Date of Publication:
February 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
Hydrogeomorphology, floristics, classification and conservation values
of the little-known montane mires of the upper Cudgegong River
catchment, Central Tablelands, New South Wales
Ian R. C. Baird 12 and Doug Benson 3
‘School of Health and Science, Western Sydney University, Locked Bag 1797, Penrith, NSW 2751, AUSTRALIA.
2 Current address: 3 Waimea St, Katoomba, NSW 2780, AUSTRALIA, petalurids@gmail.com
3 Honorary Research Associate, National Herbarium of New South Wales, Botanic Gardens & Domain Trust, Sydney 2000,
AUS TRALIA. D oug. B enson@rbg syd. nsw. gov. au
Abstract. Mires or peat swamps have a restricted distribution in Australia and are limited to areas where hydrological
inputs exceed evapotranspiration. In NSW, mires are restricted to the coast, adjacent ranges or tablelands, and along
the Great Dividing Range; most are listed as threatened ecological communities under State or Commonwealth
legislation. Due primarily to the relatively high rainfall and suitable geology, the Blue Mountains region includes
a number of such threatened mire ecological communities. Most of these mire types are largely included within the
Greater Blue Mountains World Heritage Area, although there are notable exceptions, such as the endangered Newnes
Plateau Shrub Swamps.
This paper reports on a little-known group of diverse, relatively isolated and largely unprotected mires, in a relatively
low rainfall area in the upper Cudgegong River catchment, east of Rylstone in the NSW Central Tablelands, and of
their floristic, hydrogeomorphic and typological relationship with other mires of the Blue Mountains. They can be
broadly divided into montane bogs, montane fens and hanging swamps. Particular attention is focussed on the largest
and most diverse one, Rollen Creek swamp, which contains all three types. It is hoped that highlighting this hitherto
unrecognised group of high conservation-value mires will contribute to their improved conservation and encourage
further research into mires of eastern NSW.
Key words: peat swamp, wetlands, groundwater-dependent ecosystems, Eucalyptus camphor a., Petahira gigantea ,
National Heritage, Greater Blue Mountains World Heritage Area.
Cunninghamia (2018) 18: 001-021
doi: 10.775 l/cunninghamia.2018.18.001
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
2 Cunninghamia 18: 2018
Introduction
Wetlands encompass a range of vegetated ecosystems,
including those referred to as mires, bogs, fens, swamps and
marshes (see Gore, 1983; Mitsch & Gosselink, 2007; Mitsch
et al, 2009). Wetlands are characterised by a diversity of
hydrological regimes, and may be permanently, seasonally or
intermittently inundated or saturated (DECCW, 2010). Even in
permanent wetlands, water table depth can vary considerably
within a particular wetland and between different wetland
types. Such spatial heterogeneity, even within a permanent
wetland complex, may result in considerable heterogeneity
in substrates, vegetation associations and habitat, often
across small spatial scales (e.g., Brown et al., 1982; Keith
& Myerscough, 1993; Keith et al., 2006). For example,
wetland complexes are often characterised by a complex
intergrading of fens, bogs, swamps or marshes (Hajek et al.,
2006; Kirkpatrick & Bridle, 1998; Rydin & Jeglum, 2013;
Wheeler & Proctor, 2000; Yabe & Onimaru, 1997).
Hydrology and water balance (evapotranspiration compared
to precipitation and other hydrological inputs) are the
critical factors in determining the development of peaty or
organic-rich wetland sediments. A basic requirement for
peat formation is that plant biomass production (carbon
production) exceeds decomposition (ecosystem respiration
or carbon output). Consistently high water tables and
a relatively anoxic environment generally provide the
necessary conditions for peat accumulation in wetlands.
Conditions of seasonal drying or widely fluctuating water
tables, and/or negative water balance, result in oxidisation
and bacterial decomposition of organic matter, and are not
conducive to accumulation of peat (Gore, 1983; Rydin &
Jeglum, 2013).
In the international context the term mire refers to peat¬
forming wetlands, and includes bogs and fens (Gore, 1983;
Joosten et al., 2017; Rydin & Jeglum, 2013), although the
terms may be applied somewhat differently in Australia (see
Whinam & Hope, 2005). In the Australian context, bogs are
typically low nutrient, acidic, dominated by sclerophyllous
sedges and shrubs, and mosses, and may have large parts
of the surface raised above the water table, while fens are
usually more nutrient-rich, less acidic or alkaline, dominated
by softer herbaceous and graminoid vegetation, and usually
with a near surface water table (Keith, 2004; Whinam &
Hope, 2005). Additionally, in Australia, the term swamp , as
used in this paper, is a generic term which may refer to a wide
range of wetland types, including peat-forming wetlands
(such as bogs and fens), and wetlands with primarily mineral
substrates with environmental conditions unsuitable for peat
development, in contrast for example, to its more specific
application in the USA (Cowardin et al., 1979) or Europe
(Rydin & Jeglum, 2013). While they receive hydrological
inputs from precipitation, surface flow and groundwater in
varying proportions, most mires (peat swamps) in Australia,
including the diverse mire communities across the Blue
Mountains of NSW, are considered groundwater-dependent
ecosystems (see NSW Government, 2002; Serov et al., 2012;
Whinam & Hope, 2005).
Baird & Benson, Montane mires, upper Cudgegong River catchment
Mire ecosystems in the Blue Mountains are captured within
the Montane Bogs and Fens and Coastal Heath Swamps
vegetation classes of NSW and the ACT of Keith (2004).
These are represented by a number of mire vegetation types,
including Blue Mountains Sedge Swamps (BMSS), Newnes
Plateau Shrub Swamps (NPSS), Coxs River Swamps
and Boyd Plateau Bogs (Benson & Keith, 1990; Keith &
Benson, 1988). These are characterised by considerable
spatial heterogeneity across a number of environmental
gradients, within and between individual swamps and
swamp types (e.g., Holland et al., 1992a; Holland et al.,
1992b). Variation in vegetation across the hydrological
gradient (from ephemeral to permanent saturation) is
particularly evident within the upland mires developed on
sandstone geology (e.g., Benson & Baird, 2012; Holland
et al., 1992a). The term montane , as used in this paper,
follows its usage in various vegetation publications in
NSW (e.g., Bell et al., 2008; Hunter & Bell, 2007; Keith,
2004; NSWSC, 2004). In describing the Montane Bogs and
Fens vegetation class, Keith (2004) indicated an elevation
range of 600-1500 m, which includes the elevation range
of the study area. Following Keith (2004), bogs and fens
identified in this study are thus referred to as montane
bogs and montane fens. In their classification of the native
vegetation of southeast NSW, Tozer et al. (2010) attributed
higher elevation mire vegetation in the southern half of the
Blue Mountains (the northern extent of their study area
does not extend beyond Fithgow) to the Tableland Bogs,
Tableland Swamp Meadows (in the Montane Bogs and
Fens vegetation class) and Blue Mountains-Shoalhaven
Hanging Swamps (in the Coastal Heath Swamps vegetation
class) types.
The upper Cudgegong River catchment (east of Rylstone)
on the NSW Central Tablelands drains an area surrounded
by the Great Dividing Range (at elevations over 1000 m) and
forms part of the inland flowing Macquarie River catchment
(Fig. 1). Part of the upper Cudgegong River catchment (in
the Dunns Swamp area) is included in Wollemi National
Park in the Greater Blue Mountains World Heritage Area
(GBMWHA) (Fig. 2). No comprehensive vegetation
or wetland mapping has been undertaken across this
catchment, although previous reports have identified and
briefly described a range of wetland vegetation types in the
Dunns Swamp area (Bell, 1998a, b; Tame, 1997). During
doctoral thesis fieldwork on the biology of the endangered
mire-dwelling Giant Dragonfly, Petalura gigantea , Baird
(2012) identified a number of previously undocumented
and relatively isolated mires, with disjunct flora and
fauna occurrences, in and adjoining the upper Cudgegong
River catchment. Baird and Benson (2017) subsequently
highlighted the value of one of these mire systems, Rollen
Creek swamp, which occurs largely in Coricudgy State
Forest, in support of a proposal to add Coricudgy State
Forest to the National Heritage list; ultimately a candidate
for potential addition to the GBMWHA (GBMWHAAC,
2015). The added inclusion of specific areas on the edges of
the GBMWHA, including Coricudgy State Forest, would
significantly enhance the currently established biodiversity
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
3
values of the GBMWHA, and has been recommended
(Benson & Smith, 2015).
The aim of this paper is to draw attention to the overlooked
mires of the upper Cudgegong River catchment, with a
particular focus on the Rollen Creek mire system, the largest
and most diverse, to contextualise them in relation to other
mires of the Blue Mountains, to contribute to improving their
conservation, and to encourage further research on mires in
southeastern Australia.
Upper Cudgegong River study area
Location and physiography
The study area is the upper Cudgegong River catchment
between the Great Dividing Range and Rylstone, centred
approximately on Dunns Swamp, Wollemi National Park
(32° 50' 04" S, 150° 12' 21" E). The area is located in the
NSW Central Tablelands, in the north-western part of
the Blue Mountains and of the Sydney Basin bioregion
(Thackway & Cresswell, 1995). The Great Dividing Range
forms the watershed of the upper Cudgegong River, which
drains inland to the Macquarie River (Fig. 1 & 2). Drainages
to the north, east and south of the upper Cudgegong River
watershed drain to the coast. Elevation decreases from east
to west; from 1256 m on the basalt peak of Mt Coricudgy on
the Great Dividing Range, westward to 571 m at Rylstone.
Additional high elevation areas (above 1000 m) along the
catchment watershed, and associated with the Great Dividing
Range, include the remnant basalt-capped plateau of Nullo
Mountain, and the isolated basalt-capped peaks of Mt Darcy
and Mt Coorongooba.
The geology of the study area consists of underlying Permian
sandstone, conglomerate, shale and siltstone; Triassic
Narrabeen Sandstone, shale and conglomerate, forming
characteristic ‘pagoda country’; and Tertiary basalt-capped
peaks and remnant plateaus. The Capertee, Collingwood,
Inglewood, Lees Pinch, Munghorn Plateau, Nullo Mountain
and Coricudgy soil landscapes dominate the study area (Kovac
& Lawrie, 1991). The Inglewood (Yellow Earths), Lees Pinch
(Shallow Soils) and Munghorn Plateau (Siliceous Sands)
soil landscapes have developed primarily in association
with the Triassic Narrabeen sandstone geology, while Nullo
Mountain and Coricudgy soil landscapes (Krasnozems) have
developed in association with the basalt-capped peaks and
remnant plateaus, where in situ weathered Tertiary basalt and
basalt colluviums overlie the Narrabeen sandstone. Capertee
(Yellow Podzolics) and Collingwood (Red Podzolics) soil
landscapes have developed in association with the Permian
geology (Kovac & Lawrie, 1991). The peaty sediments of
the mires in the area were not differentiated in the coarse-
scale soil landscape mapping of Kovac and Lawrie (1991).
The study area includes the Dunns Swamp area of Wollemi
National Park, a popular tourist site, and parts of Nullo
Mountain and Coricudgy State Forests. With the exception
of logging of better quality timber from the slopes of Mt
Coricudgy, there is no evidence of recent logging activity in
the immediate Coricudgy State Forest area which is mainly
essentially undisturbed shrubby woodland with little timber
value. Variably cleared and farmed freehold lands are mainly
concentrated along the Cudgegong River valley and its
tributaries, such as Coxs Creek, and in the Nullo Mountain
area. The lower Cudgegong valley around Rylstone was
explored by the botanist Allan Cunningham in the 1820s,
and the area taken up for pastoral settlement soon after.
Occupational licences in the upper Cudgegong in the Parish
of Kelgoola were offered for sale in 1843 (,Sydney Morning
Herald 18/1/1843).
Climate
Rainfall across the study area is characterised by high inter¬
annual variability, with average annual rainfall decreasing
with elevation and from east to west. Highest mean monthly
rainfalls occur between November and March (late spring to
early autumn), with lowest mean monthly rainfalls from July
to September (winter to early spring). Mean annual rainfall at
the Nullo Mountain AWS (1130 m elev.) was 955 mm/annum
(1994-2017) (www.bom.gov.au). Mean maximum monthly
temperature at Nullo Mountain was 16.7° C (9.2-24.0° C) and
mean minimum monthly temperature was 8.0° C (2.5-13.6°
C). Mean Annual rainfall at the property “Kelgoola” (747 m
elev.) in the upper Cudgegong River valley near Rollen Creek
swamp and Mt Coricudgy, was 809 mm (1963-2006), also
with high inter-annual variability (486-1235 mm/annum).
The higher elevation areas along the Divide, including Mt
Coricudgy, Mt Coorongooba and Mt Darcy, can also be
expected to have higher rainfall than that of the nearby upper
Cudgegong valley at “Kelgoola” due to an orographic effect,
and similar to that of Nullo Mountain. Mean annual rainfall
at Rylstone (Ilford Rd AWS, 605 m elev.), at the western
edge of the study area, is lower, with 655 mm/annum. Mean
maximum monthly temperature at Rylstone was 22.6° C
(14.4—30.7° C) and mean minimum monthly temperature was
8.2° C (1.2-15.9° C) (www.bom.gov.au).
Previous vegetation studies in the upper Cudg¬
egong
Previous vegetation studies in the upper Cudgegong are
mainly confined to the Dunns Swamp area (Bell, 1998a; Gellie
& McRae, 1985; Tame, 1997). Gellie and McRae (1985)
referred more broadly to the Cudgegong Swamps. Building
on a brief report by Tame (1997), and with limited additional
fieldwork, Bell (1998a) described Cudgegong River Swamp
Grassland, Upper Cudgegong Alluvial Sedgeland, Upper
Cudgegong Alluvial Reedland, Upper Cudgegong Alluvial
Shrub-swamp and Upper Cudgegong Sphagnum Bog in the
Dunns Swamp area of Wollemi National Park.
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
Fig. 1: Map of the Greater Blue Mountains World Heritage Area showing the location of the upper Cudgegong River catchment east of
Rylstone (inset box).
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
5
, Nullo Moun
National Park
CM10
CM08
Nullo\^V
Mountain
State Forest
\\\\ WOA
CM03
Dunns
Swamp'
:one
Mt Darcy'
Fig. 2. Map of the Upper Cudgegong River catchment study area and locations of surveyed mires. The site codes in this diagram are
explained in Table 1.
6 Cunninghamia 18: 2018
Although Bell (1998a, b) did not identify the Carex fen
vegetation along Never Never Creek and other similar sites,
he noted the need for additional survey of wetland vegetation
communities across the upper Cudgegong area to clarify their
floristic relationships and distributions. Bell (1998a) noted
that the Cudgegong River Swamp Grassland vegetation
occurred only in waterlogged peaty alluviums and the Upper
Cudgegong Alluvial Sedgeland occupied “grassy bogs”.
Tame (1997) suggested the substrate of this vegetation type
was probably a shallow sandy peat overlying impermeable
clays. Bell (1998b) noted that Cudgegong Sphagnum Bogs
only occurred in limited areas on heavy soils with poor
drainage, along creek-lines or in small localised patches, but
also referred to it occurring on peaty sands.
Bell (1998a) also noted the mention by Ford (1989) of an
area of swamp near Mt Darcy (“along Rollen Creek”)
with apparent similarities to his Cudgegong River Swamp
Grassland. This is the first known reference to a swamp
system along Rollen Creek, later described by Baird (2012)
and Baird and Benson (2017).
Methods
Reconnaissance fieldwork (by IRCB) was carried out across
the upper Cudgegong River catchment and adjoining areas
in 2007-2009, searching for evidence of the endangered
Petalura gigantea and to identify mires with potential
breeding habitat for this obligate, groundwater-dependent,
mire-dwelling dragonfly (Baird, 2012, 2014). Identification
of upper Cudgegong mire sites has been based on a range of
sources; aerial photograph interpretation (API, using Google
Earth); previous vegetation reports from the Dunns Swamp
area (Bell, 1998a, b; Tame, 1997); additional information
provided by former NPWS ranger Chris Pavich (pers.
comm.); data gathered by Baird (2012) (which included
recording mires along 4 creek systems), and the results of a
vegetation survey of the swamp system along Rollen Creek
in February 2017 (Baird & Benson, 2017); and additional
vegetation surveys across the study area in October 2017.
Swamps were surveyed across Coxs, Ganguddy, Kings
Swamp, Never Never and Rollen creeks (Fig. 2). Swamps
on the remaining creek systems are on private property.
Vegetation surveys consisted of walking transects targeted
to cover as much local variation as possible and to record
all readily identifiable plant species (and any evidence of
threatened faunal species) across the range of identified
floristic and hydrogeomorphic variability within the swamps
surveyed. The presence of localised seepage areas or of
an emergent water table provided further confirmation of
the presence of suitable hydrological conditions for the
development of organic-rich substrates. A sediment probe
(1.8m long x 8 mm diameter steel rod) was used to measure
depth, and qualitatively assess characteristics, of organic-
rich sediments in lower Rollen Creek swamp.
Baird & Benson, Montane mires, upper Cudgegong River catchment
Results
Mires of the upper Cudgegong River catchment
Swamp vegetation in the upper Cudgegong River catchment
was identified along the Cudgegong River and its tributary
creeks - Coxs, Dairy Swamp, Ganguddy, Gavins, Kings
Swamp, Mill, Never Never, Rollen, Sugarloaf and Swampy
creeks (Fig. 2). Mire vegetation communities (with persistent
high water tables and organic-rich or peaty sediments) were
confirmed by survey along Coxs, Ganguddy, Never Never
and Rollen creeks. Mire vegetation was identified along the
main Cudgegong River valley immediately upstream and
downstream of the confluence with Rollen Creek, but was not
surveyed due to its location on private property. The presence
of mire was not confirmed along the highly degraded Kings
Swamp drainage. Details of mires surveyed are shown in
Table 1. Occasional small isolated patches of seepage-fed
hanging swamp on valley sides are also scattered across the
upper study area on Narrabeen Sandstone, particularly in
the Rollen Creek catchment. Additional small mire patches
can be expected to occur elsewhere across the study area in
locations that are difficult to access or identify on Google
Earth imagery.
The mires of the upper Cudgegong River and its tributaries
above Dunns Swamp, and along upper Coxs Creek, occur
primarily in the Munghom Plateau soil landscape on Triassic
geology (see Kovac & Lawrie, 1991). Mire sediments
included organic sands, sandy peats, sapric peats and fibrous
peats, depending on the hydrological regime and topographic
context. Some patches of swamp with sedgeland-heath
vegetation, with floristic similarities to the wetter sedgeland-
heath bogs, occur on heavier clayey soils (e.g., along lower
Ganguddy Creek above Dunns Swamp) and cannot be
considered mires. Depths of soft, saturated, organic-rich/peaty
sediments throughout the main valley floor along lower Rollen
Creek swamp (CM01) were generally >1.8 m, frequently with
no major differences in density or texture observed across
that depth. Dense sandy-gravel material was encountered,
however, below 1.6 m in several locations. In one area near
the confluence with a tributary near the Coricudgy State
Forest boundary, some narrow bands of sandy-gravel material
at different depths indicated historical deposition events,
probably associated with post-fire erosion, upstream and/or
upslope. Predictably, depth of organic-rich swamp sediment
progressively decreased, and sand content increased, towards
swamp margins on lower valley side slopes.
Considerable hydrogeomorphic and floristic diversity
was observed in the mires across the study area, including
montane bogs and montane fens, broadly referable to the
Tableland Bogs and Tableland Swamp Meadows typologies,
respectively, of Tozer et al. (2010), within the Montane Bogs
and Fens vegetation class of Keith (2004), and hanging
swamps on valley sides, with similarities to the Blue
Mountains-Shoalhaven Hanging Swamps typology of Tozer
et al. (2010), within the Coastal Heath Swamps vegetation
class of Keith (2004). The rare, swamp-dwelling small tree,
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
7
Eucalyptus camphor a subsp. camphora, is widely distributed
across the upper Cudgegong catchment within each of these
mire variants, along associated upstream drainage lines,
and in some other swampy areas on heavier soils, including
along lower Ganguddy Creek.
The main mire systems
Rollen Creek swamp, most of which occurs in Coricudgy
State Forest, is one of only two extensive and relatively
undisturbed (based on API, using Google Earth) mire
systems in the upper Cudgegong catchment. The Rollen
Creek swamp system appears to include the full range of
hydrogeomorphic and floristic variation observed within
the identified mires in the study area, including montane
bogs (Tableland Bogs) and montane fens (Tableland Swamp
Meadows) on valley floors, and hanging swamps on valley
sides (Figures 3-11). Due to its outstanding diversity, size,
condition and its representativeness of the range of mires in
the study area, the Rollen Creek swamp system is described
in detail below.
In addition to the extensive area of valley-floor sedgeland-
heath bog (Tableland Bog) along Rollen Creek, several
isolated and very small patches of variably degraded,
seepage-fed sedgeland-heath bog (CM08-10) were also
identified adjacent to, and intergrading with, the alluvial
Q/rev-dominated fen (Tableland Swamp Meadow) (CM07)
along upper Coxs Creek (Fig. 13). A narrow band of hanging
swamp (CM04) of sedgeland-heath and Sphagnum , with
conspicuous springs, also occurs just above Ganguddy
Creek, near to and upstream of the Coricudgy Road bridge.
The other large mire system in relatively good condition is
along Never Never Creek and its lower tributaries above
Dunns Swamp (Kandos Weir impoundment) in Wollemi
National Park. The Never Never Creek mire system is
dominated by Carex fen (Tableland Swamp Meadow) along
the main valley floor (CM03) with scattered Leptospermum
shrubs (Fig. 12), although floristic variants occur in some
marginal areas and in slightly higher gradient tributary
swamp (CM02), in association with Eucalyptus camphora.
The Rollen Creek and Never Never Creek mire systems
are surrounded by high quality, dry sclerophyll eucalypt-
dominated woodland on adjoining slopes.
The swamps along the main valley floors of the Cudgegong
River (above and below Dunns Swamp) and its tributaries,
including Coxs Creek (Fig. 13) are generally low gradient,
alluvial swamp meadows. They are generally located on
cleared freehold land, with a history of grazing, nutrient
enrichment and weed invasion, and mostly range in
condition from moderately- to highly-degraded. They range
from Ctf/^x-dominated fens with organic-rich alluvium in
wetter areas (Tableland Swamp Meadows) which may be
variably inundated with shallow water, to seasonally wet
tussock grasslands on largely mineral soils at the drier end
of the hydrological gradient. Some of the swamp meadows
include small pools along the main drainage line (possible
chain-of-ponds system) with reedland (marsh) vegetation
of Typha sp., Phragmites australis and/or Eleocharis
sphacelata. Along the main Cudgegong River valley above
Dunns Swamp, around the confluence with Rollen Creek,
areas of highly degraded Tableland Swamp Meadow
sometimes also appear to have small remnant patches of
sedgeland-heath bog (Tableland Bog) on their margins.
Based on API (using Google Earth), additional areas of
Tableland Swamp Meadow with possible Tableland Bog
were identified along Gavins and Sugarloaf creeks, upper
tributaries of the Cudgegong River; these swamps on private
properties are mostly surrounded by cleared grazing land and
were not visited.
A localised part of the upper Kings Swamp drainage includes
some plants which occur in mires elsewhere in the area (e.g.,
Callistemon citrinus , Eucalyptus camphora , Leptospermum
obovatum ), with some groundwater seepage evident,
but it is highly degraded and occurs on cleared grazing
land. Regardless of its potential pre-disturbance state, it
is not treated as mire here. A degraded and grazed narrow
swamp with Callistemon citrinus and Eucalyptus camphora
subsp. camphora also occurs along upper Mill Creek. This
small swamp patch on otherwise cleared, private property
was observed but not surveyed and it was not possible to
determine its substrate or hydrological characteristics. It is
also not treated as mire here.
Another large swamp system is in Jones Hole, on Jones
Hole Creek (a headwater tributary of Coricudgy Creek), to
the northwest of Mt Coricudgy (Fig. 2), in the coastward
flowing Hunter River catchment, not the Cudgegong River.
It is mostly in Wollemi National Park though partly in
Coricudgy State Forest. The vegetation of this difficult-to-
access valley-floor swamp system (32° 48' 16" S, 150° 20'
11" E, 720-850 m elev., -3250 m length) is unknown, but it
is reasonable to assume, due to its elevation and proximity
to the higher rainfall area of adjacent Mt Coricudgy,
that it is characterised by a high water table and variably
peaty sediments. API (using Google Earth) suggests that
it is predominantly a Carex-dominated fen, with areas of
intergrading sedgeland-heath bog. The swamps in Jones
Hole and along Never Never Creek have some history of
cattle grazing prior to National Park gazettal (C. Pavich, pers.
comm.). Conspicuous, abundant and structural plant species
recorded for each of the identified and surveyed mires in
the study area are included in Table 2. This is an incomplete
list as some grasses and monocots were not identified, and
some seasonal species may not have been evident at the time.
Only two conspicuous bryophytes are listed, but the total
bryophyte richness is likely to be high.
Other swamps
The study area includes a diversity of swamp vegetation
types distributed across the hydrological gradient; much
of the swamp area is characterised by only seasonally or
intennittently wet sandy alluviums or heavier clayey soils
along valley floors and cannot be considered mires. This
includes seasonally wet tussock grasslands on largely mineral
8 Cunninghamia 18: 2018
soils at the drier end of the hydrological gradient, which
generally occur in the lower elevation and lower rainfall
parts of the catchment. Most of these seasonally wet tussock
grasslands are surrounded by cleared grazing areas and occur
along the Cudgegong River and its tributaries below Dunns
Swamp, and along lower Coxs Creek; they have a history
of grazing, nutrient enrichment and weed invasion; and are
moderately- to highly-degraded.
Some shrubby swamps on sandy alluviums, including in the
Dunns Swamp area and upper Ganguddy Creek, have a scrub
of Leptospermum polygalifolium (often colonising disturbed
sites), and are probably referable to the Upper Cudgegong
Alluvial Shrub-swamps of Bell (1998a). Upper Cudgegong
Alluvial Reedland (fringing vegetation around the Kandos
Weir impoundment) and Upper Cudgegong Alluvial
Shrub-swamp of Bell (1998a) in the Dunns Swamp area
are not mires and were not considered potential habitat for
Baird & Benson, Montane mires, upper Cudgegong River catchment
P. gigantea by Baird (2012). Other reedland (marsh) patches,
however, occur in association with the Carex fens and can be
considered part of the mire systems. Swamps in the lower
parts of the Cudgegong River catchment (east of Rylstone)
below Dunns Swamp and along lower Coxs Creek, occur in
the Capertee soil landscape on Permian geology (Kovac &
Lawrie, 1991).
Additional seasonally wet swamps occur to the north of the
catchment watershed in headwaters of the northward flowing
Growee River catchment (e.g., Spring Log Swamp on Spring
Log Creek), but these are on primarily mineral sediments
and are not mires. Growee Swamp (500 m elevation), which
occurs further downstream on the Growee River to the north,
was not visited, but it occurs on cleared grazing land in a
lower rainfall area, and is also assumed to be on seasonally
drying alluvial sediments or heavier clayey soils.
Table 1: Identified mires of the upper Cudgegong River catchment (arranged by decreasing elevation) showing - site code, location,
coordinates, approximate elevation range (upper to lower), approximate length or area, mire type and condition. An unsurveyed
Cudgegong River mire above Dunns Swamp has been included. Mire types - Tableland Bog (TB), Tableland Swamp Meadow
(TSM), and Hanging Swamp (HS). The locations of these mires are indicated in Fig. 2.
Mire site
Location
Coordinates
(Google Earth)
Elevation (m)
(Google Earth)
Length or area
Mire type
Condition
CM01
Rollen Creek
32° 52' 47" S,
150° 17' 01"E
745-700
3 km
TB, HS, TSM
Good
CM07
Coxs Creek
32° 44' 27" S,
150° 09' 56" E
740-660
10 km
TSM
Poor-Good
CM08
Coxs Creek margins
adjoining CM07
32° 44' 43" S,
150° 10' 18" E
724
20 x 20 m
TB
Poor
CM09
Coxs Creek margins
adjoining CM07
32° 44' 37" S,
150° 10' 07" E
722
40 x 40 m
TB
Poor
CM10
Coxs Creek margins
adjoining CM07
32° 44' 35" S,
150° 09' 48" E
718
150 x 40 m
TB
Poor-Moderate
CM04
Above Ganguddy
Creek
32° 51'22" S,
150° 12' 52" E
680-678
100 x 5 m
HS
Good
CM05
Ganguddy Creek
below and adjoining
CM04
32° 51' 22" S,
150° 12’ 52" E
678
12 x 10m
TB
Moderate -Good
CM02
Never Never Creek
tributary
32° 49' 26" S,
150° 12' 49" E
677-665
200 m
TB/TSM
Good
CM06
Ganguddy Creek
tributary at junction
with Ganguddy Creek
32° 51' 07" S,
150° 12' 51" E
674
15 x 10m
TB
Moderate
CM03
Never Never Creek
32° 49’ 32" S,
150° 12' 46" E
672-656
1.7 km
TSM
Good
Not surveyed
Cudgegong River
above Dunns Swamp,
upstream and
downstream of the
confluence with Rollen
Creek
32° 05' 55" S,
150° 17' 04" E (indicative
location)
741-719
TSM, TB
Very poor
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
9
Table 2: Native plant species (including mosses) recorded for each of the mires surveyed in the upper Cudgegong River catchment.
Refer to Table 1 for details.
SPECIES
Family
CM01
CM02
CM03
CM04
CMOS
CM06
CM07
CM08-10
Asperula gunnii
Rubiaceae
1
Baeckea utilis
Myrtaceae
1
Boloskion fimbriatus
Restionaceae
1
1
Baloskion australe
Restionaeeae
1
Baumea sp.
Cyperaceae
1
1
1
1
1
1
Blechnum nudum
Blechnaceae
1
1
1
Bulbine bulbosa
Liliaceae
1
Callistemon citrinus
Myrtaceae
1
1
1
1
1
1
Corex appressa
Cyperaceae
1
Carex gaudichaudiana
Cyperaceae
1
1
1
1
1
1
1
1
Centella osiotico (or eordifolio )
Apiaceae
1
1
Centrolepis?
Centrolepidaceae
1
Comesperma retusum
Polygalaceae
1
Drosera binato
Droseraceae
1
1
1
Drosera peltata/auriculata
Droseraceae
1
Drosero spothulato
Droseraceae
1
Eleocharis sphacelata
Cyperaceae
1
1
Em pod ism a minus
Restionaceae
1
1
Epacris microphyllo
Ericaceae
1
1
Epacris paludosa
Ericaceae
1
1
1
Eriocaulon scoriosum
Eriocaulaceae
1
Eucalyptus camphora subsp. camphora
Myrtaceae
1
1
1
1
1
1
1
Eucalyptus pauciflora
Myrtaceae
1
Gahnia sieberiana
Cyperaceae
1
Geranium neglectum
Geraniaceae
1
1
1
Gleichenia dicarpa
Gleicheniaceae
1
1
1
Gonocarpus micrantha
Haloragaceae
1
Goodenia sp.
Goodeniaceae
1
1
Gratiola peruviana
Scrophulariaceae
1
Gymnoschoenus sphaerocephalus
Cyperaceae
1
Hakea microcarpa
Proteaceae
1
1
Hybanthus?
Violaceae
1
Hydrocotyle sp. (sibthorpioides ?)
Apiaceae
1
1
Hypericum sp. (gramineum ?)
Hypericaceae
1
Isachne globosa
Poaceae
1
Isotoma fluviatilis?= Pratia surrepens
Campanulaceae
1
Juncus spp.
Juncaceae
1
1
1
1
Lepidosperma limicola
Cyperaceae
1
Leptospermum continentale
Myrtaceae
1
1
1
1
Leptospermum grandifolium
Myrtaceae
1
1
1
Leptospermum obovatum
Myrtaceae
1
1
1
1
1
1
Lepyrodia sp.
Restionaceae
1
1
Lythrum salicaria
Lythraceae
1
Patersonia fragilis
Iridaceae
1
Phragmites australis
Poaceae
1
1
1
Polytrichum sp. (Dawsonia sp. ?)
Polytrichaceae
1
Pultenea divaricate
Fabaceae
1
Pultenea sp.
Fabaceae
1
Ranunculus sp.
Ranunculaceae
1
Senecio sp.
Asteraceae
1
Sphagnum cristatum
Sphagnaceae
1
1
1
1
1
Spiranthes australis
Orchidaceae
1
Stylidium graminifolium
Stylidiaceae
1
1
Tetrarrhena juncea
Poaceae
1
Typha sp. (orientalis ?)
Typhaceae
1
Utricularia dichotoma
Lentibulariaceae
1
1
Viola caleyana
Violaceae
1
1
Xyris ustulata
Xyridaceae
1
Xyris gracilis
Xyridaceae
1
10 Cunninghamia 18: 2018
Rollen Creek mire system: location and physiography
Because of its size, range of mire variation and good
condition, the mire system along Rollen Creek deserves
particular attention. Apparently once known as Rotten
Creek (G. Summers pers. comm.), Rollen Creek is an
upper Cudgegong River tributary about 13 km east of
Olinda. The upper swamp is located in Coricudgy State
Forest, with the contiguous downstream section of good
quality swamp on freehold land (“Inglewood”) extending
to where the Coricudgy Road crosses (32° 52' 18" S, 150°
16' 59"E). Though cattle have historically had some access
to at least parts of the swamp upstream of the Coricudgy
Road crossing, it is generally unsuitable for cattle, largely
due to its high water table and soft deep peaty soils (G.
Summers pers. comm.). A somewhat degraded area of
sedgeland-heath bog remnant also occurs downstream of
the Coricudgy Road crossing (between the road and cleared
grazing land) towards the Cudgegong River junction, but has
been impacted by grazing, fire and a previous timber mill
with associated sawdust dump which was located below the
crossing (G. Summers pers. comm.).
The Rollen Creek mire system (CM01) is narrow and
elongate, oriented roughly NNW-SSE along the creek
valley, about 3 km long (about 2.5 km in the State Forest)
and varies in width from about 10 to 100 m, but is mostly
<40 m wide. The catchment is Narrabeen Sandstone with
the exception of the isolated peaks of Mt Coorongooba
(-1060 m) and Mt Darcy (1079 m), with their residual
basalt, Coricudgy soil landscapes. Approximately 20%
of the slopes of Mt Coorongooba, and almost 50% of the
slopes of Mt Darcy are in the Rollen Creek catchment. The
area of the catchment is approximately 18 km 2 , but with
the complexity of the hydrogeology of these Narrabeen
sandstones, aquifers may be collecting water from a larger
area. There is a conspicuous seepage/spring at the head
of the main creek and several small drainage lines enter
the swamp system along its length; some also have areas
of groundwater-fed peaty swamps with sedgeland-heath
vegetation along their lower sections where they join
Rollen Creek.
The head of Rollen Creek mire is about 2 km from the
Divide (to the south); Mt Darcy and Mt Coorongooba are
both within 3 km of the head of the mire. Elevation ranges
from about 745 m at the source of the mire to 700 m at the
road crossing, giving a low overall gradient of 1.5% (15
m per km) and within the range of Newnes Plateau Shrub
Swamps (NPSS) (Benson & Baird, 2012). The downstream
end lacks a nick-point waterfall, characteristic of NPSS, but
grades into a Carex gaudichaudiana- dominated fen above
the Coricudgy Road crossing, and transitions to cleared
grazing land downstream of the Coricudgy Road crossing,
with the exception of the patch of remnant sedgeland-heath
bog below the Coricudgy Road. The extent of the fen area
is likely to have been increased (with conversion from
sedgeland-heath bog) by damming associated with the road
Baird & Benson, Montane mires, upper Cudgegong River catchment
crossing. The mire system is surrounded by dry eucalypt
woodland on poor sandy soils on adjoining slopes.
Rollen Creek mire system: vegetation structure and
composition
The Rollen Creek mire system is dominated by extensive,
valley-floor, sedgeland-heath bog, referable to the Tableland
Bogs typology. In addition to the seepage/spring at the head
of the main creek, there are small seepage/spring-fed mire
patches adjacent to the main valley-floor mire in at least two
other locations further downstream. There is also an extensive
valley-side hanging swamp dominated by Gymnoschoenus
sphaerocephalus.
The Tableland Bog along the main Rollen Creek valley
floor includes areas of closed and open sedgeland, with a
variable proportion of shrubs forming either sedgeland-
heath or with a mallee or tall shrub canopy (Figures 3-10).
The swamp is essentially treeless, though much of it is
dominated by areas of mallee or multi-stemmed, shrubby
Eucalyptus camphora subsp. camphora plants. Towards the
head of the mire system, where the valley narrows, older
Eucalyptus camphora occasionally reach 6-8 m high. In
this area, this mallee/tall shrub canopy (sometimes low
open woodland) generally forms dense scrub associated
with Callistemon citrinus and Leptospermum shrubs up
to 5 m high. Along the lower parts of the swamp system,
however, Eucalyptus camphora plants are much smaller
and often less than 2 m high. Plants with a range of sizes/
ages indicate continual recruitment of these species. Parts
of the upper swamp system with a canopy of Eucalyptus
camphora and other large shrubs also have a wet meadow/
forb-land ground-layer, which includes various herbs, Carex
gaudichaudiana , Sphagnum cristatum and unidentified
grasses (Fig. 3). A similar vegetation variant was also
recorded along a small tributary (CM02) adjoining the
Carex fen along Never Never Creek (CM03), along lower
Ganguddy Creek near the Coricudgy Road bridge (CM05,
CM06), and in association with small patches of sedgeland-
heath bog (CM08-10) along margins of the Carex fen along
upper Coxs Creek (CM07).
Throughout the sedgeland-heath dominated valley-
floor bog there may be an open or sparse, small-medium
height shrub layer, including Baeckea utilis , Comesperma
retusum , Epacris microphylla, Epacris paludosa , Hakea
microcarpa , Leptospermum obovatum and Pultenea
divaricata. The groundcover is generally very dense,
mostly >90% cover, predominantly of Sphagnum moss,
sedges and other graminoids, smaller shrubs and herbs,
including carnivorous species such as Drosera binata and
Utricularia dichotoma. Herbs and ferns include Centella
asiatica, Geranium sp., Gleichenia dicarpa , Goodenia sp.,
Isachne globosa , Spiranthes australis and Viola caleyana.
Sedgeland species within the sedgeland-heath include
Baumea sp., Baloskion australe , Empodisma minus ,
Gymnoschoenus sphaerocephalus , Juncus sp., Lepyrodia
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
11
sp., Tetrarrhena juncea andXyris ustnlata. There are patches
with Sphagnum hummocks up to 1 m high and the soil
surface is spongy on flat and steeper side slopes (Figures 3,
10). There may also be localised areas of Sphagnum bog in
shadier, wetter and more fire-protected sites along tributary
drainage lines.
