (0.3) 3 (0.3) 815

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820 The period of time between sample collection did not exert a significant influence on either 821 total number of invertebrates sampled (Tables 3.2 and 3.3, Fig 3.4) or total diversity (Tables 822 3.4 and 3.5, Fig 3.5). All other factors except time of sampling had a significant impact on 823 invertebrate abundance. The best fit model for total invertebrates captured included site, 824 method, vegetation and a site interaction with vegetation, with a negative binomial error 825 distribution (Table 3.2). All other models with delta AIC <5 included trapping method as a 826 significant term, but no other model was considered equally likely as the next best model had 827 delta AIC > 2 (Table 3.2).

828

829 Table 3.2. Generalised linear model results for total Invertebrate abundance. S=site, 830 M=Method (df 1), V=vegetation type (df 1), T=time (df 2). Only models with ∆ AIC < 5 are 831 shown. Subscripts denote d.f.; NS: p > 0.05; *: p < 0.05; ** p < 0.01; *** p < 0.001. 832 833

Model Likelihood Wald Χ2 tests for best model effects Best fit models Model ∆ AIC Ratio Χ2 Intercept S M V 1 1 1 interaction terms SM(0.001)NS SV(7.680)** S,M,V full factorial 4.973 26.5397*** 1093.54*** 5.458* 10.056** 4.665* MV(0.012)NS SMV(0.989)NS SM(0.001)NS S, M, V, SM, SV, 3.949 25.563 *** 1125.52*** 5.894* 10.377** 4.227* SV(7.769)** MV 6 MV(0.051)NS

S, M, V, SV 0 25.5124*** 1140.29*** 5.991* 10.531** 4.621* SV(7.861)** 834

835

836

837 Figure 3.4. Total number of invertebrates caught in pitfall traps (P) vs Sancta Aranearum (S)

838 in pasture (G) vs forest (F) over a three-month period at the a) Waitati and b) Taieri sites.

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843 Figure 3.5. Total diversity of invertebrate families caught in pitfall traps (P) vs Sancta

844 Aranearum (S) in pasture (G) vs forest (F) over a three-month period at the a) Waitati and b)

845 Taieri sites.

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849 Total arachnids captured was also modelled best by a negative binomial distribution, with 850 trapping method a significant term in all models with delta AIC < 5 (Table 3.3). The best fit 851 model included only trapping method and vegetation type, but an equally valid alternative 852 model included method, vegetation type and the non-significant interaction between them 853 (delta AIC < 2; Table 3.3).

854

855 Table 3.3. Generalised linear model results for total spider abundance. S=site, M=Method (df 856 1), V=vegetation type (df 1), T=time (df 2). Only models with ∆ AIC < 5 are shown. 857 Subscripts denote d.f.; NS: p > 0.05; *: p < 0.05; ** p < 0.01; *** p < 0.001. 858

859 Model Likelihood Wald X2 tests for best model effects

Best fit models 2 Intercept Model interaction ratio X S1 M1 V1 ∆ AIC terms SM(0.009)NS SV(1.036)NS S, M, V full factorial 4.014 36.756 *** 450.85*** 1.739NS 13.210** 12.798*** 7 MV(1.339)NS SMV(3.266)NS

NS NS M, V, MV 1.628 31.1433*** 501.26*** - 12.766*** 0.374 MV(0.374) M, V only 0 30.770 *** 506.64*** - 12.837*** 14.092*** - 860 2

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870 Total diversity of invertebrate families trapped best fitted a Poisson distribution. All models 871 with delta AIC < 5 excluded trapping method and inferred a strong impact of vegetation on 872 diversity (Table 3.4). Site interaction with sampling period and site interaction with 873 vegetation were both significant with the site/vegetation interaction have the strongest overall 874 impact on invertebrate family diversity (Table 3.4). A greater number of families were 875 present at the Taieri site (eight) compared with the Waitati site (six).

876

877 Table 3.4. Generalised linear model results for total Invertebrate family diversity. S, site; M, 878 capture method; V, vegetation type; T, sampling period. Only models with ∆ AIC < 5 are 879 shown. Subscripts denote d.f.; NS: p > 0.05; *: p < 0.05; ** p < 0.01; *** p < 0.001. 880

Model Likelihood Wald Χ2 tests for best model effects Best fit models Model ∆ AIC Ratio Χ2 Intercept S M V 1 1 1 interaction terms ST(17.879)*** NS NS S, V, T, SV, ST, VT 2.482 49.6959*** 2077.33*** 3.353 5.057 8.413** SV(15.600)*** TV(1.521)NS

ST(18.033)** S, V, T, SV, ST 0 48.177 *** 2098.17*** 3.384NS 4.428NS 7.762** 7 SV(16.671)*** 881

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892 Total arachnid species diversity also differed between sites, but in addition was affected by 893 trapping method, with a higher number of arachnid families caught in Sancta Aranearum than 894 pitfall traps. Of the four equally valid alternative models (delta AIC < 2), three included 895 trapping method but only two found it to have a significant effect (Table 3.5). The strongest 896 consistent effect on spider species diversity in samples was the difference between sites.

897

898 Table 3.5. Generalised linear model results for total spider diversity. S, site; M, capture 899 method; V, vegetation type; T, sampling period. Only models with ∆ AIC < 5 are shown. 900 Subscripts denote d.f.; NS: p > 0.05; *: p < 0.05; ** p < 0.01; *** p < 0.001. 901

Model Wald Χ2 tests for best model effects Likelihood Model Best fit models 2 ∆ AIC Ratio Χ Intercept S1 M1 V1 interaction terms

Site only 1.838 7.1231** 215.51*** 7.081** - -

NS NS S, M, SM 0.408 12.5533** 205.762*** 7.810** 3.453 - SM(2.453)

SM(2.455)NS S, M, V, SM, SV, NS 0.209 18.7526** 203.706*** 9.099** 4.414* 1.562 SV(3.992)* MV MV(1.866)NS

SM(3.030)NS

S, M, V full NS SV(4.490)* 0 20.9627** 181.812*** 9.036** 4.695* 2.135 factorial MV(2.305)NS SMV(2.167)NS 902

903

904 Smaller bodied spiders, such as the families theridiidae and linyphiidae, appeared in the Sancta in 905 higher numbers than in the pitfall traps (Appendix 2). Total linyphiidae sampled by Sancta were 404 906 individual spiders, compared to 21 via pitfall traps. Theridiidae numbers were also higher in the 907 Sancta, 154 individual spiders compared to 21 sampled via pitfall traps.

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912 3.4 Discussion

913 The pitfall trap has been arguably the most common method of invertebrate sampling for 30 914 years (Brown & Matthews, 2016) with historical usage dating back further (Williams, 1958). 915 Its limitations are well understood and inherent biases can be accommodated in quantitative 916 analyses (Hancock & Legg, 2012). It is clear from this proof-of-concept study that the 917 Sanctum Aranearum does not collect the same data as a pitfall trap; the Sancta Aranearum I 918 deployed in grassland and forest collected a greater number of invertebrates, and in some 919 situations greater taxonomic diversity, and especially increased the yield of small-bodied 920 arachnids, when compared to the paired pitfall traps. Similar to pitfall traps, the Sanctum 921 Aranearum does not provide an estimate of true abundance. Pitfall traps, due to the 922 destructive nature of the sampling method, invoke bias based on invertebrate activity 923 (Hancock & Legg, 2012). The Sanctum Aranearum invokes bias as it actively lures 924 invertebrates by generating a favourable habitat. Though the animals are not truly trapped, 925 there is also no reason for them to vacate the Sanctum Aranearum and this may falsely 926 increase estimates of invertebrate abundance during the sample processing. This bias is likely 927 to be most pronounced in areas of high invertebrate abundance or where safe sites are 928 limiting. Furthermore, as there is a finite number of habitable areas within the Sanctum 929 Aranearum, reaching maximum capacity (as noted with A. hilaris, below) can prevent any 930 further influx of invertebrates into Sancta Aranearum left in place for long periods of time.

931 The attractiveness of the habitat provided by the Sanctum Aranearum was made clear in this 932 proof of concept study by the multitude of shed spider exoskeletons (not included in counts 933 during this study) that were located within the Sancta Aranearum. Moulting is a lengthy 934 process and if the spider survives moulting, it must also wait for the exoskeleton to harden 935 before resuming normal activity (Uhl, et al., 2015). There is some evidence that humidity and 936 temperature can negatively impact key life stages in spiders, including moulting (Jones, 937 1941), thus the success of this process suggests that the Sanctum Aranearum design 938 successfully generated viable spider habitat.

939 The main key advantage of the Sanctum Aranearum over pitfall trapping is that invertebrates 940 within it can be counted and identified by inspection in situ, without the need to kill 941 individuals. This makes the Sanctum Aranearum a much lower impact sampling tool for use 942 within ecologically sensitive areas. I found that the Sanctum Aranearum was not able to 943 avoid bycatch – taxonomic diversity was higher than in matched pitfall traps - but it does

944 provide a non-kill option for inadvertently captured ground invertebrates as well as small 945 vertebrates, the death of which can be an issue with pitfall traps (Lange, Gossner & Weisser, 946 2011).

