Seed biology seems to be one of the most important factors contributing to the invasion success of Australian acacias (Milton & Hall, 1981; Richardson & Kluge, 2008). Seed biology syndromes in many Acacia species are largely shaped by fire-driven ecosystems that are present throughout much of Australia and introduced Mediterranean-type climate regions. Fire-adaptive traits include: production of large quantities of hard-coated, heat-tolerant and long-lived seeds with the capacity for long dormancy; stimulation of germination by heat and/or smoke; seed dispersal and burial by ants; and the ability to resprout (Berg, 1975; Bell et al., 1993; Specht & Specht, 1999), all of which are likely essential for the persistence and invasive success of Australian acacias (see Fig. 4 for photographs of seed biology traits associated with invasiveness).
Figure 4. Important seed biology traits associated with invasiveness in Australian acacias. (a) Seed production of Acacia saligna in South Africa during the early 1980s, prior to the introduction of the rust fungus Uromycladium uromyces, which has since greatly reduced seed production (photograph: D.M. Richardson). (b) Seed production of A. longifolia in its native range in Australia (photograph: C. Harris). Seeds that fall to the ground can remain viable for 50+ years, making their eradication nearly impossible. (c) A. cyclops seeds remain in the tree canopy longer than those of species that are typically ant-dispersed; the bright red aril attracts birds that disperse the seeds (photograph: A.M. Rogers). (d) A. longifolia seeds are typically ant-dispersed in the native range, although bird-dispersal is predicted based on aril attributes; they are attached to the seed pod by an elaiosome that attracts ants (photograph: C. Harris). (e) Invasive species, such as A. saligna pictured here, have a greater tendency to resprout following a disturbance event than non-invasive species (photograph: D.M. Richardson). (f) The mass germination of Acacia seeds after fire, as in A. pycnantha in South Africa shown here, is a major hurdle to control efforts (photograph: D.M. Richardson).
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Dispersal is a crucial aspect of progression from ‘naturalized’ to ‘invasive’ status when recruitment occurs at considerable distances from parent plants (Richardson et al., 2000a,b). Australian acacias possess seed adaptations for dispersal by birds and ants (Davidson & Morton, 1984; O’Dowd & Gill, 1986), although passive dispersal via water, wind and gravity is also common.
Broadly, biotic seed dispersal in Acacia falls into two syndromes based on features of arils: a ‘bird-dispersal syndrome’ and an ‘ant-dispersal syndrome’ (O’Dowd & Gill, 1986). The fleshy arillate appendages (in bird-dispersed seeds) and an elaiosome (in ant-dispersed seeds) attach the seed to the seed pod lining and make them accessible to a range of bird and ant species across multiple foraging types. Such generalization of morphological traits associated with dispersal makes limitation of a seed dispersal agent in the introduced range unlikely (see Glyphis et al., 1981; Holmes, 1990a; Richardson et al., 2000a; Underhill & Hofmeyr, 2007). Furthermore, these traits may be evolutionarily labile since A. ligulata reportedly displays both syndromes (Davidson & Morton, 1984), each of which has its own advantages. Birds are important agents in that they aid in longer distance dispersal (Holmes, 1990a) and, through ingesting the seeds, are able to aid in the germination of Acacia species requiring chemical scarification (e.g. A. cyclops, A. melanoxylon) (Glyphis et al., 1981; Richardson & Kluge, 2008). Ants rapidly remove and bury Acacia seeds in subterranean nests and so contribute to dispersal on a local scale (Holmes, 1990a). Species noted as having a ‘bird-dispersal syndrome’ are likely also dispersed vertically by ants, as myrmecochory accounts for much of the movement of seed from the litter layer into the seed bank (Richardson & Kluge, 2008). Dispersal by birds of an ‘ant-dispersal syndrome’ species appears less likely (O’Dowd & Gill, 1986).
Importantly, seed morphology and dispersal agents in the native range of Australian acacias are not always accurate predictors of dispersal agents in introduced ranges. For example, in Portugal, South Africa and Florida, invasive Acacia seeds are effectively dispersed by a wide range of opportunistic agents besides those that one would consider functional equivalents of dispersal agents in the native range. These include baboons, domestic and wild ungulates and humans (Ridley & Moss, 1930; Middlemiss, 1963; Kull & Rangan, 2008). In the Western Cape of South Africa, primarily insectivorous barn swallows ingest seeds and act as effective dispersal agents of A. cyclops (Underhill & Hofmeyr, 2007), and other granivorous, ground-dwelling birds disperse Acacia seeds (Duckworth & Richardson, 1988; Knight & Macdonald, 1991). In New Zealand, most native avian seed dispersers are now extinct (Anderson et al., 2006), and the ant fauna is relatively depauperate and limited in distribution (Don, 2007), with only three ant species including seeds in their diet. Despite these limitations, at least eight Australian Acacia species have become invasive in New Zealand (Richardson & Rejmánek, 2011) with A. baileyana showing evidence of long-distance dispersal although the dispersal agent is not known (E.M. Wandrag, unpublished data). Furthermore, in many human-dominated systems, long-distance dispersal of introduced species is mostly human mediated (Trakhtenbrot et al., 2005), so this distinction is likely less important in determining spread rates than may be predicted.