Carex gaudichaudiana is scattered throughout the valley
floor mire in localised patches of fen amongst sedgeland-
heath bog in wetter low gradient areas along the drainage
lines (Fig. 6), and as a discrete patch of Carex-dominated fen
referable to Tableland Swamp Meadow at the downstream
end near the Coricudgy Road crossing (Fig. 11).
In hanging swamps on side slopes open patches of
sedgeland with few or no shrubs may have Gymnoschoenus
sphaerocephalus and Gleichenia dicarpa dominant, with
Empodisma minus , Lepidosperma limicola and Xyris ustulata
less abundant (Figures 8-10). Baeckea utilis generally
occurs as scattered individuals across these Gymnoschoenus-
dominated hanging swamp areas, with occasional other
shrubs present. Upper hanging swamp margins are relatively
distinct with a band of Gleichenia along the edge of the dry
eucalypt woodland; Eucalyptus camphora also occurs on
swamp margins.
The main woodland dominants adjoining the swamps are
Eucalyptus dives , Eucalyptus radiata , Eucalyptus rossii
and Eucalyptus dalrympleana. In one place (32° 53' 24" S,
150° IT 24" E), the geographically restricted Eucalyptus
corticosa was recorded in woodland adjacent to the upper
margin of the hanging swamp on the eastern side of the
Rollen Creek mire. Eucalyptus pauciflora may occur on
woodland margins sometimes edging into the swamp, as it
does in the higher elevation mires of the Newnes and Boyd
plateaus further south.
About 48 native species, including most of the conspicuous,
abundant and structural species, were recorded in Rollen
Creek mire (Table 2). This is an incomplete list as some
grasses and monocots were not identified, and some seasonal
species may not have been evident at the time. Appendix 1
includes the list of species recorded in Rollen Creek swamp
with reference to species also recorded in NPSS (Benson &
Baird, 2012; Benson & Keith, 1990) and Boyd Plateau Bogs
(Keith & Benson, 1988; Kodela et al., 1996).
Rollen Creek mire system: hydrology
Given the relatively low average annual rainfall compared
to the nearby basalt-capped mounts and the higher elevation
parts of the Blue Mountains further south, and its high inter¬
annual variability at the property “Kelgoola”, the presence
of the well-developed mire system along Rollen Creek
indicates a strong groundwater influence, in addition to
rainfall input. This is exemplified by the presence of springs
and hanging swamps with groundwater seepage adjacent
to the valley floor mire. The Narrabeen Sandstone of the
upper Blue Mountains is also characterised by complex
groundwater hydrology and the presence of aquifers which
support a diversity of groundwater-dependent swamp
communities (e.g., Baird, 2012; DLWC, 1999a, b; Marshall,
2005). The expected higher rainfall in the higher elevation
areas above 1000 m, around the catchment watershed in
the headwaters of the upper Cudgegong River catchment,
is likely to contribute significantly to the presence of the
groundwater-dependent mires through both direct runoff,
and infiltration into the groundwater system in the relatively
permeable geology.
The characteristic swamp vegetation and extensive
occurrence of peaty/organic-rich mire sediments indicates
a permanently high water table in Rollen Creek mire.
Peaty sediments include fibrous peats and more highly
decomposed sapric peats, which require such hydrological
conditions to develop. Depth and characteristics of the peaty/
organic-rich sediments can be expected to vary considerably
across the swamp system, as occurs in NPSS for example,
but the depths recorded in the downstream part of Rollen
Creek swamp (>1.8 m) typically exceeded depths recorded
in NPSS by Benson and Baird (2012). They are, however,
similar to depths recorded in the somewhat similar, low
gradient montane bog and fen complex at the head of Long
Swamp (part of the Coxs River Swamps of Benson & Keith,
1990) in Ben Bullen State Forest (Baird, 2014; IRCB,
unpubl. obs.; Martin, 2017).
The hanging swamps on the valley side, dominated by
Gymnoschoenus sphaerocephalus , and contiguous with
the main valley floor mire (Figures 8-10), are similar to
those which occur elsewhere in the Blue Mountains on
sandstone geology. These hanging swamps are typically
associated with the presence of a low-permeability aquitard
(e.g., claystone layer). They develop where groundwater is
redirected downslope following the slope of the aquitard to
emerge on the valley side as seepage. These hanging swamps
along Rollen Creek include groundwater seeps and springs
with associated groundwater-dependent species such as
burrowing crayfish ( Euastacus sp.) (Fig. 10).
The predominance of Sphagnum in the Rollen Creek mire
system is indicative of persistent or permanent wetness. The
hummocks are slow growing and the large hummocks were
associated with large shrubs with a cluster of basal stems,
or large tussocks of sedgeland vegetation, which appear
to provide support for the hummock growth and provide
moderate levels of shade. Larger Sphagnum hummocks
in Boyd Plateau Bogs are similarly associated with multi¬
stemmed shrubs and sedgeland tussocks (IRCB unpubl. obs.).
The Sphagnum is vulnerable to fire during drought, when the
hummocks may dry out (see Hope et al., 2009; Whinam,
1995; Whinam & Chilcott, 2002; Whinam et al., 1989).
12
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
Fig. 3: Dense Sphagnum cover and hummocks in Carex and grass-
dominated area in the upper section of Rollen Creek swamp, with
Eucalyptus camphora subsp. camphora. Photo by Ian Baird
Fig. 4: Lignotuberous Eucalyptus camphora subsp. camphora
resprouting after fire along a drainage line through a
Gynmoschoenus-domimtQd patch of swamp in the upper Rollen
Creek. Photo by Ian Baird
Fig. 6: A patch of Carex gaudichaudiana fen amongst sedgeland-
heath in a low gradient section of Rollen Creek swamp. Note the
emergent water table (bottom centre) and Eucalyptus camphora
subsp. camphora in front of adjoining woodland at rear. Photo by
Ian Baird
Fig. 5: Dense montane sedgeland-heath bog along the Rollen
Creek valley floor. Photo by Ian Baird
Fig. 7: Valley floor bog with a patch of Empodisma minus and
Tetrarrhena ///wcea-dom mated ground-layer, with low Eucalyptus
camphora subsp. camphora (rear and right). Photo by Ian Baird
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
13
Fig. 8: Gymnoschoenus- dominated hanging swamp in the upper
section of Rollen Creek. The fringing low Eucalyptus camphora
subsp. camphora is in front of adjoining upslope woodland at rear.
Photo by lan Baird
Fig. 10: Seepage in a Gymnoschoenus-domimtQd section of
a hanging swamp in the upper Rollen Creek with Sphagnum ,
Gleichenia dicarpa and scattered shrubs of Baeckea utilis evident.
A Euastacus sp. crayfish burrow is visible in the pool. Photo by
Ian Baird
Fig. 9: Resprouting Gymnoschoenus sphaerocephalus tussocks
post-fire in a hanging swamp in the upper Rollen Creek swamp.
Note the large size (and age) of the bases of tussocks, fire-killed
shrubs and fringing resprouting Eucalyptus camphora subsp.
camphora (rear). Photo by Ian Baird
Fig. 11: Carex gaudichaudiana -dominated montane fen (centre left)
in a low gradient area of the downstream end of the Rollen Creek
swamp, with fringing sedgeland-heath bog. Photo by Ian Baird
14 Cunninghamia 18: 2018
Fig. 12: Valley-floor Carex fen along Never Never Creek above
Dunns Swamp, with scattered Eucalyptus camphora subsp.
camphora and Leptospermum shrubs. Photo by Ian Baird
Fig. 13: Valley-floor Carex fen with flowering Lythrum salicaria
along Coxs Creek (looking upstream) showing cleared grazing land
on private property on right-hand (southern) side and good quality
bushland on the left-hand side. Small patches of fringing sedgeland-
heath bog, with some Eucalyptus camphora, occur in embayments
along fen margins at right. Photo by Ian Baird
Discussion
Comparison of montane mires of the upper Cudgegong
River catchment with other montane mires of the Blue
Mountains
We found a range of mires with considerable floristic and
hydrogeomorphic diversity across the upper Cudgegong
River catchment. These mires can be broadly divided into
montane bogs, montane fens and hanging swamps; all
groundwater dependant ecosystems typically found in
high rainfall areas. Keith (2004) indicated a rainfall range
of 800-1500 mm/year for the Montane Bogs and Fens
vegetation class and 1000-1500 mm/year for the Coastal
Heath Swamps vegetation class. The Cudgegong mires
are at the extreme lower limits of these ranges. The upper
Baird & Benson, Montane mires, upper Cudgegong River catchment
Cudgegong montane bogs (Tableland Bogs) and montane
fens (Tableland Swamp Meadows) occur in a lower rainfall
area, and at relatively lower elevation, compared to similar
Blue Mountains mires further south. Mires in Ben Bullen
State Forest and on the Newnes Plateau are above 900 m
(but most above 1000 m), compared to those at 660-745 m
in the upper Cudgegong (Table 1). Mean annual rainfall for
these central Blue Mountain mires generally exceeds 1000
mm, and on the Boyd Plateau and nearby areas (above 1100
m elevation) exceeds 1100 mm.
The hanging swamps of the upper Cudgegong River
catchment, particularly along upper Rollen Creek, also
occur in a lower rainfall area than similar hanging swamps
on sandstone in the central Blue Mountains (BMSS),
particularly those on the eastern side of the Blue Mountains
at similar elevation, where mean annual rainfall also exceeds
1100 mm, compared to mean annual rainfall of 809 mm at
“Kelgoola”, near Rollen Creek mire. In upper parts of the
Blue Mountains, where the best developed hanging swamps
on sandstone occur, mean annual rainfall may exceed 1300
mm. In lower elevation areas with considerably lower
rainfall and higher evapotranspiration in the lower reaches
of the upper Cudgegong River catchment east of Rylstone
(mean annual rainfall <750 mm), to the north of the upper
Cudgegong River catchment in the Growee River catchment,
and to the south in the Capertee River catchment, suitable
conditions for mire development are evidentially absent.
Floristically, the Rollen Creek mire is similar to NPSS with
about 65% of the 48 recorded native species in common,
particularly shrub species (see Appendix 1), though the
predominance of Eucalyptus camphora , restricted mainly
to these swamps, is a conspicuous point of difference.
Eucalyptus camphora subsp. camphora is geographically
restricted to Nullo Mountain and the upper Cudgegong River
catchment, with a localised disjunct population in swamps in
the Megalong Creek valley near Katoomba. It is also of note
in being restricted mainly to swamp habitat; indeed there
are very few Eucalyptus species that are found in wet or
poorly drained sites and Eucalyptus camphora appears to be
confined to this habitat. The only similar swamp-inhabiting
eucalypt in the Sydney Basin bioregion (there are almost 100
eucalypt species in the the GBMWHA; Benson & Smith,
2015) is the closely related Eucalyptus aquatica, which is
found in analogous montane mires in the Southern Highlands
(Penrose State Forest) (Shepherd & Keyzer, 2014).
The abundance of Callistemon citrinus along Rollen
Creek is another conspicuous difference. This species does
not occur in NPSS or Boyd Plateau Bogs, but is the only
Callistemon found in BMSS (and some transitional to NPSS).
Interestingly, Callistemonpityoides, the typical montane bog
Callistemon (found in Boyd Plateau Bogs, some NPSS, and
montane bogs elsewhere), was not recorded anywhere in the
upper Cudgegong swamps, nor does it occur in BMSS. Other
conspicuous NPSS species not so far recorded are Celmisia sp.
(aff. longifolia), Boronia deanei subsp. deanei and Grevillea
acanthifolia subsp. acanthifolia. Celmisia and Boronia
deanei also occur in Boyd Plateau Bogs. These species may
be more drought-sensitive and given the much less extensive
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
15
swamp areas on the upper Cudgegong, in comparison with
the main Blue Mountains area further south, could have been
extirpated by past periods of extensive drought, assuming
they have been present in the past.
Although Leptospermum grandifolium was recorded
in the small patch of hanging swamp above Ganguddy
Creek (CM04) and a nearby small tributary bog (near and
downstream of the bridge on the Coricudgy Road) (CM06),
the apparent lack of the species in Rollen Creek swamp and
other surveyed sites is noteworthy, considering that it is a
common shrub in NPSS, BMSS and Boyd Plateau Bogs
(interestingly, single plants of Epacris paludosa , which was
otherwise only recorded from Rollen Creek swamp, were
also recorded in CM05 and the nearby CM06). There is
also a general lack of dominance by Leptospermum species,
compared to their abundance in many NPSS. The only
species recorded in Rollen Creek swamp was Leptospermum
obovatum, which is common in Rollen Creek swamp, occurs
in some NPSS, and is common in Boyd Plateau Bogs.
Leptospermum continentalis also occurs on heavier soils on
drier margins of swamps. Baeckea utilis, the only Baeckea
species present, and widely distributed in Rollen Creek
swamp, is a typical montane bog and wet heath species
which occurs in and around Boyd Plateau Bogs and some
higher elevation Newnes Plateau swamps, often on heavier
soils, but does not occur in BMSS. Baeckea utilis is typically
replaced by Baeckea linifolia in most NPSS and in the
higher elevation Coastal Heath Swamps of the region such
as BMSS.
Most other abundant species are similar to those occurring
in NPSS, including Lepidosperma limicola, Empodisma
minus, Epacris paludosa and Leptospermum obovatum.
Although Gymnoschoenus sphaerocephalus occurs in small
patches within the main valley floor bog and in upstream
areas where the valley narrows, it is mostly restricted to
the hanging swamps and seepage/spring-fed mire patches
adjacent to the main valley floor mire. Rollen Creek was
the only mire in the upper Cudgegong catchment where
Gymnoschoenus was recorded, although it has been recorded
just east of Mt Coricudgy, in a previously undocumented
small hanging swamp (990 m elev.) where Petalura gigantea
was recorded (Baird, 2012: Appendix 1, site PMC01).
Gymnoschoenus is widely distributed in mires on sandstone
elsewhere in the Blue Mountains (in NPSS and BMSS) and
the Southern Highlands, and montane bogs of the Northern
and Southern Tablelands (mostly on granite), but it appears
to be absent from the granite-based Boyd Plateau Bogs. The
only occurrence in Tableland Bog on granite in the Blue
Mountains region which is known to one of the authors
(IRCB) is in a small montane bog south of the Kowmung
River near Trailers Mountain, although it may also occur in
several small unsurveyed bog patches nearby. Floristically
about 58% of the 48 recorded native species in Rollen Creek
swamp are shared with the Boyd Plateau Bogs, including
herbaceous species such as Centella asiatica, Geranium sp.,
Hypericum sp., and Viola caleyana (Appendix 1).
Areas of hanging swamp dominated by Gymnoschoenus
sphaerocephalus (Figures 8-10) are not referrable to any
of the swamp vegetation types in the upper Cudgegong
area identified by Bell (1998a). With the exception of the
occurrence of Eucalyptus camphora and Baeckea utilis ,
these hanging swamps are more similar floristically and
hydrogeomorphically to those occurring elsewhere on
Narrabeen Sandstone in the Blue Mountains, such as the
BMSS of Keith and Benson (1988).
The presence of extensive areas of Sphagnum cristatum
along Rollen Creek, often forming large hummocks within
the sedges and shrubs, was noteworthy. In comparison with
Rollen Creek swamp, Sphagnum is either very restricted or
absent from NPSS and BMSS (but note the extensive and
unusual Sphagnum cover in the fen/marsh in Goochs Crater
in the upper Wollangambe River catchment and in a BMSS in
McCrae’s Paddock in Katoomba). The much more extensive
Sphagnum cover in Rollen Creek swamp (and elsewhere in
mires and other parts of the upper Cudgegong catchment) is
surprising considering the lower rainfall recorded at nearby
“Kelgoola”, compared to NPSS and other central Blue
Mountains mires. Its persistence is likely to be the result, at
least in part, of a sustained groundwater influence and the
nearby presence of higher elevation peaks upstream around
the catchment watershed contributing additional rainfall
inputs as a result of an orographic effect. According to Chris
Pavich (pers. comm.), Sphagnum was more abundant in
the upper Cudgegong before the drought of the 1930-40s,
with a subsequent increase in fire and grazing probably
contributing to its further disappearance (C. Pavich pers.
comm.). Rollen Creek swamp, however, has not been subject
to heavy grazing or frequent anthropogenic fire (G. Summers
pers. comm.), which may have helped prevent loss of
Sphagnum. Sphagnum is generally much more abundant in
montane bogs on granite, such as those of the Boyd Plateau,
and nearby areas on metasedimentary geology (e.g., near Mt
Werong and in Jenolan State Forest), than in the swamps
developed on sandstone in the Blue Mountains, such as the
NPSS and BMSS (also see Downing et al., 2007; Whinam
& Chilcott, 2002). Differences in fire history and climatic
factors between these areas are likely to be contributing
factors.
While there are small patches of Sphagnum bog in the upper
Cudgegong which are broadly consistent with Bell (1998a)’s
Cudgegong Sphagnum Bogs, these are treated here as part of
a variable upper Cudgegong montane bog type. Sphagnum
also occurs along various drainage lines (often on heavy
soils) and in localised patches within the other swamp
types (including areas transitional between Carex fen and
sedgeland-heath bog) where suitable conditions occur. The
Cudgegong River Swamp Grassland and Upper Cudgegong
Alluvial Sedgeland of Bell (1998a) are also treated as part
of the montane fens identified in this study. As a result of
more extensive survey across the study area, and based on
their broadly similar floristics and potential organic-rich
substrates, as described by Bell (1998a) in the Dunns Swamp
area, Baird (2012) noted that there is considerable gradation
between these swamp vegetation types.
In summary, the Rollen Creek swamp includes extensive
valley floor mire of intergrading bog and fen, with some
16 Cunninghamia 18: 2018
very small valley-side seepage/spring areas, and extensive
valley-side hanging swamp dominated by Gymnoschoenus.
The valley-side seeps, springs and hanging swamps are most
similarto other such groundwater-dependent mire expressions
developed on Narrabeen Sandstone in the Blue Mountains,
such as the NPSS and BMSS. While these seeps and hanging
swamps are hydrogeomorphically similar to similar mires
developed on sandstone geology elsewhere in the Blue
Mountains, they are characterised by a somewhat distinctive
floristic assemblage. BMSS form part of the Coastal Heath
Swamp vegetation class (NSWSC, 2007), although most are
above the nominal 600 m upper elevation range indicated
by Keith (2004) for the Coastal Heath Swamps. NPSS are
considered transitional between the Coastal Heath Swamps
and Montane Bogs and Fens vegetation classes (NSWSC,
2005b), with that transition occurring between Bell and
Clarence. The valley floor mire along Rollen Creek,
however, with areas of bog and fen, shows greater affinity,
respectively, to the Tableland Bogs and Tableland Swamp
Meadows of Tozer et al. (2010), within the Montane Bogs
and Fens vegetation class. The valley-floor bogs along Rollen
Creek and elsewhere in the upper Cudgegong have floristic
and hydrogeomorphic similarities with montane bogs of the
Southern, Central and Northern Tablelands (e.g., Hunter &
Bell, 2007; Tozer et al., 2010). The Carex fens along Coxs
Creek and Never Never Creek, at the downstream end of
Rollen Creek swamp, and similar areas across the upper
Cudgegong River catchment, have affinities with the Carex
fen vegetation of Northern NSW (Hunter, 2013; Hunter &
Bell, 2009), the Carex-Poa fen vegetation of Long Swamp
in Ben Bullen State Forest and nearby areas (part of the
Coxs River Swamps of Benson & Keith, 1990), the swampy
meadows of the Central Tablelands (see Mactaggart et al.,
2008; Mactaggart, 2008) and the Mountain Hollow Grassy
Fens (DEC, 2006). This complexity highlights the value of
this and other geographically isolated mires of the upper
Cudgegong River catchment.
Conservation value of Rollen Creek swamp and other
remaining mires of the upper Cudgegong River catchment
Mires are geographically restricted ecosystems in Australia
and their extent and health have been considerably reduced
and degraded since European settlement through urban and
transport infrastructure development, agriculture, drainage,
grazing, and more recently through mining impacts,
particularly the impacts of subsidence from longwall coal
mining, a Key Threatening Process in NSW (CoA, 2005,
2010; NSWSC, 2005a). Climate change and fire are also
recognised as significant threats to these ecosystems (e.g.,
Baird & Burgin, 2016; CoA, 2010; Keith et al., 2014;
Pemberton, 2005).
In the Sydney Basin bioregion, Temperate Highland Peat
Swamps on Sandstone (THPSS) is listed as an Endangered
Ecological Community (EEC) under the Commonwealth
Environment Protection and Biodiversity Conservation Act
1999 (CoA, 2005). THPSS also includes NPSS and BMSS.
Under the NSW Biodiversity Conservation Act 2016 (which
replaced the NSW Threatened Species Conservation Act
Baird & Benson, Montane mires, upper Cudgegong River catchment
1995) NPSS is also listed as an EEC (NSWSC, 2005b), and
Blue Mountains Swamps (including the Blue Mountains
Sedge Swamps of Keith and Benson (1988)) is listed as
a Vulnerable Ecological Community (NSWSC, 2007).
Although not specifically included in the description of
THPSS, the authors consider that some of the mires of the
upper Cudgegong, particularly the Rollen Creek valley
floor mire and hanging swamps, are clearly referrable to
the THPSS EEC, and have been simply overlooked because
of lack of documentation. For a similar reason, none of
the valley-floor bogs and fens in the upper Cudgegong are
specifically identified in the Montane Peatlands and Swamps
of the New England Tableland, NSW North Coast, Sydney
Basin, South East Corner, South Eastern Highlands and
Australian Alps bioregions EEC determination (NSWSC,
2004). In circumscribing the Montane Peatlands and Swamps
EEC, the NSWSC (2004) referred to swamps above 400-
600 m elevation. The montane bogs and fens of the upper
Cudgegong are also referrable to this EEC and are clearly an
important part of the complex of endangered montane mire
communities distributed across the tablelands and adjacent
ranges of NSW.
Habitat for rare species- flora
The mires of the upper Cudgegong River catchment are
isolated from other montane mires. The nearest montane
mires to the north occur at high elevation at Barrington Tops
(Mort, 1983), although somewhat drier and more floristically
impoverished wet heath vegetation, with some floristic
affinity to the wet heath in the upper Cudgegong mires,
occurs on heavier soils on Coolah Tops (Binns, 1997). The
nearest to the south occur on the Newnes Plateau and the
adjacent Ben Bullen State Forest.
This isolation contributes to the significance of some rare and
restricted plant species in the mires and adjacent woodlands
of the upper Cudgegong. For example, the distribution of the
mallee Eucalyptus camphora subsp. camphora , illustrates the
disjunct biogeography of many montane species restricted to
specialised habitats. This taxon is geographically restricted
to higher elevation mires and swampy drainage lines along
this part of the Divide, with populations at Nullo Mountain
and in the upper Cudgegong River catchment (Rollen Creek
swamp in particular has a large population), and a highly
localised and disjunct population 90 km further south in the
Megalong Creek Valley near Katoomba, where it occurs in
swamp patches with different floristics on heavier clayey
soil and colluviums on Permian geology (600 m elev.). This
locality in the Megalong Creek valley is also noteworthy for
the occurrence of the rare and locally endemic Callistemon
megalongensis (Craven, 2009; Udovicic & Spencer, 2012)
and Callistemonpurpurascens (Douglas & Wilson, 2015).
Another rare eucalypt Eucalyptus corticosa, a locally
endemic tree species restricted to the upper Cudgegong
valley east of Rylstone, occurs in eucalypt woodland on
shallow infertile soils on sandstone ridges and was recorded
on the outer edge of the Gymnoschoenus-dommatQd hanging
swamp along Rollen Creek. Veronica blakelyi, a small shrub
restricted to the western Blue Mountains, near Clarence,
Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
17
near Mt Horrible, on Nullo Mountain and in the Coricudgy
Range, is listed as Vulnerable under the NSW Biodiversity
Conservation Act 2016 and was recorded in woodland
adjoining Rollen Creek swamp. Additional rare plant species
are likely to occur in the woodlands adjoining these swamps,
including in Coricudgy State Forest.
Habitat for rare species- fauna
The mires of the upper Cudgegong, particularly Rollen
Creek, may also provide habitat for specialist mire fauna.
Baird (2012) considered that the upper Cudgegong mires,
particularly the valley-floor bogs and Gymnoschoenus-
dominated hanging swamps, provide potential habitat for
the endangered dragonfly, Petalura gigantea. The record
of a single Petalura in a small hanging swamp (one of
several such previously undocumented swamps in close
proximity) east of Mt Coricudgy in Wollemi National Park,
a considerable distance from the nearest known populations
on the Newnes Plateau (Baird, 2012), strongly suggests that
a population occurs somewhere within the complex of mires
in the upper Cudgegong. Rollen Creek swamp is the most
likely habitat, although surveys on three occasions have not
recorded the species.
Rollen Creek swamp also appears to provide suitable
habitat for the endangered Blue Mountains Swamp Skink,
Eulamprus leuraensis, which is only known from mid-upper
elevation mires in the central Blue Mountains (BMSS) and
the Newnes Plateau (NPSS). This groundwater-dependent
species has not been recorded further north than the Newnes
Plateau (Gorissen, 2016; LeBreton, 1996). Evidence of
foraging and tunnelling activity of Swamp Rats, Rattus
lutreolus, a common species of BMSS and NPSS, was
also observed during fieldwork. Additional obligate mire-
dwelling fauna are likely to occur here (e.g., skinks and
invertebrate stygofauna), but more detailed surveys will be
required to identify them.
A burrowing spiny crayfish, Euastacus australasiensis,
is widely distributed in Blue Mountains mires (including
hanging swamps), and Euastacus burrows and partial
remains were observed by the authors in the Rollen Creek
mire system (including hanging swamps) and (by IRCB)
in the small hanging swamp east of Mt Coricudgy where
Petalura gigantea was recorded. While these are likely to
be Euastacus australasiensis (see McCormack, 2012), the
possibility exists that an unidentified swamp-dwelling taxon
is involved. In the forested headwaters of the Cudgegong
River in Coricudgy State Forest, below Mt Coricudgy
and upstream of Rollen Creek, a large, stream-dwelling,
burrowing spiny crayfish species, Euastacus vesper , closely
related to Euastacus spinifer of eastern drainages, has
recently been described (McCormack & Ahyong, 2017).
The discovery of this apparently highly-localised species
provides further evidence of the high conservation value of
Coricudgy State Forest and of the opportunities for further
research in this area.
Value of mires for palaeo-ecological, evolutionary and
climate studies
Previous palaeoecological studies of Blue Mountains mires,
based upon pollen and charcoal analysis of radiometrically-
dated sediment cores, have greatly increased our
understanding of the developmental history of these mires
and their past climates and vegetation, particularly since
the Last Glacial Maximum (~21 000 years BP). While the
oldest sediment core ages suggest some mires may have
commenced development around 13 000 years BP, sediment
cores from other swamps have maximum ages within the
Holocene (e.g., Black et al., 2008; Chalson & Martin, 2009;
Fryirs et al., 2014; Martin, 2017). The mires of the upper
Cudgegong River catchment can be expected to be a similarly
rich source of knowledge related to species distributions, the
developmental history of these mires and of climatic change
since the Last Glacial Maximum, and expand our existing
understanding of vegetation change across southeastern
Australia and the Sydney Basin during this period.
The biogeography of the mallee Eucalyptus camphora
subsp. camphora, geographically restricted to localised
high elevation patches of suitable habitat along the Divide,
is similar to that of many montane species with disjunct
populations restricted to specialised habitats. Such disjunct
distributions are likely to reflect, or be the result, of past
climatic fluctuations to some extent, either during the Last
Glacial Maximum or as a result of previous glacial/interglacial
cycles. For Eucalyptus camphora , two other subspecies have
also been recognised; subsp. relicta , found further north at
Guyra and in Queensland, and subsp. humeana , occurring
from Wee Jasper south into Victoria. The distribution of
these taxa, with their greater morphological and geographic
variation, presumably indicates older geographic separation
than that within subsp. camphora, and provides evidence
of past climate-related divergence. Other montane species
restricted to specialised habitats, such as mires, have similar
biogeographic patterns; understanding these patterns is of
considerable scientific interest, particularly in the context of
a rapidly changing climate.
Conclusion
The upper Cudgegong River catchment includes a complex
of endangered peat swamp or mire types which are
geographically disjunct from their nearest neighbours and
characterised by some distinctive floristic assemblages.
These include areas of montane bog, montane fen and
hanging swamp. The presence of rare species such as the
mallee Eucalyptus camphora subsp. camphora , and the
potential for endangered fauna such as Petalura gigantea
to be present, further highlights their value. In the context
of the relatively low rainfall where these groundwater-
dependent mires occur (and the upper Cudgegong is at their
climatic limits), there is a surprising hydrogeomorphic and
floristic diversity across these different mire types. With
their relatively low rainfall, these mires may be particularly
vulnerable to climate change.
18 Cunninghamia 18: 2018
Historically there has been considerable loss and degradation
of the mires in the upper Cudgegong River catchment through
land clearing and agriculture, beginning in the 1840s, and
evidence of this damage provides a strong imperative to
protect those examples that have survived. Conservation of
these mires and their associated flora and fauna will benefit
from further survey, mapping and biodiversity census. In
addition to improved management of identified mires on
private lands, including the possible use of biodiversity
conservation covenants and incentives, improved recognition
of the inherent values of these mires wifi be fundamental to
their long-term conservation. Rollen Creek swamp is unique
in the area, in terms of its size, floristic and hydrogeomorphic
diversity, and good condition, and its conservation must be a
priority. National Heritage listing of Coricudgy State Forest
would provide a substantial foundation to highlight the values
of this mire system and its surrounding woodland landscape,
including the biodiversity associated with the significant,
higher rainfall, basalt-capped island peaks adjoining the
GBMWHA. Such listing is a prerequisite for nomination
of this and other high biodiversity areas for addition to the
GBMWHA and is recommended.
Acknowledgements
Gay Summers, owner of “Inglewood”, generously provided
access to Rollen Creek swamp on various occasions and
shared her local knowledge. Chris Pavich (NPWS) shared
his extensive local knowledge and provided assistance with
property access for IRCB during earlier fieldwork. Ahamad
Sherieff (NSW Office of Environment and Heritage)
kindly prepared the line map of the study area. Huw Evans
(Local Land Services) provided assistance during an earlier
fieldtrip. Rob McCormack reviewed the text on spiny
crayfish and Scott Mooney and an anonymous reviewer are
thanked for their useful comments which have improved the
paper. Initial fieldwork by IRCB was supported by a Western
Sydney University Post-graduate Research Award.
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Cunninghamia 18: 2018
Baird & Benson, Montane mires, upper Cudgegong River catchment
21
Appendix 1: Native plant species (including mosses) recorded in Rollen Creek mire (TB=Tableland Bog, TSM=Tableland Swamp
Meadow, HS= Hanging Swamp), and whether those species have also been recorded in Newnes Plateau Shrub Swamps (NPSS)
(Benson & Baird, 2012) or Boyd Plateau Bogs (Kodela et al., 1996).
PLANT SPECIES
Family
Rollen Creek mire
NPSS
Boyd Plateau Bogs
Asperula gunnii
Rubiaceae
TB
NPSS
BPB
Baeckea utilis
Myrtaceae
TB, HS
NPSS
BPB
Baloskion austrole
Restionaceae
TB, HS
NPSS
BPB
Baumeo sp.
Cyperaceae
TB, HS
NPSS
BPB
Blechnum nudum
Blechnaceae
TB
NPSS
BPB
Bulbine bulbosa
Liliaceae
TB
Callistemon citrinus
Myrtaceae
TB, HS
Carex goudichaudiana
Cyperaceae
TB, TSM
BPB
Centella asiatica (or cordifolia )
Apiaceae
TB
BPB
Centrolepis?
Centrolepiaceae
TB
Comesperma retusum
Polygalaceae
TB
BPB
Drosera binata
Droseraceae
TB, HS
NPSS
BPB
Drosera spathulata
Droseraceae
TB
NPSS
Empodisma minus
Restionaceae
TB, HS
NPSS
BPB
Epacris microphylla
Ericaceae
TB
NPSS
BPB
Epacris paludoso
Ericaceae
TB, HS
NPSS
BPB
Epilobium gunnianum
Onagraceae
BPB
Eriocoulon scariosum
Eriocaulaceae
TB
Eucalyptus camphora subsp. camphora
Myrtaceae
TB, HS
Eucalyptus pauciflora
Myrtaceae
TB
NPSS
BPB
Gahnia sieberiana
Cyperaceae
NPSS
Geranium neglectum
Geraniaceae
TB
BPB
Gleichenia dicarpa
Gleicheniaceae
TB, HS
NPSS
BPB
Gonocarpus micrantha
Haioragaceae
TB
NPSS
BPB
Goo den ia sp. (bellidifolia)
Goodeniaceae
TB
NPSS
Gymnoschoenus sphaerocephalus
Cyperaceae
TB, HS
NPSS
Hakea microcarpa
Proteaceae
TB
NPSS
BPB
Hybanthus?
Violaceae
TB
Hydrocotle sp. ( sibthorpioides ?)
Apiaceae
TB
NPSS
BPB
Hypericum sp. (gramineum ?)
Hypericaceae
TB
NPSS
Isachne globosa
Poaceae
TB
Isotoma fluviatilis?- Pratia surrepens
Campanulaceae
TB
BPB
Juncus spp.
Juncaceae
TB, TSM
NPSS
BPB
Lepidosperma limicola
Cyperaceae
TB, HS
NPSS
Leptospermum continentale
Myrtaceae
TB
NPSS
BPB
Leptospermum myrtifolium
Myrtaceae
NPSS
BPB
Leptospermum obovatum
Myrtaceae
TB, HS
NPSS
BPB
Lepyrodia spp.
Restionaceae
TB
NPSS
Patersonia fragilis
Iridaceae
TB
NPSS
BPB
Pultenea di\/aricata
Fabaceae
TB, HS
NPSS
Pultenea sp.
Fabaceae
TB
Ranunculus sp.
Ranunculaceae
TB
Scirpus polystachyus
Cyperaceae
BPB
Senecio sp.
Asteraceae
TB
Sphagnum cristatum
Sphagnaceae
TB, HS
BPB
Spiranthes australis
Orchidaceae
TB
BPB
Stylidium graminifolium
Stylidiaceae
TB
NPSS
BPB
Tetrarrhena juncea
Poaceae
TB
NPSS
Utricularia dichotoma
Lentibulariaceae
TB
NPSS
BPB
Viola caleyana
Violaceae
TB
NPSS
Xyris ustulata
Xyridaceae
TB, HS
NPSS
BPB
Xyris gracilis
Xyridaceae
TB
NPSS
TOTAL
48
33
31
Number RC shared with NPSS and/or BPB
31
28
Percentage RC shared with NPSS and/or BPB
65%
58%
Date of Publication:
August 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
Insects associated with flowering of Rhodomyrtus psidioides (Myrtaceae):
Is this a Myrtle Rust (Austropuccinia /w/r///)-in duced Plant-pollinator
interaction Extinction Event?
Geoff Williams
Lorien Wildlife Refuge and Conservation Area, Lansdowne via Taree, NSW 2430. Honorary Research Associate,
Australian Museum, College Street, Sydney, NSW 2000 AUSTRALIA.
Abstract : The threatened Australian endemic rainforest tree Rhodomyrtus psidioides (Myrtaceae) is visited and
pollinated by a taxonomically diverse assemblage of mainly small, ecologically unspecialised, insects. Flower structure
suggests that it may also be adapted for wind-pollination. However, the recent (2010) invasion by the aggressive
fungal pathogen Myrtle rust ( Austropucciniapsidii) has resulted in the local extinction of both the floral resource and
associated plant-insect relationships. Here I table observed insect visitors to the flowers of Rhodomyrtus psidioides
made before the impact of Myrtle rust - no other records appear to have been published.
KeyWords: Rhodomyrtus psidioides, Myrtaceae, threatened plants, extinction, subtropical rainforest, Lorien Wildlife
Refuge, anthophilous insects, wind-pollination.
Cunninghamia (2018) 18: 023-027
doi: 10.7751 / cunninghamia.2018.002
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
24
Cunninghamia 18: 2018
Williams , a Myrtle Rust Extinction Event?
Introduction
In the 1990s, as part of my PhD investigation of the pollination
ecology of lowland subtropical rainforests in northern New
South Wales (Williams 1995), I undertook the observation
and hand collection of insects that visited the flowers of the
small myrtaceous tree Rhodomyrtus psidioides (G. Don)
Benth.). This was one of numerous mass-flowering tree
and shrub species I investigated over a three-season period
(1990-1993); owing to time constraints observations of
Rhodomyrtus psidioides were random rather than following a
rigorous experimental protocol. Nevertheless, the proximity
of several plants at one site (Lorien Wildlife Refuge) allowed
opportunity for frequent casual observation, almost on a
daily basis during seasonal flowering events in that period.
Rhodomyrtus psidioides in currently proposed for listing
in New South Wales as critically endangered (Preliminary
Determination, NSW Scientific Committee 2017) as it is
severely threatened from infection from Austropucciniapsidii
(Myrtle rust) an introduced pathogen first noted in NSW in
2010. Plants are extremely susceptible with all parts of the
plant being affected and populations are threatened with
extinction. Quantitative findings of recent very large declines
in Rhodomyrtus psidioides populations due to Austropuccinia
psidii infection reported in Carnegie etal. (2016) are supported
by field botanists who have encountered the species during
routine botanical surveys and seed collecting over multiple
years (B. Makinson in litt. April 2016).
Rhodomyrtus psidioides is a large shrub or small tree endemic
to Australia, distributed from Gosford on the central coast
of New South Wales to Gympie in southeastern Queensland
(Harden, 1991). Populations flower synchronously but
flowers on individual trees open sequentially, however, not all
regional populations flower each year (Williams 1995). The
flowers are large (1.4cm) and usually clustered, individually
last for 3-7 days, are bisexual but self-incompatible (Adam
& Williams, 2001), fragrant, creamish-white in colour, and
with yellow, slightly sticky pollen that is readily expelled
from dehiscent anthers (Williams, 1995). Little nectar was
evident in the flowers that were microscopically examined
(Williams, 1995). The anthers are brush-like, with the stigma
extending slightly above, the stigmatic surface being broad
and laterally flattened apically (Fig. 1) (Williams, 1995).
Flower morphology partly agrees with the wind-pollinated
(anemophilous) syndrome (Faegri & van der Pijl, 1979)
such that in addition to being adapted for pollination by
biotic vectors the floral structure suggests flowers are also
facultatively wind-pollinated (Williams & Adam, 2010). The
exine sculpture is indistinct — smooth (Williams & Adam,
1999), indicating no special modification for biotic dispersal.