947 As noted in Seldon & Beggs (2015), there is no way to control for predation within non-kill 948 traps and predatory behaviour and cannibalism were noted within Sancta Aranearum. 949 Anoteropsis hilaris L. Koch, 1877, was a prominent cannibal during field observations. 950 However, this behaviour was only observed in one Sanctum Aranearum during sampling and 951 appeared to be a result of overcrowding; over 50 live individuals were taken during this 952 sampling event, fragmentary remains indicating over 100 individuals had been in the 953 Sanctum Aranearum during that sampling month. Despite this, predatory behaviour was not 954 sufficient to reduce the capture effectiveness of the Sanctum Aranearum, as more individuals 955 were observed in Sancta than were caught and killed in pitfall traps. The potential arena for 956 interspecific interactions provided by the Sanctum Aranearum also creates opportunities to 957 study trophic ecology as well as competitive behaviour. For example, craneflies were noted 958 to climb into the Sancta Aranearum to pupate; predatory Cryptachaea veruculata Urquhart, 959 1885, was noted to take advantage of this. Other interesting observations made included the 960 cohabitation in one tube of two individual spiders of the species Neoramia otagoa (Forster & 961 Wilton, 1973) (Fig 3.6).

962

963

964 Figure 3.6. Two adult, female Neoramia otagoa cohabitating a tube from a Sanctum

965 Aranearum. Note the silken partition.

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968

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972

973 In terms of protecting bycatch species, the Sanctum Aranearum was a success. Though the 974 samples collected in this project were ultimately preserved in wet storage for future reference, 975 the potential of the Sanctum Aranearum for use as a non-kill trap is clearly evident. 976 Furthermore, while the Sanctum Aranearum proved to be an effective specialist method for 977 collecting arachnid samples, it was still suitable for surveying invertebrates in comparison 978 with the classic pitfall trap without killing the collected invertebrates. Use of the Sanctum 979 Aranearum is financially viable due to simple design and low production costs, so can be 980 used in remote as well as ecologically sensitive areas without sacrificing efficiency. While 981 this project made use of some materials with limited availability such as centrifuge tubes, 982 these can be substituted for any type of metal or plastic tube or tubing structure provided the 983 material will not spoil during field deployment.

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1000 Chapter 4. The impact of vegetation type and predatory regime on

1001 Araneae abundance and diversity.

1002

1003 4.1 Introduction

1004 New Zealand is renowned for the extensive protection programs for the multitude of native 1005 threatened birds throughout the country (Miskelly, et al., 2008). The lack of understanding 1006 regarding inverterbates, including spiders, in New Zealand ecosystems indicates a dangerous 1007 blind spot in wildlife management. The ecological role of New Zealand’s spider populations has 1008 not been fully explored; thus, the potential impact of spider decline on New Zealand ecosystems 1009 as a whole is unknown. Spiders are known to play key roles elsewhere, both as predators and as 1010 prey (Eggs & Sanders, 2013: Wimp, et al., 2013; Meehan, et al., 2009; Rypstra & Marshall, 1011 2005; Collier, et al., 2002; Théry & Casas, 2002; Chen & Wise, 1999; Poinar Jr & Early, 1990; 1012 Gunnarsson, 1983; Moulder & Reichle, 1972). Birds are notably a key predator of spiders, 1013 particularly during the winter period in Sweden (Gunnarsson, 1983). If spiders also play an 1014 important role in winter energy banks for New Zealand birds, then it is crucial that steps are 1015 taken to protect spider populations. Spiders also share an extensive and mutually beneficial 1016 relationship with plants, having a positive correlation in abundance and diversity with vegetation 1017 structural complexity alongside spider predation of invertebrate herbivores (Vasconcellos-Neto, 1018 et al., 2017). Adverse impacts on spider populations may have unexpected cascade effects on 1019 plant populations (Rypstra & Buddle, 2013).

1020 Further complicating this issue of spider population preservation, the impact of introduced 1021 mammals, while thoroughly explored and understood in relation to birds (Bradley & Marzluff, 1022 2003), is not so readily understood in the context of spiders. Rodent impact on invertebrates is 1023 still an emergent field of research in New Zealand, despite the extensive study of mammalian 1024 impact on New Zealand ecosystems (St Clair, 2011). As explored in St Clair (2011), studies of 1025 mice impacts are often observed only with regards to singular areas (predominately within New 1026 Zealand, notably offshore New Zealand islands and the North island) and specific invertebrate 1027 groupings. Reviewing this material currently offers the conclusion that medium to large spiders 1028 experience declines with the introduction of rodents while smaller spiders are less affected

1029 (Green, 2002; Ruscoe, 2001). There is evidence that New Zealand invertebrates, including 1030 spiders, are also developing or already possess behavioural responses that increase the likelihood 1031 of survival after encountering rodents (Bremner, et al., 1989). A range of observed spider 1032 predatory behaviours can allow spiders to reverse the predator/prey interaction with mice, though 1033 this is primarily related to either very large spiders (Sirvid , n.d.; Zimmermann & Spence, 1989) 1034 or spiders with venom that is effective at quickly dispatching mammals (Nyffeler & Vetter, 1035 2018). It should be noted that even though some spiders can survive rodent encounters, they are 1036 still vulnerable to mouse predation (Laing, 1975).

1037 Thus, the current consensus on rodent impact on spiders has failed to offer a satisfactory 1038 conclusion. There is not enough information to make any definitive statement on the effect of 1039 rodents on spiders in New Zealand. One reoccurring observation in the literature indicates that 1040 spiders with a body length under 4mm do not appear to be impacted by rodent presence (St Clair, 1041 2011). The aforementioned defensive behaviours may have slowed the decline compared to the 1042 more immediate effect observed with New Zealand native vertebrates (Diamond & Veitch, 1043 1981), but population declines are still observed (St Clair, 2011).

1044 As introduced mice are ubiquitous in New Zealand, there are very few opportunities to quantify 1045 their impact in isolation from that of other rodents. Mammal excluding ecosanctuaries provide an 1046 opportunity to examine the impact of mammals within a New Zealand ecosystem. This chapter 1047 makes use of an opportunity presented by Orokonui Ecosanctuary. Orokonui is a 307 hectare 1048 area of coastal forest located between Waitati and Purakaunui, 20km north of Dunedin.

1049 The primary method utilised at Orokonui to exclude mammals is a nine kilometre long perimeter 1050 mesh fence that extends underground to prevent rabbit intrusion and includes a metal hood that 1051 prevents climbing mammals from circumventing the border. This has protected several key New 1052 Zealand habitats of most of the introduced mammalian pest species, allowing native flora and 1053 fauna to regenerate with human assistance. As of 2019, the only known mechanisms of regular 1054 breaching include birds carrying live animals in (per com. Elton Smith, Orokonui Ecosanctuary 1055 manager, 2019), mammals gripping on settling snow and ice on the hood (per com. Tahu 1056 McKenzie, 2019) and juvenile mice passing through the mesh (per com. E. Smith, 2019). To 1057 counter these incursions, regular tracking and trapping regimes are in place alongside controlled

1058 poison use. The effect of these controls has been to reduce mammal presence within the fence 1059 line to primarily mice and human activity.

1060 Mice are a consistent problem for all New Zealand ecosanctuaries, despite being heavily 1061 controlled within the fenced zone. At Orokonui Ecosanctuary mice are monitored via ink tunnels 1062 and traps are used in and outside of the predator free fence to control internal populations and 1063 provide an outer buffer from both mice and other mammals that consume mice. A poison 1064 operation was initiated in 2019 using bait stations to lower mouse density within the 1065 Ecosanctuary (pers. comm. E. Smith, 2019). The aim of this chapter is to use the opportunity 1066 offered by Orokonui to separate the impact of mice on spider populations from that of other 1067 predators, within the context of different vegetation types. As rodent control operations at 1068 Orokonui are intermittent, sampling before, during and after control can highlight the impact of 1069 reduced mammalian pressure on spiders in the area. This can be contrasted with changes in 1070 spider populations in adjacent areas with lower mammalian suppression.

1071 The inclusion of study areas outside the ecosanctuary also allowed for consideration of the 1072 impact of increased predation pressure on spiders from native birds and reptiles inside the 1073 ecosanctuary. The primary function of the ecosanctuary is to provide sanctuary for charismatic 1074 New Zealand fauna, primarily tuatara, various lizards and birds. A lot of these animals are noted 1075 predators of spiders. The presence of mice within the sanctuary alongside the success of these 1076 animals within the sanctuary is expected to have an impact on the local spider population. This is 1077 particularly important for Orokonui as an ecosanctuary with a goal of ecological restoration; 1078 spiders may play a crucial role in the ecosystem as prey for the many endangered birds protected 1079 there (notably, the Haast Kiwi). As Orokonui has reoccurring mouse incursions, there is potential 1080 for rodents to compete with native birds for spiders as a food resource.

