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Three lines of evidence to link outbreaks of the crown-of-thorns seastar Acanthaster planci to the release of larval food limitation

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Abstract and Figures

Population outbreaks of the coral-eating crown-of-thorns seastar, Acanthaster planci, continue to kill more coral on Indo-Pacific coral reefs than other disturbances, but the causes of these outbreaks have not been resolved. In this study, we combine (1) results from laboratory experiments where larvae were reared on natural phytoplankton, (2) large-scale and long-term field data of river floods, chlorophyll concentrations and A. planci outbreaks on the Great Barrier Reef (GBR), and (3) results from A. planci—coral population model simulations that investigated the relationship between the frequency of outbreaks and larval food availability. The experiments show that the odds of A. planci larvae completing development increases ~8-fold with every doubling of chlorophyll concentrations up to 3μgl−1. Field data and the population model show that river floods and regional differences in phytoplankton availability are strongly related to spatial and temporal patterns in A. planci outbreaks on the GBR. The model also shows that, given plausible historic increases in river nutrient loads over the last 200years, the frequency of A. planci outbreaks on the GBR has likely increased from one in 50–80years to one every 15years, and that current coral cover of reefs in the central GBR may be 30–40% of its potential value. This study adds new and strong empirical support to the hypothesis that primary A. planci outbreaks are predominantly controlled by phytoplankton availability. KeywordsCrown-of-thorns starfish-Seastar-Trophic limitation-Great Barrier Reef- Acanthaster planci -Eutrophication-Phytoplankton-Chlorophyll
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Three lines of evidence to link outbreaks of the crown-of-
thorns seastar Acanthaster planci to the release of larval food
K. E. Fabricius
*, K. Okaji
, G. De’ath
1 Australian Institute of Marine Science, PMB No. 3, Townsville MC, Qld 4810, Australia.
2 Coralquest Inc., Asahicho 1-34-10, Atsugi, Kanagawa 243-0014, Japan
* Corresponding author: , phone: +61-747534412, fax +61-747725852.
Running title: A. planci outbreaks and phytoplankton
Key words: Crown-of-thorns starfish, seastar, trophic limitation, Great Barrier Reef, Acanthaster planci,
eutrophication, phytoplankton, chlorophyll.
Population outbreaks of the coral-eating crown-of-thorns seastar, Acanthaster planci, continue to kill
more coral on Indo-Pacific coral reefs than other disturbances, but the causes of these outbreaks have
not been resolved. In this study we combine (1) results from laboratory experiments where larvae were
reared on natural phytoplankton, (2) large-scale and long-term field data of river floods, chlorophyll
concentrations and A. planci outbreaks on the Great Barrier Reef (GBR), and (3) results from A. planci
coral population model simulations that investigated the relationship between the frequency of
outbreaks and larval food availability. The experiments show that the odds of A. planci larvae
completing development increases ~8-fold with every doubling of chlorophyll concentrations up to 3
µg L
. Field data and the population model show that river floods and regional differences in
phytoplankton availability are strongly related to spatial and temporal patterns in A. planci outbreaks
on the GBR. The model also shows that, given plausible historic increases in river nutrient loads over
the last 200 years, the frequency of A. planci outbreaks on the GBR has likely increased from one in
50 80 years to one every 15 years, and that current coral cover of reefs in the central GBR may be 30
40% of its potential value. This study adds new and strong empirical support to the hypothesis that
primary A. planci outbreaks are predominantly controlled by phytoplankton availability.
On most Indo-Pacific coral reefs, including the Great Barrier Reef (GBR), coral cover has
been declining at rates of 0.2 1.5% per year since the 1960s (Bruno and Selig 2007). To
date, predation of coral by crown-of-thorns seastar (Acanthaster planci) accounts for a large
proportion of the observed decline in coral cover on the GBR. Between 1985 and 1997,
population outbreak of A. planci were observed on ~32% of monitored reefs on the GBR,
with their coral cover averaging 9% one year after the outbreak, compared with a mean of
28% coral cover on reefs that had not experienced an outbreak in the same period (Lourey et
al. 2000). These figures suggest a GBR-wide reduction in coral cover of 0.5% yr
due to A.
planci alone in this 12 years period. Population outbreaks of A. planci, i.e., the sudden
emergence of a large population after a period of relative rarity (Moran 1986), were first
recorded throughout the Indo-Pacific in the 1960s. The abrupt population increase by orders
of magnitude from a small parent population is called a ‘primary outbreak’ (Birkeland and
Lucas 1990), and questions concerning the cause(s) of such primary outbreaks, and whether
or not human activities have changed their frequency, have to date remained unresolved. In
contrast, secondary outbreaks are simply the consequence of the large numbers of gametes
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
produced upstream by a primary outbreak or another secondary outbreak population. Such
secondary outbreaks have been reconstructed using hydrodynamic models (Moran 1986;
Dight et al. 1990).