Thus the species is considered cryptically ambophilous,
a previously poorly recognized biotic-abiotic pollination
strategy now considered to be expressed by rainforest
angiosperms more widely (Bullock, 1994, Williams &
Adam, 2010).
Floral visitor observations were undertaken principally
of plants growing on the margin of a subtropical lowland
rainforest at Lorien Wildlife Refuge, approximately
3km north of Lansdowne (31°45'00"S, 152°32'30"E).
Observations were carried out there during November 1990
and November and December 1992 (occasional observations
were continued in subsequent years). At the nearby
Lansdowne Nature Reserve (31°47'30"S, 152°32'30"E), a
small floodplain rainforest remnant, a single day (19 Nov.
1990) of observations was additionally undertaken. At both
sites a small number of mass-netted insect samples were also
collected (Williams 1995). These gave indications of the
nature of visitor assemblages at particular moments, but are
insufficient to allow any statistical analysis.
Rainforest at Lorien Wildlife Refuge and Lansdowne
Nature Reserve represent vegetation communities listed as
endangered ecological communities (respectively ‘Lowland
Rainforest in the NSW North Coast and Sydney Basin
Bioregions’ and ‘Lowland Rainforest on Floodplain in the
New South Wales North Coast Bioregion’) originally under
the NSW Threatened Species Conservation Act 1995\ this
now supplanted by the Biodiversity Conservation Act 2016.
Both formations are also listed as critically endangered under
Federal legislation (see ‘Lowland Rainforest of Subtropical
Australia’, Environmental Protection and Conservation Act
1999 ) because of the extent of past agricultural clearing and
their now limited extent. Conservation has been a major
consideration in their recent management. However, the
unforeseen and widespread entry of the South American
fungal pathogen ‘Myrtle rust’ (Austropuccinia psidii )
(Invasive Species Council, 2011, Makinson, 2018) into the
region around 2010 has resulted in the death of all mature
Rhodomyrtus psidioides trees.
Figure 1: Rhodomyrtus psidioides flower showing extended stigma
and expanded stigma surface.
Cunninghamia 18: 2018
Williams , a Myrtle Rust Extinction Event?
25
Table 1. Insect taxa recorded visiting the flowers of Rhoclomyrtus psiclioides (1990-1992)
(Insects determined to family- Lorien Wildlife Refuge records cited first, Lansdowne Nature Reserve indicated with an asterisk
multiple species given in parentheses; ‘sp./spp. = number of species uncertain).
COLEOPTERA-beetles
Aderidae
Aderus sp.
Boganiidae
Athertonium sp., *Athertonium sp.
Cerambycidae
Sy 11 it us sp.
Chrysomelidae
Crepidodera sp., Ditropoda spp. (2), Monolepta australia, M. Iminuscula, *Crepidodera sp.,
*Monolepta sp.
Cleridae
Scrobiger splendidus
Coccinellidae
Harmonia testudinaria, Rhizobius sp., Scymnus sp.
Corylophidae
Sericoderus spp. (2), * Sericoderus sp.
Curculionidae
Cytallia sydneyensis , undetermined spp. (4), *Cytallia sydneyensis, *Orthorhinus sp.,
* undetermined spp. (2)
Dermestidae
Anthrenus sp.
Elateridae
Megapenthes futilis, Microdesmes collaris, *Drymelater sp., *Megapenthes futilis
Latridiidae
Cortinicara sp.
Melyridae
Helcogaster spp. (2), Neocarphurus ,} august i basis
Mordellidae
Mordella inusitata, Mordella sp., Mordellistena sp., * Mordellistena sp
Nitidulidae
Notobrachypterus sp.
Oedemeridae
llschnomera spp. (2), Pseudolychus spp. (2)
Phalacridae
‘lOlibroporus sp.
Ptilidae
Acrotrichis sp.
Scarabaeidae
Diphucephala Ipygmaea, Phyllotocus scutellaris, *Diphucephala Ipygmaea
DIPTERA-flies
Bombyliidae
Gerou spp. (2)
Calliphoridae
ICalliphora sp., Stomorhina sp.
Dolichopodidae
Amblypsilopus Ibrouleusis, Diaphorus sp.
Drosophilidae
Drosophila spp. (2)
Empididae
undetermined sp.
Lauxaniidae
Melanina sp., Stegauopsis melanogaster
Scatopsidae
undetermined sp.
Tachinidae
undetermined spp.
HEMIPTERA-bugs
?Jassidae
undetermined spp.
Miridae
undetermined sp.
Psyllidae
undetermined sp.
HYMENOPTERA-wasps and ants
Braconidae
undetermined sp./spp.
Encyrtidae
undetermined sp./spp.
Eulophidae
undetermined sp./spp.
Formicidae
ICamponotus sp., *Crematogaster sp.
Pergidae
INeoeurys sp.
Pteromalidae
undetermined sp./spp.
Vespidae
Polistes humilis
HYMENOPTERA/Apiformes-bees
Apidae
Amegilla 1 pule hr a, Apis mellifera, Trigonula carbonaria
Colletidae
Hylaeus lofarrelli, Hylaeus sp., Leioproctus sp., *Heterapoides sp.
PSOCOPTERA-book lice, bark lice
*Caeciliidae
*Caecilius ‘A in eat us
Ectopsocidae
Ectopsocus sp. near meridionalis
THY S ANOPTERA-thrips
Phlaeothripidae
Haplothrips sp., * 1 lap loth rips sp.
Thripidae
Heliothrips haemorroidalis, Thrips setipennis, Thrips sp., *lThrips sp.
26
Cunninghamia 18: 2018
Williams , a Myrtle Rust Extinction Event?
Table 2. Numbers of individual visiting insects and taxa recorded from selected single Rhodomyrtus psidioides tree sampling events
(1990-1992)
Samples collected only during the morning at each site (from Williams, 1995)
No. of individuals
No. of taxa
No. of individuals
No. of taxa
Lorien Wildlife Refuge Nov. 1990
tree 1
tree 2
total Coleoptera
47
24
14
11
total Diptera
36
14
51
19
total all Hymenoptera
4
3
6
6
s/total bees
0
0
0
0
total visitors
107
47
103
45
Total <6mm in size
101
103
Lorien Wildlife Refuge, Nov. 1992
tree 1
total Coleoptera
38
13
total Diptera
23
17
total all Elymenoptera
9
9
s/total bees
0
0
total visitors
87
total <6mm
86
Lansdowne Nature Reserve, Nov. 1990
tree 1
total Coleoptera
467
12
total Diptera
13
7
total all Elymenoptera
2
2
s/total bees
1
1
total visitors
534
32
total <6mm in size
532
Results and Discussion
Sampling and observation results show insect taxa recorded
visiting the flowers of Rhodomyrtus psidioides in the 1990-
1992 period (Table 1) and numbers of individual visiting
insects and taxa recorded from selected single Rhodomyrtus
psidioides tree sampling events (1990-1992) (Table 2). No
vertebrates were seen visiting flowers. Table 1 underestimates
the numbers of insect species owing to difficulties in
identifying to family groups such as small Diptera and
microhymenoptera. Insect visitors were predominantly
(99%) in the <6mm size class (Table 2, Williams, 1995).
Only Amegilla Ipulchra and the introduced ‘honey bee’
Apis mellifera constituted notable size exceptions. Although
individual temporally-discrete sampling events can result in
seemingly large numbers of individuals and taxa (Table 2),
in general over the period of the study, insects were often few
in number at any one time of observation; no taxon exhibited
mass attraction responses to open blossoms, most individual
blossoms were devoid of insects when observations were
made, and even the otherwise ubiquitous Apis mellifera was
usually absent. This seemed counter-intuitive given that
blossoms were massed, conspicuous and fragrant, but might
be explained by the small quantity of nectar that individual
flowers seemed to offer. No specific visitation patterns were
observed. Rather, insects appeared to recruit randomly to
flowers throughout each day.
The single day of observations, and the single netted sample
collected at Lansdowne Nature Reserve (Table 2), is too
small to establish an understanding of the possible full suite
of visitors there, however, observations indicated that the
assemblage likely mirrors that recorded at Lorien Wildlife
Refuge; this being a mixture of ‘incidental’ visitors (e.g.,
Dolichopodidae, Psyllidae, microhymenoptera) and potential
pollinators dominated by small species, ecologically
unspecialised for pollination, that are commonly encountered
on a range of mass-flowering rainforest trees and large shrubs
elsewhere in the region (Williams, 1995, G. Williams unpubl.
records). The few species with specialised morphological
adaptations to a floricolous habit were represented by
Mordellidae, the scarab Phyllotocus scutellaris, apid bees,
and the bombyliid fly genus Geron. All visitors, regardless
of their degree of adaptation to feeding upon floral resources,
have the potential to transport pollen loads. In the case of
thrips and other minute insects, only single or small numbers
of grains are anticipated to be transported, and movements
are largely restricted to adjacent flowers and plants; their
contribution to out-crossing thus being individually small,
though cumulatively over time potentially significant.
Only Amegilla Ipulchra and Apis mellifera undertake
relatively frequent or long distance foraging movements
between dispersed individual plants, with the former known
to regularly exhibit ‘trap-lining’ foraging strategies (see
Williams & Adam, 2010, Willmer, 2011).
Myrtle rust has spread globally and was first detected in
New South Wales in April 2010. Its infection results in
crown dieback, branch death and mortality of sensitive
members of the Myrtaceae (Invasive Species Council, 2011,
Carnegie et al., 2016, Pegg, 2017, Makinson, 2018). Myrtle
rust has the potential to fundamentally alter the ecology
Cunninghamia 18: 2018
Williams , a Myrtle Rust Extinction Event?
27
of Australia’s vegetation communities. At Lorien Wildlife
Refuge Rhodomyrtus psidioides and the related and highly
sensitive Rhodamnia rubescens (Benth.) Miq. have been
severely attacked but species of the related Myrtaceae genera
Archirhodomyrtus, Corymbia, Eucalyptus, Lophostemon,
Syncarpia, Syzygium and Tristaniopsis have been seemingly
unaffected. All these genera recruit a taxonomically wide
assemblage of putative pollinators, most of which are small
in size, that frequent a diversity of mass-flowering shrubs and
trees with open insect-adapted floral structures (G. Williams
pers. obs.).
Although Rhodomyrtus psidioides is able to resprout from
root stock, vegetative regrowth over the ensuing years is
constantly re-infected, causing gross leaf deformity and tip
mortality (Fig. 2). Consequently, sucker growth, persistent
as it has been, has not been able to successfully progress
to reproductive maturity. This scenario is exhibited by
numerous other populations in the region, for example in
littoral rainforest at Harrington (G. Williams pers. obs.)
and ornamental plantings at Diamond Beach (T. Wright
pers. comm.). Although none of the insect visitors recorded
constitute species with known obligate plant dependencies,
Rhodomyrtus psidioides flower - putative pollinator
interactions are extinct, at least locally. But should the species
be able later to re-establish viable populations, a resident
assemblage of pollinators would be present to reconstitute
its pollination suite.
Acknowledgements
Paul Adam (University of New South Wales) offered
numerous insights over the years into the nature and
conservation of Australia’s rainforests, Dan Bickel, David
McAlpine, Courtenay Smithers (Australian Museum),
Laurence Mound and Ian Naumann (CSIRO) kindly assisted
with the identification of Diptera, Psocoptera, Thysanoptera
and microhymenoptera. Kim McKay (Director and CEO),
Rebecca Johnson (Director, AMRI) and Derek Smith
(Collection Manager) are thanked for facilitating access to
the resources of the Australian Museum. Dan Bickel is also
thanked for comments on an original draft of the manuscript
and Terry Wright (Diamond Beach) kindly provided
observations on a population he had planted near Taree.
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suspected? Biotropica, 31, 520-524.
Williams, G. & Adam, P. (2010) The Flowering of Australia’s
Rainforests. CSIRO Publishing, Collingwood.
Willmer, P. (2011). Pollination and Floral Ecology. Princeton
University Press, Princeton.
Manuscript accepted 26 July 2018
Figure 2: Root sucker tip mortality on Rhodomyrtus psidioides
resulting from Myrtle rust attack.
Date of Publication:
October 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
A new classification of savanna plant communities on the igneous rock
lowlands and Tertiary sandy plain landscapes of Cape York Peninsula
bioregion
Eda Addicott’ 1 ' 2,3 , Mark Newton 1,2 , Susan Laurance 3 , John Neldner 1 , Melinda Laidlaw 1
and Don Butler 1
'Queensland Herbarium, Mt. Coot-tha Road, Toowong, Department of Environment & Science,
Queensland Government, QLD 4066, AUSTRALIA
Australian Tropical Herbarium, James Cook University, Cairns, QLD 4870, AUSTRALIA
3 Centre for Tropical Environmental & Sustainability Science (TESS) and College of Science & Engineering,
James Cook University, RO. Box 6811, Cairns, QLD 4870, AUSTRALIA
"corresponding author, eda.addicott@des.qld.gov.au,
Abstract : Classifying and mapping landscapes are tools to simplify complex systems into the discreet subsets widely
used in landscape management. In 1999, the Queensland Government adopted a Regional Ecosystems approach
as a state-wide landscape classification scheme. Lor the Cape York Peninsula bioregion in north-eastern Australia,
Regional Ecosystems (RE) were initially recognised based on a pre-existing vegetation map and classification for the
bioregion. The classification had been developed using expert-techniques based on extensive field plot data. Here,
we use numerical analyses to classify the field plot data and identify savanna plant communities associated with two
widespread landform groups in the bioregion (the old loamy and sandy plains (land zone 5) and the hills and lowlands
on igneous rocks (land zone 12). Communities were identified at the plant association level, using species importance
values calculated from foliage cover and vegetation height at each plot. We developed a descriptive-framework
for each community using statistically based characterising species and biophysical attributes. We recognise 57
communities compared with 110 that had been previously identified using expert-techniques. This classification is
used to recommend refined Regional Ecosystems under the government’s regulations. The descriptive-framework
supported consistent descriptions of communities and assignment of new sites to the classification. We conclude that
incorporating quantitative methods in classifying and describing plant communities will improve the robustness and
defensibility of Regional Ecosystems and their use in landscape management across Queensland.
Cunninghamia ( 2018 ) 18 : 029-072
doi: 10.7751 / cunninghamia.2018.18.003
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
30
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
Introduction
Vegetation classification is a globally used tool for land
management and for investigating ecological diversity at
multiple scales. Consistent vegetation classification schemes
which cross geographical and administrative boundaries are
therefore highly desirable (ESCAVI 2003; Rodwell 2006;
Jennings et al. 2009; De Caceres et al. 2015). Recognising
this, the Queensland government adopted the Regional
Ecosystem (RE) approach as a state-wide classification
scheme in 1999. This is a triple-tiered hierarchy with the
first division being based on the Interim Biogeographical
Regions of Australia (Thackway & Cresswell 1995). The
second division of the hierarchy is ‘land zone’; a concept
that involves broad geological divisions with consideration
of geomorphological processes and soils (Wilson & Taylor
2012). Examples of land zones include ‘alluvial river and
creek flats’, ‘coastal dunes’ or ‘hills and lowlands on granitic
rocks’. The third level of the classification scheme is termed
‘vegetation community’ and is a plant community, recognised
at the plant association level (Figure 1). A Regional
Ecosystem is therefore defined as “ a vegetation community,
or communities, in a bioregion that are consistently associated
with a particular combination of geology, landform and soil ”
(Sattler & Williams 1999). REs can therefore contain one
or more vegetation communities. REs are mappable, with
a distinctive signature recognisable from remote sensing
imagery at the landscape scale of 1:100,000. REs are revised
and updated when new data is supplied. To this end, each
bioregion has a technical committee to review and implement
proposed changes based on appropriate data. This technical
review committee performs the same function as similar
panels in other Australian and international jurisdictions
(EVSWG 2017; OEH 2018; USNVC 2018).
Figure 1: Regional Ecosystem classification scheme. Regional Ecosystems are a triple-tiered hierarchy. The first tier is biogeographical
regions based on the Interim Biogeographical Regions of Australia. The second tier is broad geological / geomorphological groups (labelled
land zones). The third tier are plant communities recognised at the association level (labelled vegetation communities)
For Cape York Peninsula (the Peninsula) a vegetation map
and qualitatively-based classification at the plant association
level was developed as part of the Cape York Peninsula
Land Use Study (CYPLUS) carried out in the early 1990s
(Neldner & Clarkson 1995). With the adoption of the RE
framework, the CYPLUS vegetation classification was
converted to a Regional Ecosystem classification using
qualitative methods. The vegetation map was also revised in
the context of a state-wide RE mapping program at a scale of
1:100,000, an exercise that ultimately necessitated a revision
of the RE classification of the Peninsula.
A best-practice framework for vegetation classification is
centred around standardised methods of data collection and
classification techniques (De Caceres etal. 2015). Following
this best-practice, the RE classification framework has
accompanying documentation describing a standardised
survey and mapping methodology for Queensland (the
methodology) (Neldner et al. 2017c). It outlines a consistent
set of classification protocols for defining vegetation
communities which align with both the Beadle (1981)
definition of a plant association and the necessary emphasis
on canopy species used in classifications for vegetation
mapping. These protocols identify the pre-dominant
layer within a vegetation’s structure as that contributing
most to the above-ground biomass (Neldner et al. 2017c).
Communities are then defined using the height, cover and
dominant species in the pre-dominant layer, with sub¬
ordinate consideration given to associated species in other
layers (Neldner et al. 2017c). Plant associations are thus
defined as a community where the pre-dominant layer has
a uniform floristic composition and exhibits a uniform
structure. This forms the basis for mapping and survey
projects at all scales across the state and is embedded in
legislation. Currently however, implementation of these
classification protocols relies on qualitative techniques and
subjective sorting of plot data into similar groups. The use of
Cunninghamia 18: 2018
Addicott et al., New plant community classification, Cape York, Queensland
31
qualitative techniques is widespread and common in remote
areas with limited researchers such as in Queensland, but they
have acknowledged problems based primarily around their
lack of transparency, repeatability and consistency between
researchers (Mucina 1997; Kent 2012; Oliver et al. 2012). A
good outcome from such processes is heavily dependent on
a researcher’s knowledge of the vegetation of the area and
the biases introduced by their assumptions of the ecological
and biophysical processes important to landscape function
and biodiversity. Consequently, qualitative methods do not
produce communities which are statistically comparable
(Harris & Kitchener 2005; Kent 2012; Oliver et al. 2012).
Using quantitative techniques in the classification process
can help to overcome some of these problems allowing
consistent, statistical information to be produced about
community composition and structure.
A classification scheme has widest applicability if it can
perform two major tasks: firstly, determine communities
with transparent and repeatable techniques, and secondly
provide consistent and reliable assignment of new sites
to the classification scheme (De Caceres & Wiser 2012).
The aim of this study is to address these requirements by
incorporating quantitative analyses into the classification
of vegetation communities within the RE framework.
Specifically, we aim to classify the savanna communities
of two land zones on the Peninsula at the association level,
assess the adequacy of the preferential sampling design used
and develop a descriptive-framework which incorporates
statistically derived characterising species for assigning new
site data into these communities. We use this framework to
describe REs suitable as distinct vegetation mapping units.
Methods
The Cape York Study area
Cape York Peninsula bioregion covers 120,000 km 2 in
the monsoon tropics of north-eastern Australia and lies
between 10 and 16 degrees south (Figure 2). Elevations
range from sea level to approximately 800 m. The annual
average rainfall varies between 1000-2000 mm with 80%
falling in the wet season between December and March
(Horn 1995). Temperatures range from an average annual
monthly minimum of 14 °C in winter (July) to an average
monthly maximum of 35 °C in summer (December) (BoM.
2016). Our study encompasses the savanna communities
on two of the ten land zones on the Peninsula (Neldner
1999); the old loamy and sandy plains (land zone 5) and the
hills and lowlands on igneous rocks (land zone 12). These
communities on land zone 5 cover 45,000 km 2 (40% of
the bioregion) and on land zone 12, 6,500 km 2 (5% of the
bioregion). Land zone 5 is distributed across the full extent
of the bioregion while land zone 12 occurs primarily along a
north-south spine associated with the Great Dividing Range
(Figure 2).
Figure 2: Distribution of the two land zones on Cape York Peninsula
classified in this study.
Data Collation
During the mapping process two major types of vegetation
data were collected; observational sites and vegetation
plot data. These were sampled from 1990 to 2015,
with the majority between 1992 and 1996 as part of the
original mapping project (Neldner & Clarkson 1995). The
observational sites were collected in large numbers as rapid
records made during field traverses of the mapping area.
They include records of geolocation, dominant species in
the pre-dominant layer and vegetation structure. The survey
design for locating vegetation plot data was preferential, with
locations chosen based on either interpreted photo-patterns
from air photos and ease of access, or on plant assemblages
identified during the collection of observational sites.
Observational site data were extracted from GIS coverages
associated with the mapping project and vegetation plot data
from the Queensland government ‘CORVEG’ database. The
latter were categorised as either ‘detailed’ plots, containing
data appropriate for use in determining the vegetation
classification, or ‘non-detailed’ plots, containing incomplete
data or data collected using different methods. Detailed plots
contained data on percent foliage projected cover (%FPC)
for each species in each woody vegetation layer recorded
32
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
along a 50m transect using the line intercept method (Neldner
et al. 2017c). The average height of each layer was also
recorded. The ground layer had species abundance recorded
as an estimate of %FPC in 1 m 2 quadrats at 10 m intervals
along the 50 m transect (five quadrats in total) and averaged.
There were 192 detailed, 38 non-detailed plots and 4,670
observational sites on land zone 5 and 96 detailed, 45 non-
detailed sites and 1,424 observational sites on land zone 12.
Vegetation communities in which the pre-dominant canopy
was the ground layer we refer to collectively as grasslands,
but this group includes sedgelands and rock pavements with
scattered herbs and forbs as well as true grasslands (Neldner
et al. 2017c).
It was necessary for our quantitative analysis to accord with
the classification protocols and principles outlined in the
methodology as they are embedded in all current regional
ecosystem mapping relied upon for regulatory purposes. To this
end, previous research identified transformations to detailed
plot data suited to incorporating vegetation structure into the
classification of plant associations (Addicott et al. 2018). As
a result, %FPC for each species was multiplied by the height
of the layer in which it occurred prior to data analysis. This
generated a species importance value for every species in each
plot. The protocols also specify using dominant species and
removing species which have low occurrence or abundance
(here termed ‘sparse’) from a dataset is recommended in
general scientific practice when exploring ecological patterns
in data. Sparse species may mask relationships of interest at
the relevant scale, their occurrence and identification may
be dependent on survey design or their distribution may be
spatially and temporally inconsistent (Grime 1998; McCune &
Grace 2002; Kent 2012). To standardise the removal of sparse
species across plots we excluded those whose contribution to
total foliage cover was never >1%. For example, if a species
contributed <1% in one plot and >1% in others it was retained
in all plots in which it occurred. A species whose contribution
to total foliage cover was <1% in every plot was excluded.
This resulted in removing 175 taxa from the land zone 12 and
254 taxa from the land zone 5 analyses. Standardising the
removal of sparse species in this way provides a consistent
method across the dataset which does not delete infrequent
but dominant taxa (Field et al. 1982), eliminates most of the
unidentified taxa in a site without affecting the identification
of communities at the association level (Pos et al. 2014)
and improves the recognition of communities identifiable
at landscape mapping scales (Addicott et al. 2018). We
acknowledge that sparse species and unidentified taxa may
be new, rare and/or endangered species and hence of high
conservation significance. However, for this study they are
not critical to vegetation classification at landscape scales
(Addicott et al 2018). Non-native species were also excluded,
and any remaining taxa not reliably identified to species were
amalgamated to genus level. The ground layer at wooded plots
was excluded to identify communities suitable as mapping
units (Neldner & Howitt 1991; Archibald & Scholes 2007;
Mucina & Daniel 2013). In the final analyses there were
351 species with 241 occurring on land zone 5, 258 on land
zone 12, and 148 shared between the two land zones. Plant
nomenclature follows Bostock & Holland (2015).
Identifying plant communities
Plots were allocated to either land zone 5 or 12 based on
field observations as well as geology, regolith and / or soils
mapping available at plot each location. We analysed the
detailed plots in each land zone to look for groups of co¬
occurring plant species using agglomerative hierarchical
clustering and the software package PRIMER-E v6 (Clarke
& Gorley 2006). We produced a similarity matrix (square-
root transformation, Bray-Curtis coefficient) and ran the
CLUSTER routine, using unweighted pair group mean
averaging, to form clusters. To choose the level of cluster
division for identifying plant communities we used a
combination of three evaluation methods: 1) the SIMPROF
routine which determines clusters significantly different to
each other (Field et al. 1982), 2) Indicator Species Analysis
(Dufrene & Legendre 1997) (in the Tabdsv’ R package
(Roberts 2013)) which determines clusters maximising
species occurrence and 3) generalised linear models (GLM)
in a mulitvariate framework (Lyons et al. 2016) (available
in the ‘optimus’ R package (Lyons 2018) to estimate
the relative performance of differing cluster divisions in
predicting species foliage cover. This last method uses
GLMs and Akaike’s Information Criteria (AIC). AIC is
summed across individual species, and the final sum-of-AIC
score is used as a measure of how well the cluster division
predicts species cover. A lower sum-of-AIC score indicates
a better prediction. In situations where the three evaluators
produced differing results, we formed a subset of plots and
tested cluster divisions within the subset.
Assigning plant communities to the Regional Ecosystem
framework
Our final plant communities were evaluated by the technical
review committee for regional ecosystems of the Cape
York Peninsula bioregion whose role was to evaluate and
give effect to proposals to modify Regional Ecosystems
classifications. During this process the committee assigned
plant communities to regional ecosystems based on expert-
judgement of non-floristic variables as outlined by the
methodology (Neldner et al 2017c), potentially producing
REs containing communities with different dominant species
and low floristic similarity to each other. For example,
communities which did not have predictable or mappable
occurrences or were <100 ha in total area of distribution
were grouped with those on closely associated landforms
and similar ecological niches. Communities recognised
as successional temporal variants, or condition states,
of a climax association were also grouped into one RE.
Where the committee requested more evidence to support
proposed changes, we used the classification protocols as
a guide for conducting further analyses. Consequently, we
tested for floristic differences between sites on different
geomorphological areas and soil types (using the ANOSIM
routine), for differences in canopy height (using an unpaired
t-test) and investigated whether differences in the ground
layers of sites were coincident with geomorphological
areas or soil divisions (using wMDS ordination and GIS
overlay). One additional role of the committee was to
Cunninghamia 18: 2018
Addicott et al., New plant community classification, Cape York, Queensland
33
identify communities not represented in the analyses but
recognisable from aerial photo interpretation, non-detailed
plot data and observational sites. There were therefore two
types of communities in the final classification scheme;
those identified through quantitative analysis and those
identified by expert-techniques. The latter communities
will be reviewed when further detailed sampling data and
quantitative analyses are available.
Creating community descriptions and assigning new sites
An important aspect of a vegetation classification scheme is to
allow description and identification of its plant communities
(De Caceres & Wiser 2012). To this end we compiled a
descriptive-framework based on characterising species,
vegetation structure and landform, including geographical
distribution when it aided identification. Characterising
species were those used to describe the floristic and
structural composition of a community (De Caceres et al.
2015) and were identified for the quantitatively defined
communities using each species’ frequency, average cover
and strength of association with a community. To determine
the strength of each species’ association with a community,
we calculated a phi-coefficient of association (Chytry
et al. 2002) based on cover, using the JUICE software
package (Tichy 2002). Each group was standardised to
equal number of plots. A phi-coefficient of 100 means
a species occurs only in that community, while values
approaching zero indicate the species is equally abundant
in several communities. The phi-coefficient values were
also used to identify species with a significant association
to a community using Fisher’s exact test (p<0.05) (Chytry
& Tichy 2003). We listed species frequency and average
cover using the technical-description routine within the
CORVEG database, which also allowed identification of
vegetation structure. We defined characterising species as
those with a phi-coefficient of association >6 or occurring in
>70% of sites. A phi-coefficient of >6 was chosen to ensure
a minimum of one statistically associated species with each
community. Landform and additional vegetation structure
information was taken from plot sheets and observational
data where available. Geographical distribution came from
the final mapping. Where communities were represented by
fewer than three sites in analyses we used non-detailed or
observational sites for additional information. To describe
qualitatively determined communities we used species,
structure and landform information from non-detailed plots
and observational sites, and, where it was diagnostic, mapped
distribution. These community descriptions are necessarily
less robust but allow indicative recognition in the held.
The ease and certainty with which new sites can be reliably
allocated into a classification scheme outside of an analysis
process is important (De Caceres & Wiser 2012) and we
expected our descriptive-framework to enable this. To test
this, we used the ‘non-detailed’ plots previously excluded
from analysis as ‘new’ sites. We matched the information
available from each plot to that in the descriptive-framework,
subjectively assigning it to a vegetation community and
rating its level of ht-to-community as high or low. These
non-detailed sites had a variety of vegetation information
available ranging from a community label with or without
a limited species list (and sometimes growth form) to
complete species lists with alternative abundance measure
such as classes, stem density or basal area and an indication
of which layer species occurred in. In sites which had only
a label (or label and a species list) we took the label as an
indicator of dominance and structure. We also used landform
infonnation where it was provided on the site pro-fonna.
Along with defining a classification via consistent analytical
techniques, labelling communities using consistent naming
conventions is important (De Caceres & Wiser 2012).
Neldner et al. (2017c) outlines these for the RE framework.
In this, a limited number of characterising species are
listed in order of dominance, with punctuation to indicate
relative abundance and frequency, followed by the structural
formation. Associated habitat characteristics, such as
landform or soil descriptors are included in labels where they
are diagnostic. We followed these conventions to develop
community labels.
Assessing sampling adequacy
Knowledge of bias in a sampling design allows an
understanding of the strength and weaknesses of results.
We reduced bias by using plots with standardised plot size,
collection methods, data attributes, data quality and season
of survey. The standard plot size of 500 m 2 has been shown
to adequately capture the species diversity at the plot level
in savanna and woodland communities (Neldner & Butler
2008). Data collection methods follow the standard survey
methods outlined in Neldner et al. (2017c). Seasonality is
an issue in the ground layer as many species occur only in
the wet and early dry season. In sites dominated by woody
vegetation, excluding the ground layer removed this potential
bias. Plots dominated by the ground layer were surveyed
between May and August (the early dry season). Despite
standardising these aspects of survey design however, we
expected some bias due to preferential rather than random
selection of plot locations. Therefore, we assessed how
well the field sampling captures firstly the environmental
variability across the landscape, and secondly the community
and species richness.
To test how well the environmental variability was sampled,
we followed the convention of testing those variables
expected to limit plant species growth, dividing them into
climate and soil themes. We used four climate variables,
two temperature variables (average annual temperature,
and the coefficient of variation of temperature seasonality)
and two rainfall variables (annual average rainfall, and the
mean moisture index of the lowest quarter), available as
ANUCLIM datasets (Xu & Hutchinson 2013). The soils
variables were grouped in to soil nutrients (organic carbon
content, and phosphorus) and soil structure (available plant
water capacity, permeability, drainage, and slope) (Lyons
et al. 2017; Neldner et al. 2017a). All soil datasets came
from Australian Soil and Resource Information System
(McKenzie et al. 2012; ASRIS 2014), with the slope
derived from the digital elevation model for the Peninsula
34
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
(GeoScience Australia et al. 2009). In addition to these
climate and soil variables, we assessed how well the
survey sampled variation in vegetation structure by using
a maximum persistent greenness GIS coverage (JRSRP
2017). This coverage is derived from LANDSAT imagery
classification and, on the Peninsula, equates to density of
woody vegetation layers, with a higher greenness index
indicating denser woody vegetation. While density of
woody vegetation is significantly correlated with the climate
variables (r = 0.6, p <0.0001), the R 2 value of the 4-way
multiple regression is 34% indicating the predictability
of density of woody vegetation using these variables
is relatively low (Appendix 6, Figures A6.1-A6.4). We
are therefore confident that assessing woody vegetation
density will provide useful additional information on bias
in sampling of vegetation structure. All these datasets
were accessed as raster coverages. Using the DOMAIN
software program (Carpenter et al. 1993) we calculated the
similarity of the environmental envelope at any grid point
to that at any plot or observational site. DOMAIN uses the
input variables to create an environmental envelope for
each grid cell and then calculates the similarity between
each grid cell and any site in a Euclidean p-dimensional
space using the Gower metric. The similarity is bounded
in one direction, with values close to 100% for maximum
congruence, and can be displayed spatially. Because
observational sites assist in identifying the assemblages
recorded in the detailed plots we investigated the amount
of environmental variability captured by both types of data.
To assess the community and species richness surveyed by
the detailed plots we estimated total population richness,
and calculated the proportion captured by our sampling.
To assess the species richness surveyed we used the full
species dataset (with weeds removed), as our classification
analyses used only a subset of species surveyed. To
estimate population richness from our samples we derived
1,000 model-populations using bootstrap techniques.
We then calculated an unbiased population estimate of
richness by 1) estimating the bias, by subtracting the
sample richness from the mean richness of the model-
populations, and 2) subtracting this bias from the sample
richness. Using the bootstrap model-populations we also
defined 90% confidence intervals (using the 0.05 and 0.95
quantiles around the mean of the 1,000 model-populations).
All calculations were done in the R environment
(R Development Core Team 2014) using the ‘bootstrap’
package (Efron & Tibshirani 1993).
Results
Assessing sampling adequacy-Environmental variability
The survey design comprehensively sampled the full
environmental variability in each land zone. Between 99 and
100% of the total area of each land zone was >90%-similar to
any observational site for all variables. Results were similar
for detailed plots for climate, vegetation structure and soil
nutrient variables. Between 99 and 100% of the total area of
land zone 5 and 98% of land zone 12 was >90%-similar to
any detailed plot (in the respective land zone). These results
were slightly lower for soil structure, with 98.6% of land
zone 5 and 95% of land zone 12 >90%-similar to any site.
Appendices 3a and 3b have detailed tables and indicative
maps of areas of lower similarity to sites. The detailed GIS
coverages of these areas are available from the senior author
if more detail is required.
Community Richness
We found the survey design reliably sampled the community
richness of land zone 5 but not that of land zone 12. On land
zone 5 it captured 95% of the estimated total community
richness. Nineteen of an estimated 20 communities were
sampled in detailed sites, within the 90% Cl (19 - 21). On
land zone 12 the survey captured 89% of the community
richness (24 of an estimated 27 communities), outside the
90% Cl of 25 - 29 (Table 1).
Species Richness
The survey did not reliably capture the full species richness
on either land zone, with the number of species sampled
lying outside of the 90% CIs (Table 1). There were 775
species sampled on land zone 5 and 673 on land zone 12,
representing 86% of the estimated species richness on either
land zone (Table 1).
Table 1: Sampled and expected community and species richness.
The expected number of communities and species and the 90%
confidence intervals (Cl) are calculated from bias corrected
estimates of 1,000 bootstrap model-populations.
Number
sampled
Number
expected
90% Cl
Community richness
Land zone 5
19
20
19-21
Land zone 12
24
27
25-29
Species richness
Land zone 5
775
904
889 - 920
Land zone 12
673
785
771 - 798
Plant Communities
There were 57 communities in our study’s final classification,
27 on land zone 5 and 30 on land zone 12. Seventy-five
percent of these were identified by quantitative methods and
25% by qualitative techniques and less detailed plot data
(Table 2). Two communities were recognised after additional
analyses requested by the technical review committee
(Appendix 4). Incorporating quantitative analysis resulted in
fewer communities on both land zones than the expert-based
classification with an overall reduction of 49%. Individually,
the reduction was higher on land zone 5 (54%) than land
zone 12 (42%), driven by the larger decrease in the number
of woodlands and shrublands identified (Table 2). Whilst
most of the final REs consisted of one plant community,
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
35
in 11 instances, the review committee assigned several
communities to individual REs. The 27 communities on land
zone 5 were assigned to 21 REs, and the 30 on land zone
12 to 23 forming some REs with more than one community
(appendix 2). Because the detailed descriptions, conservation
status and ecological notes for individual REs and their
communities are available on-line we have not included it
in this manuscript (http://www.qld.gov.au/environment/
plants-animals/plants/ecosystems). However, to portray the
communities and REs recognised, we have included the
short label descriptions, mapped areas and notes for the
REs in Appendix 2. To illustrate the floristic relationships
between the communities and REs on each land zone we
formed community dendrograms and ordination plots from
the detailed plot data (Appendix 5).
Table 2: The number of communities in each formation on each
land zone. The quantitative analysis resulted in a reduction in
the number of vegetation communities, ‘a priori’ classification
= vegetation communities in the pre-existing, qualitatively
derived, classification.
Grasslands Shrublands Woodlands
Land zone 5 (45,000 km 2 )
806 ha
1,904 km 2
46,089
km 2
Quantitatively derived
1
1
17
Qualitatively derived
1
1
6
Total after review (no. of REs)
2(1)
2(2)
23 (18)
a priori classification
4
7
48
Land zone 12 (5,500 km 2 )
154 km 2
110 km 2
5,236 km 2
Quantitatively derived
5
3
16
Qualitatively derived
1
1
4
Total after review (no. of REs)
6(5)
4(4)
20 (14)
a priori classification
7
6
38
WOODLANDS
Bamaga •
Figure 3: Distribution of the vegetation formations across Cape York Peninsula bioregion included in this study.
Summary ofplant communities and formations of land zone
5 (old loamy and sandy plains)
Grasslands are of limited extent on land zone 5 (0.01% of the
land zone) and contain two communities. One occurs only
on islands in the Torres Strait and the other in southern Cape
York Peninsula (Figure 3). Shrubland communities cover
4% of the land zone (Figure 3), the most extensive of which
(1,900 km 2 ) occur on the deep sand plains in the north-east
and east of the bioregion. The second occurs only on the
Torres Strait islands. Woodlands dominate land zone 5 (95%
of the area) (Figure 3) and can be broadly categorised into
four groups; 1) Eucalyptus tetrodonta dominated woodlands,
2) other Eucalypt and Corymbia dominated woodlands, 3)
Melaleuca dominated woodlands, and 4) Asteromyrtus
dominated woodlands. The Eucalyptus tetrodonta woodlands
dominate the landscape, covering 42,870 km 2 . Melaleuca
dominated woodlands cover the next largest area of 2,825
km 2 , the Asteromyrtus dominated woodlands 1,044 km 2
and Eucalypts and Corymbia species other than Eucalyptus
tetrodonta cover the smallest area (528 km 2 ).