Abiotic dispersal in water and soil is important in many regions (Milton & Hall, 1981). There is a strong association between A. dealbata invasions and watercourses in Chile and Portugal (H. Marchante, unpublished data; Pauchard et al., 2008). Movement of soil for road building is also a major dispersal route of A. dealbata and A. longifolia in Portugal (H. Marchante, unpublished data). Similarly in South Africa, rivers and soil movement aid in the dispersal of acacias that invade riparian areas, such as A. mearnsii (de Wit et al., 2001).
Seed mass in Acacia was found to be positively correlated with invasiveness in a recent study (Castro-Díez et al., 2011) but did not consistently differ in our study nor in a multi-species study comparing seed mass between native and introduced ranges (C. Harris et al., unpublished data). These results contradict findings for Pinus where smaller seed size is positively associated with invasiveness, as small seeds are more suitable for long-distance dispersal by wind (Richardson, 2006). The difference between pines and acacias in this regard is not surprising. Unlike pines, most acacias are animal dispersed, and dispersal by wind is of trivial importance. Factors other than size contribute to dispersibility, and seed size plays an entirely different role as mediator of colonization and establishment success.
Dispersal traits associated with a bird-dispersed syndrome in Australian acacias clearly predispose these species to spread rapidly in a new environment (see discussion of this for A. cyclops in South African fynbos by Higgins et al., 2001) because of the importance of long-distance dispersal events in driving invasions (Trakhtenbrot et al., 2005). However, of the 23 species of Australian Acacia considered invasive (sensuPyšek et al., 2004; Richardson & Rejmánek, 2011), only eight species are known to be bird-dispersed or possess typical bird-dispersed seed traits (Davidson & Morton, 1984; O’Dowd & Gill, 1986; Langeland & Burks, 1998; Stanley & Lill, 2002): Acacia auriculiformis, A. cyclops, A. holosericea, A. implexa, A. longifolia, A. mangium, A. melanoxylon and A. salicina (see Table S1). Additionally, our analysis found that seed dispersal by birds was not significantly correlated with invasiveness. In Portugal, two of the most invasive and widespread Acacia species (A. dealbata and A. longifolia) are ant-dispersed (Marchante et al., 2010), as are A. saligna and A. mearnsii in South Africa (French & Major, 2001; Richardson & Kluge, 2008). Thus, the contribution of different dispersal agents to invasiveness remains unclear but further suggests a role of human-mediated dispersal and interactions with environmental factors.
Seed bank dynamics
A reproductive trait that strongly influences invasiveness of Australian acacias is their capacity to form extensive and persistent soil seed banks (Richardson & Kluge, 2008). Accumulation times differ depending on the species (see Table 2 of Richardson & Kluge, 2008), and the average shortest time frame is roughly eight years. The seeds of some Acacia species that have become invasive can remain dormant for 50–100 years or more (Farrell & Ashton, 1978; New, 1984). Richardson & Kluge (2008) list four main factors that contribute to the size of soil-stored seed banks in Australian acacias in South Africa: the annual seed rain; the age of the stand; stand density or canopy cover; and distance from the canopy. Additional factors include level of granivory, decay and germination (Marchante et al., 2010). Biological control agents that negatively affect flower, flower bud or pod production, such as Melanterius weevils (Dennill & Donnelly, 1991; Impson et al., 2004) that directly feed on acacia seeds, can reduce annual seed rain. The rate of seed accumulation in the soil increases until the stand is about 30 years old, and denser stands produce more seeds, so control efforts to reduce seed production should focus on younger, denser Acacia stands (Milton & Hall, 1981; Holmes, 1990b). Seed density in the soil is highest under the tree canopy and decreases sharply with distance (see Zenni et al., 2009; Marchante et al., 2010), although Marchante et al. (2010) found a few seeds of A. longifolia up to 7 m from the edge of invaded stands.