There are many hypotheses that relate to the control of A. planci populations (reviewed in
Birkeland and Lucas 1990; Brodie et al. 2005). Echinoderms that release large numbers of
planktotrophic larvae such as A. planci have a propensity to population fluctuations (Uthicke
et al. 2009), and some primary A. planci outbreaks have been recorded even on coral reefs
that remained relatively unaltered by human activities (Birkeland 1982, Birkeland and Lucas
1990). However, the observed widespread decline in coral cover since the 1960s suggest that
the recently observed frequency of primary outbreaks of once in ~15-years is unsustainable
(Seymour and Bradbury 1999). Even before mass bleaching started to inflict additional
mortality, most reefs were estimated to take 10 25 years for full recovery of coral cover,
with 25% of reefs showing no signs of recovery from A. planci coral mortality (Lourey et al.
2000). Two hypotheses that specifically address the apparent increase in the frequency of
primary outbreaks have been widely debated. They are: (1) the ‘terrestrial runoff hypothesis’
aka ‘larval starvation hypothesis’ that argues that nutrient limited survival of the pelagic
planktotrophic larvae of A. planci controls population outbreaks (Birkeland 1982; Lucas
1982; Brodie et al. 2005), and (2) the ‘predator removal hypothesis’, which postulates that
more juveniles survive to maturity due to the removal of fish predators through human
exploitation (reviewed in Birkeland and Lucas 1990). Both hypotheses are based more on
circumstantial than empirical support.
The terrestrial runoff hypothesis has strong correlative evidence (reviewed in Brodie et al.
2005). More outbreaks occur on reefs near high Pacific islands or continental coasts from
which terrestrial runoff occurs, compared to low atoll islands without terrestrial runoff, and
most outbreaks follow large or drought-breaking floods that carry high nutrient and sediment
loads (Birkeland 1982). The apparent increase in A. planci outbreak frequencies is attributed
to increased nutrient levels resulting from the terrestrial runoff of fertilizers, sewage and
eroding soils, as recorded throughout the Indo-Pacific in modern times (Brodie et al. 2005).
The planktotrophic larvae of A. planci feed on nano- and microphytoplankton (>3 µm cells)
(Okaji et al. 1997) that multiply at high nutrient levels. Earlier feeding experiments with
cultured microalgae showed that larval development was optimal at 2 6.5 µg L
chlorophyll, while few larvae completed their development at <0.6 µg L
chlorophyll (Lucas
1982). However, the types of cultured microalgae also determined the developmental success
of A. planci larvae (Lucas 1982), limiting inferences from these experiments about larval
development in natural phytoplankton communities.
The predator removal hypothesis states that seastar populations are largely controlled by
predation, and that increased human exploitation of fish predators has resulted in increased
numbers of seastars surviving to maturity (McCallum 1987). Dulvy et al. (2004) have
suggested a relationship between A. planci outbreaks and human population density (but not
fish predator densities) on 7 of 13 investigated reefs in Fiji, and Sweatman (2008) has shown
a relationship between reef protection status and A. planci outbreaks on pairs of reefs open
and closed to fishing in the GBR. However, the more commonly fished large predatory fish
don’t usually prey upon A. planci (Sweatman 1995) and to date, no fish predator has been
identified that can effectively regulate A. planci populations, although predation on juveniles
undoubtedly occurs (Keesing et al. 1996). It is argued (but difficult to show empirically) that
the removal of large predators may (1) suppress prey switching behavior, i.e., the fewer large
predatory fish stop eating the less preferred A. planci, and/or (2) have caused some complex
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
trophic cascades, eventually resulting in fewer bottom-dwelling invertebrates that eat juvenile
seastar (Keesing et al. 1996; Sweatman 2008).
In this study we re-examine the contribution of terrestrial runoff to limiting the frequency of
primary outbreaks, by combining new empirical evidence from A. planci larval feeding
experiments, large-scale and long-term river flood and chlorophyll monitoring data for the
GBR, and model-based simulations of the population dynamics of A. planci and its main
drivers. Extensive long-term and large-scale coral reef, water quality and river monitoring
data are now available for the GBR (Brodie et al. 2007; Sweatman et al. 2008), thereby
providing a new opportunity to resolve the A. planci issue that affects the health of many
Indo-Pacific coral reefs.