Summary ofplant communities and formations of land zone
12 (hills and lowlands on granitic rocks)
Grasslands are again of limited extent on land zone 12,
covering 2% (Figure 3). The most widespread of these was
the rock pavements with scattered herbs and forbs associated
with the tops of the major mountain chains on the mainland
and the Torres Strait islands (66 Ion 2 ). The remaining five are
36
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
all dominated by Poaceae species. Shrublands cover 12% of
the land zone (Figure 3), with three of the four communities
dominated by Melaleuca species. The fourth, covering the
largest area (57 km 2 ), is dominated by an endemic species,
Leptospermum purpurascens. Despite having the largest area,
its range is restricted to the hills and mountains associated with
Iron Range in the centre of the bioregion. Woodlands are again
the most widespread fonnation (75% of land zone) (Figure
3). These are dominated by Eucalyptus tetrodonta woodlands
(41% of woodland area) and ironbark woodlands (Eucalyptus
cullenii and Eucalyptus crebra) (28%). Other Eucalypt and
Corymbia dominated woodlands cover 21%. Melaleuca
woodlands cover 3% of the land zone, a much smaller area
than on land zone 5. The remaining area is covered by one
mixed species low woodland and two Acacia communities
(both of which occur only in the Torres Strait islands).
Assigning new sites into the classification scheme
Figure 4: RE 3.5.19 Asteromyrtus lysicephala and Neofabricia
myrtifolia open heath to shrubland on sand sheets.
Using the descriptive-framework ( Appendix 1) we were able
to incorporate all 83 non-detailed sites into the classification
scheme. The characterising species provide the most useful
information; strength of association allowed us to rank
characterising species in importance for a community. The
species information in the non-detailed sites could then be
matched to this, even when not all characterising species
were recorded at a site. While the characterising species was
the most useful individual piece of information, the most
powerful tool for assigning sites in to the classification was
the combination of characterising species plus vegetation
structure information. Landform became diagnostic where
the characterising species overlapped (particularly the
Eucalyptus tetrodonta woodlands). We could assign 66% of
sites (55) with a high level-of-fit to community. These were
the sites that contained quantitative abundance and structure
data collected using different methods. The sites assigned
with a low level-of-fit to community were those with only a
community label to indicate abundance and structure.
Inclusion of results in mapping
Vegetation mapping and classification are two separate
processes often accompanying each other (Franklin 2013).
In this survey the process was iterative, with the mapping
(and accessibility) driving the choice of transects, and the
outcomes feeding back to change the qualitative classification
depicted in the mapping. Continuing this process, the
results of our classification analyses were used to revise
the Regional Ecosystem mapping to reflect the updated
vegetation communities and REs. As part of the mapping,
individual mapped areas (i.e. polygons) are also assigned
levels of reliability for attributes and locational accuracy.
Polygons which contained detailed plots were given a high
reliability in the mapping, as were areas containing non-
detailed plots assigned in to the classification with a high fit-
to-community. Polygons containing non-detailed plots with
low fit-to-community were mapped with a low reliability
and identified as requiring further survey.
Figure 5: RE3.5.36a Eucalyptus tetrodonta and Corymbia
nesophila woodland on undulating plains.
Figure 6: RE 3.5.41b Melaleuca viridiflora low open woodland
+/- Petalostigma banksii on plains.
Cunninghamia 18: 2018
Addicott et al., New plant community classification, Cape York, Queensland
37
Figure 7: RE 3.12.10a Eucalyptus cullenii +/- Corymbia
clarksoniana woodland on granite hills and footslopes.
Figure 8: RE 3.12.28 Leptospermum purpurascens tall shrubland
on igneous hills.
Figure 9: RE 3.12.48a Heteropogon triticeus dominated grasslands
on igneous headlands and offshore islands.
Discussion
We present, for the first time, a bioregional scale classification
of vegetation communities, within the Regional Ecosystem
framework, incorporating quantitative analyses. After
initial assignment of sites to land zones, we allocated sites
to communities using 1) numerical classification based on
floristic attributes, and 2) statistical analysis of vegetation
structure and environmental factors. These communities
were incorporated into the RE framework by an expert
panel peer-review process. We developed a descriptive-
framework to characterise the vegetation communities (using
statistically derived floristic attributes and non-statistically
derived abiotic variables), and used this to assign new sites
to the classification. In so doing we addressed the two main
tasks of a classification scheme (as outlined by De Caceres
& Wiser -2012) - to determine vegetation communities
using transparent and repeatable techniques, and to provide
a framework for consistent and reliable assignment of new
sites into the classification scheme.
While our classification incorporates as much quantitative
analysis as available data allows, 25% of communities were
still identified using expert-based techniques. This was
done using plots with different data collection methods, or
observational data from helicopter flights over inaccessible
areas of the bioregion, meaning the data could not be used
in the analyses. Communities identified by expert-based
techniques therefore represent ‘known unknowns’ and
provide a targeted direction for future data collection.
A notable outcome of the quantitative analysis was the
49% reduction in the number of communities recognised,
compared to the expert-driven process. Quantitative analysis
allows experts to test their interpretation of the factors
influencing landscape function; in this case, unquantified
floristic and biophysical attributes. One question our analysis
asks is, ‘Does the floristic composition of the landscape reflect
the divisions chosen by experts, based on their assumptions
about the importance of these attributes?’ The 49% reduction
suggests that, in this case, it does not. Quantifying the
differences between the expert and quantitatively derived
communities is beyond the scope of this paper, but is the
focus of ongoing work. However, one function of quantitative
analysis is to help gain consensus among experts about the
species driving vegetation community differences.
Preferential-sampling designs are biased in several ways
compared to stratified random-sampling designs (Diekmann
et al. 2007; Hedl 2007; Michalcova et al. 2011). It is well
recognised that the statistical power of preferential-sampling
designs is lower (Lajer 2007), but much of the aim of
vegetation survey and mapping is to distinguish patterns using
descriptive procedures rather than to produce inferential results
from null hypothesis significance testing (De Caceres et al.
2015). Rolecek et al. (2007) found that preferential sampling
designs cover a greater range of environmental extremes
than random sampling designs for the same level of survey
effort; our results appear to agree with this. Despite an initial
perception that 51,500 km 2 would not be adequately sampled
with 288 detailed plots, this survey covered the environmental
38
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
variability and community richness comprehensively on land
zone 5, and adequately on land zone 12. We suggest this is due
to the two-tiered system of data collection, with large numbers
of rapid observational sites augmented by detailed vegetation
plots in representative locations. The small difference in
the sampling adequacy between land zones is likely due to
accessibility. Whilst the landscapes of the old loamy and sandy
plains of land zone 5 are relatively well traversed by roads,
allowing access for detailed plot collections, the landscapes
of the igneous rocks of land zone 12 have mountainous terrain
with few roads providing limited access.
Although our sample design adequately surveyed landscape
variability and community richness, our analyses show this
is not so for species richness. This differs from other studies
that found preferential-sampling had a higher likelihood of
sampling the full species richness than stratified-random
sampling, as researchers tend to choose sample locations
with higher species richness (Michalcova et al. 2011). In this
survey, however, locations were chosen on a perception of
representativeness of distinctive communities, rather than
species richness, potentially explaining the difference to
other studies. Our survey’s design of detailed plot locations
evidently captures the communities present, but not the full
floristic variability within those communities. This result
agrees with the intuitive assessment that sampling such a
large area with so few sites would not provide comprehensive
coverage; and with Lawson et al. (2010) who found high
levels of floristic heterogeneity within regional ecosystems
in south-east Queensland.
A major function of a classification scheme is to allow new
site data to be assigned to it (De Caceres & Wiser 2012).
In the authors’ experience, an important issue when using a
qualitatively-derived classification for this task, is ambiguity in
allocating new sites into the scheme. A descriptive-framework
based on quantitative data helped overcome this by allowing
us to allocate sites with different data collection methods to the
classification scheme with a high level-of-fit to community,
enhancing the repeatability of allocating new sites. This, in
turn, increases the classification’s applicability by allowing 1)
easier recognition of community types, 2) greater confidence
in identifying sites from communities new to the classification,
and 3) the classification to become a dynamic scheme
responsive to new infonnation. Our descriptive-framework
does not fit the definition of membership rules outlined by De
Caceres & Wiser (2012), (in that the same rules used to define
communities are not used to allocate new sites into it) but it
performs a similar function.
A potential benefit of incorporating quantitative analyses
in the Regional Ecosystem framework is to allow a display
of relationships between communities not obvious in a
qualitative classification. An area with many similar REs,
may have less diversity than an area with fewer dissimilar
REs. For instance, a result of the committee process of
allocating communities to REs, based on non-floristic
variables, is that REs can contain communities dominated
by different species with low similarity to each other.
Dendrograms, scatter plots and similarity matrices produced
by quantitative analyses provide a visualisation and measure
of the similarities between REs and their vegetation
communities (Appendix 5). For example RE 3.12.18 has two
communities ‘a’ and ‘b’ (Appendix 5, figs 5.2 and 5.4). RE
3.12.18b is found in small patches scattered through larger
areas of 3.12.18a, on the same landform, and not predictable
enough to be reliably mapped at 1:100,000 scale. Displaying
these relationships between communities may be useful in
conservation planning, for example.
Incorporating quantitative analyses in the Regional
Ecosystem framework will enhance its already wide use.
As well as the current comparisons of spatial and temporal
change of REs (Accad et al. 2017), statistical comparisons
between vegetation communities at a cross-bioregion
scale will become possible (Goodall 1973). We anticipate
quantitatively-based vegetation communities wifi aid
investigations into questions such as the assumptions
behind their use as surrogates for biodiversity (Sattler &
Williams 1999), the environmental drivers of the patterns
of community distribution, and the phylogenetic diversity
of communities. Importantly it will provide statistically-
backed base-line data against which to measure the effects
of future changes, such as climate and land use. REs are
used by a wide cross-section of the public and form part of
legislation at multiple tiers of government. With vegetation
communities (the base-line level of the RE hierarchy) based
on quantitative analyses, REs are more robust and readily
defensible, providing legislators and users with greater
confidence in the classification scheme.
Conclusion
To standardise classification procedures across large
geographic areas and multiple administrative boundaries is
one of the globally-recognised goals of vegetation science
(Jennings et al. 2009; Walker et al. 2013; De Caceres et
al. 2015). These procedures are generally described as
standardised data collection methods, classification schemes
and quantitative classification techniques. In Australia, most
state governments have adopted approaches which work
towards achieving these goals (Sun et al. 1997; Gellie et al.
2017). In Queensland this is well advanced. As well as having
state-wide Regional Ecosystem mapping at 1:100,000 scale,
there is a standardised classification scheme, data collection
methods and qualitative classification techniques. Extending
our quantitative classification approach to the Regional
Ecosystem framework across the remainder of Cape York
Peninsula and other bioregions in Queensland, wifi further
the achievement of these globally recognised goals.
Acknowledgements
This work was carried with the support of the Queensland
Herbarium, Department of Environment and Science,
Queensland Government. We thank Peter Bannink for the map
figures. We particularly thank the 18 members of the expert
panel for their time and commitment in attending the technical
review committee workshop in Caims, Queensland in 2015.
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
39
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Manuscript accepted 20 July 2018
Appendix 1: Descriptive-framework for quantitatively derived vegetation communities on land zone 5 and 12 in Cape York Peninsula bioregion.
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
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Addicott et al.. New plant community classification, Cape York, Queensland
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Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
Appendix 3a. Assessment of the sampling adequacy of survey design on land zone 5 and 12, Cape York
Peninsula bioregion
Sampling adequacy of landscape variability
Table 3.1: Total area of land zone 5 and 12 at different similarity levels to any site for each environmental variable. For
example, 818 ha of land zone 5 is between 75 - 89% similar in climate to any observational site. This represents 0.01% of
the total area of land zone 5. The minimum similarity in climate of any grid cell to any observational site is 81%. Figures are
rounded to the nearest hectare or km 2 .
LZ 5
Observational sites
Analysis Sites
% Similarity
Class
ha
km 2
% total area
Minimum
%similarity
%
Similarity
Class
ha
km 2
% total
area
Minimum
%similarity
Climate
<75%
0
0
0%
81
<75%
205
2
0.003%
70
75 - 89%
818
8
0.01%
75 - 89%
26076
261
0.41%
90-95%
3283
33
0.1%
90-95%
400620
4006
6.32%
>95%
6333855
63339
99.9%
>95%
5911054
59111
93.26%
Vegetation Density
<75%
241
2
0.004%
63
<75%
3657
37
0.06%
12
75 - 89%
659
7
0.01%
75 - 89%
6360
64
0.10%
90-95%
673
7
0.01%
90-95%
40744
407
0.64%
>95%
6336315
63363
99.98%
>95%
6287275
62873
99.2%
Soil Nutrient
<75%
76
1
0.001%
66
<75%
2229
22
0.04%
0
75 - 89%
53
1
0.001%
75 - 89%
261
3
0.004%
90-95%
197
2
0.003%
90-95%
20946
209
0.33%
>95%
6331552
63316
99.99%
>95%
6278650
62787
99.16%
Soil Structure
<75%
0
0
0
84
<75%
21888
219
0.35%
0
75 - 89%
426
4
0.01%
75 - 89%
65998
660
1.04%
90-95%
6601
66
0.10%
90-95%
493978
4940
7.80%
>95%
6324721
63247
99.89%
>95%
5749884
57499
90.81%
LZ 12
Observational sites
Analysis sites
% Similarity
Class
ha
km 2
% total area
Minimum
%similarity
%
Similarity
Class
ha
km 2
% total
area
Minimum
%similarity
Climate
<75%
0
0
0.0%
84
<75%
149
1
0.02%
60
75 - 89%
2903
29
0.3%
75 - 89%
17770
178
1.9%
90-95%
19622
196
2.1%
90-95%
178764
1788
19.5%
>95%
894385
8944
97.5%
>95%
720226
7202
78.5%
Vegetation density
<75%
64
1
0.01%
59
<75%
524
5
0.06%
5
75 - 89%
167
2
0.02%
75 - 89%
12768
128
1.4%
90-95%
1147
11
0.1%
90-95%
14875
149
1.6%
>95%
915365
9154
99.9%
>95%
888575
8886
96.9%
Soil nutrient
<75%
651
7
0.1%
35
<75%
6879
69
0.8%
27
75 - 89%
2615
26
0.3%
75 - 89%
12865
129
1.4%
90-95%
14165
142
1.6%
90-95%
21919
219
2.4%
>95%
884381
8844
98.1%
>95%
860150
8602
95.4%
Soil structure
<75%
20
0.2
0.002%
68
<75%
1773
18
0.2%
48
75 - 89%
5135
51
0.6%
75 - 89%
42218
422
4.7%
90-95%
28428
284
3.2%
90-95%
116632
1166
12.9%
>95%
868186
8682
96.3%
>95%
741159
7412
82.2%
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
57
Appendix 3b: Areas of low sampling adequacy by survey design on land zone 5 and 12, Cape York
Peninsula bioregion
Figure 3.1: Distribution of areas of land zone 5 and land zone 12 which are <90%-similar to any site for each environmental
variable. Because such large areas of both land zones were >90%-similar to any site, for display purposes we show only areas
with <90%-similarity. Areas on land zone 12 correspond largely with areas of rainforest which are not included in this study.
These maps are indicative only. GIS layers are available from the first author if more detail is required
Figure 3.1a: Climate.
58
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
Bamaga
Weipa.
Coen
Pormpuraaw,
Cooktown
Laura 1
Lakeland
Legend
Towns
Main Roads
I I Cape York Bioregion
■ LZ 5 Similarity < 90%
■ LZ 12 Similarity < 90%
0 20 40 80
Kilometres
Coral Sea
Lockhardt River
10 y S
120
14 0 S
- 16°S
Figure 3.1b: Woody vegetation density (represented by maximum persistent greenness)
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
59
Figure 3.1c: Soil nutrient
60
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
Bamaga
Weipa
Coen
Pormpuraaw
Gooktown
Laura
Lakeland
- 10°S
Legend
• Towns
r-^ Main Roads
I I Cape York Bioregion
■ LZ 5 Similarity < 90%
■ LZ 12 Similarity < 90%
0 20 40 80
Kilometres
Coral Sea
120
14 S
16° S
Figure 3. Id: Soil structure
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
61
Appendix 4: Additional analysis requested by the technical review committee and recommendations.
The expert panel queried two communities identified by the numerical analysis, requesting further analysis. These were the
Eucalyptus tetrodonta, Corymbia nesophila woodlands and the Eucalyptus tetrodonta, Corymbia stockeri woodlands, both
distributed across the extent of land zone 5. The final recommendations are discussed below.
Methods
We carried out the initial investigations with the Eucalyptus tetrodonta, Corymbia nesophila woodlands, testing for differences
in three attributes; canopy heights of the tallest layer, and floristic differences in the woody and ground layer vegetation
(separately). We tested each attribute for differences between landform (Tertiary remnant plateaus and sand plains), soil colour
(red, yellow, brown) and soil texture (sand and earth) as recorded on site pro-formas. We used the ANOSIM routine (Clarke
and Gorley 2006) which has two outputs; an R statistic and a significance value. The R statistic generally lies between 0 (there
is no difference between the groups) and 1 (there is no similarity between the groups) but negative values indicate the within
group variation is larger than the between group variation. In the ground layer we firstly looked for distinct species assemblages
using 77 MDS and visually assessed whether these were coincident with different landform, soil colours or soil texture using GIS
overlay. To test for differences in canopy height we also used an unpaired t-test as well as the ANOSIM routine. Due to the
results of these investigations in the Eucalyptus tetrodonta, Corymbia nesophila woodlands, analysis requested by review panel
for the Eucalyptus tetrodonta, Corymbia stockeri woodlands was limited to differences in canopy height between landform
(again Tertiary remnant plateaus and sand plains) and soil colour (red earths versus all other colours).
Of the 50 sites in the Eucalyptus tetrodonta, Corymbia nesophila woodlands, 32 contained data useful for ground layer
analysis and 49 for soil analysis. There were 3 additional sites not included in the original dataset which contained enough
information for testing canopy heights. This resulted in 53 sites in the canopy height analysis. There were 31 sites in the
Eucalyptus tetrodonta, Corymbia stockeri woodlands.
Results
Eucalyptus tetrodonta, Corymbia nesophila woodlands
Floristic differences in woody vegetation layers.
There was no floristic difference between soil textures ( R = -0.05, p = 0.75), soil colours ( R = 0.08, p = 0.14 ) or landform {R
= 0.01, p = 0.44). The negative R value for soil texture indicates that the floristic differences individually on the sandy soils
and on the earth soils is greater than the floristic differences between these two soil types.
Floristic differences in ground layer vegetation.
The two-dimensional «MDS ordination showed two ground layer species assemblages, one dominated by Heteropogon
triticeus and the other by Schizachyrium species (figure 4.1), but with a lot of variability as evidenced by the high stress level
(0.2). However, these assemblages were not significantly associated with either different soil textures (R = 0.02, p = 0.40),
different soil colours (R = -0.08, p = 0.71) or different landforms (R = 0.04, p = 0.33). This was also supported by the GIS
overlay where there was no clear alignment of these assemblages with different soils or landforms.
62
Cunninghamia 18: 2018
Addicott et al. , New plant community classification, Cape York, Queensland
Figure 4.1: Bubble plot showing two species assemblages in the ground layer of the Eucalyptus tetrodonta, Corymbia nesophila woodlands
- one dominated by Schizachyrium spp, the other by Heteropogon triticeus. Abundances are standardised.
Canopy height differences
There was no difference in the canopy heights on different soil textures ( t(47) = 1.1, p = 0.28) and the ANOSIM results
indicated the variability of heights within individual soil textures was greater than between the soil textures (R = -0.04).
Differences in canopy height on different coloured soils was not straight forward. There was a distinct, but not significant
difference between the heights of trees on red earths versus brown earths (R = 0.86, p = 0.06), and an indistinct, but significant
difference between the heights of trees on red earths versus yellow earths (R = 0.18, p = 0.03). The differences in canopy
heights between landforms, however, was highly significant ( t(51) = 5. 7 , p<0.0001), with the average height of trees on the
Tertiary remnant plateaus being 5.2m taller than those on sand plains. We confirmed these results by running two different
ANOSIM analysis. Firstly, we included all sites; 13 on the plateaus and 40 on the plains. These results showed a significant
difference (p = 0.01), but a large overlap in height (R = 0.25). We then ran ANOSIM with an equal number of sites (13)
in both landforms (sites from the plains were chosen randomly). The difference in height was again significant (p = 0.1)
however there was a small overlap in height (R = 0.84).
Eucalyptus tetrodonta, Corymbia stockeri woodlands
There was a significant difference in the canopy heights of trees on both different landforms and different soil colours. The
average height difference between trees on Tertiary remnant plateaus and on sand plains was 7.5m (t(29) = 7.0, p<0.0001)
and on red earths versus all other coloured soil was 7.2m (t(29) = 6.4, p<0.0001). The ANOSIM results show that there is
overlap in tree height on both landform (R = 0.63) and soil colour (R = 0.52).
Discussion
There are no differences in the floristics of the woody vegetation of the Eucalyptus tetrodonta, Corymbia nesophila woodlands
across land zone 5. There is a difference in the floristics of the ground layer, but it is not relatable to differences in soil types
or landform and it is possible that the different assemblages are due to disturbance history (Kutt and Woinarski 2007, Miller
and Murphy 2017).
There were significant differences in the canopy height of both Eucalyptus tetrodonta, Corymbia nesophila woodlands and
Eucalyptus tetrodonta, Corymbia stockeri woodlands on different landforms and soil colour. The red earths, which are most
common on the remnant plateaus, grow significantly taller woodlands than other coloured soils, which are most common
on the sand plains. From this it is not surprising that the woodlands on the Tertiary remnant plateaus are significantly taller,
however, as our ANOSIM results indicate there are areas on sand plains and on yellow earths where woodlands are also tall.
This leads us to conclude that the height of woodlands on sand plains is variable, but woodlands on the remnant plateaus are
consistently taller.
Recommendation
The classification protocols used in Queensland (Neldner et al. 2017) specify that woodlands with the same dominant species,
but with a consistent height difference of 5m, can be split into separate communities. Despite having no consistent floristic
differences, the Eucalyptus tetrodonta, Corymbia nesophila woodlands and the Eucalyptus tetrodonta, Corymbia stockeri
woodlands on the Tertiary remnant plateaus are consistently >=5m taller than those on sand plains. However, there is an
overlap in height between the plateaus and the sand plains. We therefore recommend the woodlands on the remnant plateaus
are recognised as vegetation communities within the appropriate floristically defined regional ecosystem.
References
Clarke, K.R. and R N. Gorley. 2006. PRIMER v6: User Manual/Tutorial. PRIMER-E, Plymouth.
Kutt, A.S. and J.C.Z. Woinarski. 2007. The effects of grazing and fire on vegetation and the vertebrate assemblage in a tropical savanna
woodland in north-eastern Australia. Journal of Tropical Ecology 23 (1): 95-106.
Miller, B.P and B.P Murphy. 2017. Fire and Australian Vegetation. In: D. Keith (eds), Australian Vegetation. Cambridge University Press,
Cambridge. 113-134.
Neldner, V.J., B.A. Wilson, H.A. Dilleward, T.S. Ryan and D.W. Butler. 2017. Methodology for Survey and Mapping of Regional
Ecosystems and Vegetation Communities in Queensland, version 4. Queensland Herbarium, Queensland Department of Science,
Information Technology and Innovation, Brisbane, https.//publications.qld.gov.au/dataset/redd/resource/6dee78ab-cl2c-4692-9842-
b7257c2511e4, accessed 1st June 2017.
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
63
Appendix 5: Floristic similarities between communities on land zone 5 and land zone 12 in Cape York
Peninsula bioregion.
Plot data in each community was averaged. The dendrogram was formed using the CLUSTER routine and scatter plots using
wMDS ordination in PRIMER v6 (Clarke & Gorley 2006).
Eucalyptus leptophleba woodland on plains
Melaleuca stenostachya+l-Eucalyptus chlorophylla woodland
Melaleuca viridiflora+t- Corymbia clarksonlana woodland
Asteromyrtus lysicephainaodfor Neofabncia myrtifolia open heath
tetrodonta+l- C.sfocker/woodland with a Me 'aleuca spp shrub layer ^ ^
Corymbia novoguinensis+l- C. tessellaris woodland
Eucalyptus tetrodonta , Corymbia stocked +/- E. culleni woodland
Eucalyptus tetrodonta+l- Corymbia stocked woodland
Eucalyptus tetrodonta, Corymbia stockeri+l- C. setosa woodland
E. tetrodontaandC. nesophifa open forest on remnan
E. tetrodonta and
Eucalyptus tetrodonta W- Go
E. tetrodonta, C. n
JW etateuca
C. nesoohilawood\ani on undul ating plains
rymbia e/ar/csomanawoodland
plateaus
Corymbia nesophila open forest
3.5.25
3.5.40
3 5.41b
Asteromyrtus brassHand /or Neofabricia myrtifolia low open forest _ 3 5 42
3.5.19
3.5.5
3.5,38a
3.5.37a
3.5.9
3.5.36b
3 5.36a
3.5.39
esophilawooz land with a heathy understory 3 5 35
3.5.34
viridiflora +1- Corymbia clarksonlana woodland _ 3 5 41a
Eucalyptus phoenicea woodland _ 3 5 6
Dapsilanthus spathaceus open sedgeland _ 3 & 15b
20
40
% similarity
60
80
Figure 5.1: Dendrogram showing hierarchical relationships of communities on land zone 5 identified by quantitative analyses.
64
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
Eucalyptus leptophleb a+l- Corymbia clarksoniana woodland 3 12 18 a
Eucalyptus cullen fy +/- Corymbia clarksoniana woodland 3 ^ -|Q a
Eucalyptus chlorophyllawoodland
Eucalyptus tetrodonta woodland
Corymbia nesophila+l- El
Eucalyptus tetrodon
We
calyptus etrodont awoodland
a woodland +/- heath species
Melaleuca viridiflora low woodland
chiodendron longivale and Acacia brassii lowwoodland
Corymbia stockeri+l - Welchiodendron longivalve woodland
Corymbia disjuncta+l- C. c/arfrson/anawoodland
feromyrti/s iysicephata+l- Allocasuarina littoralis mixed lowwoodland
Lep ospermum purpurascens tall shrubiand
Asteromyrtus lysicephala, Choriceras fwcorne,dwarf s hrubland
Melaleuca citrolen slow open woodland
Schizachryium spp. +/- Rhynchosiaspp . grasslands
Corymbia tessellaris +/- Welchiodendron longivalve open forest
Deciduous to semi deciduous vine thicket
lentous rock pavements associated with mountains and some offshore islands
Heteropogon triticeus+l- Sarga plumosum grasslands
3,12,10b
Zorymbia clarksoniana, Eucalyptus brassiana open forest 3127
3,12.42
3.12.40
3.12.41
3.12.45
3.12.43a
3.12.11
3.12,18b
3.12.47a
3.12.28
3.12,47b
3.12.44
3.12.32
3.12.9
Melal q uca viridiflora and Welchiodendron longivalesh rub land 3 ^ 43b
3,12.21
Schizachryium spp., Aristida spp. grasslands 3 12 4gc
3.12.48a
3.12.34
I mperata cylindrica +1- Heteropgon contortus closed tussock grassland 3 12 30
Lophostemon suaveolens lowopen forest
Eucalyptus crebra+l- Corymbia hyfandii lowwoodland
3.12.39b
3.12.39a
T
40
T
60
0
20
% Similarity
Figure 5.2: Dendrogram showing hierarchical relationships of communities identified by quantitative analyses on land zone 12.
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
65
■ Eucalyptus tetrodonta dominated woodland ♦ Other Eucalypt and Corymbia dominated woodlands
Figure 5.3: Land zone 5 scatter plot showing relative similarity of communities to each other in two-dimensions. Communities close
together are more similar to each other. The greater clumping of communities than on land zone 12 scatter plot (Fig. 5. 4) indicates a higher
level of similarity of communities on land zone 5 than those on land zone 12.
3.12.18b
♦ 3.1Z10b
3.12.44
A
342.32
X
2D Stress: 0.22
3,12.39a
3.12.18a
3.12.7 ▲
▲
^ 3.12.10a 3.12.45
A
3 1247b
•
342.34
+
342.30
X
3.12.39b
♦
3.12.28
31 ^ 42 3.12.43a
3 12 40 " * 342.47a
♦ #
3.12.41
■
3 12.11
♦
3.12.48c
X
342.48a
X
3.12 43b
•
3 12 9
♦
A Melaleuca dominated woodland
■ Eucalyptus tetrodonta woodland
Ironbart; dominated woodland
X Grassland Rock pavements
Other Eucalypt and Corymbia dominated woodlands
• Shrubland ^ Acacia dominated woodland
Figure 5.4: Land zone 12 scatter plot showing relative similarity of communities to each other. Communities close together are more
similar to each other. The more scattered spread of communities on land zone 12 when compared to the land zone 5 (Fig. 5.3) indicates a
lower level of similarity between communities than land zone 5.
66 Cunninghamia 18: 2018 Addicott et al., New plant community classification, Cape York, Queensland
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Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
67
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Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
References
Clarke, K.R. & Gorley, R.N. (2006) PRIMER v6: User Manual/Tutorial. PRIMER-E, Plymouth.
Neldner, V.J., Niehus, R.E., Wilson, B.A., McDonald, W.J.F., Ford, A.J. & Accad, A. (2017) Vegetation of Queensland. Descriptions of
Broad Vegetation Groups. Queensland Herbarium, Department of Science, Information Technology and Innovation.
Cunninghamia 18: 2018
Addicott et al., New plant community classification, Cape York, Queensland
69
Appendix 6. Investigation into the correlations
between persistent greenness index and climate
variables.
We tested for correlations between climate variables and
woody vegetation density using a 4-way ANOVA in the
EXCEL stats package. Woody vegetation is represented
by a maximum persistent greenness index (JRSRP 2017).
The line-fit plots of woody vegetation density against each
climate variable (Fig 6.1 - 6.4) provide a visualisation of the
strength of correlation and the low predictability for woody
vegetation. While there is a significant correlation between
woody density and climate, the spread of actual woody
vegetation values compared to expected values portrays the
low predictability of woody vegetation density by climate
(R 2 = 0.34)
Table A6.1: 4-way ANOVA of woody vegetation density against
climate variables.
Regression Statistics
Multiple R
0.58
R Square
0.34
Adjusted R Square
0.34
Standard Error
13.00
Observations
1000
ANOVA
df
SS
MS
F Significance F
Regression
4
86136.3
21534.1
127.3 8.13988E-88
Residual
995
168254.7
169.1
Total
999
254391
Coefficients
Standard Error
t Stat
P-value
Lower 95%
Upper 95%
Intercept
203.77
26.98
7.55
9.67745E-14
150.82
256.71
Temperature
seasonality (C of V%)
-0.33
0.09
-3.51
0.0005
-0.51
-0.15
Mean moisture index of
lowest quarter
1.75
0.39
4.50
7.63064E-06
0.99
2.52
Annual precipitation (mm)
0.02
0.01
3.23
0.0013
0.01
0.03
Annual mean
temperature (°C)
-2.76
0.88
-3.12
0.0018
-4.49
-1.02
70
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
x
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-1 - 1 - 1 - 1 - r~
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Annual mean temperature (deg Celsius)
27
Figure A6.1: Line-fit plots of annual mean temperature against woody vegetation density (represented by maximum persistent greenness
index). ♦ = actual maximum persistent greenness index at each observation point, ■ = predicted maximum persistent greenness index at
each observation point
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Average annual precipitation (mm)
2000
Cunninghamia 18: 2018
Addicott et al.. New plant community classification, Cape York, Queensland
71
Figure A6.2: Line-fit plots of average annual precipitation against woody vegetation density (represented by maximum persistent greenness
index). ♦ = actual maximum persistent greenness index at each observation point, ■ = predicted maximum persistent greenness index at
each observation point
210
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i-r-1-i-i-1-1
40 50 60 70 80 90 100
Temperature Seasonality (C of V%)
Figure A6.3: Line-fit plots of temperature seasonality against woody vegetation density (represented by maximum persistent greenness
index). ♦ = actual maximum persistent greenness index at each observation point, ■ = predicted maximum persistent greenness index at
each observation point
X
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Figure A6.4: Line-fit plots of mean moisture index of the lowest quarter against woody vegetation density (represented by maximum
persistent greenness index). ♦ = actual maximum persistent greenness index at each observation point, ■ = predicted maximum persistent
greenness index at each observation point
72
Cunninghamia 18: 2018
Addicott et al ., New plant community classification, Cape York, Queensland
References.
JRSRP. (2017) Seasonal fractional vegetation cover for Queensland derived from USGS Landsat images. Joint Remote Sensing Research
Project. Department of Science, Information Technology and Innovation, Brisbane.
Date of Publication:
October 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
Angophora subvelutina (Myrtaceae) on atypical diatreme habitat
at Glenbrook : an addition to the eucalypt list for the Greater Blue
Mountains World Heritage Area
Judy Smith 1,2 , Peter Smith 1 and Doug Benson 2,3
'P & J Smith Ecological Consultants, 44 Hawkins Pde, Blaxland, NSW 2774, AUSTRALIA, smitheco@ozemail.com.au
2 Sometime member GBMWEIA Advisory Committee.
3 Honorary Research Associate, National Herbarium of New South Wales, Botanic Gardens & Domain Trust,
Sydney NSW 2000, AUSTRALIA.
Abstract : The Greater Blue Mountains World Heritage Area (GBMWHA), a natural area of about one million hectares
immediately west of Sydney, Australia, is significant for its biodiversity, and particularly for its richness of eucalypt
species (species of Eucalyptus , Angophora and Corymbia in the family Myrtaceae), numbered at 96 species in
2010. This paper describes the finding of a previously unlisted Angophora species in the GBMWHA, and makes a
conservation assessment of the population. A population of the Broad-leaved Apple Angophora subvelutina L. Muell.
occurs at Euroka Clearing south of Glenbrook just within the eastern edge of Blue Mountains National Park, one
of the eight conservation reserves that make up the GBMWHA. The population numbers over 200 plants and there
is evidence that the species has been present at the site since before European settlement. The population includes
a mixture of age classes and is considered viable, although substantial intergradation is occurring with the closely
related species Angophora floribunda. Elsewhere in the Sydney area, the species is relatively uncommon and has been
extensively cleared from its relatively fertile habitats. The population in the GBMWHA noted here has conservation
significance for its size and long history at the site, and for the unusual ecological conditions of the Euroka diatreme,
which is an atypical habitat for the species.
Cunninghamia (2018) 18: 073-078
doi: 10.775 l/cunninghamia.2018.18.004
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
74
Cunninghamia 18: 2018
Smith et al, Angophora subvelutina at Glenbrook
Introduction
The Greater Blue Mountains was inscribed on the World
Heritage List in 2000 for its outstanding natural values, a
major component of which is the high number of eucalypt
species and eucalypt-dominated communities (the term
‘eucalypt’ refers to the closely related genera Eucalyptus,
Angophora and Corymbia of the family Myrtaceae). In
2000,91 eucalypt species were known from the Greater Blue
Mountains World Heritage Area (GBMWHA). A subsequent
assessment of the eucalypts in the eight conservation reserves
which make up the GBMWHA (Blue Mountains, Gardens of
Stone, Kanangra-Boyd, Nattai, Thirlmere Lakes, Wollemi
and Yengo National Parks and Jenolan Karst Conservation
Reserve) listed 96 eucalypt species (Hager & Benson 2010).
The number of eucalypt species recognised in the area is
likely to fluctuate given the somewhat equivocal nature of
systematic description and changes in the state of scientific
knowledge, particularly with the increased application of
genetic research. For example, Rutherford et al. (2018), in a
genetic study, suggested that at least one of the eight green¬
leaved ashes (. Eucalyptus cunninghamii ) in the GBMWHA
(Hager & Benson 2010) showed distinct genetic variation
between populations warranting recognition of a new
undescribed species, while two other species {Eucalyptus
laophila and Eucalyptus stricta ) could not be distinguished
from each other.
A different situation applies to the finding of a population of
a well-accepted existing species, not previously formally
recorded in the GBMWHA. Such species may be found in areas
that have remained inadequately explored botanically or within
lands that are subsequently added to the GBMWHA. The
species may also be very rare in the GBMWHA, with only one
or two obscure previous records that have gone unrecognised.
In 2017 Peter Smith noticed that Angophora subvelutina
F.Muell. (Broad-leaved Apple) was not included in the list
of GBMWHA eucalypts in the foyer of the World Heritage
Exhibition at the Blue Mountains Cultural Centre, Katoomba.
The species is also not included in the GBMWHA eucalypt
list of Hager & Benson (2010). Peter and Judy Smith
recalled that Angophora subvelutina occurred at Euroka
Clearing, Glenbrook, just inside the eastern boundary of
Blue Mountains National Park and hence the GBMWHA.
Here, we confirm the presence of a previously unlisted (Benson
& Hager 2010) eucalypt species {Angophora subvelutina)
in the GBMWHA. We describe and assess the Angophora
subvelutina population at Euroka Clearing, Glenbrook, and
consider the long term viability of this population.
Angophora subvelutina usually grows on deep alluvial soils
and may be locally abundant. It occurs at scattered locations,
mainly east of the Great Dividing Range, from south-eastern
Queensland south to the Bega district in southern NSW. In
NSW it has been recorded in the North Coast (NC), Northern
Tablelands (NT), North Western Slopes (NWS), Central
Coast (CC), Central Tablelands (CT), Central Western Slopes
(CWS) and South Coast (SC) botanical subdivisions, south
to the Araluen district (PlantNET NSW Flora Online 2018).
Intergrades with the closely related species Angophora
floribunda are known from the NWS, CC, CWS and SC,
and also occur beyond the known distribution of Angophora
subvelutina in the Bega district of the far SC, and in the North
Western Plains (NWP) (PlantNET NSW Flora Online 2018).
MUSWELLBROOK
SINGLETON
HOWES VALLEY
CESSNOCK
KANDOS
LITHGOW
KATOOMBA
PICTON
BOWRAL
• iiijf sydney
NEWCASTLE
30 km
WOLLONGONG
Fig. 1: Map of Angophora subvelutina records in the vicinity of
the Greater Blue Mountains World Heritage Area. Records from
Australasian Virtual Herbarium and NSW BioNet databases,
extracted 27 July 2018. Red circles, records from 2000 or later;
blue circles, records before 2000; purple star, Euroka Clearing;
black star, Sun Valley. Records with inexact locations have not been
mapped.
Distribution of Angophora subvelutina
Angophora subvelutina was first described by Ferdinand
Mueller in 1858. It is a tree that typically grows to about
20 m high with persistent, grey, fibrous-flaky bark and adult
leaves which are relatively broad, more or less sessile, and
cordate at the base.