1081 This chapter aims to determine if mice are affecting ground-dwelling spider populations in 1082 Orokonui Ecosanctuary by surveying vertebrate numbers before and after a mouse control 1083 operation. During this period of time, any fluctuations in the spider population that deviate from 1084 the general observed trend may be attributable to the mouse control. Parallel surveying at a 1085 control site with no mammal control would account for other potential bias such as seasonal 1086 shifts. The primary issue with invertebrate surveying at Orokonui is the often-fatal methods 1087 employed, such as pitfall trapping. Pitfall traps allow for long term collection of invertebrates,

1088 but they are not specific and can result in bycatch of small vertebrates (Lange, et al., 2011). This 1089 is unacceptable for an ecosanctuary that houses endangered reptiles, such as Orokonui that 1090 currently has jewelled geckos, numerous skink species and tuatara. To circumvent this issue, a 1091 non-kill method was developed specifically targeting arachnids, the Sanctum Araneae. As 1092 discussed in Vasconcellos-Neto, et al., (2017), spiders positively correlate with structural 1093 complexity. Artificially inflating the structural complexity within a protected space, mimicking 1094 plant vegetation in the process, would allow spiders to accumulate within a “trap” over time 1095 (proven successful in Chapter 2 with relation to mistletoes on deciduous trees). The detail of the 1096 Sanctum Araneae trap designed for this project has been provided in Chapter 3, along with the 1097 results of a verification experiment. As discussed in Chapter 3 this non-kill invertebrate sampling 1098 method proved successful and was utilised in Orokonui ecosanctuary to examine the impact of a 1099 mouse control operation on spider populations.

1100

1101 4.2 Methods

1102 4.2.1 Orokonui mouse data collection for 2019

1103 Orokonui has an extensive and robust tracking and trapping system in operation within the 1104 perimeter fence, with a 50m by 50m grid of ink tunnel traps that are active on a quarterly basis 1105 (per com. Elton Smith, 2019). A list of mouse population recordings from this network starting 1106 in 2011 and running until the start of this project was obtained from the Ecosanctuary manager. 1107 The mouse control operation relevant to this study started on the 4th of June, 2019 except for the 1108 Tui track study area (Fig. 4.1), which began on the 1st of July. Mice were specifically targeted for 1109 poison baiting with Brodifacoum within a strict time schedule, and a decline in the mouse 1110 population was observed using ink tunnel traps.

1111 4.2.2 Sampling regime

1112 Sampling within Orokonui Ecosanctuary and at the control site was conducted before and after 1113 the mouse control operation using the Sanctum Araneae live capture method (Chapter 3). This 1114 method allowed for safe capture of invertebrates with minimal risk of harm to native lizards and 1115 other non-target organisms. Four broad vegetation types were defined at Orokonui Ecosanctuary 1116 to control for the impact of vegetation structural complexity on spider abundance. These

1117 vegetation types were grass and light shrub (Butterfly track), gorse and shrub (Aviary), open 1118 forest floor (Aviary track) and fern dominated forest floor (Tui track) (Fig. 4.1).

1119

1120

1121 Figure 4.1. Locations for placement of Sancta Aranearum within Orokonui. White crosses 1122 indicate immediate location of Sancta, thin white outlines indicate areas surveyed for Sanctum 1123 placement.

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1141 Figure 4.2. Locations of Sanctum locations at Waitati QE II trust covenant site. 1 indicates forest 1142 with fern ground cover, 2 indicates open ground forest, 3 indicates gorse shrub and 4 indicates 1143 grassy meadow (https://www.google.co.nz/maps/@- 1144 45.7476906,170.5658809,195m/data=!3m1!1e3).

1145

1146 Figure 4.3. Google maps capture of distance between Orokonui ecosanctuary and the Waitati 1147 QE II trust sites (https://goo.gl/maps/TPoPDVb2Umdq6WdT7).

1148

1149

1150

1151

1152 The control site was a Queen Elizabeth II reserve located in the Waitati area (Fig. 4.2) 4.20km 1153 NE of Orokonui ecosanctuary (Fig. 4.3). This site contained the same four vegetation types 1154 identified at in Orokonui but lacked major mammal control methods, in particular the mammal 1155 exclusion fence. In each vegetation type in both Orokonui and the control site, 5m x 5m plots 1156 were defined and two sanctums placed at opposing corners of each plot to minimise interference 1157 but maintain standardisation of vegetation texture types. Two ink tunnel traps were placed and 1158 baited with peanut butter at the QE II trust site. This was combined with personal observations 1159 by the landowner to record mammalian presence.

1160

1161 Sancta were placed in position (Fig. 4.1) within Orokonui and the sites were sampled on 1162 alternating weeks starting on the 22/04/2019 at Orokonui and ending on the 23/09/2019 at the 1163 QE II trust. Sanctums were in position for a five-day period each fortnight at each site (Monday 1164 – Friday) between the 22nd of April, 2019 and the 23rd of September, 2019, with sampling weeks 1165 at Orokonui alternating with sampling weeks at the control site. This alternating sampling regime 1166 was implemented to allow time for sample processing between sampling periods. Each Friday, 1167 the currently active sanctums would be checked and cleared of non-target organisms then the 1168 entire sanctum with the remaining target organisms was frozen for two days at -18°C. After this 1169 the sanctum contents were hand emptied and all invertebrates preserved in 70% ethanol. Insects, 1170 crustaceans, myriapods and soft bodied invertebrates were identified to family level and spiders 1171 were identified to species level or as close to species level as possible.

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1176

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1178

1179 4.2.3 Spiders as a component of mouse diet

1180 To verify that mice at Orokonui were predators of spiders, mice were gathered from within 1181 Orokonui near bait stations or from traps and stored in onsite freezers. These were later dissected 1182 for gut content analysis, isolating invertebrate fragments.

1183 4.2.4 Statistical analyses

1184 Chi squared analysis was used to compare the abundance of invertebrates in sanctums between 1185 Orokonui and the control site. Generalised linear models using a Poisson error distribution were 1186 used to explore possible relationships between spiders specifically and the following variables, 1187 Location (2 levels representing Orokonui and Waitati), Vegetation (4 levels, grass/scrub, gorse, 1188 open forest floor, fern) and Time period (a binary variable that switched from 1 to 0 during July 1189 coinciding with mouse control at Orokonui). An interaction between the Time variable and 1190 Location would indicate mouse control at Orokonui was having an effect. Alternative models 1191 incorporating subsets of variables and interactions were compared using log likelihood and 1192 Akaike’s Information criterion. Statistical analysis was undertaken using SPSS v.24 (IBM 1193 Corporation, 2016).

1194

1195 4.3 Results

1196 Chi squared analysis found that the number of invertebrates captured at Orokonui and the control 1197 site were significantly different (χ2 (df= 6, N= 106) = 61.90, p = 0.000). Cell chi-squared values 1198 indicated this was primarily due to two spider species, the native Cycloctenus fugax (Goyen, 1199 1890) and the invasive Diplocephalus cristatus (Blackwall, 1833). D. cristatus was only 1200 observed at the control site, while C. fugax was found at both locations but most abundant within 1201 the Orokonui ecosanctuary (Fig. 4.4).

1202

1203 Figure 4.4. Total number of invertebrates sampled at Orokonui ecosanctuary and the Waitati QE 1204 II trust site. Sorted by invertebrate families.

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1208

1209 Mouse gut analysis indicated that invertebrates, including spiders, were consumed by mice 1210 within Orokonui. Thirty-five dissected mice yielded two leg fragments belonging to two species 1211 of spider and one wing from an insect. One of the spider leg fragments appeared similar to C. 1212 fugax specimens. Ink tunnel traps only revealed the presence of possums at the QE II trust site. 1213 Personal observation by landowner Paul Guy indicated the presence of both mice and rats on the 1214 QE II trust site (per com. Paul Guy, 2019).

1215 The majority of models tested failed to converge on a solution; a frequent issue with invertebrate 1216 data comprising many zero values (Spence & Niemelä, 1994). The poisson model also did not 1217 indicate location as a sole statistically significant effect on spider population (Table 4.1). The 1218 poisson (Orokonui data only) model explored the impacts on spider population exclusively at 1219 Orokonui, removing the location effect. AIC and Log likelihood were halved while retaining 1220 convergence and vegetation and time period effect were strongly indicated to have statistically 1221 significant impact on spider populations at this point.

1222 The differences between spiders found at Orokonui and spiders found in the QE II trust could not 1223 be separated from seasonal effects through statistical analysis. While the Chi squared test did not 1224 test this, the second linear model removed location as a significant reason for the observed 1225 distributions (Table 4.1). Vegetation and location show a significant effect on spider abundance 1226 (Table 4.1) which is attributable to the difference in spider species found between the locations. 1227 C. fugax can be found in all four vegetation types but D. cristatus was only found in grassy and 1228 gorse shrub habitats. D. cristatus was not found in Orokonui at all and this shifted the number of 1229 individual spiders found in grassy and gorse shrub habitats to higher in the QE II field site when 1230 compared to the Orokonui site. The vegetation versus mouse control effect observed in linear 1231 model three can also be attributed to D. cristatus. There was no major change in the D. cristatus 1232 population during this time and there was no mammal control at the QE II site.

1233 Table 4.1. Generalized linear models for Orokonui and control spider abundance collected in 1234 2019. Note that “time” indicates the period of time when poison baiting was occurring within 1235 Orokonui as a binary variable that switched from 1 to 0 during July. Eight models have been 1236 omitted due to failure to converge.

1237

Model Likelihood Model Interaction Best fit models ∆ AIC Ratio Χ2 Convergent terms X2 df p Time 17.95 1 0 Poisson 173.18 -150 Yes Location*Vegetation 20.91 2 0 Vegetation*Time 9.389 2 0.009

Poisson Vegetation 24.18 3 0 0 -71.138 Yes (Orokonui data only) Time 15.3 1 0.000 1238

1239 4.4 Discussion

1240 The evidence from the mouse gut analysis indicates that mice are consuming invertebrates. Le 1241 Roux, et al., (2002) found that an average of 25 mouse guts from a population can offer the best 1242 coverage of invertebrate prey so the sample size used here, 35 mice, is within acceptable 1243 parameters. The small number of invertebrate fragments may be due to the abundance of poison 1244 bait available to the mice (per com. Elton Smith, 2019). Ideally mouse gut contents would have 1245 been analysed some weeks prior to the control operation but mouse samples were not able to be 1246 obtained at that time. Time limitations also prevented more thorough gut content analysis, 1247 allowing only large fragments to be utilised for analysis.

