Solanum mauritianum is an evergreen invasive shrub (Olckers, 1998; Olckers, 1999; Fig. 1.3a); native to Argentina, Brazil, Uruguay, and Paraguay (Cowie et al., 2018). It was accidentally introduced globally during the 16th Century by a ship of the Portuguese trade routes (Olckers, 2011); therefore, becoming problematic in Australia, Madagascar, New Zealand and South Africa (Singh and Olckers, 2017). It was first recorded in KwaZulu-Natal, South Africa, in the 1860s, as an ornamental plant (Singh and Olckers, 2017). It has been estimated to have invaded approximately 80 500-ha (Cowie et al., 2018), but mainly flourishing in higher rainfall regions of South Africa (Olckers, 1998). It is a transformer species, thus converts the structure of plant 22 communities by outcompeting native plant species (Olckers, 2011). Moreover, its fast-growing nature shades out surrounding vegetation (Bromilow, 2010; Cowie et al., 2018). Mechanical, chemical and biological management options are used to control S. mauritianum, but copious seed production and high seed survival in seed banks promote an aggressive return post-management (Campbell and van Staden, 1983; Goodall et al., 2017). South Africa is the only country with two released biocontrol agents against S. mauritianum (Olckers, 2011; Cowie et al., 2018), with New Zealand only recently implementing a biocontrol programme (Singh and Olckers, 2017). Although many biocontrol experiments have been conducted to combat invasive plants such as S. mauritianum (Olckers, 2011), none have been entirely effective and invasive shrubs remains highly problematic. Unripe fruits of S. mauritianum are toxic (Henderson, 2001; Cowie et al., 2018). However, ripe fruits serve as food sources for frugivorous avian species (Jordaan et al., 2011a). A dietary shift by native frugivorous avian species with a preference for (Witkowski and Garner, 2008) and inclusion of S. mauritianum more than native fleshy-fruited plant species have been observed (Mokotjomela et al., 2013), potentially dispersing the invasive shrub. In addition to avian species, mammalian species such as bats (Olckers, 2011; Jordaan et al., 2011a) and potentially ungulates may serve as seed vectors of S. mauritianum. The invasive tree grows between 2-10 m and sometimes higher (Henderson, 2001; Singh and Olckers, 2017), mainly fruiting from September to March, and sometimes until June, with some reports of fruiting all-year-round (Goodall et al., 2017). On average, annual seed production ranges between 100,000-20,000 seeds/plant (Olckers, 2011; Singh and Olckers, 2017) of fruits with a mean fruit size of 13.7 mm (Thabethe et al., 2015b; Bitani et al., 2020). Much consideration has been given to the germination of S. mauritianum after ingestion by avian species (Campbell and van Staden, 1983; Jordaan et al., 2011a; Thabethe et al., 2015b), which has shown high germination seed viability patterns (Campbell and van Staden, 1983; Witkowski and Garner, 2008). However, these are influenced by seasonality and seed age (Campbell et al., 1992; Goodall et al., 2017). Nevertheless, research is required on the influences of seed dispersal by mammalian species such as ungulate species.
Melia azedarach (syringa) is an invasive deciduous tree (Botha and Penrith, 2009; Voigt et al., 2011; Fig. 1.4a), originating from southern Asia to northern Australia (Tourn et al., 1999; Tilney et al., 2018). Within the native range, the South African cultivar is native to India (Voigt et al., 2011). In South Africa, it was introduced for ornamental purposes (Henderson, 2001; Stavarache et al., 2008) and first recorded in Cape Town in the 1800s (Voigt et al., 2011). It has highly invaded the margins of forests, roadsides and riparian zones (Holmes et al., 2005; Voigt et al., 2011), especially in the savannah biome (Tilney et al., 2018). Due to the invasion, native ecosystems have been transformed by converting the structure of plant communities (Henderson, 2001). Although M. azedarach have a relatively short lateral root system, their tap root is sent deep underground for water (Tourn et al., 1999), therefore possibly outcompeting indigenous plants for water resources. Moreover, the invasive tree is highly defensive against insects and pathogens, therefore, more advantageous than native plant species (Stavarache et al., 2008). All parts of the plant are poisonous (Ntalli et al., 2010; Lungu et al., 2011), especially the fruits (Stavarache et al., 2008). The toxicity of fruits is influenced by environmental variables, with some trees not being toxic at all (Botha and Penrith, 2009). The invasive tree grows between 12 m (Corlett, 2005; Bitani et al., 2020) and up to 23 m (Henderson, 2001), and fruits during the dry season in South Africa (Voigt et al., 2011). The single-seeded fleshy drupes are characterized by a mean fruit size of 11.9 mm (Corlett, 2005; Bitani et al., 2020; Fig. 1.4d; Fig. 1.4c). Fruit bats (Jordaan et al., 2011b) and avian species largely source the ripe fruits for food (Wirminghaus et al., 2002; Voigt et al., 2011). Therefore, further investigation and understanding of the seed dispersal pathways of M. azedarach are important for effective management strategies.