In the Sydney area, specimen records from the Australasian
Virtual Herbarium (2018) and sightings records from the NSW
BioNet Atlas (2018) indicate that Angophora subvelutina
is mainly associated with river and creek systems on the
Cumberland Plain (Fig. 1). It is found in floodplain forest
and on creek banks on deep fertile alluvial soils, but may also
Cunninghamia 18: 2018
Smith et al, Angophora subvelutina at Glenbrook
75
be associated with shale-derived soils with medium to high
nutrient levels. In floodplain forest, associated tree species
may include Eucalyptus baueriana, Eucalyptus tereticornis,
Eucalyptus amplifolia and Eucalyptus botryoides x saligna
(Benson & McDougall 1998). Most of its original habitat on
the floodplains has now been cleared or severely degraded;
James et al. (1999) regarded its regional conservation status
in western Sydney as Vulnerable. Angophora subvelutina
is a characteristic species of ‘River-Flat Eucalypt Forest
on Coastal Floodplains of the NSW North Coast, Sydney
Basin and South East Corner Bioregions’, which is listed
as an Endangered Ecological Community under the NSW
Biodiversity Conservation Act 2016 (NSW Scientific
Committee 2011).
Study Area
Our study was undertaken at Euroka Clearing in Blue Mountains
National Park, 2.5 km south of the township of Glenbrook.
Euroka Clearing is a large semi-cleared area in a circular
valley on a volcanic diatreme. Euroka Creek and its tributaries
traverse the valley and drain to the Nepean River, some 800 m
to the east. It is a popular camping area and is managed as such
by NSW National Parks and Wildlife Service. Tree species in
the remnant forest at Euroka Clearing include Allocasuarina
torulosa, Casuarina cunninghamiana, Angophora bakeri,
Angophora floribunda , Angophora subvelutina , Corymbia
eximia. Eucalyptus agglomerata, Eucalyptus beyeriana,
Eucalyptus deanei , Eucalyptus eugenioides, Eucalyptus
fibrosa , Eucalyptus punctata , Eucalyptus saligna , Eucalyptus
tereticornis and Syncarpia glomulifera. The vegetation on this
diatreme is notable for the diversity of tree species and for its
Cumberland Plain influences.
Methods
On 27 June 2017, Peter and Judy Smith collected a specimen
(including adult leaves and fruiting capsules) of a tree at
Euroka Clearing that they had identified as Angophora
subvelutina (Broad-leaved Apple). The specimen tree
(Fig. 2) was growing on the cleared slope on the eastern
side of Euroka Creek, upstream of where the creek makes a
right-angle bend. Its height was measured with a clinometer
as 15 m on 8 February 2018 and it appeared healthy. The
GPS coordinates of the tree were -33.799102, 150.618715
(GDA94 datum). The elevation of the tree was about 85 m asl
(determined from a 10 m GIS contour layer). The specimen
was given to the National Herbarium of NSW, Royal Botanic
Gardens Sydney for identification.
On 8 February and 28 March 2018, Peter and Judy Smith
carried out a series of field observations to provide an
estimate of the size and age class structure of the Angophora
subvelutina population at Euroka Clearing. It was soon noted
that the population of Angophora subvelutina was intermixed
with a population of Angophorafloribunda and that a number
of plants were intergrades between the two species. A random
sample of 100 rough-barked Angophora plants was selected
and each was identified to species {Angophora subvelutina,
Angophora floribunda or intergrade) and classified as a tall
tree, low tree (mature tree less than two-thirds the height of
the tallest trees) or sapling. Intergrades were identified on
leaf characters: leaves that were intermediate between the
cordate, virtually sessile leaves of Angophora subvelutina
and the cuneate, petiolate leaves of Angophora floribunda , or
else leaves that were a mixture of the two leaf types. A rough
estimate was made of the total number of rough-barked
Angophora plants at Euroka Clearing.
On 27 July 2018, Peter Smith collated and mapped previous
Angophora subvelutina records in and around the Greater
Blue Mountains World Heritage Area. Records were
obtained from the specimen database of the Australasian
Virtual Herbarium (2018) and the sightings database of the
NSW Bionet Atlas (2018).
Fig. 2: Angophora subvelutina tree (closest to camera) at Euroka
Clearing, Blue Mountains National Park, from which a specimen
has been lodged at the National Herbarium of NSW.
Results
The National Herbarium of NSW confirmed that the
specimen collected at Euroka Clearing in June 2017 was
indeed Angophora subvelutina.
The database searches revealed only a single previous record
that definitely came from within the GBMWHA: a specimen
at the National Herbarium of NSW that was collected at
Euroka Clearing by T.M. Whaite in 1952. This specimen
was from a “tree 20 ft [6 m], bark stiffly fibrous .... on
breccia by creek”. There is also a specimen at the N.C.W.
Beadle Herbarium, University of New England, collected at
“Glenbrook” by T.J. Hawkeswood in 1975, which may have
come from Euroka Clearing.
76
Cunninghamia 18: 2018
Smith et al, Angophora subvelutina at Glenbrook
Another record of interest is a 2005 sighting in the Bionet
Atlas with coordinates (no description of the location but
coordinate accuracy reported as within 30 m) that place it
in the Putty Road corridor where it passes through Yengo
National Park just north of the Putty Valley Road turnoff, near
the head of Snakes Valley Creek. This record warrants further
investigation but, if correct, then Angophora subvelutina
likely occurs within the GBMWHA at this location.
There is also a cluster of four records (one undated specimen
and three 2006 BioNet sightings) along the Macdonald River
near St Albans, close to the boundary of Yengo National
Park. However, the Angophora subvelutina population along
the river may be restricted to the floodplain outside the
GBMWHA.
Another sighting of Angophora subvelutina in the BioNet
Atlas is from the railway corridor near Bullaburra station
in 2011, close to Blue Mountains National Park. This is an
unlikely location for the species and, if correct, is probably
not a natural occurrence. There are also duplicate Angophora
subvelutina specimens at the National Herbarium of NSW
and the National Herbarium of Victoria that were collected
at “Mt Victoria” by J.H. Maiden in 1901. This is another
unlikely location for the species but is probably either
an error or refers to the general Mt Victoria district rather
than the immediate vicinity of the township itself (which is
surrounded by Blue Mountains National Park). The specimen
was possibly collected in the nearby Kanimbla Valley or
Hartley Valley, outside the GBMWHA.
From our fieldwork in February and March 2018, we
estimated that the combined population of Angophora
subvelutina , Angophora floribunda and intergrades at
Euroka Clearing numbered over 500 individuals, with
most individuals found along or near Euroka Creek and its
tributaries. The three entities were intennixed and we could
see no obvious habitat differences between them.
Based on our sample of 100 plants, the rough-barked
Angophora composition at Euroka Clearing consisted of
43% Angophora subvelutina , 32% Angophora floribunda
and 25% intergrades (Table 1, Fig. 3). On this basis, we
estimate the size of the Angophora subvelutina population
at Euroka Clearing at over 215 plants. The Angophora
subvelutina population was of mixed sizes and 19% of
the population consisted of saplings of various ages. The
population also included some very large old trees, and one
notable example had a height of 37 m and a trunk diameter
of 1.65 m (Fig. 4). There was pronounced intergradation
between the Angophora subvelutina and Angophora
floribunda populations, with about three intergrade plants to
every five Angophora subvelutina plants.
Table 1: Rough-barked Angophora composition at Euroka
Clearing in February-March 2018, based on a sample of 100
plants.
Species
Tall trees
Low trees
Saplings
Total
Angophora subvelutina
21
14
8
43
Angophora floribunda
17
13
2
32
Intergrades
13
5
7
25
Total
51
32
17
100
<L>
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45 -
40 -
35
30 -
25 -
20 -
15
10
5
0
pTall trees llj-ow treesjp Saplings
Angophora subvelutina Angophora floribunda
Intergrades
Fig. 3: Rough-barked Angophora composition at Euroka Clearing
in February-March 2018, based on a sample of 100 plants.
Fig. 4: Large old Angophora subvelutina tree at Apple Tree Flat,
Euroka Clearing, Blue Mountains National Park. The tree is about
37 m tall, with a diameter at breast height of about 1.65 m. It likely
pre-dates European settlement.
Cunninghamia 18: 2018
Smith et al, Angophora subvelutina at Glenbrook
77
Discussion
We have confirmed the presence of a population of Angophora
subvelutina at Euroka Clearing, Glenbrook in Blue Mountains
National Park and thus the Greater Blue Mountains World
Heritage Area. The Angophora subvelutina population at
Euroka Clearing numbers over 200 plants and has clearly been
there for a long time. A specimen was collected from the site
in 1952 and the population includes a number of trees which
are likely to pre-date European settlement, such as the tree in
Fig. 4. The long history of the species at the site, the size of the
population in February and March 2018, and the range of age
classes present, indicate that it is a viable population, despite
substantial local hybridisation with another species at the site,
Angophora floribunda.
This species should now be added to the formal list of
eucalypts in the GBMWHA. Angophora subvelutina is
a well known and accepted species. Its late addition to
the GBMWHA eucalypt list highlights the fact that the
GBMWHA, an area that is internationally renowned for its
biodiversity, remains inadequately explored and documented
botanically. Far too few botanists have been afforded the
opportunity to work in this area.
The diatreme at Euroka Clearing is an unusual ecological site
for Angophora subvelutina. It has a creek system but is not
a floodplain. The Nepean River is nearby but runs through
a gorge and has only a narrow floodplain in this vicinity.
The slopes around the diatreme are on sandstone geology
and there is a remnant shale cap on the surrounding ridge.
The high nutrient soils of the diatreme, together with the
influence of the surrounding sandstone and shale geology,
appear to provide suitable conditions for both Angophora
subvelutina and Angophora floribunda , as well as for an
unusually large number of other eucalypt species that occur
at the site. We have also observed Angophora subvelutina
and Angophora floribunda growing together at Campbells
Ford beside the Nepean River in Gulguer Nature Reserve
east of the GBMWHA.
Angophora subvelutina also occurs on another large diatreme
at Sun Valley, about 10 km north of Euroka Clearing, outside
the GBMWHA (Fig. 1). Similarly to Euroka Clearing,
both Angophora subvelutina and Angophora floribunda
are present at Sun Valley, with evidence of intergradation
between them (Andrew Orme, pers. comm.). The Sun Valley
diatreme is a semi-rural area with many houses. The remnant
forest on this diatreme, which is dominated by Eucalyptus
amplifolia (Cabbage Gum), is listed as an Endangered
Ecological Community, ‘Sun Valley Cabbage Gum Forest in
the Sydney Basin Bioregion’, under the NSW Biodiversity
Conservation Act 2016 (NSW Scientific Committee 2001).
There is another, smaller diatreme, Machins Crater, about
5 km south-west of Euroka Clearing, within Blue Mountains
National Park. Judy and Peter Smith inspected this diatreme
on 12 September 2018. We found a single Angophora
floribunda tree but no Angophora subvelutina. The diatreme
supports relatively undisturbed forest dominated by
Eucalyptus deanei (Mountain Blue Gum).
We conclude that the population of Angophora subvelutina
at Euroka Clearing makes a valuable contribution to the
biodiversity of the GBMWHA, as well as to conservation
of the species in the general Sydney area. This is the only
currently known population in the GBMWHA. Local
populations in the Sydney area outside the GBMWHA are
considered vulnerable as most floodplain forests have been
cleared or severely degraded (Benson & McDougall 1998).
We recommend that the Euroka Clearing population be
monitored and managed to ensure its long-term viability.
Although located within a national park that forms part of
a World Heritage Area, the site where the population occurs
is managed as a camping ground and day-use area, which
may conflict with conservation management of Angophora
subvelutina. Potential future threats to the population
include climatic changes, inappropriate fire regime, lack of
adequate regeneration, vegetation clearing, diseases such as
the recently introduced Myrtle Rust, and genetic swamping
through hybridisation with Angophora floribunda.
Acknowledgements
We gratefully acknowledge the assistance that we obtained
during this study from Kristina McColl and Andrew Orme
of the National Herbarium of NSW, Jacqueline Reid of
NSW Office of Environment and Heritage, and local
botanist Margaret Baker. Dr Peter Wilson of the National
Herbarium of NSW confirmed the identification of the
Angophora subvelutina specimen. Roger Lembit reviewed
the manuscript and gave us helpful feedback. The Greater
Blue Mountains World Heritage Advisory Committee was
supportive of the study.
References
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Benson, D.H. (1992). The natural vegetation of the Penrith
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Benson, D. & McDougall, L. (1998). Ecology of Sydney plant
species Part 6: Dicotyledon family Myrtaceae. Cunninghamia ,
5, 808-987.
Hager, T. & Benson, D. (2010). The eucalypts of the Greater Blue
Mountains World Heritage Area: distribution, classification
and habitats of the species of Eucalyptus, Angophora and
Corymbia (family Myrtaceae) recorded in its eight conservation
reserves. Cunninghamia , 11, 425-444.
James, T., McDougall, L. & Benson, D. (1999) Rare Bushland
Plants of Western Sydney. Royal Botanic Gardens Sydney.
NSW BioNet Atlas (2018). Accessed 27 July 2018. http://www.
bionet.nsw.gov.au
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list the Sun Valley Cabbage Gum Forest in the Sydney
Basin Bioregion as an Endangered Ecological Community
on Part 3 of Schedule 1 of the Threatened Species
Conservation Act, November 2001. Available from
https://www.environment.nsw.gov.au/determinations/
SunValleyCabbageGumForestSydneyEndComListing.htm
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Smith et al, Angophora subvelutina at Glenbrook
NSW Scientific Committee (2011). River-Flat Eucalypt Forest on
Coastal Floodplains of the N SW North Coast, Sydney Basin and
South East Corner Bioregions - Determination to make a minor
amendment to Part 3 of Schedule 1 of the Threatened Species
Conservation Act, July 2011. Available from https://www.
environment.nsw.gov.au/determinations/riverflat36a.htm
PlantNET NSW Flora Online (2018). Accessed 30 July 2018.
http://plantnet.rbgsyd.nsw.gov.au/cgi-bin/NSWfl.pl?page=nsw
fl&lvl=sp&name=Angophora~subvelutina
Rutherford, S., Rossetto, M., Bragg, J.G., McPherson, FL, Benson,
D., Bonser, P. & Wilson, PG. (2018). Speciation in the presence
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Manuscript accepted 1 October 2018
Date of Publication:
November 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
Fate of a rare flowering event in an endangered population
of Acacia pendula (Weeping Myall) from the Hunter Valley,
New South Wales
Stephen A .J. Bell
School of Environmental and Life Sciences, University of Newcastle, University Drive, Callaghan,
NSW 2308, AUSTRALIA, stephen.bell@newcastle.edu.au
Abstract. A rare flowering event in a stand of Acacia pendula (Weeping Myall) (family Labaceae, Mimosoideae)
from the Hunter Valley of New South Wales is documented. This species flowers poorly in the region and (with
the exception of horticultural specimens) has not been observed to fruit and develop viable seed for over a decade.
One stand of this threatened Hunter Valley population of Acacia pendula was monitored over a seven month period
(January to July 2018) to investigate this poor reproductive output. Despite copious bud production in January and
Lebruary, the extent and condition of these, and all subsequent flowers rapidly declined, and none progressed to fruit.
Primary reasons for reproductive failure were postulated to be a combination of mass desiccation of capitula following
extended dry conditions, infestation by native flower- and phyllode-galling midges and thrips (Asphondylia sp.,
Dasineura glomerata, Kladothrips rngosus ), fungal galls ( Uromycladium sp.), caterpillars ( Ochrogaster lunifer ), and
mistletoe ( Amyema quandang). Collectively, these stressors appear to be eliminating seed production from the study
population; survival is maintained only by the copious root-suckering observed around most plants, particularly after
the pressure from stock grazing (cattle, sheep) has been removed. The age of trees studied, based on measures of girth
and comparison with growth rates reported for other semi-arid Acacia , was inferred to be between 50 and 150 years.
The level of Amyema quandang (mistletoe) infestation on Acacia trees was independent of tree size, and there was no
evidence to suggest that mistletoe density alone influenced flowering progress.
Consequences of these observations on future management of Acacia pendula in the Hunter Valley are briefly
discussed.
Key words: Acacia pendula , Hunter Valley, endangered population, flowering fate, health
Cunninghamia ( 2018 ) 18 : 079-088
doi: 10.7751 / cunninghamia.2018.18.005
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
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Bell , Flowering fate in Acaciapendula. Hunter Valley
Background
The tree Acacia pendula Cunn. ex Don (Weeping Myall)
(family Fabaceae, Mimosoideae) in the New South Wales
Hunter Valley region is a threatened species protected under
three elements of legislation. Nationally, the Environment
Protection and Biodiversity Conservation Act 1999 includes
Acacia pendula as a key species in the Critically Endangered
Hunter Valley Weeping Myall (Acacia pendula) Woodland.
Under the NSW Biodiversity Conservation Act 2016
(BC Act), an Endangered Population of Acacia pendula is
listed for the Hunter Valley, and the species also forms a
key component of the Critically Endangered Hunter Valley
Weeping Myall Woodland in the Sydney Basin Bioregion.
For the highly modified Hunter region, Weeping Myall and
its habitat is one of the most protected plant entities and
subsequently presents a range of conservation management
challenges wherever it occurs.
Recent debate in the literature over whether or not Acacia
pendula populations in the Hunter Valley were present
prior to European settlement is difficult to fully resolve
without detailed cross-population genetic studies (Bell &
Driscoll 2014, 2016; Tozer & Chalmers 2015, 2016). As
a consequence, the NSW Threatened Species Scientific
Committee resolved to retain Acacia pendula within its lists
of threatened entities under the Biodiversity Conservation
Act 2016 until such clarifying evidence emerges. That being
the case, continuing research into the ecology of Hunter
Valley populations of Acacia pendula is desirable to better
understand the taxon and to inform its management, so
that government policies and conservation actions can be
effectively implemented.
An opportunity arose to study one stand of the Hunter
Valley population of Acacia pendula in detail following an
observation of flower buds on plants in early 2018, a stage
that few stands of the species in the region experience (Bell
et al 2007). Importantly, successful fruiting and development
of seed has never been observed in Acacia pendula in the
Hunter Valley (critical for conclusive identification),
promoting the hypothesis that plants here may be neotonous
(retaining juvenile features in the adult phase) or may have
lost the capacity for seed production (dispersing instead
through vegetative suckers) in response to unfavourable
habitat (Bell & Driscoll 2014). This paper documents the
fate of a flowering event in Acacia pendula over a seven
month period in this population, and examines the current
age structure and health of individuals within the stand.
Conclusions reached on the reproductive output and general
health of these plants are considered in the context of future
management.
Study Population
Location and habitat
The study population lies near Broke (32° 45' 0.4" S, 151°
6' 7" E) in the Hunter Valley of New South Wales (Fig. 1).
This land has been established as part of the Weeping Myall
Management Area (WMMA) by Glencore (Bulga Coal), with
the central aim of conserving Acacia pendula and its habitat.
A monitoring program has been established to inform the
management of these plants to ensure that impacts associated
with nearby coal mining activities do not denigrate the site.
The 3.8 hectare WMMA was fenced and cattle removed in
March 2015, and only resident macropods (mostly Grey
Kangaroos, Macropus giganteus ) have grazed the site since.
At the time of fencing, twelve individuals of Acacia pendula
were known from the WMMA.
Currently, ten live individuals of Acacia pendula remain
within the WMMA (Fig. 2). Seven of the ten individuals
are old, well established trees, two are of medium size, and
one is represented only by young suckering plants. Most of
the older trees have fallen (some comprising two or more
trunks) but persist as living plants, their heavy limbs now
supported by the ground (Fig. 3). Two of three senescing
plants are represented only by vigorous suckering from
rootstock following trunk collapse, the third has shown no
such suckering and appears dead.
Fig. 1 : Location of the Weeping Myall Management Area
(WMMA) near Broke in the Hunter Valley, showing local relief
(contour interval is 10 m) and extent of landscape clearing.
The WMMA lies in largely cleared, undulating country at a
mid-slope position (110 m ASL), on Permian aged geology
(Fig. 4). Wollombi Brook, a major feeder stream to the
Hunter River, lies c. 2.5 km to the west and is separated by an
elevated Jurassic-aged basalt ridgeline housing the historic
Milbrodale trig station (c. 170 m ASL). Prior to European
settlement, the original vegetation across the study site, as
determined by a census of the larger remaining ‘paddock’
trees within a radius of 500 m, likely consisted of a grassy
woodland of Eucalyptus moluccana , with occasional
Eucalyptus crebra. There is conjecture as to the origins of
the Acacia pendula individuals on this site (and elsewhere
within the Hunter Valley region), given that their presence
in a grassy eucalypt woodland such as this runs contrary to
their occupied habitat elsewhere in inland eastern Australia
(see Bell & Driscoll 2014).
Cunninghamia 18: 2018
Bell , Flowering fate in Acaciapendula , Hunter Valley
81
Fig. 2: The study population of Acacia pendula in the Weeping
Myall Management Area, showing inspection locations on
individual trees.
Acacia pendula within the WMMA conform to morpho-
type B of Bell and Driscoll (2014), represented by plants
with green foliage, slightly pendulous branches on older
specimens, flowering irregularly but rarely if ever proceeding
to the fruiting stage, and commonly root-suckering.
Monitoring of these plants following the exclusion of cattle
grazing in March 2015 has shown an eruption of new shoots
emanating from roots (‘root-suckers’, commonly mistaken
by some observers as new recruits). Over the course of just
two years, the number of stems of Acacia pendula rose from
12 in 2013 to 685 in 2015, a 57-fold increase following
fencing of the WMMA (visible in Figs 3 and 16). In time,
these root-suckers develop into a dense thicket of vegetation
shading out a large proportion of native grasses and herbs,
and is currently the subject of a separate study.
Land use history
Prior to establishment as a reserve for the protection of
Acacia pendula , the area formed part of an extensively
cleared and modified agricultural landscape. As early as
1821, cattle agistment was granted by land owner Benjamin
Singleton for the wider Patrick Plains area, with grazing by
cattle and sheep centred on the nearby township of Broke
(7 km to the south-east). In the 1850s, subdivision of the
land fronting Wollombi Brook began, with partial clearing
to accommodate grazing and dairying enterprises (Umwelt
2012). These pursuits remained the sole use of the land for
the next 150 years, whereupon properties were purchased for
coal mining or biodiversity offsets.
Fig. 3: Two fallen stems of a single individual of Acacia pendula
(rooted in the centre at the position of observer), showing the
canopy of each at extreme left and right.
Climatic conditions
Rainfall leading up to the flowering event in January 2018
was well below average for an extended period of time
(Fig. 5). Apart from above average falls in the March and
October of 2017, little rain fell for the thirteen months prior
to flowering (December 2016 to December 2017). Over
the course of monitoring (January to July 2018), rainfall
remained below the long-term average.
. Moving Average (1960 - 2018)
300
2015 2016 2017 2018
Fig. 5: Rainfall received at Bulga (3.5 km to the west) over the
two years prior to flowering. Budding was first observed in January
2018 (* Jan.), following a prolonged dry period. Data source:
Bureau of Meteorology (2018).
Fig. 4: Landscape context of the Weeping Myall Management
Area (fenced area in middle distance, below and within remnant
Eucalyptus moluccana woodland).
82
Cunninghamia 18: 2018
Bell , Flowering fate in Acaciapendula. Hunter Valley
Methods
Flowering inspections
Eight individuals of Acacia pendula were selected for
monitoring. For each monitored plant, two observation
points were designated that were accessible and where most
flower buds were evident at the commencement of the study.
As far as possible, inspection points strived to include one
receiving high sun exposure and a second receiving low sun
exposure, but this was not always possible and was dictated
by the extent of flowering on each individual (see Fig. 2).
Inspections commenced in late February 2018 and continued
monthly until the end of July 2018.
At each monthly inspection, general observations were made
pertaining to the proportion and health of buds and flowers,
and the presence or otherwise of developing pods. Tagging of
specific inflorescences for more regimented monitoring was
not undertaken as previous experience had shown high failure
rates during flowering in this species, and observations of a
more general nature were more likely to gather useful data.
Additionally, notes were also made on the extent of flower
and leaf galls, and activities of ants and other invertebrates.
The presence of buds, flowers and fruits at each inspection
point were assigned to one of four numerical categories:
0 (none present), 1 (few present, < 25 visible), 2 (many
present, 25-100 visible), 3 (numerous present, >100 visible).
Buds and flowers were considered viable and healthy if they
were yellow and not dry and ‘crispy’, with no visible signs
of galling or flower desiccation. Categorical data on bud and
flower presence were averaged across the sixteen inspection
points to graphically summarise the progress of flowering
over the monitoring period.
Acacia age and health
In the absence of more definitive, non-destructive methods,
the assessment of the age of individual Acacia pendula trees
used stem diameter as a surrogate. The diameter-at-breast
height was consequently measured on all Acacia plants within
the study population (n=12, incorporating both live and dead
individuals). In cases where more than one trunk was evident,
all were measured but only the largest was used in analyses.
For collapsed individuals that lay across the ground surface
but remained alive, diameter was measured at approximately
1.7 m above the rooted point of the main trunk. Root suckers
were too numerous to measure, and were ignored.
The presence of aerial mistletoe shrubs can impact on the
general health and vigour of host species (Reid et al 1994;
Watson 2011). In the case of Acacia pendula , the number of
mistletoe clumps ( Amyema quandang ) was counted on each
study plant to allow general observations on whether or not
their presence appeared to influence the progress of flowering.
Results
Flowering phenology
Following initial observations of flower buds in early 2018
(Fig. 6), anthesis occurred from March (Fig. 7) but rapidly
declined. There was a steady decline in both the number and
health of inflorescences over the subsequent six months to
July, where no active buds or flowers were evident (Fig. 8).
In June, a small number of fresh buds were observed on some
trees, suggesting that a second flush of flowering may occur
but subsequent observations revealed otherwise. Flowering
(open buds) peaked in March but then also underwent a
decline to June, and none were present in July. No flowers
were observed to progress to the fruiting stage, and no pods
were recorded on any monitored tree. Rainfall during this six
month period was below the long-term average, with April
and May particularly well below average.
All monitored trees displayed at least some bud and flower
development over the course of the study. Representative
flowers sampled for microscopic examination appeared
healthy and properly developed (Fig. 9), but over time these
either senesced due to ongoing dry conditions (Figs. 10 &
11), or transitioned into galls. The majority of galls were
found to be the result of infestation by the Common Flower
Galler ( Dasineura glomerata ) (Fig. 12), and represents the
first time Acacia pendula has been recorded as a host for this
species (R Kolesik pers. comm.). Previously documented
host species include Acacia deanei, Acacia elata, Acacia
hakeoides, Acacia mearnsii, Acacia melanoxylon, Acacia
pycnantha, Acacia retinoides and Acacia schinoides
(Kolesik et al 2005). Other galls present on inflorescences
were attributable to bud galler (Asphondylia sp.) (Fig. 13),
although these appeared less prevalent than Dasineura. On
some flowers, woody, bulbous structures attributable to
fungal gall ( Uromycladium sp.) were also observed (Fig. 14).
It is unknown if any individual flowers were successfully
pollinated during this flowering event, but if so none
proceeded to develop pods.
Fig. 6: Budding Acacia pendula (photographed 28 February 2018).
83
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Bell , Flowering fate in Acaciapendula , Hunter Valley
Fig. 7: Flowering Acacia pendula , approaching anthesis
(photographed 27 March 2018).
3
Feb-18 Mar-18 Apr-18 May-18 Jun-18 Jul-18
Buds — — Flowers Fruit
Fig. 8: Schematic summary of flowering fate of eight monitored
trees over a six month period in 2018. No seed or fruit was
produced. Reproductive Stage: 0 = none present, 1 = few present
(< 25 visible), 2 = many present (25-100 visible), 3 = numerous
present (>100 visible).
Fig. 9: Capitulum of Acacia pendula at anthesis, showing healthy
stamens (photographed 28 February 2018).
Fig. 10 Acacia pendula inflorescence, showing desiccating
capitula and partial dislodgement of stamens (upper capitulum)
(photographed 31 May 2018).
Fig. 11: Acacia pendula inflorescence, showing complete
dislodgement of stamens from each capitulum, and no development
of pods (photographed 31 May 2018).
Fig. 12 Acacia pendula capitula freshly infected by galls of
Dasineura glomerata , showing remnants of anthers and filaments
between individual gall chambers (photographed 24 April 2018).
84
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Bell , Flowering fate in Acaciapendula. Hunter Valley
Fig. 13: Acacia pendula inflorescence, showing capitulum infected
by bud galler ( Asphondylia sp.) (upper left) and newer, healthy
capitula (right) (photographed 31 May 2018).
Fig. 14: Fungal gall ( Uromycladium sp.) on Acacia pendula
capitulum, showing its woody texture (photographed 24 April
2018).
Acacia age and health
Across the study population, the average size of Acacia
pendula trees was 50 cm DBH (diameter-at-breast height),
with a standard deviation of 16.2 cm (n=12). The smallest
tree was 23.2 cm DBH and the largest 82.8 cm DBH
(Fig. 15). Three of the twelve individuals (including live and
dead plants) possessed two trunks, while a further three had
completely fallen trunks (two with copious root suckering)
and lay across the ground (Fig. 16). One individual comprised
a fallen trunk only with no root suckers and has presumably
died, while another showed post-collapse development of
roots from its trunk where it lay along the ground (Fig. 17).
The large girth of trees within the study area is of some
interest, as Boland et al (2006) described Acacia pendula
with a diameter-at-breast height of “up to 30 cm”, nearly one
third of the size of the largest specimen measured here. The
large size of Acacia stems within the study area may explain
why many of them have fallen over but continue to grow
while supported on the ground.
Based on reported growth rates of the related Acacia salicina
elsewhere (Grigg & Mulligan 1999; Jeddi & Chaieb 2012),
the estimated age of individual Acacia pendula trees is
likely to be between 50 and 150 years. The inferred age of
individuals did not appear to influence the extent and success
(albeit limited) of flowering, as buds and flowers were
observed across all eight study trees ranging between 23 and
67 cm DBH.
90
DBH .Mean
123456789 ABC
Acacia pendula individual
Fig. 15 Diameter-at-breast height (DBH) of Acacia pendula
individuals within the study area. Only individuals 1-8 were the
subject of flower monitoring; individual #9 supported an elevated
canopy and was not monitored, while individuals A-C (also not
monitored) were collapsed plants with vigorous (#A-B) or no (#C)
root suckering.
Fig. 16 Aerial view of Acacia pendula (same individual as Fig. 3),
showing fallen but still alive trunks (crowns at far left and far right),
and copious root suckering in and around the centre.
Fig. 17: New root development mid-way along the collapsed trunk
of Acacia pendula.
Cunninghamia 18: 2018
Bell , Flowering fate in Acaciapendula , Hunter Valley
85
The extent of mistletoe ( Amyema quandang) growth on
Acacia pendula plants ranged from 2-38 clumps/tree with
a high degree of variance (n=9, median of 5, mean of 10.8,
SD of 11.7). There was no correlation between the size of
Acacia trees and the number of mistletoe clumps supported
on them. However, general observations suggest that those
trees with higher densities of mistletoe appeared in poorer
overall health than those with few mistletoes (Fig. 18).
Many trees also displayed evidence of attack by both Bag
Shelter Moth ( Ochrogaster lunifer, family Thaumetopoeidae)
and galling-thrips (family Phlaeothripidae). Larvae of
Ochrogaster lunifer feed on Acacia phyllodes and, in some
cases, can completely defoliate a tree (Floater 1996). Large
silk nests are formed in the canopy (Fig. 19), comprising
Acacia phyllodes and silk produced by the larvae, and are used
for resting during daylight hours. Galling-thrips also attack
the phyllodes of Acacia , producing galls (Fig. 20) which
extensively modify the shape and form of phyllodes (Crespi
& Worobey 1998; Morris & Mound 2002). Galls present on
Acacia within the study population appear attributable to
Kladothrips rugosus , and although not extensive are present
on most trees. All of these invertebrates are native Australian
species and form part of the natural ecosystem in which
Acacia pendula occurs.
Fig. 18 Mistletoe infested Acacia pendula showing signs of stress
and death of limbs.
Fig. 19: Bag shelter produced by larvae of Ochrogaster lunifer in
the branches of Acacia pendula (photographed 24 April 2018).
Fig. 20: Gall produced by Kladothrips rugosus on the phyllodes of
Acacia pendula (photographed 27 March 2018).
Discussion
A rare flowering event in a population of Acacia pendula in the
NSW Hunter Valley failed to progress to fruiting, suggesting
that at least in the short-term persistence at this location is
reliant on asexual reproduction. The 2018 flowering event
was the first in that population since at least 2015 (when
monitoring began), and such irregularity is reportedly a trait
consistent with many other stands of the species in the region
and throughout its range (Tame 1992; Boland et al 2006; Bell
et al 2007). Fencing and the cessation of cattle grazing at
the site in 2015 has been followed by copious emergence
of root-suckers from nearly all individuals, yet evidence of
successful seed production and subsequent new recruitment
remains absent. As with other Hunter Valley stands of this
species, long-term survival is likely to be contingent on the
appropriate management of stock grazing pressures.
What events lead to the failure of fruit production in Acacia
pendulad Within the study population during 2018 this
appears primarily attributable to infestations of galling
insects, accompanied by flower desiccation due to dry
conditions. Gall-forming midges of the Dasineura and
Asphondylia genera (family Cecidomyiidae) deposit eggs in
the open flowers of Acacia , typically within the perianth tube
near the ovary. On hatching, larvae then induce the ovary to
evaginate and form a number of chambers, so that in some
cases entire flower heads can transform into clusters of galls
(Kolesik et al 2005; Kolesik 2015). This process results in
the loss of flowering material, and hence reproduction in that
inflorescence has effectively ceased. In other areas of New
South Wales and South Australia, Dasineura glauca (Grey
Fluted Galler) reportedly often occurs at such high densities
that seed production is completely prevented in entire Acacia
pendula trees (Kolesik et al 2005).
Dasineura glomerata (Common Flower Galler) is prevalent
within the study population and appears likely to persist
there permanently while ever the host plant remains.
Dasineura glomerata has not been recorded infecting Acacia
pendula previously, and represents a new host tree record
for the species (P. Kolesik pers. comm.). Other known hosts
for Dasineura glomerata included several Acacia species
distributed mainly in coastal and near-coastal locations,
86
Cunninghamia 18: 2018
Bell , Flowering fate in Acaciapendula. Hunter Valley
including Acacia deanei, Acacia data, Acacia hakeoides,
Acacia mearnsii, Acacia melanoxylon, Acacia pycnantha,
Acacia retinoides and Acacia schinoides. None of these
occur in the immediate locality of the study area, although
Acacia deanei and Acacia hakeoides are present further
west in the upper Hunter Valley (c. 60 to 90 km away from
the study area), and Acacia data, Acacia melanoxylon
and Acacia schinoides occur in the adjacent mountainous
districts. Only Dasineura glauca is known to infest Acacia
pendula (Kolesik et al 2010), with other hosts for this
species including the closely related Acacia omalophylla. A
similar but undescribed gall-midge occurs on other semi-arid
Acacia species, such as Acacia aneura and Acacia ramulosa
(Kolesik et al 2005).
Individuals of Acacia pendula in the study population are
also infected (to a lesser degree) by an undetermined species
of a second gall-midge, Asphondylia sp., and a fungal
gall, Uromydadium sp. (Pileolariaceae). In some Acacia
populations, rust disease caused by Uromydadium poses a
severe threat to the health and survival of infected individuals
(e.g. McTaggart et al 2015), although at present this does not
appear to be the case within the study population. Phyllodes
are similarly attacked by the larvae of Ochrogaster lunifer
(Thaumetopoeidae) and the galling-thrip Kladothrips
rugosus (Phlaeothripidae). All of these invertebrates are
native Australian species and form part of the natural
ecosystem in which Acacia pendula occurs. When host
plant species are under stress, such as brought about through
habitat modification, infestations can severely impact nonnal
growth and reproduction. Where Australian Acacia species
have become invasive in other parts of the world, deliberate
introduction of similar insects has been trialled as a biological
control to limit spread (e.g. Impson et al 2008).
Flower desiccation due to dry conditions is a common reason
for failure to reproduce in any one season (e.g. Anjum et al
2011). This phenomenon was also suspected to be occurring in
the study population of Acacia pendula which was regularly
under water stress with below average rainfall, despite
reasonable falls in February, March and June. These falls did
not, however, ensure the retention of flowering material on
branches, and for those inflorescences not affected by galls the
shedding of stamens to leave ‘bald’ capitulas soon followed.
It was not possible to quantify the extent to which flower
desiccation affected the overall potential for pollination and
seed production, but this is suspected to be high. In a Western
Australian study, Gaol and Fox (2002) noted that good winter
rainfall was necessary to induce flowering in several Acacia
species, but that further rain after flowering promoted pod
development and seed production. For the Acacia pendula
plants under study in the Hunter Valley, the abortion of
flowering and the lack of pod production occurred despite
rainfall in February, March and June.
Although plausible, an absence of pollinators is difficult to
advance as a primary cause of flower failure. Most Acacia
species are self-incompatible, and the transfer of pollen
between individuals and populations via pollinating vectors
is crucial for outcrossing and seed set (Stone et al 2003).
For the bulk of Acacia species, this involves unspecialised,
generalist insects (Tybirk 1997). Pollinators of Acacia
pendula are thought to comprise small native flies, bees and
wasps (Bernhardt 1987), all of which are likely to travel over
considerable distances visiting multiple stands of flowering
plants. Given that the landscape surrounding the study
population has been heavily cleared of native vegetation for
at least 150 years (now fitting the fragmented or relictual
states of McIntyre & Hobbs 1999), it is possible that the
necessary pollinating invertebrates have also declined or
disappeared (Kearns et al 1998). Many co-occurring Acacia
species flower simultaneously, and in such cases such an
event serves to attract a number of pollinators which are
shared between species. In heavily modified landscapes,
co-occurring species are often absent leading to a lack of
co-flowering between species, and the threshold needed to
attract pollinators may therefore not be reached. Apart from a
single individual of Acacia salicina, there are no co-occurring
Acacia within the study population, nor in the immediate
vicinity (although good stands of Acacia filicifolia do occur
1 km to the north). Stone et al (2003) noted that populations
of Acacia reduced to relict populations may have already
lost their pollinator networks, resulting in lower seed set and
dependence on opportunist pollinators. This scenario could
also be extended to the study population of Acacia pendula,
but this requires further investigation.
Recruitment failure as a result of grazing pressure has been
documented for other arid-zone Acacia species (e.g. Batty
& Parsons 1992; Auld 1995), although for the study area
Acacia pendula impacts from grazing have affected the
regeneration of root-suckers. Where recruitment failure is
ongoing due to an absence of seed production, there can be
important implications for conservation and management.