1248 D. cristatus is noted to be under the estimated size of invertebrates that are regularly taken as 1249 mouse prey as noted in St Clair, 2011 (4-5mm body length, D. cristatus rarely exceeds 2mm). D. 1250 cristatus is also noted to be of European origin (Rozicka, 1995), suggesting the spider evolved 1251 contemporary to rodents such as mice and may be better adapted to survival in shared habitat 1252 with rodents. C. fugax, as found at both sites but with higher abundance within Orokonui, has a 1253 typical body length between eight and twelve millimetres. The success of the larger spider within 1254 the ecosanctuary could reflect the control of rodents within Orokonui. Despite the impact of the 1255 presence of D. cristatus, the effect of time period at Orokonui was still shown to be statistically 1256 significant. This included the apparent radiation of C. fugax at Orokonui after the poison baiting 1257 program was initiated. Prior to controlling the mouse population, C. fugax was only found within 1258 the open ground forest region. After the baiting, the spiders started to appear in the fern forest as 1259 well as the gorse shrub area. This may indicate long distance dispersal was supplementing 1260 depleted spider populations in these areas. With the removal of most of the mice, the spiders 1261 quickly repopulated these areas before the winter seasonal effect could hit in earnest. It may also 1262 be an example of the aforementioned behavioural adaptations New Zealand invertebrates have 1263 developed (Bremner, et al., 1989) indicating less local dispersal when mouse populations are 1264 high.

1265 Poor fitting for models is most likely attributable to low invertebrate numbers. Seasonal effects 1266 are likely responsible for this, as the sampling primarily occurred during the autumn and winter 1267 period. Additional samples were taken at the QE II trust site in the spring and summer period 1268 later in 2019 in order to better establish the effectiveness of the Sanctum Aranearum sampling

1269 method. This resampling revealed a significant difference in invertebrate abundance and 1270 diversity during this time period (Chapter 3).

1271 The actual impact of rodents on spider populations at either site is still unknown. The native C. 1272 fugax had a higher abundance at Orokonui than at the QE II trust site but the cause of this 1273 observation is not readily apparent in the statistical analysis. Dolomedes minor (L. Koch, 1876) 1274 was caught once at the QE II site but was also noted to be abundant at Orokonui and low catch 1275 numbers may be a combination of the seasonal timing of the project and the sedentary nature of 1276 D. minor (Williams, 1979). The invasive D. cristatus was only found at the QE II trust site. 1277 These observations are in line with current expectations of rodent impact on spider populations, 1278 but the data collected does not offer enough power to make any conclusions. The generalized 1279 linear models have indicated that time period has had a negative impact on the spider 1280 populations. It is difficult to separate this effect of seasonal change from the impact of rodents.

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1295 Chapter 5. General Discussion and Conclusions

1296

1297 The aim of this thesis was to research the impacts of rodents and mistletoe plants on spider 1298 abundance and diversity. Both fields of study are poorly understood and the presence of 1299 Orokonui ecosanctuary alongside the abundance of the green mistletoe, Ileostylus micranthus, 1300 found in the Otago region made for an opportune chance to cover both topics.

1301 5.1 Impacts of vegetation structural complexity and mammalian predators on New Zealand 1302 spiders.

1303 Spider abundances are positively correlated with plant structural complexity (Vasconcellos-Neto, 1304 et al., 2017), and deciduous plants that seasonally shed leaves, naturally lower seasonally 1305 available vegetation bound structural complexity. In Chapter 2, I showed that the reduction in 1306 structural complexity of deciduous plants during the winter period had a significant negative 1307 impact on both spider abundance and diversity. The results of that mistletoe and host sampling 1308 also highlighted a largely unexplored aspect of introduced deciduous plants in New Zealand. 1309 While mistletoe and evergreen hosts did not exhibit any significant differences in abundance and 1310 diversity, deciduous plants that had undergone leaf fall during the winter were found to have 1311 significantly less invertebrates and invertebrate species. The implications include the potential 1312 for reduced ability for invertebrate populations to survive or recover from the winter period 1313 within New Zealand in areas with extensive deciduous plantings. Alternatively, there is potential 1314 for agricultural research regarding how pest species may make use of these winter refugia in crop 1315 near ornamental and horticultural trees that play host to mistletoes. With regards to Araneae, the 1316 results are in line with the literature. Spider abundance and diversity are repeatedly shown to 1317 positively correlate with available vegetative structural complexity (Vasconcellos-Neto, et al., 1318 2017; Halaj, et al., 2000; Marc, et al., 1999; Greenstone, 1984). This relationship with structural 1319 complexity is common between spiders and invertebrates in general and has similar implications 1320 for the general invertebrate population as it does for the spider population.

1321 The impact of mammalian predators on spiders was less pronounced. The literature offered 1322 mixed opinions on how invertebrates might be impacted by rodent predators, essentially 1323 covering all potential outcomes from no impact to mild and finally serious impacts (St Clair, 1324 2011). The results presented in Chapter 4, sampling invertebrates before and after rodent control 1325 was not able to separate the impact of seasonal change from the impacts of rodent control. This 1326 has made it difficult to comment on how rodents have impacted invertebrates found at Orokonui. 1327 The literature has noted that smaller invertebrates are less impacted by rodent activity. 1328 Invertebrates with body lengths under five millimetres are noted in the literature to avoid rodent 1329 predation more effectively. The primary arachnid recorded at the Queen Elizabeth II trust site 1330 was D. cristatus which has a typical body length of 1.5 mm. In contrast, Orokonui was 1331 dominated by the much larger C. fugax, with a body length more commonly found to between 1332 eight and twelve millimetres. This could imply that the extensive rodent control practised at 1333 Orokonui has favoured the success of a larger spider species. More research will be required to 1334 make any further inferences.

1335

1336 5.2 Significance of this thesis.

1337 This project has further reiterated the importance of mistletoes as a keystone ecological resource. 1338 Mistletoes are already known as an important food resource for New Zealand birds. Now there is 1339 evidence suggesting that lack of host specificity exhibited by I. micranthus enables invertebrates 1340 to over winter due to this mistletoe parasitizing deciduous plant hosts. This project is amongst 1341 the first attempts to examine how spiders interact with mistletoe parasitism of host plants via the 1342 increased structural complexity provided by mistletoes.

1343 The novel, non-kill trap design used in Chapters 3 and 4 of this project has proved successful in 1344 capturing invertebrates non-destructively. The simple and automated nature of the design 1345 alongside cheap production costs offer enough potential to warrant further exploration of the 1346 design in field use. The non-kill, passive collection method was proven a successful sampling 1347 method for use in ecologically sensitive areas, being deployed in Orokonui ecosanctuary without 1348 any vertebrate by catch.

1349 Clear differences in spider populations existed between Orokonui ecosantuary and the QEII 1350 reserve and differences also existed between vegetation types, however the experiment described 1351 in Chapter 4 was not able to detect an effect of the mouse control operation at Orokonui. Mouse 1352 gut samples did provide evidence of rodent predation of invertebrates within Orokonui but was 1353 unable to offer sufficient evidence regarding quantity of invertebrates consumed. This was 1354 probably due to the use bait stations, with poison bait dominating the mouse gut contents, 1355 collection of mice prior to the use of bait stations may have offered more accurate data regarding 1356 mouse dietary habits within the ecosanctuary. Future work will require sampling of mice caught 1357 by traps earlier in the year, outside of the poison bait operational phase. The experimental design 1358 was not ideal for researching the impact of rodents on spiders as there was only one treatment 1359 site (Orokonui ecosanctuary). While ecosanctuaries do offer a unique chance to explore mammal 1360 exclusion in New Zealand, individual projects such as this work only contribute singular data 1361 sets and would require regular data collection to compile a stronger conclusion. Projects being 1362 diversified to include other ecosanctuaries would also improve the strength of the data. This 1363 project is simply the first step towards exploring spider and mammal interactions via 1364 ecosanctuaries.

1365

1366 5.3 Next endeavors.

1367 Future work regarding invertebrate populations within mistletoes, with a wider focus of mistletoe 1368 diversity within New Zealand, would help provide a better idea of how important mistletoes are 1369 to invertebrate ecology. This project focused solely on the common I. micranthus, but there are a 1370 further seven species of mistletoe (De Lange, et al., 2006) in New Zealand that require similar 1371 research. Exploration of mistletoes in more isolated habitats is also required to further examine 1372 how invertebrates and mistletoes interact. All samples collected in this project came from either 1373 urban or rural environments, excluding the important forest habitat from analysis due to time 1374 constraints. This project also highlighted the importance of future work regarding New Zealand 1375 invertebrate populations and introduced flora. There is a potentially unknown threat to New 1376 Zealand biodiversity with the influx of deciduous trees that have invaded New Zealand via 1377 human vector. The data collected reaffirmed that reducing vegetation structural complexity

1378 would negatively impact invertebrate populations. It also suggested that mistletoe parasitizing 1379 could offset this negative impact with mistletoes acting as over winter refugia.