Table 2. Seed rain density (SRD), seed bank density (SBD) and seed viability (SV) for Australian acacias in native and introduced ranges.
|Acacia species||Seed rain density per m2per year (SRD)||Seed bank density per m2 (SBD)||Seed viability (SV)||Region||References||Observations|
|A. baileyana||19559||–||–||Australia (native range)||17||SRD – maximum #seed/tree|
|A. baileyana||1824 (3010)||–||–||New Zealand||26||SRD – average # seeds per m2 averaged over 7-day period|
|A. cyclops||–||1430–5140 (142 –281)||46–95.3%||South Africa||10|| |
|A. cyclops||–||2832–7792 (402–1019)||99.2%||South Africa|| 8||SBD – range of four different blocks|
|A. cyclops||1197 [1373–3019*]||2031||87%||South Africa||15||SRD –*estimated #seed per m2 projected canopy|
|A. cyclops||540 (710)||–||–||Australia (introduced range)|| 6||SRD – estimated from reproductive output data (determined by dividing total mass of seeds removed from pods by mass per individual seed)|
|A. cyclops||1900 (1930)||–||–||Australia (native range)|| 6|| |
|A. dealbata||–||10000||90%||Chile||25|| |
|A. dealbata||2553 (3244)||–||–||New Zealand||26||SRD – average # seeds per m2 averaged over 7-day period|
|A. dealbata||–||ca. 22500||30%||Portugal||13||SV: probably underestimated (seeds heated to 50°C without scarification)|
|A. elata||–||–||50%||–||22||SV – final germination after scarification|
|A. holosericea||–||–||>95%||Australia (native range)|| 7|| |
|A. longifolia||2000–12000||500–1500||>85%||Portugal||14||SRD – 2000: smaller trees next to the ocean (windward); 12000: bigger trees leeward|
|A. longifolia||–||–||>88%||Portugal||16|| |
|A. longifolia||11500||34000||–||South Africa||19||SRD – maximum number|
|A. longifolia||–||2078–3473 (488–498)||99%||South Africa||21|| |
|A. longifolia||2923||7646||97%||South Africa||15|| |
|A. longifolia||–||4528 (1075)||99%||South Africa|| 4||After introduction of biological control agent, max numbers|
|A. longifolia||2530 (3430)||–||–||Australia (introduced range)|| 6||SRD – estimated from reproductive output data (determined by dividing total mass of seeds removed from pods by mass per individual seed)|
|A. longifolia||810 (1180)||–||–||Australia (native range)|| 6|| |
|A. mangium||410||–||–||Indonesia||23||SRD – estimated from seed production in kg per ha per year|
|A. mearnsii||–||5314/696||–||South Africa||20||SBD- maximum number/average|
|A. mearnsii||–||38340||–||South Africa||15|| |
|A. mearnsii||–||–||>83.4%||South Africa||12|| |
|A. melanoxylon||3218||48739||70%||South Africa||15||SRD & SBD: Donald, 1959 cited by Milton & Hall, 1981|
|A. melanoxylon||–||–||85–91%||Australia (native range)|| 2|| |
|A. melanoxylon||740 (800)||–||–||Australia (introduced range)|| 6||SRD – estimated from reproductive output data (determined by dividing total mass of seeds removed from pods by mass per individual seed)|
|A. melanoxylon||1160 (1810)||–||–||Australia (native range)|| 6|| |
|A. paradoxa||–||1000||–||South Africa||28|| |
|A. paradoxa||58#||–||–||Australia (native range)|| 1||SRD – #firm seed production per plant|
|A. pycnantha||31#||–||99%||Australia (native range)|| 1|| |
|A. saligna||–||7920–45800 (560–3220)||>86%||South Africa||10|| |
|A. saligna||2645–13472||–||–||South Africa||27||SRD – measured in 1989, ca. 2 years after introduction of biocontrol agent|
|A. saligna||446–3035||–||–||South Africa||27||SRD – measured in 2004, ca. 18 years after introduction of biocontrol agent|
|A. saligna||5443 [10562*]||11920||83%||South Africa||15||SRD – #seed/tree based on few trees; * estimated seed per m2 projected canopy|
|A. saligna||–||715–8097||–||South Africa|| 9||SBD – after introduction of biological control agent; values estimated from 4 places and 3 depths|
|A. saligna||–||–||>90%||Israel|| 3|| |
|A. saligna||–||2000–189000 (53333)||–||South Africa||18||After introduction of biological control agent; average from 8 sites, samplings during 6 years|
|A. saligna||–||1389–3600 (207–279)||–||Australia, New South Wales (introduced range)||24|| |
|A. saligna||–||–||73%||–||22||SV – final germination after scarification|
|A. saligna||–||3158–38714 (1194–4006)||>65%||South Africa||11||SBD – range of 4 sites, at 0–15 cm|
|A. saligna||760 (750)||–||–||Australia (introduced range)|| 6||SRD – estimated from reproductive output data (determined by dividing total mass of seeds removed from pods by mass per individual seed)|
|A. saligna||540 (650)||–||–||Australia (native range)|| 6|| |
|A. salicina||–||–||77%||–||22||SV – final germination after scarification|
|A. victoriae||–||50–3900||80%||Australia (native range)|| 5|| |
The main drivers of seed bank persistence and maintenance appear to be ants, although gravity and water may be the dominant drivers where ants are absent. Once seeds have dropped to the ground, ants bury many of them in their nests to allow them to exploit arils (Milton & Hall, 1981). In doing so, they often account for the majority of vertical seed movement into the upper seed bank. Acacia seeds gain a threefold advantage through protection from above-ground seed predators, protection from fire and incorporation into the seed bank (Gill, 1985; Holmes, 1990a). In South Africa, ants may play a critical role in accumulating seed banks of Australian acacias and aiding in their invasiveness (Holmes, 1990c; Richardson et al., 2000a).