For Acacia carneorum, Roberts et al (2017) found this
species to be almost entirely reliant on asexual reproduction
for persistence in an area, and that relatively few genetically
distinct individuals were present across its range despite the
often many thousands of stems in a stand. In that case, land
managers were encouraged to protect both vegetative root-
suckers and true seedlings from threats, as well as to use the
few stands that did produce viable seed to augment existing
populations through translocations. The lack of seed-
producing stands of Acacia pendula in the Hunter Valley
suggest that a similar recommendation for propagation
and translocation of local provenance material cannot be
promulgated unless genets originating from outside the
region are used. Such an action is not recommended given
uncertainty over plant origin in the Hunter Valley (Bell &
Driscoll 2014).
Forrest (2016) related flowering events and prolonged
recruitment failure from grazing impacts to rainfall
patterns for several arid-zone Acacia. He found successful
reproduction occurred in at least one of the two consecutive
years following a La Nina wet period for the arid zone
species Acacia melvillei, Acacia homalophylla and Acacia
loderi. However, although these wet periods initiated sexual
reproduction in these species, other factors appeared to limit
success. Gaol and Fox (2002) earlier suggested that a wet
winter period was required to induce flowering in some
Acacia, and that follow up falls were necessary to ensure
Cunninghamia 18: 2018
seed production. For the study population of Acacia pendula ,
the 2018 flowering event occurred two years after a very wet
three-month period from November 2015 to January 2016.
South-eastern Australia at this time was in the grip of an El
Nino event, and this wet period contrasted strongly with the
below average falls received at other times in 2015 and 2016
(refer Fig. 5). Flowering in 2018 was therefore potentially a
response to the wet period two years earlier, although without
additional data on flowering phenology prior to 2015 this
remains conjecture. Apart from this event, examination of
rainfall data in the period leading up to flowering shows no
clear pattern or spike in rainfall that may have triggered the
2018 flowering event. Winter rainfall was below average in
2017 prior to the documented flowering event, but largely
above average in 2016 where no flowering was observed.
High mistletoe density on some Acacia pendula within the
study area is impacting on the health and vigour of these
plants, but desiccation and gall-infestation of flowers was
consistent across all study trees, irrespective of the number
of mistletoe clumps. However, some trees appear to have
suffered branch death as a result of high mistletoe densities.
Modification to landscapes associated with agricultural
activities are known to increase the density of mistletoes
(e.g. Bowen et al 2009; Watson 2011), as the availability of
perches for avian vectors becomes greatly reduced. In other
studies, mistletoes have been implicated in rapid turnover
and increased mortality of host trees (e.g. Reid et al 1994;
Reid & Stafford Smith 2000), although susceptibility is not
universal (e.g. MacRaild et al 2009).
The general poor health and flowering displayed by Acacia
pendula within the study population and elsewhere in the
Hunter Valley are perhaps symptomatic of wider implications
following extensive landscape modification. The study
population lies on land that has been largely cleared for
grazing purposes for many decades. Henry Hangar’s 1828
map of the Hunter River area shows the WMMA to be “open
forest country, deep loam soils occasionally intersected by
scrubs & ill watered ’ (Umwelt 2012), but by 1850 subdivision
and clearing of the land for grazing purposes had begun.
Pastoralism was the first industry established in this part of
the Hunter Valley, and in the nineteenth century the Broke
area was a centre of pastoral interests based on sheep and
cattle grazing (Umwelt 2012). Progressive removal of canopy
and shrub species would have occurred during this period to
increase the carrying capacity of the land for agriculture and
grazing. Such modification to landscapes, with the inherent
fragmentation of habitats that ensues, often leads to extinction
cascades when the loss of key species in an ecosystem triggers
the loss of other species (Fischer & Lindenmayer 2007). It
is possible that such removal of key structural and floristic
components of the fonner Eucalyptus moluccana woodland
over an extended period of time may have led to the poor
health and reduced sexual reproduction currently evident in
the Acacia pendula population at the WMMA.
In any case, Acacia pendula trees within the study population
are evidently subject to a number of stressors which affect
successful and ongoing recruitment. These include but are not
limited to infection by various flower- and phyllode-galling
Bell , Flowering fate in Acacia pendula , Hunter Valley 87
midges and thrips {Asphondylia sp., Dasineura glomerata,
Kladothrips rugosus), fungal galls ( Uromycladium sp.),
caterpillars ( Ochrogaster lunifer ), and mistletoe ( Amyema
quandang), together with stress brought about through
drought and other climatic extremes. Pollinator absence or
decline may also be imposing a different stress on the trees,
but as of yet there is no data to confirm this. The absence of
any old seed pods beneath all ten of the study trees suggests
that these stressors have been operating on and limiting
recruitment in them for many years, and that persistence in
the area relies solely on asexual reproduction. Arguably, all
of these stressors are a result of, or are exacerbated by, a
highly modified and cleared landscape, and their collective
impacts raise serious questions over how the species can
remain viable in such a habitat into the future. Exclusion of
stock grazing from Acacia populations may be feasible in
the short-term at some locations, but management of grazing
pressures for the benefit of Acacia is uncertain in the long¬
term, particularly during times of drought when all lands are
subject to increased pressure to feed hungry stock.
Such a predicament for Acacia pendula has serious
implications for conservation management, both here and
in the wider Hunter Valley region if the patterns observed
in the study population are repeated at other stands. This is
particularly so in regard to conservation actions that require
the augmentation of existing stands through translocation
or supplementary planting. With no seed produced,
augmentation planting can only rely on propagation from
cutting material which re-distributes the existing poor
genetic base. Alternatively, propagation using seed sourced
from horticultural specimens (morpho-type A in Bell and
Driscoll 2014) will introduce new genetic material into the
region, a situation that is unfavourable given conjecture over
the origin of existing plants. If Acacia pendula is ultimately
shown through genetic studies to be a natural component
of the contemporary Hunter Valley landscape, it remains
unclear why such a disjunct population of the species occurs
and persists in seemingly inhospitable habitat well east of its
accepted geographical range. Hypotheses around its presence
as a relict population from a previous drier climate regime
(e.g. DEWHA 2009; OEH 2013), which may help explain
the root-suckering habit, require further investigation.
Acknowledgements
Thanks to Glencore (Bulga Coal) for financial support of this
study, Peter Kolesik (Bionomics) and Robin Adair (Australis
Biological) for assistance with the identification of galls, and
Tom Scott (Glencore) for review of an earlier version of this
paper. Review comments from Doug Benson, Phillip Kodela
and the publication committee improved the manuscript and
are greatly acknowledged. Remotely Piloted Aircraft use
was undertaken by Phil Lamrock (Cormal Environmental)
in accordance with part 101 of the Civil Aviation Safety
Regulation and following notification to the Civil Aviation
Safety Authority. Permission to capture aerial imagery on
their land for this study was granted by Glencore.
88
Cunninghamia 18: 2018
Bell , Flowering fate in Acaciapendula. Hunter Valley
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Manuscript accepted 5 October 2018
Date of Publication:
November 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
Managing Persoonia (Proteaceae) species in the landscape through a
better understanding of their seed biology and ecology
Nathan J. Emery 1 * and Catherine A. Offord‘
1 The Australian PlantBank, the Royal Botanic Gardens and Domain Trust,
the Australian Botanic Garden Mount Annan, NSW 2567 AUSTRALIA.
* Corresponding author nathan.emery@rbgsyd.nsw.gov.au
Abstract : Persoonia (family Proteaceae) is a diverse genus of 99 species, mainly of woody shrubs and small trees,
that are endemic to Australia. The fleshy fruits that characterise these plants are an important resource in otherwise
resource-depauperate ecological communities. However, this genus is highly under-represented in restoration and
conservation programs, as its species are notoriously difficult to propagate and transplant in the wild. Understanding
the mechanisms that control seed production, viability, dormancy and germination will hasten progress on Persoonia
propagation. Here we review Persoonia studies to reveal the nature of, and variability within, the genus. We identify
key factors that need to be addressed; specifically, those affecting fruit set, endocarp degradation and subsequent
propagation of Persoonia. This synthesis of current knowledge provides important material to inform management of
this taxon in the landscape, and outlines several important priorities for future seed biology research on the genus. We
outline several important priorities for future seed biology research on the genus.
Keywords: Endocarp, germination, plant ecology, propagation, restoration, seed production
Cunninghamia (2018) 18: 089-107
doi:10.7751/ctmninghamia.2018.18.006
Cunninghamia : a journal of plant ecology for eastern Australia
www. rbgsy d. ns w. gov. au/science/sc lentific-publ ications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
90
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
Introduction
Success in many natural revegetation and restoration
projects depends on the establishment of a wide range
of species, but is often limited to those that are easy to
collect, propagate and establish. Proteaceae are important
keystone species in many restoration projects (Roche et
al., 1997; Koch, 2007b; Stingemore & Krauss, 2013), be it
at the landscape scale, with species of Banksia, Grevillea
and Hakea being commonly used, or the translocation of
a single threatened species (e.g. Persoonia panciflora ).
Seed production in woody-fruited Proteaceae can vary
significantly, from zero to tens of thousands of seeds on
a single plant in any given season (Groom & Lamont,
1998). Seeds of some Proteaceae species are relatively
easy to germinate but because many species occur in fire-
prone habitats, germination may be cued to fire, and can
be difficult to germinate due to specific conditions being
required to break the complex dormancy mechanisms (Van
Staden & Brown, 1977; Morris, 2000; Morris et al., 2000;
Arnolds et al., 2015; Chia et al., 2016).
Persoonia is one genus within the Proteaceae that has very
complex dormancy mechanisms. This genus includes many
species that are the subject of restoration or conservation
projects and include Persoonia longifolia (southern Western
Australia), Persoonia panciflora (Hunter Valley, NSW),
Persoonia hindii (Newnes Plateau, NSW) and Persoonia
hirsnta (Sydney Basin, NSW). However, their inclusion
is hampered by consistently poor propagation success
(Cambecedes & Balmer, 1995; Ketelhohn et al., 1996; Bauer
& Johnston, 1999). For example, Persoonia longifolia plants
were present in areas of Western Australian jarrah woodland
prior to bauxite mining in the 1960s (Koch 2007a; 2007b).
The mine site restoration plan for these areas included a
target of restoring the ecosystem to a state comparable with
pre-mining, but although Persoonia longifolia had viable
seeds, they could not be germinated and were absent from
rehabilitation projects (Koch, 2007b). Recent research on the
seed ecology of this species (Chia et al., 2015), including
how seasonality and fire affect in situ fruit set, dormancy
release and germination, identified the seasonal conditions
and length of time required for donnancy to break and
germination to occur in the soil (Chia et al., 2016).
We see a focus on plant ecology to be of great value for
progress in conservation and restoration. It is evident that
re-establishing new populations, or augmenting extant
populations through translocation, requires a detailed
understanding of how the plants interact in the natural
environment, if they are to have long-term success. We argue
that progress on Persoonia seed production and propagation
will be facilitated by understanding the mechanisms in
nature that control seed production, viability, donnancy and
germination. This review aims to identify major research
priorities and develop a logical framework to guide future
investigations towards a more systematic approach to resolve
species persistence in the landscape. Such an approach could
also have benefits for other Proteaceae with a similar seed
biology to Persoonia , and may be relevant to other families
with similarly deep, intractably dormant seeds.
The genus Persoonia
Species of Persoonia (family Proteaceae) range from low
prostrate or spreading shrubs to small trees (Appendix 1);
they are characterised by light green foliage with high
morphological variability across species, and yellow
hermaphroditic flowers (Weston 2003; Fig. 1). A maturing
ovule forms a fleshy drupaceous fruit comprising a single
hard woody stone containing either one or two seeds (Fig. 2),
and an embryo can have up to nine cotyledons (Weston,
2003). All 99 species are endemic to Australia, and together
the genus occupies 64 of the 87 national bioregions (IBRA7;
Appendix 1), but is largely absent from central arid regions.
Including subspecies there are 115 taxa, 72 endemic to
eastern Australia, 42 to Western Australia, and one across
northern Australia. Nine species are listed as threatened,
endangered or critically endangered under the Australian
Environment Protection and Biodiversity Conservation
Act, 1999 (eight species occur in NSW; Appendix 1), being
impacted by several anthropogenic factors, including land
clearing, mining, habitat fragmentation, grazing, slashing
and predation. There is a distinct lack of ecological and
seed biology data on Persoonia ; much of this research has
been conducted on eastern Australian Persoonia species
(Table 1).
Fig. 1. Morphological variation within the Persoonia genus.
A- Persoonia myrtilloides plant; B- Persoonia fevis plant;
C- Persoonia hirsuta plant, and; D- Persoonia panciflora plant
(Photos by N. Emery).
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
91
Table 1. Persoonia species that have been studied for breeding system, fruit set and/or seed germination. Several species have had
multiple independent studies that measured one or more component of the reproductive niche (see Appendix 2 for References)
Species
Distribution (state)
Habitat 1
Rarity 2
Fire
response 3
Breeding
System 4
Fruit set 5
Germination 6
Persoonia bargoensis
NSW
DSF, DW
En*, Vuf
1
SC
Persoonia elliptica
WA
DSF, DW
NL
3
39%
Persoonia glancescens
NSW
DSF
En*, Vuf
1
NC
18%; 86%
Persoonia juniperina
NSW, SA, TAS, VIC
DSF, H
NL
2
SC
40%; 30—41.4%
Persoonia lanceolata
NSW
DSF
NL
1
NC
41%; 88%; 97%
-10.0%
Persoonia levis
NSW, VIC
DSF, DW
NL
3
NC
5—55%; 52%
50.0%
Persoonia longifolia
WA
DSF, DW
NL
3
10%; 98.3%
31.8—94.7%
Persoonia mollis subsp. maxima
NSW
DSF
En*f
1
NC
18%; 89%
Persoonia mollis subsp. nectens
NSW
DSF
NL
1
NC
35%; 91%
-40.0%
Persoonia myrtilloides
NSW
DSF, DW
NL
1
NC
10—70%
Persoonia rigida
NSW, VIC
DSF
NL
1
SC
67%
Persoonia sericea
QLD, NSW
DSF, WSF
NL
2
87.5%
Persoonia virgata
QLD, NSW
DSF
NL
1
NC
36.1—41.6%;
48.9%
50%; 58.8—
87.5%; 100%
1 DSF = dry sclerophyll forest, DW
= dry woodland, H =
heath, WSF =
wet sclerophyll forest
2 En = endangered, NL = not listed under state or national legislation, Vu = vulnerable; * threatened status listed under state/territory
legislation; t threatened status listed under the national EPBC Act
3 Ability to resprout following fire: 1 = cannot resprout and reliant on seeds, 2 = can resprout from base only. 3 = can resprout from base and
stems; Rymer (2006)
4 NC = non-compatible breeding system, SC = self-compatible breeding system; reference list available in Appendix 2
5 Fruit set from outcrossed pollination reported in the literature as of 23/04/2018; reference list available in Appendix 2
6 Germination results reported in the literature as of 23/04/2018; reference list available in Appendix 2
Field ecology behaviour
Some Persoonia species occur across multiple climatic
zones, but others are more localised including several of
the rarer obligate-seeding species in the Sydney region, and
these species often establish alongside roads and tracks. For
example, Persoonia hirsuta plants occur in drainage lines
along track edges with the largest populations along disturbed
road easements (N. Emery, pers. obs. 2017). Myerscough
et al. (2000) postulated that soil disturbance events might
substitute for the effects of fire, particularly in environments
with long inter-fire intervals.
Persoonia species lack the proteoid roots (characteristic
of most Proteaceae species) that aid inorganic nutrient
absorption, yet plants often occur in well-drained, nutrient-
poor acidic soils such as the sandstone and shale soils of
the Sydney region (Myerscough et al. 2000; Weston 2003).
As individual plants can thrive in their environments it is
possible that unknown mycorrhizal associations might occur
in the roots. Persoonia pauciflora plants often occur at
the base of Broad-leaved Ironbark ( Eucalyptus fibrosa) or
Spotted Gum ( Corymbia maculata ) trees (N. Emery, pers.
obs. 2017), which might indicate a possible relationship
between these species.
Flowering times
Peak flowering in most Western Australian Persoonia species
occurs over winter and spring, and eastern Australian species
predominantly flower during summer and autumn (Bernhardt
& Weston 1996; Table 2). Some species such as Persoonia
pinifolia can produce flowers for most of the year. Eastern
Australian Persoonia species experience a high frequency of
hybridisation (Myerscough et al. 2000) that could be explained
by a combination of coinciding distributions, flowering times
and/or pollinators, as well as a lack of pre-zygotic barriers for
interspecific pollen (Bernhardt & Weston 1996).
Pollination
Many Proteaceous species produce large inflorescences
with copious amounts of nectar, making them well-suited
for vertebrate pollination (Carolin, 1961; Collins & Rebelo,
1987). Persoonia species, in contrast, have small yellow
flowers that are most notably pollinated by bees and wasps
(Carolin, 1961; Collins & Rebelo, 1987; Bernhardt &
Weston, 1996). It was originally postulated that a mutualistic
relationship exists between Persoonia and small native, hairy
Leioproctus bees (Hymenoptera: Colletidae) (Bernhardt and
Weston 1996). Leioproctus have since been observed to
pollinate numerous Persoonia species including Persoonia
glancescens, Persoonia lanceolata , Persoonia mollis subsp.
maxima , Persoonia mollis subsp. nectens and Persoonia
virgata (Wallace et al., 2002; Rymer et al., 2005). Other
bee pollinators known to visit Persoonia include Exoneura
spp. (Bernhardt & Weston, 1996), Tetragonula carbonaria
(formerly Trigona) (Wallace et al., 2002), and the European
honeybee Apis mellifera (Bernhardt & Weston, 1996;
Wallace et al, 2002; Rymer et al., 2005; Chia et al, 2015).
92
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
Native bees are thought to be more effective pollinators than
the introduced Honey Bee Apis mellifera as they have been
observed to travel greater distances, and visit more flowers
across multiple plants (Rymer et af, 2005). Furthermore, Apis
mellifera may collect floral resources without pollinating the
flower (Paton, 2000).
Table 2. Peak flowering times of 115 taxa (including all 99 species) of Persoonia obtained from Benson & McDougall (2000) and
ABRS Flora of Australia Online (http://www.anbg.gov.au/abrs/online-resources/flora/). Taxa are arranged by the Australian state
or territory that the species mostly occurs in.
Peak flowering time (month)
Species
Distribution
(state)
winter
spring
summer
autumn
J
J
A
S
O
N
D
J
F
M
A
M
Persoonia acicularis
WA
Persoonia angustiflora
WA
Persoonia baeckeoides
WA
Persoonia biglanduJosa
WA
Persoonia bowgada
WA
Persoonia brachystvlis
WA
Persoonia brevirhachis
WA
Persoonia chapmaniana
WA
Persoonia comata
WA
Persoonia cordifolia
WA
Persoonia coriacea
WA
Persoonia cymbifolia
WA
Persoonia dillwynioides
WA
Persoonia elliptica
WA
Persoonia filiformis
WA
Persoonia Jlexifolia
WA
Persoonia graminea
WA
Persoonia hakeiformis
WA
Persoonia helix
WA
Persoonia hexagona
WA
Persoonia inconspicua
WA
Persoonia kararae
WA
Persoonia leucopogon
WA
Persoonia longifolia
WA
Persoonia manotricha
WA
Persoonia micranthera
WA
Persoonia papillosa
WA
Persoonia pentasticha
WA
Persoonia pertinax
WA
Persoonia pungens
WA
Persoonia quinquenervis
WA
Persoonia rudis
WA
Persoonia rufiflora
WA
Persoonia saccata
WA
Persoonia saundersiana
WA
Persoonia scabra
WA
Persoonia spathulata
WA
Persoonia striata
WA
Persoonia stricta
WA
Persoonia sulcata
WA
Persoonia teretifolia
WA
Persoonia trinervis
WA
Persoonia falcata
QLD, NT, WA
Persoonia amaliae
QLD
Persoonia iogyna
QLD
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
93
Peak flowering time (month)
Species
Distribution
(state)
winter
spring
summer
autumn
J
J
A
S
O
N
D
J
F
M
A
M
Persoonia prostrata *
QLD
Persoonia tropica
QLD
Persoonia adenantha
QLD, NSW
Persoonia cornifolia
QLD, NSW
Persoonia media
QLD, NSW
Persoonia sericea
QLD, NSW
Persoonia stradbrokensis
QLD, NSW
Persoonia tenuifolia
QLD, NSW
Persoonia terminalis subsp. recurva
QLD, NSW
Persoonia terminalis subsp. terminalis
QLD, NSW
Persoonia virgata
QLD, NSW
Persoonia volcanica
QLD, NSW
Persoonia acerosa
NSW
Persoonia acuminata
NSW
Persoonia bargoensis
NSW
Persoonia chamaepitys
NSW
Persoonia conjuncta
NSW
Persoonia curvifolia
NSW
Persoonia cuspidifera
NSW
Persoonia daphnoides
NSW
Persoonia fastigiata
NSW
Persoonia glaucescens
NSW
Persoonia hindii
NSW
Persoonia hirsuta subsp. evoluta
NSW
Persoonia hirsuta subsp. hirsuta
NSW
Persoonia isophylla
NSW
Persoonia katerae
NSW
Persoonia lanceolata
NSW
Persoonia laurina subsp. intermedia
NSW
Persoonia laurina subsp. laurina
NSW
Persoonia laurin subsp. leiogyna
NSW
Persoonia laxa*
NSW
Persoonia marginata
NSW
Persoonia microphylla
NSW
Persoonia mollis subsp. caleyi
NSW
Persoonia mollis subsp. ledifolia
NSW
Persoonia mollis subsp. leptophylla
NSW
Persoonia mollis subsp. livens
NSW
Persoonia mollis subsp. maxima
NSW
Persoonia mollis subsp. mollis
NSW
Persoonia mollis subsp. nectens
NSW
Persoonia mollis subsp. revoluta
NSW
Persoonia myrtilloides subsp. c unninghamii
NSW
Persoonia myrtilloides subsp. myrtilloides
NSW
Persoonia nutans
NSW
Persoonia oblongata
NSW
Persoonia oleoides
NSW
Persoonia oxycoccoides
NSW
Persoonia pauciflora
NSW
Persoonia pinifolia
NSW
Persoonia procumbens
NSW
Persoonia recedens
NSW
Persoonia rufa
NSW
94
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Emery & Offord, Managing Persoonia in the landscape
Peak flowering time (month)
Species
Distribution
(state)
winter
spring
summer
autumn
J
J
A
S
O
N
D
J
F
M
A
M
Persoonia subtilis
NSW
Persoonia subvelutina
ACT, NSW, VIC
Persoonia juniperina
NSW, SA, TAS, VIC
Persoonia asperula
NSW, VIC
Persoonia brevifolia
NSW, VIC
Persoonia chamaepeuce
NSW, VIC
Persoonia confertiflora
NSW, VIC
Persoonia levis
NSW, VIC
Persoonia linearis
NSW, VIC
Persoonia rigida
NSW, VIC
Persoonia silvatica
NSW, VIC
Persoonia arborea
VIC
Persoonia gunnii
TAS
Persoonia moscalii
TAS
Persoonia muelleri subsp. angustifolia
TAS
Persoonia muelleri subsp. densifolia
TAS
Persoonia muelleri subsp. muelleri
TAS
Manipulative pollination experiments suggest that Persoonia
species have a breeding system that favours out-crossing
(Krauss, 1994; Cadzow & Carthew, 2000; Wallace et al.,
2002; Table 1). In Persoonia mollis , for example, 20% of out-
crossed flowers set fruit compared to just 1% of selfed flowers
(Krauss, 1994). Furthermore, pollen tubes were not present
in the ovaries of self-pollinated flowers. The experimentally-
manipulated result for out-crossed flowers also reflected the
natural pollination level, with 17% of unmanipulated flowers
setting fruit (Krauss 1994). Similarly, Persoonia virgata
pollination experiments showed weak self-compatibility, as
fruit set was significantly lower in self-pollinated flowers
(6.6%) than cross-pollinated flowers (48.9%) (Wallace
et al., 2002). Krauss (1994) first noted the possibility of a
post-zygotic mechanism within Persoonia seeds that caused
the majority of selfed fruits to be prematurely terminated
between 4 and 30 weeks following pollination. In Persoonia
juniperina it was reported that 76% of selfed fruits terminated
during the maturation period compared to 33% for open-
pollinated fruits (Cadzow & Carthew 2000). Alternatively,
self-compatibility has been reported in Persoonia , as pollen
tubes were frequently observed in self-pollinated flowers
of Persoonia rigida, and final differences in the number of
matured fruit from self- and cross-pollinated flowers were
not statistically significant (Trueman & Wallace, 1999).
Self-compatibility has also been documented for Persoonia
juniperina and Persoonia bargoensis (Cadzow and Carthew
2000; Field et al. 2005).
Fruit set success
Flowers of Australian Proteaceae are hermaphroditic and
typically produce a very low rate of fruit set - around 5%
(Ayre & Whelan, 1989). Fruit set success varies considerably
among Persoonia species, and has been documented in
re-sprouting and obligate seeding species (Table 1). Fruit
set in Persoonia longifolia was reported to be more variable
among plants within a population than between populations,
and to be positively correlated with plant height and time since
last fire (Chia et al., 2015). The availability of carbohydrates
that could be transferred from branches to the fruits was
reported to be positively correlated with fruit size in Persoonia
rigida , and fruit set on leaf-bearing branches being 4-6 times
higher than defoliated branches (Trueman & Wallace 1999).
Minimal vegetative growth during fruit maturation was
observed on Persoonia virgata plants (Bauer et al. 2001).
The slow development of Persoonia embryos coupled
with the requirement of nutrient uptake for embryo growth,
suggests that most of the plant resources are allocated to fruit
development during the fruiting season (Strohschen, 1986).
Fruit maturation
Persoonia peak flowering and fruit set precedes a long and
highly variable period of fruit maturation reported to require
at least 2 months. In some species it can take up to a year
for fruits to fully mature and drop from the maternal plant
(Trueman & Wallace, 1999; Benson & McDougall, 2000;
Wallace et al., 2002; Weston, 2003; Rymer et al., 2005; Chia
et al., 2015). It has been reported that Persoonia pinifolia
embryo maturation is significantly slower than other
Proteaceae genera such as Macadamia , and the endospenn
is almost completely replaced by the embryo at 34 weeks
post-anthesis (Strohschen 1986).
Fruit dispersal
Seed dispersal beyond the maternal plant environment
may be limited to fruit-drop from the maternal plant (Rice
& Westoby, 1981), but the fleshy Persoonia fruits are
also likely to be consumed and dispersed by birds and
Cunninghamia 18: 2018
mammals (Weston, 2003; Auld et al, 2007). In one study,
90% of Persoonia lanceolata fruits were consumed by
Swamp Wallabies ( Wallabia bicolor) with 98% of these
still being viable after being collected from scats (Auld et
al., 2007). Persoonia longifolia fruits have been reported
to be consumed by Brush Tail Wallabies ( Macropus irma).
Western Grey Kangaroos (Macropus fuliginosus ) and
Bobtail Lizards (Tiliqua rugosa ) (Chia et al. 2015). Many
native birds consume Persoonia fruits, including the Olive-
backed Oriole (Oriolus sagittatus). Silver-eye ( Zosterops
lateralis ), Pied Currawong (Strepera graculina ), Regent
Bowerbird (Sericulus cfnysocephalus). Satin Bowerbird
(Ptilonorhynchus violaceus ), Red Wattlebird (Anthochaera
carunculata) and Lewin’s Honeyeater (Meliphaga lewinii )
(Barker & Vestjans, 1990). However, it is not known
whether these vertebrates facilitate dispersal of viable seeds
in their scats. Persoonia longifolia fruits are commonly
found in Emu ( Dromaius novaehoHandiae) scats, but the
germinability of these remains very low (Mullins et al.,
2002). Cockatoos and other parrots have also been observed
to predate on immature Persoonia fruit (Weston 2003;
K. Chia, pers. comm. 2016). The removal of Persoonia
seeds may correlate with the rarity and size of plants, as
macropods were found to remove significantly more fruits
of two common Persoonia species (Persoonia lanceolata
and Persoonia mollis subsp. maxima ) compared with two
rare species (Persoonia glaucescens and Persoonia mollis
subsp. nectens) (Rymer, 2006). Furthermore, seed removal
was significantly positively correlated with plant height in
common species only, although plant population size was
not reported to be influencing removal (Rymer, 2006).
Seed biology
Within the fleshy Persoonia exocarp and mesocarp is the
woody endocarp (Fig. 2), which restricts gennination as
a form of mechanical donnancy. In Persoonia longifolia
laboratory trials, germination only occurred when all or
half of the endocarp was removed (78% and 68% success,
respectively - Chia et al. 2016). Norman and Koch (2008)
determined that Persoonia longifolia endocarps were
permeable to water, (increasing in weight by 10-30%
following 30 hours of imbibition), but the permeability and
hardness of buried endocarps did not significantly differ
from the controls after a 2-year soil burial trial, suggesting
that endocarp weakening over time is slow. A recent study on
Persoonia longifolia noted that the removal of the endocarp
lid did not increase the rate of imbibition (Chia et al., 2016).
Coupled with the mechanical dormancy mechanism
imposed by the endocarp is the physiological dormancy of
the embryo, which may require treatment using a chemical
stimulant such as gibberellic acid (GA 3 ), or a combination of
warm and cold stratification to improve overall germination
success (Mullins et al., 2002; Chia et al., 2016). Mullins et al.
(2002) suggested that Persoonia longifolia seeds required an
unknown period of cold temperatures over winter to maximise
overall germination. By contrast, Persoonia myrtilloides and
Persoonia levis seeds showed significantly reduced and no
germination, respectively, following a chilling pre-treatment
Emery & Offord, Managing Persoonia in the landscape 95
(Nancarrow 2001). A recent study on Persoonia longifolia
suggested that the environmental conditions the endocarps
are exposed to are more important than the actual burial
time. Specifically, germination was highest when endocarps
were treated with two simulations of summer rainfall events
and a constant summer temperature of 30°C (Chia et al.
2016). Adding more complexity, a heat spike treatment
(50°C) improved germination when moisture was limiting,
but germination was significantly reduced if long wet cycles
were introduced.
It is possible that the proportion of physiologically dormant
seeds may be species-specific or vary among years depending
on conditions during fruit maturation. Some studies have
found that the addition of GA 3 made no difference or had
highly variable results, to overall germination (Ketelhohn
et al., 1996; Nancarrow, 2001; Chia et al., 2016). For
example, GA 3 increased gennination of Persoonia virgata
seeds by at least 50% (Ketelhohn et al., 1996; Bauer et al.,
2004). Similarly, it increased germination of Persoonia
levis seeds, whereas Persoonia myrtilloides seeds were
unaffected (Nancarrow, 2001). As seeds from Persoonia
myrtilloides only germinated after 4 months in storage, it
is possible that a period of after-ripening or stratification is
required to alleviate physiological dormancy. Furthermore,
as no indication of viability was given, it is also possible
that viable embryos of both species were damaged during the
removal of the endocarp.
Fig. 2. Diagrammatic cross-section of a typical Persoonia fruit
(not to scale), comprising mostly a fleshy mesocarp behind a
leathery external layer (exocarp). The mesocarp covers the woody
stone (endocarp), which protects the seed. Persoonia seeds are
predominately made up of a testa and non-endospermic embryo, and
may contain one or two seeds within the endocarp. The endocarp is
the key structure that prevents germination from occurring.
Early studies on Persooniapinifolia and Persoonia longifolia
noted that germination was negatively affected by microbial
contamination (McIntyre 1969; Mullins et al. 2002), but
the recommended disinfecting and germinating of seeds
under aseptic conditions, has produced mixed results for
germination success (Bauer et al., 2004; Chia et al., 2016).
Interestingly, contamination of Persoonia longifolia seeds
was most prevalent in those treated with the smoke stimulant
karrikinolide (Chia et al., 2016). Persooniapauciflora seeds
treated with GA, also suffered from severe contamination
despite being surface-sterilised (N. Emery, unpublished data).
Microbial growth within the seed could be promoted by both
GA 3 and karrikinolide as similar chemical derivatives have a
microbial origin (Brian et al., 1954; Fight et al., 2009).
96
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
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Fig. 3. An example of the framework for Persoonia longifolia seed
ecology research. Seeds have been tested for germination as part of
large mining rehabilitation projects (Abbott 1984; Koch & Ward
1994), conservation studies (Dixon et al. 1995; Norman & Koch
2008), and propagation studies (Mullins et al. 2002; Norman &
Koch 2008). Successful germination was not achieved until 2002,
when endocarps were physically compromised and seeds were
treated with gibberellic acid (GA 3 ). Germination remained low
(25 - 40%); however, there was an indication from direct sowing
that germination was higher following cooler temperatures. This
link with climate was then rigorously examined in recent studies on
the developmental phenology and the endocarp (Chia et al. 2015;
Chia et al. 2016). Seed germination was reported to be highest
following three years of soil burial, and, moreover, wet and dry
cycle length interacted with the rate of endocarp weakening. These
results now raise the possibility of fire or heat being used as a
management tool for endocarp weakening, and whether endocarps
degrade at a faster rate than seed viability in situ.
Soil seedbankpersistence
In the soil seedbank the endocarp is expected to decompose
over time, thereby increasing the rate of oxygen and water
reaching the embryo, and allowing the embryo to ‘push out’
of the weakened endocarp. Previous research has reported
mixed results of recruitment success in Persoonia and,
therefore, long-term persistence of seeds in the soil (Auld
et al. 2007; Ayre et al. 2009; McKenna, 2007; Chia et al.,
2015). Persoonia pinifolia fruit, for example, were estimated
to have a half-life of one year in the soil seedbank (Auld
et al., 2000). The viability of Persoonia elliptica seeds in
the soil seedbank declined from 39% to 5% after one year
(Nield et al., 2015). Persoonia longifolia seeds showed a
comparatively smaller decline in viability from 93% to 68%,
recorded after three years (Chia et al. 2016). In contrast, four
Persoonia species (two rare and two common) experienced
a significant decline in viability to around 30% following a
1-2 year soil burial (McKenna 2007). Interestingly, viability
decline also significantly varied among populations, which
could indicate local variation due to genetic or environmental
factors. However, such a decline in viability might not
adversely affect recruitment success where annual fruiting
events produce an accumulating seedbank. For example, 476
Persoonia mollis subsp. nectens seedlings emerged following
a wildfire from a population of 25 adult plants (Ayre et al.
2009). An additional 381 seedlings emerged following a
second wildfire four years later, and before the first seedling
cohort reached reproductive maturity (Ayre et al. 2009).
These results suggest that the population had a large and
persistent soil seedbank, capable of withstanding multiple
fires. Auld et al. (2007) estimated that although the number
of Persoonia lanceolata seedlings that emerged post-fire was
6-7 times greater than the pre-fire adult numbers, there was at
least 72% of available soil seedbank that did not germinate.
Natural recruitment: the role of fire and smoke
It was originally reported that seedling recruitment of
Persoonia is most likely to occur following a disturbance
such as fire (McIntyre, 1969). Fire was thought to be a crucial
factor for the recruitment of Persoonia elliptica seedlings
in jarrah woodlands in Western Australia, with only one
seedling observed in plots that had not been burnt (Nield et
al. 2015). With fire comes the risk of seeds being destroyed
by the combustion (Chia et al., 2015); fire intensity is likely
Cunninghamia 18: 2018
to be an important factor influencing seed survival, but has
not yet been investigated for Persoonia.
Auld et al. (2007) remarked that since most Persoonia
lanceolata fruits were found in the top 5 cm of soil, contact
with smoke during or shortly after a fire was a distinct
possibility. However, smoke has not been commonly used
in Persoonia gennination experiments or shown to have a
positive benefit on germination success. In a comprehensive
study of the effect of smoke water on the germination of
plants in Western Australia, Persoonia longifolia seeds did
not germinate when treated with smoke water (Dixon et al.,
1995). However, this study used fresh, 11011 -aged seed which
would not reflect the ecological priming of soil-stored seed
in nature and the post-fire germination found in this species.
A subsequent study also found no change in germination
success of Persoonia longifolia when smoke was applied
at different times prior to sowing; however, again non¬
aged seed was used (Mullins et al., 2002). Similarly, there
was no additive effect when smoke water was applied with
GA 3 . This could be due to both smoke and GA 3 having a
similar mode of catalysing germination, through opening
the respiration pathway by stimulating the conversion of
oxygen to superoxide in the seed (Sunmonu et al., 2016).
Furthermore, it is known that smoke catalyses the production
of enzymes such as amylase, mobilising starch compounds
from the endosperm to other parts of the seed (Cembrowska-
Lech & K^pczynski, 2016; Sunmonu et al., 2016). Whether
the application of smoke stimulates other mechanisms,
particularly for species with non-endospermic seeds, remains
unknown. However, this might explain why using smoke as
a pre-treatment has no additive effect on the germination
of Persoonia as the seeds lack an endosperm at maturity
(Strohschen 1986).
The variability in germination success reported in Persoonia
species means that further testing is required to determine
the environmental and population factors needed for optimal
seed collection, storage and germination conditions. We
consider that while a disturbance such as fire may be required
to break the mechanical dormancy in Persoonia seeds (i.e.
the woody endocarp), smoke water provides no greater
benefit for overall germination success of fresh seed than
GA 3 . However, if the role of smoke on germination is to be
examined in an ecologically relevant manner, then trialling
a possible interactive effect on soil-aged or seasonally-
stratified seeds would be an appropriate future study.
Discussion
Much of our knowledge of Persoonia seed biology is
derived from ex situ propagation research, be it for seed
germination, dormancy status or viability. However, we also
need to understand how the plants interact with their local
environment in situ , i.e. their ecology. Once the main factors
driving fruit set and dormancy are determined, we can use
this knowledge to include particular species in propagation
and restoration programs. An example demonstrating the
relative benefits to ‘progress’ upon changes from ex situ
germination testing to in situ ecological requirement studies
Emery & Offord, Managing Persoonia in the landscape 97
can be seen in our synopsis of Persoonia longifolia research
(Fig. 3). Following on from this and other Persoonia species
research efforts, we outline the major knowledge gaps for
restoration practice and, therefore, the research priorities for
future work on this taxon.
Climate and phenology
A major omission in many Persoonia studies to date is a
quantifiable link between phenological events and climatic
factors, namely temperature and rainfall. For example,
changes in the timing and duration of flowering can have
flow-on effects for other phenological events. In several
Proteaceae species, a decline in mean daily germination (due
to enforced seed dormancy) correlating with an increase of
1.4°C and 3.5°C during seed incubation has been reported
(Arnolds et al. 2015). Below-average rainfall was postulated
to cause the mortality of several Persoonia species
following fire (McKenna, 2007). If the flowering phenology
of Persoonia can be linked with climate, then this could
provide better predictions of the species niche, as well as
determine appropriate growing conditions and the adaptive
timing for fruit set and maturation. Since Persoonia species
have a breeding system that preferences outcrossing and are
predominately pollinated by native and exotic bee species,
the effect of the timing of flowering on interactions with
pollinators warrants further investigation.