1380 Further comparative analysis of the Sanctum Aranearum sampling method is required to 1381 properly contrast the non-kill method against the pitfall trap. Advice given by entomology and 1382 ecology experts has suggested a more standardized size comparison to remove bias and make 1383 contrast of samples collected more accurate. Future tests could utilize the basal liver pail used as 1384 the chassis of the Sancta as a pitfall trap. This would standardize trap aperture gape which is 1385 positively correlated with spider diversity and abundance caught in pitfall traps.

1386 Future work regarding rodent and invertebrate interactions will require a wider range of 1387 sampling dates to cover the spring and summer period. This will improve invertebrate counts and 1388 offer stronger statistical power in data.

1389 The combination of projects also suggests important research is required into the impact of 1390 mistletoe predators on urban invertebrate populations. There is current research in progress 1391 regarding the endangered mistletoe Tupeia antarctica including the impact mammal herbivory 1392 has on the plants overall survival and distribution (per com. Zoe Lunniss, 2019). Research 1393 conducted in this project infers that urban and rural mistletoes on deciduous hosts are important 1394 winter refugia and possum, rat and human destructive activity may negatively impact overall 1395 ecology in these habitats due to targeting mistletoes.

1396

1397

1398

1399

1400

1401

1402

1403

1404 Acknowledgements

1405

1406

1407 Thanks to Janice Lord for supervising this thesis; to John Steel and Barbara Barratt for advising 1408 on this thesis; to the Botany Department of Otago University for hosting this thesis; to the 1409 University of Otago for providing financing via scholarship funding; to Paul Guy, Maia Mistral, 1410 Heather Murrary, Tracey Frisby, The Dunedin Botanical Gardens and Orokonui ecosanctuary for 1411 providing site access and samples for research; to Jenny Jandt for providing further advice 1412 regarding methodology; to Michael Heads for Latin grammar.

1413 1414 1415 1416 1417 1418 1419 1420 1421 1422 1423 1424 1425 1426 1427 1428 1429 1430 1431

1432 References

1433

1434 Agnarsson, I., Coddington, J. A., & Kuntner, M. (2013). Systematics: progress in the study of 1435 spider diversity and evolution. Spider research in the 21st century: trends and perspectives. SSP, 1436 Manchester.

1437 Amaral, D. S., Venzon, M., dos Santos, H. H., Sujii, E. R., Schmidt, J. M., & Harwood, J. D. 1438 (2016). Non-crop plant communities conserve spider populations in chili pepper agroecosystems. 1439 Biological Control, 103, 69-77. doi.org/10.1016/j.biocontrol.2016.07.007

1440 Anderson, S. J., & Braby, M. F. (2009). Invertebrate diversity associated with tropical mistletoe 1441 in a suburban landscape from northern Australia. Northern Territory Naturalist, 21, 2-23.

1442 Angel, A., Wanless, R. M., & Cooper, J. (2009). Review of impacts of the introduced house 1443 mouse on islands in the Southern Ocean: are mice equivalent to rats? Biological Invasions, 11(7), 1444 1743-1754. doi.org/10.1007/s10530-008-9401-4

1445 Barratt, B. I., Dickinson, K. J., Freeman, C., Porter, S., Johnstone, P. D., Wing, J., & Van 1446 Heezik, Y. (2015). Biodiversity of Coleoptera and other invertebrates in urban gardens: a case 1447 study in a New Zealand city. Insect Conservation and Diversity, 8(5), 428-437. 1448 doi.org/10.1111/icad.12120

1449 Bennett, S. J., Standish, R. J., & Stringer, I. A. (2002). Effects of rodent poisoning on 1450 Powelliphanta traversi. Science for Conservation C, 195, 41-56.

1451 Bowie, M. H., & Frampton, C. M. (2004). A practical technique for non‐destructive monitoring 1452 of soil surface invertebrates for ecological restoration programmes. Ecological Management & 1453 Restoration, 5(1), 34-42. doi.org/10.1111/j.1442-8903.2004.00171.x

1454 Bradley, J. E., & Marzluff, J. M. (2003). Rodents as nest predators: influences on predatory 1455 behavior and consequences to nesting birds. The Auk, 120(4), 1180-1187. doi.org/10.1642/0004- 1456 8038(2003)120[1180:ranpio]2.0.co;2

1457 Bremner, A. G., Barratt, B. I. P., Butcher, C. F., & Patterson, G. B. (1989). The effects of 1458 mammalian predation on invertebrate behaviour in South West Fiordland. New Zealand 1459 Entomologist, 12(1), 72-75. doi.org/10.1080/00779962.1989.9722570

1460 Bremner, A. G., Butcher, C. F., & Patterson, G. B. (1984). The density of indigenous 1461 invertebrates on three islands in Breaksea Sound, Fiordland, in relation to the distribution of 1462 introduced mammals. Journal of the Royal Society of New Zealand, 14(4), 379-386. 1463 doi.org/10.1080/03036758.1984.10421738

1464 Brook, F. J. (1999). Changes in the landsnail fauna of Lady Alice Island, northeastern New 1465 Zealand. Journal of the Royal Society of New Zealand, 29(2), 135-157. 1466 doi.org/10.1080/03014223.1999.9517588

1467 Brown, G. R., & Matthews, I. M. (2016). A review of extensive variation in the design of pitfall 1468 traps and a proposal for a standard pitfall trap design for monitoring ground‐active arthropod 1469 biodiversity. Ecology and Evolution, 6(12), 3953-3964. doi.org/10.1002/ece3.2176

1470 Burney, D. A. (1995). Historical perspectives on human‐assisted biological 1471 invasions. Evolutionary Anthropology: Issues, News, and Reviews, 4(6), 216-221.

1472 Caligioni, C. S. (2009). Assessing reproductive status/stages in mice. Current Protocols in 1473 Neuroscience, 48(1), A-4I. doi.org/10.1002/0471142301.nsa04is48

1474 Cardoso, P., & Leather, S. R. (2019). Predicting a global insect apocalypse. Insect Conservation 1475 and Diversity, 12(4), 263-267. doi.org/10.1111/icad.12367

1476 Cardoso, P., Silva, I., De Oliveira, N. G., & Serrano, A. R. (2007). Seasonality of spiders 1477 (Araneae) in Mediterranean ecosystems and its implications in the optimum sampling period. 1478 Ecological Entomology, 32(5), 516-526. doi.org/10.1111/j.1365-2311.2007.00894.x

1479 Chagué-Goff, C., Nichol, S. L., Jenkinson, A. V., & Heijnis, H. (2000). Signatures of natural 1480 catastrophic events and anthropogenic impact in an estuarine environment, New Zealand. Marine 1481 Geology, 167(3-4), 285-301. doi.org/10.1016/s0025-3227(00)00035-9

1482 Chen, B., & Wise, D. H. (1999). Bottom‐up limitation of predaceous arthropods in a detritus‐ 1483 based terrestrial food web. Ecology, 80(3), 761-772. doi.org/10.1890/0012- 1484 9658(1999)080[0761:bulopa]2.0.co;2

1485 Clout, M. N., & Lowe, S. J. (2000). Invasive species and environmental changes in New 1486 Zealand. Invasive species in a changing world, 369-383.

1487 Clout, M. N., & Russell, J. C. (2006). The eradication of mammals from New Zealand 1488 islands. Assessment and Control of Biological Invasion Risks’.(Eds F. Koike, MN Clout, M. 1489 Kawamichi, M. De Poorter and K. Iwatsuki.) pp, 127-141.

1490 Collier, K. J., Bury, S., & Gibbs, M. (2002). A stable isotope study of linkages between stream 1491 and terrestrial food webs through spider predation. Freshwater Biology, 47(9), 1651-1659. 1492 doi.org/10.1046/j.1365-2427.2002.00903.x

1493 Craddock, P. (1997). The effect of rodent control on invertebrate communities in coastal forest 1494 near Auckland (Doctoral dissertation, ResearchSpace@ Auckland).

1495 De Lange, P. J., Norton, D. A. & Molloy, B. P. J. (1995) Proceedings of a workshop hosted by 1496 Threatened Species Unit, Department of Conservation, 83-104.

1497 Del-Claro, K., Stefani, V., Nahas, L., & Torezan-Silingardi, H. M. (2017). Spiders as plant 1498 partners: complementing ant services to plants with extrafloral nectaries. In Behaviour and 1499 ecology of spiders (pp. 215-226). Springer, Cham. doi.org/10.1007/978-3-319-65717-2_8

1500 Diamond, J. M., & Veitch, C. R. (1981). Extinctions and introductions in the New Zealand 1501 avifauna: cause and effect? Science, 211(4481), 499-501. doi.org/10.1126/science.211.4481.499

1502 Dias, S. C., Brescovit, A. D., Couto, E. C., & Martins, C. F. (2006). Species richness and 1503 seasonality of spiders (Arachnida, Araneae) in an urban Atlantic Forest fragment in Northeastern 1504 Brazil. Urban Ecosystems, 9(4), 323-335. doi.org/10.1007/s11252-006-0002-7

1505 Digweed, S. C. (1995). Digging out the" digging-in effect" of pitfall traps: Influences of 1506 depletion and disturbance on catches of ground beetles (Coleoptera: Carabidae). Pedobiologia, 1507 39, 561-576.