Animal visitation and fruit consumption observations were conducted simultaneously during the fruiting seasons of L. camara, S. mauritianum, M. azedarach, and O. monacantha using camera-trapping. We used simple random sampling to select each camera-trap station, determined by the presence of ripe fruits on the shrubs or trees of AIPs, treating each camera-trap station location as a random observational point. A total of 20 camera-trap (Moultrie M880i, LLC, Alabaster, USA) sites were established, six at FHE (21/05/2018 – 11/06/2018), eight at NLNP (18/07/2019 – 20/08/2019), and six at Zingela (26/10/2019 – 16/11/2019) for 21 consecutive days per camera-trap location. We standardised the camera-trap survey across all the study areas for consistency; however, we set some for photographic images and others for video recording. Before the commencement of the camera-trap installation, we conducted a field assessment survey to select camera-trap stations of areas invaded by the invasive plants, through direct observations done by driving and walking active wildlife trial transects. Although the camera-trap stations within the infested areas were not placed using a systematic grid, each camera-trap station selected was guided by an inter-trap distance of 200 m or more, depending on the size of fruiting stands of the invasive plants collectively. The chosen camera-trap stations were placed viewing and focusing on the fruiting stands of the invasive plants near active wildlife trails, guided by the presence of ungulate dung signifying habitat-use by ungulates in that sampling area. Camera-traps were installed approximately 1-3 m away from the fruiting invasive plants, and between 2-6 m in height to accommodate for the varying heights of ungulate species present in the area.
We examined all images and videos captured by the camera-traps, and independent animal visitation events of individuals were considered using the procedures described in Rivas-Romero and Soto-Shoender (2015) and Campos et al. (2018). The number of fruits consumed could not be quantified with the trap camera; therefore, the total number of fruits consumed by the ungulates was substituted and recorded as the total number of visitations with fruit consumption (the act of seed ingestion observed in the photographs and videos). The numbers of visitations with fruit consumption by ungulates to the fruiting alien invasive plants were considered as an estimate of the intensity of interactions; and total numbers of independent animal visits (with or without fruit consumption) were considered as the frequency of interactions (Schupp 1993; Campos et al. 2018). We compared both frequencies of interactions and intensity of interactions between ungulate species for each invasive plant using Chi-square tests to determine significant differences in their patterns between different ungulates. Once the community of ungulate species visiting and ingesting the alien invasive plants was determined, the overall strength of interactions (combining both frequencies of interactions and intensity of interactions) was determined. We conducted all statistical analyses using IBM SPSS Statistics (SPSS Inc, version 27, USA).
Dung depository survey
The dung deposition survey was conducted in May, July and October 2019 and June, July and August 2020, to determine dung deposition patterns and habitat occupancy by ungulate species. We conducted the surveys opportunistically using a random sampling where random transact walks along active wildlife trials were performed covering a wide range of microhabitat types (grassland, savannah, thicket, woodland, and degraded woodland). Microhabitats were described according to Kleynhans et al. (2011) and Mucina and Rutherford (2011) (Supplementary information Table S3.1). The degraded woodland microhabitat is essentially equivalent to the woodland microhabitat. However, the degraded woodland was highly eroded and invaded by O. monacantha, for this reason separately classified. During each transect walk, we recorded the geographical location of the deposited dung pile using a Global Positioning System (GPS) (Montana 650, Garmin), the ungulate species to which the dung pile belonged to, and the microhabitat type in which the dung pile was located (grassland, savannah, thicket, woodland, and degraded woodland). To avoid omitting dung samples, a 4-m buffer zone (2-m on each side of the transect) was used to include dung samples on either side of the transect line, as some ungulates do not follow trails. For our study, each dung pile was considered as a separate event, independent of other dung piles of the same species in close proximity. Overall, dung count percentage was used to determine potentially vulnerable habitats based on deposited dung frequency.