Recent work on Persoonia longifolia illustrated the
importance of climatic events on both in situ and ex situ
seed burial. For example, brief wet events over summer,
such as a thunderstorm, were reported to greatly improve
overall germination by breaking mechanical dormancy in the
endocarp (Chia et al. 2016). As Persoonia longifolia seeds
are physiologically dormant, this means that an interactive
effect, in the sense of warm and cold stratification, is also
required to alleviate dormancy in the seed. Furthermore, as
post-fire germination is not always immediate, it is possible
that heat exposure, rather than smoke, weakens or cracks
the endocarp allowing germination to commence sooner
(McKenna, 2007; Chia et al., 2015).
Rethinking the role of fire
If endocarp degradation commences in the soil after fruit
drop and follows wet and dry cycles (Chia et al., 2016),
then the timing of fire could have a significant effect on
seed germination and viability. For example, seeds from
populations of Persoonia glaucescens and Persoonia
bargoensis (both obligate seeders) that had been burnt,
declined in viability over 12 months (McKenna 2007). By
contrast, seed viability in unburnt populations did not vary
over the same time. Auld et al. (2000) reported that seed
viability in Persoonia pinifolia (an obligate seeder) declined
to 28-40% following 2 years of soil burial, and suggested
that a prescribed burn midway through the experiment may
have contributed to seed death. However, seed death did not
significantly increase post-burning when compared with
pre-burning, and it was thought that seed ageing was the
main factor contributing to viability loss (Auld et al., 2000).
98
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
Persoonia endocarps are water-permeable, and poorly-timed
fires could essentially pressure-cook any partially imbibed
seeds (Norman & Koch, 2008; Chia et al., 2016). While fire
may hasten the relaxing of mechanical donnancy, if a portion
of the soil seedbank has already experienced some level of
degradation, a fire could then scorch and kill the more water-
permeable seeds. This outcome could be the underlying
reason for significantly higher germination of Persoonia
mollis subsp. nectens seeds following a medium-intensity
burn, compared to a high-intensity burn (McKenna, 2007).
Chia et al. (2015) suggested that fire had killed Persoonia
longifolia seeds in the seedbank, as recruitment events only
occurred following post-fire fruit set. Taken together, these
results suggest a fine balance for endocarp degradation
through wet and dry cycles and fire. This raises the question
of whether endocarps require a fire and, if so, when should a
fire occur relative to fruit drop? It is also plausible that a low-
intensity burn following seed-sowing might lead to a shorter
time to recruitment.
To detennine whether either the prescribed burning of
Persoonia soil seed banks or ex situ ‘priming’ of seeds by
burning is likely to be important, the in situ seed longevity must
be known. Previous evidence indicates that seed longevity
varies among Persoonia species (Aiild et al. 2000; Nonnan
& Koch, 2008) and suggests that seed longevity and endocarp
degradation might be intimately linked. If seeds lose viability
before the endocarp breaks down in the seedbank sufficiently
to allow gennination to commence, an early controlled burn
could shorten the time for endocarp weakening.
However, some obligate-seeding Persoonia species require
an interval of at least 8 years between fires to allow juvenile
plants to reach reproductive maturity (Weston, 2003). This
part of the life-cycle is still poorly understood; the length of
the primary juvenile period for only six Persoonia species is
known (Appendix 1).
Seed production areas
The goal for any species being re-introduced to an ecosystem
is to produce self-sustaining populations. Genetic variation
in local provenances is also an important consideration for
restoration practices. For example, non-locally sourced
material could have negative consequences for persistence
due to factors such as maladaptation, where the non-local
material is selected against local genotypes, leading to
higher mortality rates (Bischoff et al., 2010). However, there
may be occasions when non-local material is required, such
as providing sufficient genetically-diverse material to buffer
rare species from future environmental change (Broadhurst
et al., 2008). It is not our intention to discuss the various
merits of local vs. non-local material. Rather, we describe a
more immediate requirement of generating a source of high
quality seed.
Seed is often sourced from the wild in large quantities for
restoration projects (Broadhurst et al. 2015) but issues
including reproductive failure, low abundance, plant age
and phenological variation can hamper the availability of
large seed collections. Rare species often have life-history
traits that can create barriers to fecundity and survival (Abeli
& Dixon, 2016; Reiter et al., 2016). Rare or threatened
Persoonia species, including Persoonia pauciflora and
Persoonia bargoensis , have poor seed-production years.
Successful translocations of rare species (and, indeed,
other restoration practices) rely on an understanding of the
ecological requirements of the species (Abeli & Dixon 2016).
Tellingly, in an analysis of249 plant species worldwide a lack
of species biology knowledge was found to be a main cause
of reintroduction failure (Godefroid et al., 2011). A lack of
data on the pollination biology was concluded to have caused
failure in previous orchid reintroduction attempts (Reiter et
al. 2016) and there is strong evidence that an understanding
of pollination ecology is also important for rare Persoonia
species (Rymer et al., 2005).
Plants from known sources can be established in seed
production ‘orchards’ to provide seed that is genetically
diverse and representative of a robust population, and as an
alternative to overharvesting wild populations (Nevill et al.
2016). This requires an agronomic approach for maintaining
and harvesting and the ecological requirements of a species
can be implemented to produce large quantities of high-
quality seed for collection. In Persoonia, for example,
manually hand-pollinating flowers to promote outcrossing
could result in a higher fruit set of large quantities of seeds
for restoration programs. Many situations may also require
short- or long-term ex situ seed storage prior to restoration
projects. In this regard, ex situ seedbanks provide an
important supportive role. High-quality collections ensure
that seeds are more robust for ex situ storage conditions.
Conclusions
There is great potential for Persoonia species to be
successfully mass-propagated from seeds and included
more widely in restoration and horticulture projects.
Research to date has added several pieces to the puzzle;
however, the focus on optimising germination success has
meant that the ecological factors affecting this process have
not been widely tested. We also stress that maximising
germination does not necessarily translate into maximum
seedling survival. Similarly, assessing one seed batch from
a population does not provide any interpretation for that
population’s health beyond the collection year. Persoonia
propagation requires integrated collaboration between
the restoration and horticulture industries with rigorous
scientific research to achieve successful reintroduction and
conservation practices. We have highlighted several key
areas for future Persoonia seed research (summarised in
Fig. 4). The ecological requirements of Persoonia, in terms
of climate, plant-pollinator interactions, and seed biology,
are important for obtaining sufficient quotas of high-quality
seed to meet the growing needs of conservation, restoration
and horticulture.
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
99
Key areas for advancing Persoonia seed germination research
Species biology
Environmental factors
Genetic factors
Seed biology
- Endocarp degradation
- Effect of smoke on
aged seeds
Breeding system
- Pollinators
- Do pollinators
facilitate seed set?
Climate
- Effect on
flowering/fruiting
phenology
Fire
- Effect on in situ
seedbank
i
Diversity
- Population(s)
- Species
Fig. 4. Key areas that must be addressed for progressing Persoonia seed research. This requires an integrative approach, where data from
environmental and genetic factors help to inform key questions surrounding seed biology and the breeding system.
Acknowledgements
We thank Katharine Catelotti for initial supportive
discussions and Kerryn Chia for helpful suggestions to
improve an earlier version of this manuscript. This work
was initially funded by the NSW Office of Environment and
Heritage, Newcastle, and the Australian Coal Association
Research Program (ACARP) from mid-2017.
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Manuscript accepted 30 October 2018
APPENDIX 1
Database table of biological and ecological species characteristics of 115 taxa (including all 99 species) of Persoonia.
Cunninghamia 18: 2018
Emery & Offord, Managing Persoonia in the landscape
101
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Cunninghamia 18: 2018
Appendix 2
References for the data on Persoonia breeding
system, fruit set and germination listed in Table 1
and Appendix 1.
Breeding System
Cadzow B. & Carthew S. M. (2000) Breeding system and
fruit development in Persoonia juniperina (Proteaceae).
Cunninghamia 6, 941 -50.
Field D. L., Ayre D. J. & Whelan R. J. (2005) The effect of local
plant density on pollinator behaviour and the breeding system
of Persoonia bargoensis (Proteaceae). International Journal of
Plant Sciences 166, 969-77.
Nancarrow C. (2006) Flybridisation in three sympatric Persoonia
species: P chamaepitys , P. myrtelloides and P. levis. Institute
for Conservation Biology and Law, Australia.
Rymer P. D., Whelan R. J., Ayre D. J., Weston P. H. & Russell K. G.
(2005) Reproductive success and pollinator effectiveness differ
in common and rare Persoonia species (Proteaceae). Biological
Conservation 123, 521-32.
Trueman S. & Wallace H. (1999) Pollination and resource
constraints on fruit set and fruit size of Persoonia rigida
(Proteaceae). Annals of Botany 83, 145-55.
Emery & Offord, Managing Persoonia in the landscape 107
Germination
Bauer L. M., Johnston M. E. & Williams R. R. (2004) Fruit
processing, seed viability and dormancy mechanisms of
Persoonia sericea A. Cunn. ex R. Br. and P. virgata R. Br.
(Proteaceae). Seed Science and Technology 32, 663-70.
Chia K. A., Sadler R., Turner S. R. & Baskin C. C. (2016)
Identification of the seasonal conditions required for dormancy
break of Persoonia longifolia (Proteaceae), a species with a
woody indehiscent endocarp. Annals of Botany 118, 331-46.
Ketelhohn L. M., Johnston M. E. & Williams R. R. (1996)
Propagation of Persoonia virgata for commercial development.
In: Third International Symposium on New Floricultural Crops
(eds J. Considine and J. Gibbs) pp. 157-64, Perth, WA.
McKenna D. J. (2007) Demographic and ecological indicators of
rarity in a suite of obligate-seeding Persoonia (Proteaceae)
shrubs. In: Biological Sciences. University of Wollongong.
Nancarrow C. D. (2001) Germination of Persoonia myrtilloides
and P. levis. In: Combined Proceedings International
Plant Propagators Society, pp. 166-70. International Plant
Propagators Society.
Fruit Set
Auld T. D., Denham A. J. & Turner K. (2007) Dispersal and
recruitment dynamics in the fleshy-fruited Persoonia lanceolata
(Proteaceae). Journal of Vegetation Science 18, 903-10.
Bauer L., Johnston M. & Williams R. (2001) Rate and timing
of vegetative growth, flowering and fruit development of
Persoonia virgata (Proteaceae). Australian Journal of Botany
49,245-51.
Cadzow B. & Carthew S. M. (2000) Breeding system and
fruit development in Persoonia juniperina (Proteaceae).
Cunninghamia 6, 941-50.
Chia K., Koch J., Sadler R. & Turner S. (2015) Developmental
phenology of Persoonia longifolia (Proteaceae, R. Br.) and the
impact of fire on these events. Australian Journal of Botany
63,415-25.
Chia K. A., Sadler R., Turner S. R. & Baskin C. C. (2016)
Identification of the seasonal conditions required for dormancy
break of Persoonia longifolia (Proteaceae), a species with a
woody indehiscent endocarp. Annals of Botany 118, 331-46.
Nancarrow C. (2006) Hybridisation in three sympatric Persoonia
species: P. chamaepitys, P. myrtelloides and P. levis. Institute
for Conservation Biology and Law, Australia.
Nield A. P, Monaco S., Birnbaum C. & Enright N. J. (2015)
Regeneration failure threatens persistence of Persoonia
elliptica (Proteaceae) in Western Australian jarrah forests.
Plant Ecology 216, 189-98.
Rymer P. D., Whelan R. J., Ayre D. J., Weston P. H. & Russell K. G.
(2005) Reproductive success and pollinator effectiveness differ
in common and rare Persoonia species (Proteaceae). Biological
Conservation 123, 521-32.
Trueman S. & Wallace H. (1999) Pollination and resource
constraints on fruit set and fruit size of Persoonia rigida
(Proteaceae). Annals of Botany 83, 145-55.
Wallace H. M., Maynard G. V. & Trueman S. J. (2002) Insect
flower visitors, foraging behaviour and their effectiveness as
pollinators of Persoonia virgata R. Br. (Proteaceae). Australian
Journal of Entomology 41, 55-9.
Date of Publication:
December 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
Washingtonia rohusta (Mexican Fan Palm) as a coloniser in an
artificial wetland at Albury, New South Wales
DirkH. R. Spennemann
Institute for Land, Water and Society; Charles Sturt University; PO Box 789; Albury NSW 2640, AUSTRALIA
dspennemann@csu .edu.au
Abstract. Washingtonia robusta (Mexican Fan palm) is endemic to the semi-arid zone of California and northern
Mexico. Dispersed globally by the horticultural trade, the species has demonstrated its ability to successfully invade
disturbed areas and urban landscapes in warm temperate climates. Once established, the plant is extremely hardy. This
paper presents the first documented instance of the successful establishment and growth of Washingtonia robusta in a
pond in continually flooded wetlands at Albury, the first record of it naturalising in New South Wales.
Keywords: Weeds in wetlands— Washingtonia robusta—Phoenix canariensis —frugivory—dispersal of exotic
palms—adaptation
Cunninghamia (2018) 18: 109-122
doi 10.7751/cimninghamia.2018.18.007
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
110 Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
Introduction
Numerous plants have shown to be adaptable to changing
and novel environmental conditions. Many of the exotics
have become invasive at the expense of local vegetation.
Birds in particular have often adapted to and taken advantage
of these exotics to broaden their feeding habits. Ornamental
palms such as Phoenix canariensis, Washingtonia filifera
and Washingtonia robusta while initially confined to
anthropogenic landscapes such as gardens, parks and streets,
have been able to spread via birds and animals well beyond
the confines of urban environments.
Introduced to the European nursery trade in the 1860s (Bailey,
1936, p. 63ff; Ishihata & Murata, 1971), and to Australia in
the 1870s and 1880s, the Mexican fan palm or Desert Palm,
Washingtonia robusta has become a major ornamental plant
on a global scale, widely planted as a feature tree in private
and public gardens, as well as an occasional street tree in many
communities with a temperate climate (Spennemann, 2018b).
Though much less invasive than other palm species, such
as Phoenix canariensis , Washingtonia robusta is known to
escape from horticultural settings and colonise natural areas.
Endemic to semi-arid regions in northwest mainland Mexico
and the southern Baja California (Cornett, 1989; McCurrach,
1960, p. 264f), Washingtonia robusta is regarded as naturalised
not only in adjacent southern California and southern Arizona
(Felger & Joyal, 1999), but also in subtropical areas such as
southern Florida (Cornett, Stewart, & Glenn, 1986; Institute
for Regional Conservation, 2016), Reunion (Indian Ocean)
(Meyer, Lavergne, & Hodel, 2008), the North Island of New
Zealand (Martin, 2009), and parts of Hawaii (Oppenheimer
& Barlett, 2002). On the Australian continent it is regarded as
naturalised in the Pilbara region, Western Australia (Pilbara
Region, Keighery, 2010); evidence from Albury (NSW,
Australia) presented in this paper demonstrates that it must
also be regarded as naturalised in NSW.
Little is known about the dispersal potential of Washingtonia
robusta. A review of dispersal vectors showed a number
of vertebrate species feed on Washingtonia drupes, with
some ingesting whole drupes and dispersing the seeds
via regurgitation or defecation, in the Australian setting
including Pied Currawongs ( Strepera graculina ) and fruit
bats {P ter opus poliocephalus) (Spennemann, 2018c, 2018d).
Washingtonia robusta have been noted as self-seeded plants
in garden settings in Albury and Sydney (pers. obs.), but
unlike Phoenix canariensis, Washingtonia do not seem to
readily invade remnant bushland or agricultural areas. It
is possible that their germination and initial establishment
success, more so than that of Phoenix canariensis , relies on
adequate moisture regularly provided in suburban gardens
but which is seasonal in bushland settings. If such moisture
is provided, it can establish in very marginal places such as
cracks in the concrete pavements (Martin, 2009; Stein, 2010)
and even in cavities in trees (Spennemann, 2018g). The
long-term success of the seedlings is dependent on stable
nutrient availability as well as lack of human intervention.
Other factors that seem to influence establishment success
in suburban gardens and peri-urban areas are the absence
of grazing animals (primarily sheep and kangaroos), more
friable and less compacted clayey soil, and property owners
who tend to be tolerant of such exotic adventitious species.
In its natural habitat, Washingtonia robusta is strictly a plant
of the semi-arid zone confined to a small area of the Sonora
desert and the southern Baja California (Bomhard, 1950;
Cornett, 1989, p. 94; Felger & Joyal, 1999; McCurrach,
1960, p. 264f; Wiggins, 1964). It occurs naturally in annually
or irregularly flooded riparian habitats and palm oases, with
the plants concentrated closest to water sources, on occasions
near, but not in, permanent water or wet soil (Felger & Joyal,
1999). In Southern California, it has colonised riparian strips
and aided by its seed shadow has crowded out other plants.
On occasion the palm thickets have become so dense that
some grew at the very stream edge with their trunks just in
the water (Kelly, 2007). Washingtonia robusta has not been
recorded as growing in the middle of ponds, lakes or flowing
bodies of water in California or Mexico. It was therefore
surprising to find plants growing in 1.5m deep water in the
middle of the eastern maturation pond of Albury’s waste
water treatment works with no evidence of such palms
growing in the adjacent bushland. This paper describes
the distribution and nature of these plants and the possible
vectors and mechanisms facilitating colonisation success.
Biology of Washingtonia robusta
Washingtonia robusta Wendl. (Mexican Fan Palm) are
monoecious, self-compatible plants that are solely propagated
by seed (Barrow, 1998). They can have up to 30+ bright green,
costapalmate fronds of 0.9 to 1.8 m with which are attached
to the tree, each with a 1.2-1.5 m long plano-convex petiole.
The red brown petiole exhibits curved spines on its entire
length (Bailey, 1936, p. 63ff). A mature Washingtonia robusta
will produce about 50 leaves annually (Brown & Brown,
2012). The trunk is usually covered with persistent dead
fronds hanging from the crown, which, unless horticulturally
removed, fonn a thick thatch (‘skirt’) surrounding the upper
section of the trunk (Moran, 1978; Morton, 1998).
Washingtonia robusta typically grows to a height of 15-20 m
with a trunk diameter of 0.6-1.2 m (up to 1.5m) (Broschat,
2017; Morton, 1998). The majority of palms will reach ages
of less than 200 years, though specimens of 500 years have
been estimated (Bullock & Heath, 2006). They reach maturity
after they have reached at least 3m in height (Cornett cited in
Martus, 2008, p. 25). The flower stalk ranges in length from
2 to 2.6 m, bearing numerous small, white bisexual flowers
in compound clusters (Felger & Joyal, 1999). These result
in infructescences (fruiting sprays) which ripen in autumn to
early winter. Each drupe is small, black and ovoid-oblong to
spherical fruit with a thin noil-oily, carbohydrate rich pericarp
and a single hemispherical seed (Brown & Brown, 2012).
The drupes measure about 7-10mm in length with an average
weight of 0.3g, while the seeds are about 4.7-6.5mm long and
about 4.5^1.9mm thick with an average weight of about 0.1 g
(Spennemann, 2018f). A single mature tree has been estimated
to produce in excess of 300,000 drupes per year.
Cunninghamia 18: 2018
Washingtonia robusta seed germinate well within 14 days
at soil temperatures of 25-35°C (Broschat, 2017; Brown &
Brown, 2012; Mifsud, 1996). The fruit are eaten by a range
of volant as well as terrestrial vectors. Passage through
the gastro-intestinal tract enhances germination success of
Washingtonia filifera, probably due to scarification by weak
acids (Noto & Romano, 1987) and this is likely to be the
case with Washingtonia robusta although no data was found
on germination success. Washingtonia filifera, seeds and
seedlings are allelopathic (Khan, 1982a, 1982b), giving the
plant seedling a competitive edge over other vegetation, and
this may also be the case with Washingtonia robusta although
data has not been found. Once established, both species
will do well with comparatively little water. Washingtonia
robusta grows at the rate of about 0.6-0.9 m per year, but if
well-watered, the palms reach annual growth rates of 1.8m,
at least in the early stages (Morton, 1998; Muirhead, 1961,
p. 41). Not well-watered Washingtonia robusta are on record
as reaching 20m height under 30 years (Proschowsky, 1921).
Albury study location
The artificial wetlands under discussion, the Kremur
Street Sewerage Treatment Plant of Albury consists of
two artificially created maturation ponds and a series of
downstream wetlands that now form part of the Horseshoe
Lagoon billabong system, located on the northern side of the
Murray River floodplain, some 2.5 km west of the centre of
Albury. While Horseshoe Lagoon and associated billabongs
are a natural system, the two maturation ponds are artificial
water bodies created by erecting retainment embankments
and flooding a paddock previously used for grazing.
After alienation from the Indigenous owners of the land
in the mid-1830s, the area first formed part of the grazing
lease ‘Bungowannah Station’ (run n° 40, Thomson, 1848),
after which the area close to Albury was reserved as part
of the Albury Temporary Common (notified on 26 June
1868, Wilson, 1868). The land covering both maturation
ponds as well as the sewerage works was then set aside for
police purposes (‘police paddock’) on 21 November 1871
(Reserve N° 853, Wilson, 1871) as revised on 6 October
1900 (Reserve N°R 31,592, Hassall, 1900a, 1900b, 1900c).
Given the existence of the wetlands, the area was converted
into an area reserved for the protection of birds in January
1916 (Black, 1916). The eastern section, which comprises
the vast majority of the first maturation pond, forms part
of land set aside in March 1917 for the Albury Sewerage
Works (Department of Lands, 1927; Strickland, 1917). The
Kremur Street sewerage works were opened in May 1919
(Anonymous, 1919), with the maturation ponds constructed
in the 1950s or early 1960s (Johnson, 2018). The plant was
modified into a biological nutrient removal plant in late 1984
or early 1985 (Johnson, 2018) with further expansion of the
scheme in the late 1990s ( Abbey, 1994).
Aerial imagery of the Kremur Street sewerage works, taken
in May 1949, shows the area as pasture with scattered
eucalypts primarily along a drainage line, as well as a market
garden in the east (Adastra Airways, 1949; Spennemann,
Spennemann, Washingtonia robusta in wetland at Albury 111
2018e). Today, evidence for the previous land use abounds in
the form of dead eucalypts in the maturation ponds. Judging
from the state of preservation of the dead trees (proportion
of large vs small limbs vs. small branches), some of the
trees were long dead well before the area was converted
into maturation ponds, but the majority died off due to
permanently water-logged soil following the flooding.
Setting aside rain water and minimal surface run-off, the
ponds are only supplied with treated, screened and filtered
wastewater from the Kremur Street plant. The water levels in
the maturation ponds, which have not been drained since their
construction, are more or less constant as they are regulated by
an outlet weir at the north-western end of maturation pond N°
2 (Johnson, 2018). Any fluctuations in water levels are limited
to a variation caused by evaporation. The bottom sections
of the ‘skirts’ of Washingtonia robusta exhibit a clean edge,
confirming a more or less steady water level.
Results
Survey of Washingtonia robusta in the Albury wetlands
In total, seven Washingtonia robusta palms were identified
in maturation pond n° 1, three in maturation pond n° 2 and
a group of palms on the southern peninsula dividing the
ponds (Figure 1) (see Spennemann, 2018e for photographic
documentation). Reference specimens were collected from a
cluster of self-seeded plants on the peninsula and deposited
in the CSU Regional Herbarium (see Spennemann, 2018a
for documentation).
The location and current appearance of the palms, as
determined by ground survey, are summarised in Table 1.
The plants range from an almost 11m tall mature plant (P2,
Figure 6) on the peninsula, to small seedlings of lm height
on the peninsula (P4) and small established plants in the
ponds (Figure 3-Figure 5). One dead palm (W2) was also
encountered in the western maturation pond (Figure 7).
The plants in the maturation ponds were documented and
photographed from the shoreline from various directions
(Spennemann, 2018e). The plants growing in the ponds
range in height from 2m to about 5m (Table 1).
Presence and dimensions over time are based on the
interpretation of historic aerial imagery of varied quality,
primarily sourced from the commercial service Nearmap™,
from Albury City’s Mapping Portal (which includes custom-
shot aerial imagery) and from GoogleEarth™. The quality of
the images depends on the resolution of the aerial or satellite
imagery, and the presence of cloud cover or sun reflections
off the waterbodies (Table 2).
The crown diameters of the palms were classed in quarter-
metre intervals. The observed size fluctuations are due to
the variations in the aerial images and the accuracy of the
embedded scales. Identification of Washingtonia on the
aerial images was carried out in two ways: i) backtracking
the presently existing palms in time, noting their presence
and any changes to diameter; and ii) the identification of
additional Washingtonia. This was possible as the foliage of
112 Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
Washingtonia stood out as a very light green tone compared
to the other small vegetation islands in the pond, and the
presence of a shadow. The latter differentiates Washingtonia
from sedge islands. The data compiled in Table 2 are
expressed as timelines in Figure 2.
The mature tree (P2) is by far the oldest. As it is growing
among eucalypts, it is not always readily distinguishable
on the aerial images, especially at its smaller sizes when it
would have been sufficiently obscured by the canopy of the
eucalypts. The palm shows above / in the canopy from about
2010 onwards, which suggests that by that time the plant
would have been about 5m tall (compare Figure 8).
Washingtonia robusta reputedly grows at about 0.6-0.9 m
per year, once established (at ca 2yrs of age). If well-watered,
the palm can, at least in the early stages, reach annual
growth rates of 1.8m (Morton, 1998; Muirhead, 1961, p. 41;
Proschowsky, 1921). Considering that the moisture regime
at the site is not subject to climatic variation, but stable due
the constantly flooded maturation ponds, the height gain of
the plant between 2010 and 2018 suggests that plant P2 grew
between 0.6 m and 0.7m per year. Projecting this growth rate
back in time suggests that the plant seeded sometime in 2001
or 2002. The other six plants for which height data and aerial
imagery exists (P1-P5, Wl), grew at rates between 0.4 m
and 0.55 m pa.
Fig. 1. Locations of the self-seeded palms shown on an aerial view of the maturation ponds of the Kremur Street Treatment Plant, Albury
NSW. Aerial image flown 8 November 2017, image courtesy Albury City.
Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
113
Table 1. Location, nature, dimensions and distance to source palm of the documented Washingtonia and Phoenix Palms at Kremur
Street wetlands, Albury
Specimen
Status
Stem
Height (m)
Crown (m)
Setting
Distance (m)
Latitude
Longitude
Washingtonia robusta
East Pond n° 1
immature
yes
3.5
2.5
ex perch
318
-36.085342
146.886768
East Pond n° 2
immature
yes
2.5
2.5
perch
259
-36.085306
146.88608
East Pond n° 3
immature
yes
3.5
2.5
perch
227
-36.084915
146.885858
East Pond n° 4a
immature
yes
4-5
2.5
low perch
195
-36.08478
146.885544
East Pond n° 4b
immature
no
2
2.5
—
194
-36.08478
146.885544
East Pond n° 5
immature
yes
3.5
2.5
perch
213
-36.085925
146.884853
East Pond n° 6a
not longer extant
perch
112
-36.084781
146.884594
East Pond n° 6b
not longer extant
perch
112
-36.084781
146.884594
East Pond n° 6c
immature
no
2.5
2.5
perch
112
-36.084781
146.884594
East Pond n° 7
immature
no
2.5
2.5
perch
123
-36.084781
146.884594
Peninsula n° 1 a
immature
no
2
indet.
palm
4
-36.084411
146.883437
Peninsula n° lb
immature
no
2
indet.
palm
4
-36.084411
146.883437
Peninsula n° 1 c
immature
no
2
indet.
palm
3
-36.084411
146.883437
Peninsula n° 1 d
immature
no
2
indet.
palm
2
-36.084411
146.883437
Peninsula n° 1 e
immature
no
2
indet.
palm
2
-36.084411
146.883437
Peninsula n° 2
mature
yes
10-11
3-4
ex-perch
0
-36.084382
146.883500
Peninsula n° 3a
immature
no
1
indet.
palm
1
-36.084353
146.883416
Peninsula n° 3b
immature
no
1
indet.
palm
1
-36.084353
146.883416
Peninsula n° 3c
immature
no
2
indet.
palm
2
-36.084353
146.883416
Peninsula n° 3d
immature
no
2
indet.
palm
2
-36.084353
146.883416
Peninsula n° 3e
immature
no
2
indet.
palm
3
-36.084353
146.883416
Peninsula n° 3f
immature
no
2
indet.
palm
3
-36.084353
146.883416
Peninsula n° 3g
immature
no
2
indet.
palm
4
-36.084353
146.883416
Peninsula n° 4
immature
no
1
1.5
perch?
11
-36.084302
146.883496
West Pond n° 1
immature
yes
2
2.5
perch
135
-36.083452
146.88245
West Pond n° 2
dead
no
0.75
2
perch
232
-36.082616
146.882024
West Pond n° 3
not longer extant
perch?
123
-36.083744
146.882298
Phoenix canariensis
Canariensis n° 1
mature, male
yes
4
5
no perch
n/a
-36.082963
146.886351
Canariensis n° 2
mature, male
yes
4
5
no perch
n/a
-36.082959
146.886461
Canariensis n° 3
immature
no
1
1
perch
n/a
-36.08282
146.888867
Canariensis n° 4
immature
no
1.5
2
perch
n/a
-36.082806
146.8888
Canariensis n° 5
immature
no
3
4
perch
n/a
-36.082242
146.88844
Fig. 2. Chronology of Washingtonia palms in the Kremur Street wetlands.
114 Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
Table 2. Crown diameters (in m) of Washingtonia in Kremur Street wetlands based on the interpretation of aerial imagery and
ground inspection (1 June 2018).
Date
El
E2
E3
E4a
E4b
E5
E6a
E6b
E6c
E7
PI
P2
P3
P4
W1
W2
W3
Image
Quality
Image
2018, June 1
3
3
3.5
3.5
2
3.5
—
—
2
2
2
3.5
2
0.5
3
dead
dead
./.
Site Visit
2018, May 1
3
3
3.5
3.5
2
3.25
—
—
2
2
invis
3.5
invis
invis
3
dead
dead
very good
NearMap
2018, Mar 11
3
3
3.5
3.5
2
3
—
—
2
1.75
invis
3.5
invis
invis
3
dead
dead
very good
NearMap
2018,Jan 20
3.5
3
3.5
3.5
1.5
3
—
—
2.5
1.75
invis
3.5
invis
invis
3
dead
dead
very good
NearMap
2017, Dec 22
obs
3
3.5
3.5
1
3
—
—
2
—
invis
3.5
invis
invis
obs
dead?
dead
partly
obscured
GoogleEarth
2017, Nov 8
3.5
3.25
4.0
3.5
1
2.75
—
—
1
1.75
invis
3.5
invis
invis
3
dead?
dead
very good
Albury City
2017, Oct 10
3
3
3.5
3.5
1
2.75
—
—
2
1.75
invis
3.5
invis
invis
2.5
2
dead
partly
obscured
NearMap
2017, Jun 13
3
3
3.5
3.5
1
2.75
—
—
1.5
1.75
invis
3.5
invis
invis
2.5
3
dead
very good
NearMap
2016, Nov 18
2.75
2.75
3.5
3.5
1
2.75
—
2
1.5
1.75
invis
pres
invis
invis
2.25
2.5
dead
very good
NearMap
2015, Nov 30
3
2.75
3.25
3.5
1
3
—
2
1.5
1.75
invis
pres
invis
invis
2
2.5
dead
partly
obscured
NearMap
2015, Oct 15
3.25
2.5
3.25
3.5
—
3
1.75
pres?
—
pres
invis
pres
invis
invis
1.5
2.5
dead
very good
AlburyCity
2015, Feb 9
2.75
2.5
2.75
3.5
—
3
1.5
1.5
—
1.5
invis
pres
invis
invis
pres
2.5
dead
very good
NearMap
2014, Aug 20
2.5
2.5
2.75
3.5
—
2.75
2
2
—
1
invis
pres
invis
invis
—
2.5
dead?
very good
NearMap
2014, Feb 22
2.5
2.25
2.5
3.5
—
2.75
2.5
2.5
—
pres
invis
pres
invis
invis
—
2
dead?
very good
(LPI, 2014)
2014, Feb 2
2.5
2.5
2.5
3.25
—
2.75
2
2.5
—
1
invis
pres
invis
invis
—
2
dead?
very good
NearMap
2013, Apr 4
2.5
pres
2.5
3.25
—
2.75
pres?
2.5
—
pres
invis
pres
invis
invis
—
2
3
very good
NearMap
2012, Oct 5
2.5
—
2.0
3.25
—
2.75
—
2.5
—
—
invis
pres
invis
invis
—
2
3
very good
NearMap
2012, May 29
2.25
—
1.75
3.25
—
2.75
—
2.5
—
—
invis
pres
invis
invis
—
1.5
2.75
very good
NearMap
2011, Oct 2
1.75
—
1.5
3.25
—
3.0
—
2.25
—
—
invis
pres
invis
invis
—
—
2.5
good
AlburyCity
2011, Apr 16
1.75
—
1.25
3.25
—
2.75
—
2
—
—
invis
pres
invis
invis
—
—
2.5
very good
NearMap
2010, Nov 17
1.5
—
obs
2.75
—
2.5
—
pres
—
—
invis
pres
invis
invis
—
—
2.5
partly
obscured
NearMap
2010, Jul 9
1
—
1.0
2.75
—
2.5
—
2.25
—
—
invis
pres
invis
invis
—
—
2
very good
NearMap
2010, Mar 13
obs
obs
obs
obs
obs
obs
obs
obs
obs
—
invis
pres
invis
invis
—
—
?
poor
GoogleEarth
2010,Jan 24
pres?
—
pres
2.5
—
1.5
—
2.25
—
—
invis
pres
invis
invis
—
—
1.5
very good
NearMap
2009, Jul 11
?
—
?
obs
obs
obs
?
?
?
—
invis
pres
invis
invis
—
—
?
very poor
GoogleEarth
2007, Oct
—
—
—
1?
—
—
—
—
—
—
—
?
—
—
—
—
?
poor
AlburyCity
2004, Nov
—
—
—
?
—
—
—
—
—
—
—
?
—
—
—
—
?
poor
AlburyCity
2003, Feb 19
—
—
—
?
—
—
—
—
—
—
—
?
—
—
—
—
?
very poor
GoogleEarth
2000, Nov 7
—
—
—
?
—
—
—
—
—
—
—
?
—
—
—
—
?
very poor
AlburyCity
Codes: invis—not visible on aerial, obscured by other vegetation; obs—obscured due to clouds or reflections; pres—present but not measurable;
115
Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
Fig. 1. Plant East n°3 as seen from south.
Fig. 2. Plant West n° 1 as seen from the northwest.
Fig. 3. Plant East n°4 as seen from north.
Fig. 4. The source palm (P2) as seen from the south.
116 Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
Fig. 5. Plant West n° 2 as seen from the northeast.
Fig. 6. The source palm (P2) as seen from the west. Note the low self-seeded clusters to the right (PI) and left (P3).
Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
117
Fig. 7. Appearance of the self-seeded Washingtonia palms (East n°3, 4, 6-9) on 22 February 2014 (top) and 8 November 2017 (bottom).
118 Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
Washingtonia dispersal and establishment processes
Washingtonia plants growing in the Kremur Street
maturation ponds and on the peninsula are all self-seeded
and started growing from 2000. Successful establishment
was dependent on the micro-graphic setting and on vectors
for dispersal of the plant seed.
The maturation pond is a flooded pastoral landscape studded
with dead eucalypt trees, most of which are now largely
branchless and act as perches for numerous bird species.
Sedges and other littoral vegetation grow at the base of
these tree trunks providing some modicum of substrate on
which Washingtonia can germinate (Figure 1, Figure 2).
Fallen trees and branches, resting semi-submerged in
comparatively shallow water, act as a traps for floating leaf
litter and develop into shallow vegetation islets. None of
these islets, however, have Washingtonia growing on them,
with the possible exception of plant E4, which may have
grown on or next to a semi-submerged log (comparing the
November 2004 and October 2007 imagery). Similarly, the
temporary establishment of plant W3 may have occurred on
a semi-submerged log.
As the plants are growing in the middle of the maturation
pond, birds, alighting on the perches provided by the dead
trees, are responsible for the deposition of the seeds. A
common feature of the successfully established plants is that
all grow in areas with little or no shading. Examination of
the shoreline, the peninsula and the larger vegetation islets
showed the presence of other bird-dispersed weeds, such as
blackberries (Rubus spp.) and figs (Ficus spp.) even in dense
patches. Washingtonia seed does not appear to germinate
or establish if it is choked by competing vegetation and
concomitant lack of sunlight.
In California, Washingtoniafilifera and Washingtonia robusta
are also dispersed by fluvial activity (Talley, Nguyen, &
Nguyen, 2012). Indeed, in the Albury setting, Washingtonia
robusta seeds tend to end up in the street gutters and from
there in the stormwater drains. As stormwater and sewerage
are discharged and treated differently, however, none of
these seeds are likely to end up in the sewerage treatment
plant, let alone in the maturation ponds (as the outgoing
water is screened and filtered). All seeds in the ponds would
have been introduced by vertebrate vectors.
Vertebrate species, both volant and terrestrial feed
on Washingtonia drupes and disperse their seeds.
In Australia the only vectors on record are the Pied
Currawong (Strepera graculina) and Grey-headed flying
fox (Pteropus poliocephalus ) (Spennemann, 2018c).
Currawongs, ingest fruit as a whole and then fly to a nearby
perch to digest the meal in the crop and regurgitate the
indigestible elements such as seeds and part of the pericarp
(Spennemann, sunpubl.).
Table 3. Potential volant vectors observed at the Kremur Street Parklands and adjacent Horseshoe Lagoon (HS)
Common name
Scientific name
consumption
perching
vector status
Greylag Goose (Domestic type)
Anser anser
probable
possible
unlikely
Spotted Dove
Spilopelia chinensis
possible
possible
possible
Crested Pigeon
Ocyphaps Jophotes
possible
probable
possible
Pacific Koel (HS)
Eudynamys orientalis
probable
unlikely
possible
Noisy Miner
Manorina melanocephala
possible
likely
probable
Red Wattlebird
Anthochaera carunculata
possible
likely
probable
White-plumed Honey eater
Lichenostomus penicillatus
probable
likely
probable
Blue-faced Honeyeater
Entomyzon cyanotis
probable
possible
possible
Little Friarbird
Philemon citreogularis
possible
likely
probable
Noisy Friarbird
Philemon corniculatus
possible
likely
probable
Australian Magpie
Cractions tibicen
possible
likely
probable
Pied Currawong
Strepera graculina
documented
likely
likely
Black-faced Cuckooshrike
Coracina novaehollandiae
probable
likely
probable
White-bellied Cuckooshrike
Coracina papuensis
probable
unlikely
possible
White-winged Triller (HS)
Lalage tricolor
possible
probable
possible
Crested Shrike-tit
Falcunculus frontatus
possible
possible
possible
Olive-backed Oriole
Oriolus sagittatus
possible
possible
possible
Australian Raven
Corvus coronoides
documented
probable
probable
White-winged Chough (HS)
Corcorax melanorhamphos
possible
possible
possible
Common Blackbird
Turdus merula
documented
unlikely
unlikely
Common Starling
Sturnus vulgaris
documented
probable
probable
Common Myna (HS)
Acridotheres tristis
documented
possible
unlikely
Cunninghamia 18: 2018
Given the small size of the drupe (diameter 9-10mm),
however, a much greater range of vectors can be inferred.