1508 Dunn, R. R. (2005). Modern insect extinctions, the neglected majority. Conservation Biology, 1509 19(4), 1030-1036. doi.org/10.1111/j.1523-1739.2005.00078.x

1510 Eggs, B., & Sanders, D. (2013). Herbivory in spiders: the importance of pollen for orb-weavers. 1511 PloS One, 8(11), e82637. doi.org/10.1371/journal.pone.0082637

1512 Forster, R., & Forster, L. (1999). Spiders of New Zealand and their worldwide kin. University of 1513 Otago Press. doi.org/10.1046/j.1440-6055.2000.0155b.x

1514 Fuchs, S. (1982). Optimality of parental investment: the influence of nursing on reproductive 1515 success of mother and female young house mice. Behavioral Ecology and Sociobiology, 10(1), 1516 39-51. doi.org/10.1007/bf00296394

1517 Gibbs, G. (2010). Do New Zealand invertebrates reflect the dominance of birds in their 1518 evolutionary history? New Zealand Journal of Ecology, 34(1), 152-157.

1519 Goddard, M. A., Dougill, A. J., & Benton, T. G. (2010). Scaling up from gardens: biodiversity 1520 conservation in urban environments. Trends in ecology & evolution, 25(2), 90-98. 1521 doi.org/10.1016/j.tree.2009.07.016.

1522 Gómez, J. E., Lohmiller, J., & Joern, A. (2016). Importance of vegetation structure to the 1523 assembly of an aerial web-building spider community in North American open grassland. The 1524 Journal of Arachnology, 44(1), 28-36. doi.org/10.1636/p14-58.1

1525 Google maps, 2020. Map of Waitati, Otago. Retrieved from URL 1526 https://goo.gl/maps/TPoPDVb2Umdq6WdT7. 1527 https://www.google.co.nz/maps/@-45.7476906,170.5658809,195m/data=!3m1!1e3

1528 Green, C.J., 2002. Turning the tide: the eradication of invasive species (proceedings of the 1529 international conference on eradication of island invasives). In: Veitch, C.R., Clout, M.N. 1530 (Eds.), International Conference on Eradication of Island Invasives.

1531 Greenstone, M. H. (1984). Determinants of web spider species diversity: vegetation structural 1532 diversity vs. prey availability. Oecologia, 62(3), 299-304. doi.org/10.1007/bf00384260

1533 Gunnarsson, B. (1983). Winter mortality of spruce-living spiders: effect of spider interactions 1534 and bird predation. Oikos, 226-233. doi.org/10.2307/3544586

1535 Halaj, J., Cady, A. B., & Uetz, G. W. (2000). Modular habitat refugia enhance generalist 1536 predators and lower plant damage in soybeans. Environmental Entomology, 29(2), 383-393. 1537 doi.org/10.1093/ee/29.2.383

1538 Halaj, J., Ross, D. W., & Moldenke, A. R. (2000). Importance of habitat structure to the 1539 arthropod food‐web in Douglas‐fir canopies. Oikos, 90(1), 139-152. doi.org/10.1034/j.1600- 1540 706.2000.900114.x

1541 Hancock, M. H., & Legg, C. J. (2012). Pitfall trapping bias and arthropod body mass. Insect 1542 Conservation and Diversity, 5(4), 312-318. doi.org/10.1111/j.1752-4598.2011.00162.x

1543 Hardouin, E. A., Chapuis, J. L., Stevens, M. I., Van Vuuren, J. B., Quillfeldt, P., Scavetta, R. J., 1544 ... & Tautz, D. (2010). House mouse colonization patterns on the sub-Antarctic Kerguelen 1545 Archipelago suggest singular primary invasions and resilience against re-invasion. BMC 1546 Evolutionary Biology, 10(1), 325. doi.org/10.1186/1471-2148-10-325

1547 Harper, G. A., & Bunbury, N. (2015). Invasive rats on tropical islands: their population biology 1548 and impacts on native species. Global Ecology and Conservation, 3, 607-627. 1549 doi.org/10.1016/j.gecco.2015.02.010

1550 Hobson, K. A., Drever, M. C., & Kaiser, G. W. (1999). Norway rats as predators of burrow- 1551 nesting seabirds: insights from stable isotope analyses. The Journal of Wildlife Management, 63, 1552 14-25. doi.org/10.2307/3802483

1553 Innes, J., & Saunders, A. (2011). Eradicating multiple pests: an overview. Island Invasives: 1554 Eradication and Management’.(Eds CR Veitch, MN Clout and DR Towns.) pp, 177-181.

1555 Jackson, R. R., & Pollard, S. D. (1990). Intraspecific interactions and the function of courtship in 1556 mygalomorph spiders: a study of Porrhothele antipodiana (Araneae: Hexathelidae) and a 1557 literature review. New Zealand Journal of Zoology, 17(4), 499-526. 1558 doi.org/10.1080/03014223.1990.10422949

1559 Jones, S. E. (1941). Influence of temperature and humidity on the life history of the spider 1560 Agelena naevia Walckenaer. Annals of the Entomological Society of America, 34(3), 557-571. 1561 doi.org/10.1093/aesa/34.3.557

1562 Kelly, D., Ladley, J. J., Robertson, A. W., & Norton, D. A. (2000). Limited forest fragmentation 1563 improves reproduction in the declining New Zealand mistletoe Peraxilla tetrapetala 1564 (Loranthaceae). Conservation Biology Series-Cambridge-, 241-252. 1565 doi.org/10.1017/cbo9780511623448.018

1566 Knauer, A. C., Bakhtiari, M., & Schiestl, F. P. (2018). Crab spiders impact floral-signal 1567 evolution indirectly through removal of florivores. Nature Communications, 9(1), 1367. 1568 doi.org/10.1038/s41467-018-03792-x

1569 Kuijt, J., & Hansen, B. (2015). Loranthaceae. In Flowering Plants. Eudicots (pp. 73-119). 1570 Springer, Cham. doi.org/10.1007/978-3-319-09296-6_14

1571 Laing, D. J. (1975). The postures of the tunnel web spider Porrhothele antipodiana: a 1572 behavioural study. Tuatara, 21(3), 108-120.

1573 Laing, D. J. (1979). Studies on populations of the tunnel web spider Porrhothele antipodiana 1574 (Mygalomorphae: Dipluridae). II. Relationship with hunting wasps (Pompilidae). Tuatara.

1575 Lange, M., Gossner, M. M., & Weisser, W. W. (2011). Effect of pitfall trap type and diameter on 1576 vertebrate by‐catches and ground beetle (Coleoptera: Carabidae) and spider (Araneae) sampling. 1577 Methods in Ecology and Evolution, 2(2), 185-190. doi.org/10.1111/j.2041-210x.2010.00062.x

1578 Le Roux, V., Chapuis, J. L., Frenot, Y., & Vernon, P. (2002). Diet of the house mouse (Mus 1579 musculus) on Guillou Island, Kerguelen archipelago, Subantarctic. Polar Biology, 25(1), 49-57. 1580 doi.org/10.1007/s003000100310

1581 Lettink, M., & Patrick, B. H. (2006). Use of artificial cover objects for detecting red katipō, 1582 Latrodectus katipo Powell (Araneae: Theridiidae). New Zealand Entomologist, 29(1), 99-102. 1583 doi.org/10.1080/00779962.2006.9722143

1584 Lewis, R. S., & Hurst, J. L. (2004). The assessment of bar chewing as an escape behaviour in 1585 laboratory mice. Animal Welfare-Potters Bar then Wheathampstead, 13(1), 19-26.

1586 Lövei, G. L., Horváth, R., Elek, Z., & Magura, T. (2019). Diversity and assemblage filtering in 1587 ground-dwelling spiders (Araneae) along an urbanisation gradient in Denmark. Urban 1588 Ecosystems, 22(2), 345-353. doi.org/10.1007/s11252-018-0819-x

1589 MacMillan, H. A., & Sinclair, B. J. (2011). Mechanisms underlying insect chill-coma. Journal of 1590 Insect Physiology, 57(1), 12-20. doi.org/10.1016/j.jinsphys.2010.10.004

1591 Marc, P., Canard, A., & Ysne, F. (1999). Spiders (Araneae) useful for pest limitation and 1592 bioindication. In Invertebrate Biodiversity as Bioindicators of Sustainable Landscapes (pp. 229- 1593 273). doi.org/10.1016/b978-0-444-50019-9.50015-7

1594 McCormick S., & Polis, G. A. (1982). Arthropods that prey on vertebrates. Biological Reviews, 1595 57(1), 29-58. doi.org/10.1111/j.1469-185x.1982.tb00363.x

1596 McGlone, M. S. (1983). Polynesian deforestation of New Zealand: a preliminary synthesis. 1597 Archaeology in oceania, 18(1), 11-25.

1598 McGlone, M. S., & Wilmshurst, J. M. (1999). Dating initial Maori environmental impact in New 1599 Zealand. Quaternary International, 59(1), 5-16. doi.org/10.1016/s1040-6182(98)00067-6

1600 McWethy, D. B., Whitlock, C., Wilmshurst, J. M., McGlone, M. S., Fromont, M., Li, X., ... & 1601 Cook, E. R. (2010). Rapid landscape transformation in South Island, New Zealand, following 1602 initial Polynesian settlement. Proceedings of the National Academy of Sciences, 107(50), 21343- 1603 21348.