The role of seed bank density in Acacia invasiveness is unclear. Both higher and lower seed bank densities have been recorded in the introduced range of various Acacia species when compared to that in the native range (Milton & Hall, 1981; Richardson & Kluge, 2008; Marchante et al., 2010). Additionally, methods of measuring seed bank and seed rain vary widely, making comparisons between introduced and native ranges problematic (see Table 2 for a summary of Australian Acacia seed data from various introduced and native regions). Prolific seed production and large accumulations of seeds in the seed bank certainly contribute to a species’ ability to invade an ecosystem but these qualities alone do not guarantee invasiveness. Buist (2003) found that closely related pairs of rare and widespread Acacia species produced similar numbers of seeds and similar-sized, persistent soil seed reserves, indicating that level of seed production does not necessarily determine abundance of a species. These traits likely need to work in concert with certain physiological and morphological traits, such as germination ability, resource utilization, rapid growth of seedlings and dispersal investment, to contribute to invasiveness.
The majority of invasive Acacia species possess seeds whose germination is stimulated by fire, but some invasive species, notably bird-dispersed taxa, may be stimulated to germinate through chemical scarification via ingestion by an appropriate dispersal agent (Glyphis et al., 1981; Fraser, 1990; Richardson & Kluge, 2008). These stimuli are required to break physical dormancy of the hard, water impermeable seed coat and allow germination of Acacia seeds, which have consistently high viability and low germinability over time. However, in Portugal, total viability and germinability were found to be significantly higher (and dormancy lower) in seeds from recently invaded soils for A. longifolia (Marchante et al., 2010).
Invasive Australian acacias tend to germinate after disturbance, although disturbance is not essential. Acacia dealbata shows high survival within native forest and in open areas in Chile where it can endure long periods of drought and shade under canopies of native trees (Fuentes-Ramírez et al., 2011). Moreover, mutualistic relationships with nitrogen-fixing bacteria are important for successful establishment of leguminous species, so the presence of compatible rhizobia is also essential for determining the colonization ability of introduced species (Parker et al., 2006; Rodríguez-Echeverría et al., 2011). Interestingly, Rodríguez-Echeverría et al. (2011) found that these bacterial symbionts are often cointroduced with their Acacia hosts from Australia, suggesting the presence of suitable soil symbionts in the introduced range may not be an important limiting factor in Acacia invasions per se.
Studies from the introduced ranges of Australian acacias report that a considerable number of seeds produced and allocated to seed rain are lost to factors such as early germination, granivory or decay (Marchante et al., 2010). However, the consistently high seed viability found in many species of Acacia appears to be fundamental to their ability to invade (see Table 2) (Richardson & Kluge, 2008; Marchante et al., 2010). Germination characters per se do not appear to be characteristic of invasiveness as invasive Australian Acacia species in South Africa can show opposing characteristics of either high dormancy, low germination and decay rates and rapid seed bank accumulation, or low dormancy, high germination and decay rates and gradual seed bank accumulation (Richardson & Kluge, 2008).
Comparisons of rare and widespread species show some association with factors that influence seed germination. The burial depth and heat-stimulation requirements of a species are important factors affecting germination that can determine how rare or widespread it is (Brown et al., 2003). Comparisons of reproductive traits in two rare acacias and their common relatives showed differences in the germination (reduced range of temperature for germination in rare species) and higher rates of predation of fruit and seed in the rare species (Buist, 2003). Seed viability and dormancy levels between invasive and non-invasive species have not been compared. It may be predicted that, because such traits are adaptations to fire-driven ecosystems, other Acacia species originating from similar regions also likely possess such germination traits.