In total 99 bird species are on record for the Kremur Street
Parklands (Cornell Lab of Ornithology, 2018b), with an
additional 26 species observed at the adjacent Horseshoe
Lagoon (Cornell Lab of Ornithology, 2018a). Of these, 101
are pure waterbirds or exclusively insectivores, which can
be ruled out as potential vectors for Washingtonia drupes/
seeds. 1 The remainder have been classified in terms of their
likelihood to act as effective vectors of the Washingtonia
growing in the maturation pond based on whether they
have been known to ingest Washingtonia or Phoenix drupes
(Spennemann, unpubl. data), or like-sized, large seeded
fruit (Barker & Vestjens, 1989) (Table 1 ‘consumption’) and
whether they are likely to perch on isolated, very exposed
dead trees or alight on small (<1 m 2 ) vegetation patches in
the maturation pond (Table 1 ‘perching’). The majority of
these will alight on isolated trees or tree trunks with good
visibility and some height to aid in emergency take off.
There they process their meal and regurgitate the indigestible
parts. The self-seeded palms on semi-submerged logs can
be explained in that birds, that land on these to drink void
indigestible elements by regurgitation before drinking. The
domestic geese originate from a property near Horseshoe
Lagoon and are spatially confined to that lagoon, away from
the area where the Washingtonia occur.
Discussion
Long-term success
At the time of on-ground inspection one of the palms (W2)
was dead (Figure 5). The remains showed at least 13 fronds,
indicating it had thrived for some time. Based on aerial
imagery, the plant existed from 2012 to 2017 and seems
to have originally established on a semi-submerged trunk.
The cause of the palm’s death is not evident as the terminal
growth bud is well above the water. It is possible that the
palm ran out of nutrients.
Another example of establishment and subsequent failure are
palms East 6a and 6b. Both grew under the perch of a dead
eucalypt. Plant 6b established in ca 2010 and was followed to
the south by plant 6a in ca 2013. Both were thriving in 2014
(Figure 7 top) but by late 2015 both had died and another
Washingtonia robusta (East 6c) had emerged closer to the
base of the dead eucalypt tree. This palm is still extant today
(Figure 7 bottom). Today nothing remains of plants 6a and
6B and it remains unclear why they died.
The plants on the peninsula are growing on clayey soil, ca.
0.5m above the water level, while the palms in the pond are
growing on an unspecified substrate. It can be surmised that
the initial establishment occurred on substrate accumulated
on a semi-submerged log or on debris trapped at the base
Spennemann, Washingtonia robusta in wetland at Albury 119
of the trunk that serves as perch. Over time the roost must
have reached the soft bottom of the maturation pond as the
stability of the palms cannot be explained otherwise.
Worth noting is the differential growth rates between
the palms growing on the peninsula (0.6-0.7m pa) and
those growing in the maturation ponds (0.4-0.5m pa).
This differential growth can be explained in terms of the
availability of soil nutrients and the inhibiting factors of
water-logged soil. Washingtonia robusta reputedly exhibited
retarded growth when over-watered or planted in wet soils
(Meerow, 1994). Clearly, the document growth rated of
1.8m per year among well-watered Washingtonia robusta
(Morton, 1998; Muirhead, 1961, p. 41) are contingent
on ample supplies of nutrients and soil substrate as water
availability alone is not a factor.
The palms growing in the maturation ponds at the Albury
Sewerage Works highlights the adaptability of the palm.
The only other mention of Washingtonia robusta growing in
water comes from Malta in the Mediterranean, where Mifsud
(1995) observed two mature palms growing in a pond in the
San Anton Gardens.
Other seif-seeded palm species nearby
Two well-established, mature Phoenix canariensis are
located in the grassed area north of maturation pond N° 1.
Both are male specimens with similar size (5 m diameter, 4m
height). Given that neither palms is located near any perch
trees or overhead wires, they are likely to have grown from
seeds that were either dropped mid-flight by avian vectors,
or, more likely, deposited in scats by terrestrial vectors such
as foxes (Spennemann, 2018d). Three additional, immature
Phoenix canariensis were noted in the cluster of eucalypt
trees to the north of maturation pond N° 1. Two of these palms
were close to the southern edge of the wooded area, while
the third was at the edge of an opening near the northern
margin (Spennemann, 2018e).
The absence of Washingtonia robusta in the areas beyond the
maturation pond is striking. The reasons are not self-evident,
as both Phoenix canariensis and Washingtonia robusta are
well represented as self-seeded plants in suburban Albury.
A similar situation was observed in open agricultural
settings at Alma Park (Southern Riverina of NSW) where
Phoenix canariensis, originating from a number of source
trees, widely established itself since the 1950s, whereas a
Washingtonia robusta, planted in 1906 (at the same time as
the homestead), produced no viable offspring.
Unlike Phoenix canariensis, Washingtonia robusta may
have trouble germinating and establishing in areas of higher
competition with groundcover grasses, especially those
which have allelopathic capacity (Downer & Hodel, 2001).
The same applies to seeds dropped under perch trees, as
1 The Spiny-cheeked Honeyeater ( Acanthagenys rufogularis ), which was identified for Horseshoe Lagoon (Cornell Lab of Ornithology,
2018a) was omitted from the table as its presence is very unlikely (pers. comm. D Watson).
120 Cunninghamia 18: 2018
Spennemann, Washingtonia robusta in wetland at Albury
similar allelopathic potential is exhibited by several species
of eucalypt (Chu et al., 2014; Zhang & Fu, 2009) as well as
Callitris (Harris, Lamb, & Erskine, 2003).
No data exist for Washingtonia robusta, but as both seeds
and seedlings of its congener, Washingtonia filifera, are
allelopathic (Khan, 1982a, 1982b), it can expected that this
is the case for Washingtonia robusta as well. The allelopathic
potential of a Washingtonia robusta seed and seedling
appears to be outcompeted by that of other plants due the
small volume of the Washingtonia robusta seed and the
small area of its seedling. In areas of non-competition it is
also possible that Washingtonia robusta may germinate but
fail to successfully establish on compacted clay soil, while
Phoenix canariensis with its larger seed mass has that ability.
Once successfully established as a seedling, the plant has a
high chance of survival unless subjected to grazing.
Implications
Setting aside a discussion on the merits of novel ecosystems
as ‘valid’ environmental states (Hobbs, Higgs, & Hall,
2013; Miller & Bestelmeyer, 2016; but see Murcia et al.,
2014), it needs to be asked from a management perspective
whether the colonisation of the wetland at Albury by
Washingtonia robusta is beneficial or detrimental. Clearly,
Washingtonia robusta do not rapidly invade the bushland
surrounding the wetlands; the failure to establish in areas
where other invaders (blackberry, fig) thrive, demonstrates
this well. Establishment seems to be random (P) and only
successful in the event of seed rain (i.e. underneath P2).
The colonisation of spaces underneath perches in open
water seems to be successful, albeit not on a large scale. No
other plant seems to successfully establish at such locations,
suggesting that Washingtonia robusta adds habitat variation
rather than detract from it.
Unlike the dense crown of pinnate fronds of Phoenix
canariensis, which provides roosting habitat for numerous
birds, marsupials and rodents, the sharp spines that line
the upper edges of the petioles of Washingtonia robusta
act as a major faunal deterrent. Consequently, colonisation
by Washingtonia robusta does not add to habitat until
such time that the plants have developed a deposit of old,
dry leaves that form the ‘skirt.’ Once this has occurred,
Washingtonia robusta provides sheltered nesting habitat for
several species.
Acknowledgements
I am indebted to Colin Johnson (formerly engineer with
Albury City) and Prof. David Watson (Institute for Land,
Water and Society, Charles Sturt University) for details on the
maturation pond and some of the bird species respectively.
Bruce Walpole (Walpole Surveying, Albury) kindly provided
access to Nearmap Imagery.
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Manuscript accepted 10 December 2018
Date of Publication:
December 2018
Cunninghamia
A journal of plant ecology for eastern Australia
ISSN 0727-9620 (print) • ISSN 2200-405X (Online)
The Royal
Botanic Garden
Sydney
The current status of exotic freshwater vascular plants in
Australia - a systematic description
Guyo Duba Gufu' and Michelle R. Leishman 1
'Department of Biological Sciences, Macquarie University
Correspondence: Guyo D. Gufu, Department of Biological Sciences, Macquarie University, NSW 2109, AUSTRALIA.
E-mail: guyo-duba.gufu@hdr.mq.edu.au
Abstract : Freshwater systems are considered particularly vulnerable to human impact, through habitat modification,
changes to water regimes and quality, invasion by exotic species and climate change. Using various records, we
conducted a descriptive analysis of the naturalised freshwater plant species in Australia. There are 63 freshwater plant
species belonging to 45 genera and 26 families naturalised in Australia with the dominant families being Cyperaceae,
Poaceae and Plantaginaceae. More than 40% of these species are categorised as either invasive or declared weeds,
the majority being perennial wetland marginal plants. They originated from all the inhabited continents with most of
the species being native to Europe, South America and North America. The greatest number of species are currently
found in New South Wales (90%), Queensland (68%) and Victoria (65%); the ornamental aquarium plant trade was
identified as the main introduction pathway. Most species are clonal plants with flexible modes of reproduction and
multiple dispersal vectors. We conclude that exotic plant species are now an important component of Australia’s
freshwater systems and that ongoing monitoring of their status, distribution and impact should be a high priority in
light of the increasing influence of anthropogenic factors including climate change.
Key words: aquatic; ecosystem; flora; invasive; native; naturalised; ornamental
Cunninghamia (2018) 18: 123-133
doi: 10.7751/cunninghamia.2018.18.008
Cunninghamia : a journal of plant ecology for eastern Australia
www.rbgsyd.nsw.gov.au/science/Scientific_publications/cunninghamia
© 2018 Royal Botanic Gardens and Domain Trust
124 Cunninghamia 18: 2018
Introduction
Freshwater ecosystems are estimated to cover about 3%
of the Earth’s land surface area (Downing et al. 2006) but
they provide habitat to a disproportionately high number of
specialised plant and animal species (Balian et al. 2008).
Globally, these ecosystems are experiencing severe declines
in biodiversity due to a mix of human-mediated threats such
as pollution, overexploitation, flow modification, habitat
degradation and invasive species (Dudgeon et al. 2006). These
declines in many cases are more pronounced compared to
terrestrial ecosystems (Sala et al. 2000); it has been argued
that freshwater ecosystems are the most threatened of global
ecosystems (Saunders et al. 2002; Dudgeon et al. 2006).
Despite freshwater ecosystems being extremely species-
rich and harbouring many threatened species (Abell et al.
2008), they do not receive the same conservation efforts and
research attention as terrestrial ecosystems (Brundu 2015).
For example, there is comparatively little information on
freshwater plants, insects, molluscs and crustaceans in most
parts of the world (Revenga et al. 2005) possibly due to the
difficulty of monitoring freshwater ecosystems (Brundu 2015).
One of the most significant threats to freshwater ecosystems
is the widespread introduction of exotic plant species into
new areas as a result of increased international human travel
and trade (Mack et al. 2000). Most species are introduced
deliberately for ornamental or agricultural purposes, others
passively find their way to new regions as contaminants of
ballast water or as hitchhikers on other species (Champion
et al. 2010). Although strengthened pre-border biosecurity
measures have slowed the rate of introductions to Australia,
the process is ongoing (Weber etal. 2008; Dodd etal. 2016),
and it is inevitable that a proportion of these introduced
species wifi become naturalised or even problematic
invaders. Currently Australia is estimated to have around
2700 naturalised plant species, about 12% of its total flora
(Randall 2007; Dodd etal. 2015).
Non-invasive naturalised species are those that establish
self-perpetuating populations in the wild without having
profound negative effects on the ecosystem (Richardson
et al. 2000). It is estimated that, with time, about 10% of
naturalised species wifi overcome reproductive and dispersal
barriers, and become invasive (Williamson & Fitter 1996;
Williams & West 2000). In future this proportion may
increase as ongoing environmental and global climatic
changes provide ecological opportunities in some regions
for some of these species to become invasive (Groves
2006; Scott et al. 2008; Duursma et al. 2013; Sorte et al.
2013; Leishman & Gallagher 2015). When this occurs, the
highly connected nature and dynamic disturbance regimes
of freshwater ecosystems will further facilitate the spread of
these species through the landscape (Dudgeon et al. 2006).
Invasive exotic freshwater plant species can exert dramatic
negative impacts on native communities and ecosystems
similar to their terrestrial counterparts (Evangelista et al.
2014). For example, a more than 50% decline in species
richness of co-occurring native freshwater plant species was
observed with increasing abundance of the invasive exotic
Gufu & Leishman , Naturalised freshwater plants in Australia
Alternanthera philoxeroides in natural ponds (Chatterjee
& Dewanji 2014), and Myriophyllum spicatum in Lake
George, New York, USA (Boylen et al. 1999). Furthermore,
this suppression of native plant communities by exotic plant
species may modify trophic interactions (Richardson & van
Wilgen 2004) by simplifying and rendering the native plant
communities a poorer food source for herbivores and higher
trophic level consumers (Havel et al. 2015). Thus, exotic
plant invasions can have detrimental ecosystem-level effects
on freshwater systems (Yarrow et al. 2009).
The naturalised flora in Australia is considered one of the
most species rich in the world (Dodd et al. 2015) and a large
effort has been made to establish a comprehensive inventory
of the entire naturalised flora (Randall 2007). In addition,
the Australian Virtual Herbarium (http://avh.chah.org.au/)
has digitised occurrence records of extant plant species
and created a publicly accessible online database (Haque
et al. 2017). These records have been useful in assessing
patterns of species endemism (Crisp et al. 2001), mapping
species threats (Evans etal. 2011), predicting plant invasions
(Duursma et al. 2013), analysing drivers responsible for
patterns of naturalisation (Dodd et al. 2015), and identifying
areas that have high richness of naturalised exotic species
(Dodd et al. 2016). However, these outcomes are broad
and generalised across different ecosystems. Therefore, it
is important for ecosystem-level descriptions of naturalised
non-invasive and invasive exotic species to be undertaken
so ecosystem-specific monitoring and management practices
can be devised. The aim of this study is to provide a
systematic description of the distribution, origin and richness
of naturalised non-invasive and invasive exotic freshwater
plant species in Australia.
Methods
Compilation of species list
We searched ISI Web of Knowledge for information on
naturalised plant species in freshwater ecosystems of
Australia using the following combinations: (invasi*) OR
(invader) OR (non-native) OR (exotic) OR (alien) OR (non-
indigenous) OR (introduced) OR (“naturalised species”)
OR (“naturalized species”) OR (biological invasion*) AND
(plant) OR (macrophyte*) AND (freshwater) OR (aquatic) OR
(river*) OR (pond*) OR (lake*) OR (dam*) OR (“farm dam”)
AND (Australia) OR (“New South Wales”) OR (“NSW”)
OR (Queensland) OR (“Northern Territory”) OR (“NT”) OR
(“Western Australia”) OR (“WA”) OR (“South Australia”)
OR (“SA”) OR (Victoria) OR (Vic) OR (Tasmania) OR
(“Australian Capital Territory”) OR (“ACT”). In addition, a
list of naturalised freshwater plant species in Australia was
compiled from existing inventories and lists (e.g. Aston 1973;
Sainty & Jacobs 2003; Randall 2007) and online databases
(e.g. http://weeds.dpi.nsw.gov.au/; http://plantnet.rbgsyd.
nsw.gov.au/; https://keyserver.lucidcentral.org/weeds/
data/media/Html/index.htm#A; https://www.business.qld.
gov.au/mdustries/farms-fishing-forestry/agriculture/land-
management/health-pests-weeds-diseases/weeds-diseases;
https://nt.gov.au/environment/weeds/weeds-in-the-nt/A-Z-
Cunninghamia 18: 2018
Gufu & Leishman , Naturalised freshwater plants in Australia
125
list-of-weeds-in-the-NT; https://florabase.dpaw.wa.gov.au/
search/advanced?current=y&alien=y; http://www.pir.sa.gov.
au/biosecurity/weeds_and_pest_animals/weeds_in_sa;
http://agriculture.vic.gov.au/agriculture/pests-diseases-and-
weeds; https://dpipwe.tas.gov.au/invasive-species/weeds/
weeds-index/declared-weeds-index).
A declared weed was defined as a plant species that has been
identified for control, eradication or prevention of entry into
an Australian jurisdiction by a legislation of that jurisdiction.
Results
We categorised plant species as ‘freshwater’ using the
following definition: “closely bound to freshwater habitats
whose vegetative parts actively grow either permanently or
periodically (for at least several weeks each year) submerged
below, floating on, or growing up through the water surface”
(Lacoul & Friedman 2006; Chambers et al. 2008; Hussner
et al. 2012). It should be noted that although we concentrate
on freshwater species, many species in Australia have wide
salinity tolerances and can occur across a wide part of the
gradient from fresh to saline in variable habitats such as
saltmarshes and estuaries.
Validity of the species names was checked using the
Australian Plant Census website (https://biodiversity.org.
au/nsl/services/APC) and species not found in the census or
with unresolved nomenclature were excluded. Any species
whose status as native or exotic was unclear according to
the Australian Plant Census was also excluded from the
analysis. We then checked the naturalisation status of each
species using a comprehensive data set of the introduced
flora of Australia - an updated version of Randall (2007)
containing unpublished data, and excluded any that was not
naturalised. We also excluded species that are associated
more with saline water than fresh water.
Plant data collation
Data on the native regions of each species, introduction
purpose, and their biology (growth habit, longevity and
dispersal mechanisms) were compiled from multiple
sources including regional floras, published literature and
the online databases (e.g. https://www.cabi.org/ISC/search;
http://ausgrass2.myspecies.mfo/content/fact-sheets; https://
npgsweb.ars-grin.gov/gringlobal/taxon/taxonomysimple.
aspx). Eight broad regions of origin were identified as follows:
Europe, North America (including Mexico), Central America
(including the Caribbean), South America, Sub-Saharan
Africa (including Madagascar), North Africa, temperate Asia
(including the Middle East), and southern and south eastern
Asia. Multiple sources of origin were assigned where a
species had a wide native geographical region. For example,
Alisma lanceolatum is native to Europe, North Africa and
temperate Asia and was counted as a species of each of these
regions. The current economic uses of the species were used
to assign their purpose of introduction (Weber et al. , 2008)
where such information was not explicitly available. We also
conducted internet searches to determine if each species is
currently available for purchase from aquarium suppliers.
The Australian Virtual Herbarium (http://avh.ala.org.au/) was
used to determine presence or absence of each species in each
of the Australian states and territories. Randall’s (2007) list
was used to categorise the species as naturalised non-invasive,
invasive (, sensu Richardson et al. 2000), or declared weeds.
Taxonomy and status
After screening 255 titles returned by the literature search, 42
papers that were studies of freshwater plants were reviewed
for collation of the naturalised species list (Appendix 1), in
addition to data derived from existing inventories and online
data sources. In total, 63 exotic species of freshwater plants
(belonging to 45 genera and 26 families) were identified
as naturalised in Australia (Figure 1; Appendix 2). Of
these, 40 species (63%) were classified as naturalised non-
invasive, 13 species (21%) were designated as invasive
and 10 (16%) as declared weeds (Figure 2; Table 1). The
plant families with the highest number of exotic naturalised
species were Poaceae (9 species), Cyperaceae (9 species)
and Plantaginaceae (5 species). Fourteen of the 27 families
were represented by only one species.
Poaceae
Cyperaceae
Plantaginaceae
Alismataceae
Onagraceae
Nymphaeaceae
Hydrocharitaceae
Brassicaceae
Pontederiaceae
Lamiaceae
Juncaceae
Acanthaceae
Typhaceae
Sparganiaceae
Salviniaceae
Ranunculaceae
Lythraceae
Hypericaceae
Haloragaceae
Cabombaceae
Asteraceae
Araliaceae
Araceae
Aponogetonaceae
Apiaceae
Amaranthaceae
0 2 4 6
Number of species (n)
10
Fig. 1. Taxonomic diversity of the naturalised freshwater plant
species in Australia.
Fig. 2. The percentage of naturalised freshwater plant species in
different categories of invasive status.
126 Cunninghamia 18: 2018
Gufu & Leishman , Naturalised freshwater plants in Australia
Table 1: Naturalised freshwater species that are considered
invasive in Australia. An asterisk (*) indicates that the species
is a Weed of National Significance (WONS) (http://www.
environment.gov.au/biodiversity/invasive/weeds/weeds/lists/
wons.html).
Species
Family
Hygrophila costata Nees
Acanthaceae
Alternatherciphiloxeroides (Mart.) Griseb*
Amaranthaceae
Gymnocoronis spilanthoides (D.Don ex
Hook. & Arn.) DC.
Asteraceae
Cabomba caroliniana A.Gray*
Cabombaceae
Myriophyllum aquatica (Veil.) Verde.
Haloragaceae
Eger ia dens a Planch.
Hydrocharitaceae
Juncus articulates L.
Juncaceae
Ludwigiaperuviana (L.) H.Hara
Onagraceae
Arundo donax L.
Poaceae
Hymenachne amplexicaulis (Rudge) Nees*
Poaceae
Eichhornia crassipes (Mart.) Solms*
Pontederiaceae
Salvinia molesta D.S.Mitch.*
Salviniaceae
Sagittaria platyphilla (Engelm.) J.G.Sm.*
Alismataceae
Growth habit
The majority of the species (94%) were perennial; 3%
were annual and the remaining 3% have annual stems but
perennial rhizomes. Most of the species were emergent
marginal wetland species (59%). The emergent plants that
grow through the water column constituted 24% while the
submerged (8%), floating leaved (6%) and free-floating (3%)
species made up the remainder.
Region of origin
Exotic naturalised freshwater species originated from a variety
of regions, with Europe, South America and North America
being the most widely represented (Figure 3). Species that are
native to southern and southeast Asia were the most poorly
represented with only two reported as naturalised in Australia.
Only eight species (12%) did not have multiple places of
origin. Of the 13 species that are classified as invasive in
Australia, 10 are native to South America.
Fig. 3. Regions of origin of the naturalised freshwater plant species
in Australia.
Distribution in Australia
New South Wales (NSW) was the most species-rich state
with 90% of the exotic naturalised freshwater species being
present, followed by Queensland (68%) and Victoria (65%)
(Figure 4). Northern Territory had the lowest number of
naturalised freshwater plant species (14 of the 63 species
or 22%). Species that were present in every state include
Cyperus eragrostis (Cyperaceae), Arundo donax (Poaceae),
and Polypogon monspeliensis (Poaceae).
State/Territory
Fig. 4. Percentage of naturalised freshwater plant species present in
the states and territories of Australia.
Introduction pathways
Almost two-thirds of the species (57%) were introduced for
aquarium and ornamental water garden purposes while a
further 25% were imported for agricultural purposes including
as vegetables, for example Alternanthera philoxeroides and
Rorippa spp., and pasture grasses. The introduction reason
for the remaining 17% of species is unknown; there is no
information available on their economic use (Figure 5).
Currently 33% of the species are available for sale within
Australia either by water garden nurseries or over the internet
(Table 2).
Purpose of introduction
Fig. 5. Purpose of introduction of naturalised freshwater plant
species
Cunninghamia 18: 2018
Gufu & Leishman , Naturalised freshwater plants in Australia
127
Table 2: Naturalised freshwater plant species available for
sale in Australia. An asterisk (*) indicates that the species is a
declared weed.
Species
Family
Aponogeton distachyus L.f.
Aponogetonaceae
Benda erecta (Huds.) Coville
Apiaceae
Bacopa caroliniana (Walter) B.L.Rob.
Plantaginaceae
Cyperus papyrus L.
Cyperaceae
Cyperus prolifer Lam.
Cyperaceae
Hydrocleys nymphoides (Humb. & Bonpl. ex
Willd.) Buchenau
Limnocharitaceae
Hygrophila polysperma (Roxb.) T. Anderson
Acanthaceae
Hypericum el odes L.
Hypericaceae
Ludwigiapalustris (L.) Elliott
Onagraceae
Ludwigia repens J.R.Forst.
Onagraceae
Mentha aquatic a L.
Lamiaceae
Menthapulegium L.*
Lamiaceae
Nymphaea caerulea Savigny
Nymphaeaceae
Nymphaea mexicana Zucc.
Nymphaeaceae
Phalaris arundinacea L.
Poaceae
Pontederia cordata L. *
Pontederiaceae
Rorippa nasturtium-aquaticuni (L.) Hayek
Brassicaceae
Rotala rotundifolia (Buch.-Ham. ex Roxb.)
Koehne
Lythraceae
Typha latifolia LA
Typhaceae
Veronica anagallis-aquatica L.
Plantaginaceae
Zantedeschia aethiopica (L.) Spreng.*
Araceae
Reproduction and dispersal
Almost half (49%) of the 63 species reproduce both
sexually and vegetatively (17 reproduce by both seeds and
fragmentation, 14 by both seeds and rhizomes). 33% of
the species (21 out of 63) reproduce exclusively by means
of seeds. The remaining 17% reproduce exclusively by
vegetative means. All the species that reproduce exclusively
vegetatively, do so by stem fragmentation. Water currents,
waterfowl, flood and watercraft were identified as the main
dispersal agents of the seeds and stem fragments.
Discussion
Our search identified 63 exotic freshwater plant species that
have become naturalised in Australia, a large proportion of
which are perennial wetland marginal species. They belong
to 26 families representing 16% of families of the naturalised
flora. The majority of the species originated from Europe,
South America and North America and are currently most
widely distributed along the eastern coastal fringes of
the country. They were mostly introduced for ornamental
purposes via the aquarium and water garden plant trade.
The majority of the species reproduce both sexually and
vegetatively, with water currents, waterfowl and watercraft
identified as their main dispersal vectors.
Given that there are about 2739 naturalised plant species in
Australia (Randall 2007), freshwater plant species represent a
very low proportion (slightly over 2%). However, despite their
seemingly small number, they may have disproportionately
strong environmental impacts as exemplified by the fact
that nearly 20% of the Weeds of National Significance are
freshwater species (6 out of 32) (http://www.environment.
gov.au/biodiversity/invasive/weeds/weeds/lists/wons.html).
This may be partly attributed to the widespread geographic
distribution of many of these naturalised freshwater species,
with most species found in multiple states within Australia.
More than 40% of the naturalised freshwater plant species
we identified are categorised as either invasive or declared
weeds in Australia. This proportion is much greater than for
the naturalised terrestrial flora of which around 14% have
become invasive (Leishman et al. 2017).
Of the naturalised freshwater species in our analysis, a large
majority are in the Poaceae, Cyperaceae and Plantaginaceae
families. These families are among the twenty most
commonly represented in the naturalised Australian flora
(Dodd et al. 2015) and reflect the Australasian (Jacobs &
Wilson 1996) and worldwide (Chambers et al. 2008) trends
where Poaceae and Cyperaceae are the most species-rich
freshwater plant families. Many plants belonging to Poaceae
and Cyperaceae are important pasture crops on which
livestock production in Australia relies heavily (Cook and
Dias 2006), which may further explain their dominance
compared to the other families.
The majority of the naturalised species we identified
are perennial, clonal plants with the ability to exploit
heterogeneous habitats. Clonality may explain the invasion
success of some of these species as it enhances persistence
and spread of plants at local scales (Santamaria 2002).
The largest proportion (57%) of the species in our analysis
were emergent species that fringe the margins of water
bodies. This may be due to water margins being suitable
for species that can withstand periodic submergence as
well as helophytes that can cope with periodic draw-downs
(Lacoul & Freedman 2006). In contrast, open water bodies
provide a narrower range of environmental conditions,
resulting in less species being suited to that habitat. The
overrepresentation of the marginal species in our analysis
may also have resulted from study biases since these
species are conspicuous and easier to sample and identify,
in contrast to, for example, submerged species. In addition,
emergent species that occur along water body margins may
be able to disperse their propagules not only by water but
also by wind, allowing them to colonise widely across the
landscape (Soomers et al. 2013).
Many naturalised freshwater plant species in Australia have
originated from Europe, South America and North America.
This is largely due to historical and trade linkages between
Australia and these continents. However, these regions of
origin are likely to have shifted through time, with invasion
success of plant species from Europe strongly linked with
European settlement in Australia (Phillips et al. 2010) and
more recent successful introductions originating from
South America now contributing the largest proportion
of naturalised freshwater plant species. The majority of
the naturalised freshwater plant species in our study had
multiple broad regions of origin. Many freshwater plant
128 Cunninghamia 18: 2018
species have broad distributions due to selective advantages
provided by asexual reproduction and long distance dispersal
of propagules (Santamaria 2002). Species with large native
ranges tend to have broad environmental tolerances and thus
may be effectively pre-adapted to their introduced range
(Pysek et al. 2009; Keller et al. 2011). This may explain why
the naturalised freshwater species of Australia are small in
number but a large proportion have spread extensively across
the continent and are now considered as species of concern.
New South Wales, Queensland and Victoria, the most densely
populated states (A.B.S. 2018), have the highest numbers of
naturalised freshwater species. This is not surprising as there
is a strong correlation between human population density
and exotic species richness, due to humans being responsible
for the deliberate or accidental introduction of exotic species
(Weber et al. 2008; Dodd et al. 2016; Haque et al. 2017).
Furthermore, a higher human population density also means
a higher number of potential aquarium keepers, representing
a greater propagule pressure (Hussner et al. 2010).
Alternatively, biases in herbarium specimen collection may
have painted a picture of relatively higher species numbers
in the densely populated states than reality (Lavoie et al.
2012; Dodd et al. 2016; Haque et al. 2017). It has been
observed that the intensity of herbarium specimen collection
in Australia, on which our species regional distribution
analysis relied, was higher in the densely populated areas
(Dodd et al. 2016).
Our analysis revealed that almost 60% of the freshwater plant
species naturalised in Australia were deliberately introduced
for ornamental water garden and aquarium purposes. This
is consistent with other studies globally reporting that
importation and trade in ornamental plants is the most
important pathway for freshwater plant introductions
(Champion et al. 2010; Strayer 2010; Keller et al. 2011).
Petroeschevsky & Champion (2008) suggested that 85% of
aquatic weeds in Australia were traded as aquarium or water
garden plants. We found that a third of Australia’s naturalised
freshwater plant species are currently available by trade for
either ornamental or agricultural purposes. Surprisingly,
among these actively traded species are four declared weeds
{Mentha pulegium, Pontederia cordata , Typha latifolia and
Zantedeschia aethiopica). Some of these species may be
traded with misspelled or incorrect scientific names that mask
their exotic status (Brunei 2009). For example, we found that
an aquarium supplier had listed Eleocharis for sale without
specifying the species; if these species escape into the wild
they could well remain undetected for a long time. There are
also reports of aquarium plant dealers who, mostly due to
ignorance, misrepresent exotic plants as similar-appearing
native ones (Kay & Hoyle 2001). A more serious practice
that may contribute to infestation of waterways is the
deliberate cultivation of exotic ornamental plants in natural
waterways, by aquarium traders in order to meet customer
demands (Petroeschevsky & Champion 2008).
Twenty-five percent of the naturalised freshwater species of
Australia have been introduced deliberately for agricultural
purposes. These include traditional vegetable species such
as Alternathera philoxeroides and Rorippa spp., and garden
Gufu & Leishman , Naturalised freshwater plants in Australia
herbs such as Mentha aquatica. However, the majority of
the agricultural species are ponded pasture plants that were
introduced for livestock grazing. Since commercial livestock
production is a major contributor to the Australian economy,
many state governments actively promoted introduction
of exotic ponded pasture species through much of the 20 th
century (Cook & Dias 2006; Cook & Grice 2013). These
species may have then spread across the broader landscape
through natural dispersal mechanisms.
Almost a fifth of the naturalised freshwater plant species in
Australia have no known economic uses and may have been
introduced inadvertently in ballast water or as contaminants
of other deliberately imported species, which is a common
occurrence (Kay & Hoyle 2001). For example, Maki and
Galatowitsch (2004) found that ten percent of freshwater
plants that they obtained commercially contained exotic plant
contaminants. Occasionally, some of these contaminants
prove attractive and easy to grow and are therefore placed on
the market. A good example is Salvinia molesta , which was
introduced initially as a contaminant of other plants but was
considered sufficiently attractive to be consequently traded
as an ornamental species in Texas, USA for several years
(Kay & Hoyle 2001).
Many of the naturalised plants in our analysis reproduce
both sexually and vegetatively, and are easily dispersed by
water currents and floods, wind, water birds and watercraft
(Santamaria 2002). As vegetative spread and multiple
dispersal vectors enhance establishment and therefore
naturalisation success (Keller et al. 2011), these factors
may also be drivers of invasion success of these naturalised
freshwater plants.
From our study, we can conclude that although naturalised
freshwater plant species form a very small proportion of
the naturalised flora, they nevertheless are an important
component of the Australian flora, being widespread across
multiple regions. In spite of the existence of many statutory
and regulatory measures to control trade in potential
weeds in Australia at local, state and federal levels, a few
declared weeds continue to be traded. A strict enforcement
of these controls is therefore necessary through monitoring
of the online aquarium market and periodically assessing
compliance by nurseries through site visits. It is also important
that we continue to assess the weed risk of naturalised species
in light of ongoing environmental and climatic changes and
to monitor potential spread of wild populations constantly
(Champion et al. 2010). Finally, accessing information on
naturalised freshwater plants ranging from the local to state
level is difficult as data is contained within disparate sites.
Therefore, a centralised system of storing data on ecology
and management of naturalised freshwater plant species
would be desirable for better knowledge sharing.
Acknowledgements
We thank Rachael Gallagher for the initial list of exotic
freshwater plant species in Australia. Anthony Manea
provided helpful comments on earlier versions of this
Cunninghamia 18: 2018
Gufu & Leishman , Naturalised freshwater plants in Australia
129
manuscript, for which we are grateful. We also thank Doug
Benson and anonymous reviewers whose inputs greatly
helped improve the manuscript.
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Cunninghamia 18: 2018
Appendix 2: List of the species included in the
analysis of naturalised freshwater plants in
Australia
Botanical name
Family
Hygrophila costata Nees
Acanthaceae
Hygrophila polysperma (Roxb.) T. Anderson
Acanthaceae
Alisma lanceolatwn With.
Alismataceae
Hydrocleys nymphoides (Humb. & Bonpl. ex
Willd.) Buchenau
Alismataceae
Limnocharis flaw a (L.) Buchenau
Alismataceae
Sagittaria calycina Engelm.
Alismataceae
Sagittariaplatyphylla (Engelm.) J.G.Sm.
Alismataceae
Alternantheraphiloxeroides (Mart.) Griseb.
Amaranthaceae
Benda erecta (Huds.) Coville
Apiaceae
Aponogeton distachyos L.f.
Aponogetonaceae
Zantedeschia aethiopica (L.) Spreng.
Araceae
Hydrocotyle ranunculoides L.f.
Araliaceae
Gynmocoronis spilanthoides (D.Don ex Hook.
& Arn.) DC.
Asteraceae
Rorippa microphylla (Boenn. ex Rchb.) Hyl.
Brassicaceae
Rorippa nasturtium-aquaticum (L.) Hayek
Brassicaceae
Rorippapalustris (L.) Besser
Brassicaceae
Cabomba carolmtana A. G ray
Cabombaceae
Cyperus eragrostis Lam.
Cyperaceae
Cyperus involucratus Rottb.
Cyperaceae
Cyperus papyrus L.
Cyperaceae
Cyperus prolifer Lam.
Cyperaceae
Eleocharis minuta Boeckeler
Cyperaceae
Eleocharis pachycarpa A%o.Desv.
Cyperaceae
Eleocharis parodii Barros
Cyperaceae
Isolepis prolifer a (Rottb.) R.Br.
Cyperaceae
Schoenoplectus califormcus (C.A.Mey.) Sojak
Cyperaceae
Myriophyllum aquaticum (Veil.) Verde.
Haloragaceae
Egeria densa Planch.
Hydrocharitaceae
Elodea canadensis Michx.
Hydrocharitaceae
Lagarosiphon major (Ridl.) Moss
Hydrocharitaceae
Gufu & Leishman , Naturalised freshwater plants in Australia 133
Botanical name
Family
Hypericum elodes L.
Hypericaceae
Juncus articulatus L.
Juncaceae
Juncus ensifolius Wikstr.
Juncaceae
Mentha aquatica L.
Lamiaceae
Mentha pulegium L.
Lamiaceae
Rotala rotundifolia (Buch.-Ham. ex Roxb.)
Koehne
Lythraceae
Nymphaea alba L.
Nymphaeaceae
Nymphaea caerulea Savigny
Nymphaeaceae
Nymphaea mexicana Zucc.
Nymphaeaceae
Ludwigia longifolia (DC.) H.Hara
Onagraceae
Ludwigia palustris (L.) Elliott
Onagraceae
Ludwigi a peruviana (L.) H.Hara
Onagraceae
Ludwigia repens J.R.Forst.
Onagraceae
Bacopa caroliniana (Walter) B.L.Rob.
Plantaginaceae
Callitriche brutia Petagna
Plantaginaceae
Callitriche stagnalis Scop.
Plantaginaceae
Veronica anagallis-aquatica L.
Plantaginaceae
Veronica catenata Pennell
Plantaginaceae
Alopecurus geniculatus L.
Poaceae
Arundo donax L.
Poaceae
Echinochloa polystachya (Kunth) Hitchc.
Poaceae
Echinochloapyramidalis (Lam.) Hitchc. &
Chase
Poaceae
Glyceria maxima (Hartm.) Holmb.
Poaceae
Hymenachne amplexicaulis (Rudge) Nees
Poaceae
Phalaris arundinacea L.
Poaceae
Polypogon monspeliensis (L.) Desf.
Poaceae
Urochloa mutica (Forssk.) T.Q.Nguyen
Poaceae
Eichhornia crassipes (Mart.) Solms
Pontederiaceae
Pontederia cor data L.
Pontederiaceae
Ranunculus sceleratus L.
Ranunculaceae
Salvinia molesta D.S.Mitch.
Salviniaceae
Sparganium erectum L.
Sparganiaceae
Typha latifolia L.
Typhaceae
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