1604 Meehan, C. J., Olson, E. J., Reudink, M. W., Kyser, T. K., & Curry, R. L. (2009). Herbivory in a 1605 spider through exploitation of an ant–plant mutualism. Current Biology, 19(19), R892-R893. 1606 doi.org/10.1016/j.cub.2009.08.049

1607 Mellanby, K. (1939). Low temperature and insect activity. Proceedings of the Royal Society of 1608 London. Series B-Biological Sciences, 127(849), 473-487.

1609 Miskelly, C. M., Dowding, J. E., Elliott, G. P., Hitchmough, R. A., Powlesland, R. G., 1610 Robertson, H. A., ... & Taylor, G. A. (2008). Conservation status of New Zealand birds, 2008. 1611 Notornis, 55(3), 117-135.

1612 Morley, E. L., & Robert, D. (2018). Electric fields elicit ballooning in spiders. Current Biology, 1613 28(14), 2324-2330. doi.org/10.1016/j.cub.2018.05.057

1614 Moulder, B. C., & Reichle, D. E. (1972). Significance of Spider Predation in the Energy 1615 Dynamics of Forest‐Floor Arthropod Communities. Ecological Monographs, 42(4), 473-498. 1616 doi.org/10.2307/1942168

1617 Müller, K., Holmes, A., Deurer, M., & Clothier, B. E. (2015). Eco-efficiency as a sustainability 1618 measure for kiwifruit production in New Zealand. Journal of Cleaner Production, 106, 333-342. 1619 doi.org/10.1016/j.jclepro.2014.07.049.

1620 Norbury, G., van den Munckhof, M., Neitzel, S., Hutcheon, A., Reardon, J., & Ludwig, K. 1621 (2014). Impacts of invasive house mice on post-release survival of translocated lizards. New 1622 Zealand Journal of Ecology, 38(2), 322-327.

1623 Nyffeler, M., & Vetter, R. S. (2018). Black widow spiders, Latrodectus spp. (Araneae: 1624 Theridiidae), and other spiders feeding on mammals. The Journal of Arachnology, 46(3), 541- 1625 549. doi.org/10.1636/joa-s-18-026.1

1626 O’Donnell, C. F. (2010). The ecology and conservation of New Zealand bats. Island bats: 1627 evolution. ecology and conservation. Chicago University Press, Chicago, 460-495.

1628 Olson, D. M. (1991). A comparison of the efficacy of litter sifting and pitfall traps for sampling 1629 leaf litter ants (Hymenoptera, Formicidae) in a tropical wet forest, Costa Rica. Biotropica, 166- 1630 172. doi.org/10.2307/2388302

1631 Paquin, P., Vink, C. J., & Dupérré, N. (2010). Spiders of New Zealand: annotated family key and 1632 species list. Manaaki Whenua Press, Landcare Research.

1633 Pitt, W. C., Sugihara, R. T., Driscoll, L. C., & Vice, D. S. (2011). Physical and behavioural 1634 abilities of commensal rodents related to the design of selective rodenticide bait 1635 stations. International Journal of Pest Management, 57(3), 189-193. 1636 doi.org/10.1080/09670874.2011.561889

1637 Poinar Jr, G. O., & Early, J. W. (1990). Aranimermis giganteus n. sp. (Mermithidae: Nematoda), 1638 a parasite of New Zealand mygalomorph spiders (Araneae: Arachnida). Revue de Nématologie, 1639 13(4), 403-410.

1640 Prenter, J., MacNeil, C., & Elwood, R. W. (2006). Sexual cannibalism and mate choice. Animal 1641 Behaviour, 71(3), 481-490. doi.org/10.1016/j.anbehav.2005.05.011

1642 Rajeswaran, J., Duraimurugan, P., & Shanmugam, P. S. (2005). Role of spiders in agriculture 1643 and horticulture ecosystem. Journal of Food Agriculture and Environment, 3(3/4), 147.

1644 Rate, S. R. (2009). Does rat control benefit forest invertebrates at Moehau, Coromandel 1645 Peninsula? Pub. Team, Department of Conservation.

1646 Régnier, C., Achaz, G., Lambert, A., Cowie, R. H., Bouchet, P., & Fontaine, B. (2015). Mass 1647 extinction in poorly known taxa. Proceedings of the National Academy of Sciences, 112(25), 1648 7761-7766. doi.org/10.1073/pnas.1502350112

1649 Rozicka, V. (1995). The spreading of Ostearius melanopygius (Araneae: Linyphiidae) through 1650 Central Europe. Short note, Institute of entomology, Branišovská.

1651 Ruscoe, W.A., 2001. Advances in New Zealand mammalogy 1990–2000: house mouse. Journal 1652 of the Royal Society of New Zealand 31, 127–134. doi.org/10.1080/03014223.2001.9517643

1653 Russell, J. C., Towns, D. R., & Clout, M. N. (2008). Review of rat invasion biology: implications 1654 for island biosecurity. Science for Conservation, (286), 53.

1655 Russell, J. C., Towns, D. R., Anderson, S. H., & Clout, M. N. (2005). Intercepting the first rat 1656 ashore. Nature, 437(7062), 1107. doi.org/10.1038/4371107a

1657 Rypstra, A. L., & Buddle, C. M. (2013). Spider silk reduces insect herbivory. Biology Letters, 1658 9(1), 20120948. doi.org/10.1098/rsbl.2012.0948

1659 Rypstra, A. L., & Marshall, S. D. (2005). Augmentation of soil detritus affects the spider 1660 community and herbivory in a soybean agroecosystem. Entomologia Experimentalis et 1661 Applicata, 116(3), 149-157. doi.org/10.1111/j.1570-7458.2005.00322.x

1662 Scott, A. G., Oxford, G. S., & Selden, P. A. (2006). Epigeic spiders as ecological indicators of 1663 conservation value for peat bogs. Biological Conservation, 127(4), 420-428. 1664 doi.org/10.1016/j.biocon.2005.09.001

1665 Seldon, D. S., & Beggs, J. R. (2010). The efficacy of baited and live capture pitfall traps in 1666 collecting large-bodied forest carabids. New Zealand Entomologist, 33(1), 30-37. 1667 doi.org/10.1080/00779962.2010.9722189

1668 Shiels, A. B., & Drake, D. R. (2011). Are introduced rats (Rattus rattus) both seed predators and 1669 dispersers in Hawaii? Biological Invasions, 13(4), 883-894. doi.org/10.1007/s10530-010-9876-7

1670 Shiels, A. B., Flores, C. A., Khamsing, A., Krushelnycky, P. D., Mosher, S. M., & Drake, D. R. 1671 (2013). Dietary niche differentiation among three species of invasive rodents (Rattus rattus, R. 1672 exulans, Mus musculus). Biological Invasions, 15(5), 1037-1048. doi.org/10.1007/s10530-012- 1673 0348-0

1674 Shiukhy, S., Raeini-Sarjaz, M., & Chalavi, V. (2015). Colored plastic mulch microclimates 1675 affect strawberry fruit yield and quality. International Journal of Biometeorology, 59(8), 1061- 1676 1066. oi.org/10.1007/s00484-014-0919-0

1677 Simpson, N. (1995). Mistletoes in Otago. Proceedings of a workshop hosted by Threatened 1678 Species Unit, Department of Conservation, 74-76.

1679 Sinclair, B. J., Lord, J. M., & Thompson, C. M. (2001). Microhabitat selection and seasonality of 1680 alpine invertebrates. Pedobiologia, 45(2), 107-120. doi.org/10.1078/0031-4056-00073

1681 Sinclair, L., McCartney, J., Godfrey, J., Pledger, S., Wakelin, M., & Sherley, G. (2005). How did 1682 invertebrates respond to eradication of rats from Kapiti Island, New Zealand? New Zealand 1683 Journal of Zoology, 32(4), 293-315. doi.org/10.1080/03014223.2005.9518421

1684 Sinkkonen, A., Somerkoski, E., Paaso, U., Holopainen, J. K., Rousi, M., & Mikola, J. (2012). 1685 Genotypic variation in yellow autumn leaf colours explains aphid load in silver birch. New 1686 Phytologist, 195(2), 461-469.

1687 Sirvid, P. J., Vink, C. J., Wakelin, M. D., Fitzgerald, B. M., Hitchmough, R. A., & Stringer, I. A. 1688 N. (2012). The conservation status of New Zealand Araneae. New Zealand Entomologist, 35(2), 1689 85-90. doi.org/10.1080/00779962.2012.686310

1690 Sirvid, P. Museum of New Zealand Te Papa Tongarewa, Wellington, NZ - Black tunnelweb 1691 spider (Porrhothele antipodiana), retrieved from htps://collections.tepapa.govt.nz/topic/9426.