Data collection procedure
Tracking sympatric ungulates using dung piles and hoof markings is particularly difficult; therefore, a pilot study was conducted, accompanied by a dung identification expert for a crash course. Thereafter, identifications were made using book guides showing images, size and shape comparisons (Walker, 1986; Stuart and Stuart, 1994, 2000; Cillé, 2000; Murray, 2011). Furthermore, the book guides shared possible misidentifications of similar pellets/piles and how to resolve these misidentifications (Stuart and Stuart, 1994, 2000). We also used samples that were correctly identified by the experts during the crash course as cross-referencing samples. The surrounding dung deposition location was scrutinised to aid in the overall dung identification, thereby developing a step-by-step identification method. Firstly, ungulate hoof markings near the dung piles were identified: to assist in the identification of the dung sample itself. Secondly, the nature of the defecated dung pile (i.e. clumped, midden or scattered) was deduced to narrow the possible species. Lastly, a fraction of pellets from the dung pile were identified using shape and size, as some dung piles looked identical but varied in individual pellet size. The greatest factor that made dung identification slightly unproblematic was that ungulate species occurring in each study site were known before the study commenced, therefore narrowing the species spectrum of possible dung samples to be found. However, we acknowledge that the process of identifying individual pellets can be difficult, especially for sympatric ungulate species (Hibert et al., 2008). Though dung collection is inexpensive because dung piles are readily available (Wrench, 1997; Smith, 2016), dung processing for seed harvesting is time-consuming. Therefore, focusing only on four invasive plant species for this study eased the processing stage. Faecal decay is problematic when conducting a dung deposition survey (Jenkins and Manly, 2008), therefore potentially obscuring the correlation of dung piles with habitat occupation. For this reason, decaying dung piles were not considered as it was difficult to identify (Hibert et al., 2008; Smith, 2016).
Seed extraction from faecal samples
Fresh dung piles were opportunistically collected simultaneously during the dung deposition survey. Before selecting a dung pile to be collected, we targeted fresh dung samples that were ‘recently defecated’. We achieved this by selecting those that were: not scorched and dried up by the sun, soft-to-touch, and in some cases still warm-to-touch and had the presence of small flying insects (indicating freshness). Fresh dung piles with the presence of beetles were not considered because of the possibility of seed predation by beetles before collection. Dung samples were collected in clear plastic bags, labelled, and weighed (wet mass, in kg) in the field, and stored at the University of KwaZulu-Natal, Pietermaritzburg in a freezer (-17 ºC) until dissected for seed extraction. Dung weight was not controlled to fully report on seed prevalence without excluding a certain portion of the sample. Therefore, the lowermost part of the dung samples was left behind and not collected to avoid contamination of fallen seeds and other debris. Dung samples were individually dissected and broken apart, placed on a fine-mesh sand sieve (0.085 mm) and carefully strained to wash off dirt, to visibility see and collect the seeds. We then harvested and identified the seeds of alien invasive Lantana camara, Solanum mauritianum, Melia azedarach and Opuntia monacantha, and properly stored the seeds in labelled glass vials.
Table of contents :
CHAPTER 1 Introduction
1.1 Alien invasive plant species
1.2 Impacts of alien invasive plant species
1.3 Role of seed dispersal in the spread of invasive alien species
1.3.1 Endozoochory seed dispersal
1.3.2 Seed dispersal effectiveness
1.3.3 Effects of fruit characteristics on invasiveness
1.3.4 Effects of seed dispersal agent behaviour
1.3.5 Effects of seed traits and gut passage on germination patterns
1.4 Ungulate species as seed dispersal agents
1.5 Alien invasive plants in South Africa
1.6 Problem statement
1.7 Aims and objectives
1.8 Study sites
1.9 Study species
1.9.1 Lantana camara
1.9.2 Solanum mauritianum
1.9.3 Melia azedarach
1.9.4 Opuntia monacantha
1.10 Structure of the thesis
CHAPTER 2 To eat or not to eat: visitation and fruit consumption of fleshy-fruited alien invasive plants by wild southern African ungulates
2.8 Supplementary information
CHAPTER 3 Seed composition in the dung of wild southern African ungulates: implications for the dispersal of fleshly-fruited invasive alien plants
3.8 Supplementary information
CHAPTER 4 Potential mean retention time and seed dispersal distance of wild southern African ungulates in KwaZulu-Natal, South Africa
4.9 Supplementary information
CHAPTER 5 Conclusions and recommendations
5.2 Visitation and fruit consumption of fleshy-fruited alien invasive plants by ungulates
5.3 Seed composition of fleshy-fruited alien invasive plants in the dung of ungulate species
5.4 Potential retention times and dispersal distances of alien invasive plants by ungulates
5.5 Conclusions and recommendations