1692 Spence, J. R., & Niemelä, J. K. (1994). Sampling carabid assemblages with pitfall traps: the 1693 madness and the method. The Canadian Entomologist, 126(3), 881-894. 1694 doi.org/10.4039/ent126881-3

1695 St Clair, J. J. (2011). The impacts of invasive rodents on island invertebrates. Biological 1696 Conservation, 144(1), 68-81. doi.org/10.1016/j.biocon.2010.10.006

1697 Stenseth, N. C., Leirs, H., Skonhoft, A., Davis, S. A., Pech, R. P., Andreassen, H. P., ... & Zhang, 1698 Z. (2003). Mice, rats, and people: the bio‐economics of agricultural rodent pests. Frontiers in 1699 Ecology and the Environment, 1(7), 367-375. 1700 doi.org/10.1890/1540-9295(2003)001[0367:mraptb]2.0.co;2

1701 Sutherland, R. J., & Dyck, R. H. (1984). Place navigation by rats in a swimming pool. Canadian 1702 Journal of Psychology/Revue canadienne de psychologie, 38(2), 322. doi.org/10.1037/h0080832

1703 Symonds, M. R., & Moussalli, A. (2011). A brief guide to model selection, multimodel inference 1704 and model averaging in behavioural ecology using Akaike’s information criterion. Behavioral 1705 Ecology and Sociobiology, 65(1), 13-21. doi.org/10.1007/s00265-010-1037-6

1706 Tahir, H. M., Khalid, N., Hamza, A., & Nadeem, J. (2019). Effect of spider silk on the foraging 1707 activity of Acrida turrita (Linnaeus, 1758)(Orthoptera: Acridinae). Oriental insects, 53(2), 151- 1708 159. doi.org/10.1080/00305316.2018.1460281

1709 Théry, M., & Casas, J. (2002). Visual systems: predator and prey views of spider camouflage. 1710 Nature, 415(6868), 133. doi.org/10.1038/415133a

1711 Uetz, G. W. (1991). Habitat structure and spider foraging. In Habitat structure (pp. 325-348). 1712 Springer, Dordrecht. doi.org/10.1007/978-94-011-3076-9_16

1713 Uhl, G., Zimmer, S. M., Renner, D., & Schneider, J. M. (2015). Exploiting a moment of 1714 weakness: male spiders escape sexual cannibalism by copulating with moulting females. 1715 Scientific Reports, 5, 16928. doi.org/10.1038/srep16928

1716 Vasconcellos-Neto, J., Messas, Y. F., da Silva Souza, H., Villanueva-Bonila, G. A., & Romero, 1717 G. Q. (2017). Spider–plant interactions: an ecological approach. In Behaviour and Ecology of 1718 Spiders (pp. 165-214). Springer, Cham. doi.org/10.1007/978-3-319-65717-2_7

1719 Venner, S., Pasquet, A., & Leborgne, R. (2000). Web-building behaviour in the orb-weaving 1720 spider Zygiella x-notata: influence of experience. Animal Behaviour, 59(3), 603-611. 1721 doi.org/10.1006/anbe.1999.1327

1722 Villanueva-Bonilla, G. A., Salomão, A. T., & Vasconcellos-Neto, J. (2017). Trunk structural 1723 traits explain habitat use of a tree-dwelling spider (Selenopidae) in a tropical forest. Acta 1724 Oecologica, 85, 108-115. doi.org/10.1016/j.actao.2017.10.004

1725 von Koenigswald, W. (1985). Evolutionary trends in the enamel of rodent incisors. In 1726 Evolutionary relationships among rodents (pp. 403-422). Springer, Boston, MA. 1727 doi.org/10.1007/978-1-4899-0539-0_15

1728 Vriend, N. M., Arran, M., Louge, M. Y., Hay, A. G., & Valance, A. (2017). On the Internal 1729 Structure of Mobile Barchan Sand Dunes due to Granular Processes. In AGU Fall Meeting 1730 Abstracts. American Geophysical Union, Washington.

1731 Waite, E. M. (2012). The role of large, isolated trees in supporting urban biodiversity (Doctoral 1732 dissertation, University of Otago).

1733 Wardle, P. (1978). Seasonality in New Zealand plants. New Zealand Entomologist, 6(4), 344- 1734 349. doi.org/10.1080/00779962.1978.9722285

1735 Watson, D. M. (2001). Mistletoe - a keystone resource in forests and woodlands worldwide. 1736 Annual Review of Ecology and Systematics, 32(1), 219-249. 1737 doi.org/10.1146/annurev.ecolsys.32.081501.114024

1738 Watson, D. M., & Herring, M. (2012). Mistletoe as a keystone resource: an experimental test. 1739 Proceedings of the Royal Society B: Biological Sciences, 279(1743), 3853-3860. 1740 doi.org/10.1098/rspb.2012.0856

1741 Wearing, C. H., Walker, J. T. S., & Suckling, D. M. (1992, August). Development of IPM for 1742 New Zealand apple production. In II International Symposium on Integrated Fruit Production, 1743 347 (pp. 277-284). doi.org/10.17660/actahortic.1993.347.32.

1744 Webb, C. J., & Kelly, D. (1993). The reproductive biology of the New Zealand flora. Trends in 1745 Ecology & Evolution, 8(12), 442-447. doi.org/10.1016/0169-5347(93)90007-c

1746 Webb, C. J., Lloyd, D. G., & Delph, L. F. (1999). Gender dimorphism in indigenous New 1747 Zealand seed plants. New Zealand Journal of Botany, 37(1), 119-130. 1748 doi.org/10.1080/0028825x.1999.9512618

1749 Weber, M. G., Porturas, L. D., & Taylor, S. A. (2016). Foliar nectar enhances plant–mite 1750 mutualisms: the effect of leaf sugar on the control of powdery mildew by domatia-inhabiting 1751 mites. Annals of Botany, 118(3), 459-466. doi.org/10.1093/aob/mcw118

1752 Williams, D. S. (1979). The feeding behaviour of New Zealand Dolomedes species (Araneae: 1753 Pisauridae). New Zealand Journal of Zoology, 6(1), 95-105. 1754 doi.org/10.1080/03014223.1979.10428352

1755 Williams, G. (1958). Mechanical time-sorting of pitfall captures. Anim. Ecol. 27, 2735. 1756 doi.org/10.2307/2171

1757 Wimp, G. M., Murphy, S. M., Lewis, D., Douglas, M. R., Ambikapathi, R., Gratton, C., & 1758 Denno, R. F. (2013). Predator hunting mode influences patterns of prey use from grazing and 1759 epigeic food webs. Oecologia, 171(2), 505-515. doi.org/10.1007/s00442-012-2435-4

1760 Yip, E. C., Clarke, S., & Rayor, L. S. (2009). Aliens among us: nestmate recognition in the social 1761 huntsman spider, Delena cancerides. Insectes Sociaux, 56(3), 223-231. doi.org/10.1007/s00040- 1762 009-0015-3

1763 Zachariah, T. T., Mitchell, M. A., Guichard, C. M., & Singh, R. S. (2007). Hemolymph 1764 biochemistry reference ranges for wild-caught goliath birdeater spiders (Theraphosa blondi) and 1765 Chilean rose spiders (Grammostola rosea). Journal of Zoo and Wildlife Medicine, 38(2), 245- 1766 251. doi.org/10.1638/1042-7260(2007)038[0245:hbrrfw]2.0.co;2

1767 Zimmermann, M., & Spence, J. R. (1989). Prey use of the fishing spider Dolomedes triton 1768 (Pisauridae, Araneae): an important predator of the neuston community. Oecologia, 80(2), 187- 1769 194. doi.org/10.1007/bf00380149

1770

1771

1772

1773

1774

1775

1776

1777

1778 Appendix 1

1779 Total invertebrate numbers sampled from mistletoe and host plants identified to family or closest 1780 taxon. Summer sampling occurred from late November 2018 to late January 2019. Winter 1781 sampling occurred throughout the month of August of 2019. Brackets indicate mites sampled 1782 from leaf domatia that were excluded from statistical analysis as an outlier. Samples were 1783 collected from Dunedin and surrounding rural areas.

Total sampled Total sampled Total sampled Total sampled Family Order in Summer in winter on Illeostylus on host plants Araneae Clubionidae 0 1 1 0 Araneae Desidae 1 0 0 1 Araneae Salticidae 2 3 0 5 Araneae Theridiidae 3 6 8 1 Araneae Thomisidae 0 5 3 2

Oribatida Galumnidae 28 (1014) 19 13 34 (1014) Oribatida Neotrichozetidae 6 0 5 1 Trombidiformes Anystidae 8 0 4 4 Trombidiformes Erythraeidae 19 (101) 0 6 13 (101) Trombidiformes Tetranychidae 20 0 10 10

Coleoptera Coccinellidae 0 1 0 1 Coleoptera Curculionoidea 0 3 1 2 Diptera Chironomidae 0 1 0 1 Diptera Syrphidae 1 0 0 1 Hemiptera Aphididae 6 1 3 4 Hemiptera Acanthosomatidae 0 3 3 0 Hemiptera Cicadellidae 0 1 1 0 Hemiptera Diaspididae 12 0 12 0 Hemiptera Pseudococcidae 18 0 4 14 Hymenoptera Formicidae 1 0 1 0 Hymenoptera Ichneumonidae 1 0 1 0 Lepidoptera Geometridae 7 0 5 2 Lepidoptera Tortricidae 0 8 8 0 Phasmatodea Phasmatidae 1 1 2 0 Psocoptera Psocoptera 60 8 62 6 Entomobryomorpha Entomobryomorpha 1 0 1 0 Symphypleona Symphypleona 1 0 1 0 1784 Tubulifera Phlaeothripidae 1 0 1 0

1785

1786 Appendix 2

1787 Total Araneae numbers sampled by species caught in pitfall traps (P) vs Sancta Aranearum (S) in 1788 pasture (G) vs forest (F) over a three-month period at the Waitati (A) and Taieri (B) field sites.

1789

1790 A

1791

1792

1793

1794

1795

1796

1797 B