Laboratory Experiments
Eight separate laboratory experiments were used to quantify rates of development and
survival of larvae reared at different concentrations of natural phytoplankton as detailed in
Okaji (1996). Seawater was collected from the ocean and filtered through a 25 µm mesh to
remove large zooplankton and detrital matter. This coarsely filtered seawater was used to
create fresh batches daily of the following treatments:
(a) 0.45-FSW: filtration through a 0.45 µm GF/B filter;
(b) 2-FSW: filtration through a 2 µm polycarbonate membrane filter;
(c) 25-FSW: no further treatment of the coarsely filtered seawater;
(d) NES: nutrient-enriched seawater: 2 ml of Guillard’s f/2 nutrient solution (Guillard 1975)
were added to 20 L of freshly collected and coarsely filtered seawater. This nutrient-
enriched seawater was incubated in a 500 L water bath outdoors without shading for 2
days prior to use to develop the phytoplankton communities. Nutrient enrichment
increased the concentration of eukaryotic phytoplankton cells and the concentration of
chlorophyll on average 20-fold in NES compared with 25-FSW, while the concentration
of cyanobacteria cells increased to a lesser extent (2- to 3-fold, Table 1).
The treatment levels of the 8 experiments are listed in Table 1. Each treatment was done in
triplicates for Experiments 1 to 6, and in duplicates for Experiments 7 and 8. Experiments 1,
2, 5 and 6 were conducted at Lizard Island Research Station in the northern Great Barrier
Reef (GBR), which is surrounded by clear offshore waters. Experiments 3 and 4 were
conducted at the University of the Ryukyus (Okinawa, Japan), using water sampled from the
front of Chatan Reef, Okinawa, from often clear coastal waters. Experiments 7 and 8 were
conducted at the Australian Institute of Marine Science (central GBR), with the often turbid
coastal seawater collected off Cape Bowling Green once a week and stored outdoors in
aerated 500 L tanks. For Experiments 1 to 4, duplicate seawater samples were taken from
each container every day, and for Experiments 5 and 6 every second day. The chlorophyll a
concentration of these samples was determined using fluorometry, and the densities of
eukaryotic algal and cyanobacteria cells were counted with an epifluorescence microscope
(Table 1). In Experiments 7 and 8, chlorophyll a of the NES was determined fluorometrically
before use, and NES was diluted with 0.45-FSW to obtain the final chlorophyll
concentrations. The use of total chlorophyll a concentrations overestimates food availability
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
Table 1: Mean percentage of surviving A. planci larvae that completed their development (late
brachiolarian stage or metamorphosed to juveniles) at age 22 days, in treatments with
contrasting concentrations of natural phytoplankton. Treatments: 0.45-FSW, 2-FSE, and 25
FSE = seawater filtered using 0.45, 2 and 25 µm filters, respectively; NES = nutrient enriched
Chl. a
g L
± SD)
Eukaryote density
± SD)
± SD)
(% of survivors ±
0.07 ± 0.03
(not detected)
25 ± 16
0 ± 0
0.17 ± 0.10
0.004 ± 0.006
55 ± 36
0 ± 0
0.40 ± 0.20
0.214 ± 0.112
65 ± 36
0 ± 0
0.08 ± 0.03
(not detected)
31 ± 16
0 ± 0
0.25 ± 0.11
0.004 ± 0.004
75 ± 34
0 ± 0
0.52 ± 0.21
0.234 ± 0.086
83 ± 30
0 ± 0
0.08 ± 0.03
0.163 ± 0.125
6.4 ± 5.9
0 ± 0
0.29 ± 0.10
0.437 ± 0.222
7.1 ± 6.1
18.7 ± 8.5
0.28 ± 0.08
0.385 ± 0.178
6.7 ± 5.1
0 ± 0
0.19 ± 0.10
0.004 ± 0.002
56 ± 40
0 ± 0
0.28 ± 0.10
0.207 ± 0.077
62 ± 41
88.3 ± 8.6
50% NES
2.91 ± 1.35
2.435 ± 0.564
142 ± 90
100 ± 0
100% NES
5.25 ± 2.32
4.441 ± 0.989
202 ± 157
100 ± 0
0.19 ± 0.10
0.004 ± 0.002
56 ± 40
0 ± 0
0.28 ± 0.10
0.207 ± 0.077
62 ± 41
97.2 ± 1.7
50% NES
2.91 ± 1.35
2.435 ± 0.564
142 ± 90
99 ± 0.7
5.25 ± 2.32
4.441 ± 0.989
202 ± 157
100 ± 0
0 ± 0
0 ± 0
0 ± 0
32.2 ± 0.5
50.2 ± 26.6
0 ± 0
0 ± 0
6.8 ± 0.9
38.6 ± 4.1
61.6 ± 4.4
for A. planci, because the contribution of nano- and microphytoplankton, the preferred food of
A. planci larvae, is typically <50% of chlorophyll a in Indo-Pacific waters (Charpy and
Blanchot 1999; Crosbie and Furnas 2001). The rest is picoplankton (<3 µm cells) that
constitutes <10% of the diet of A. planci larvae (Okaji et al. 1997). However, the use of
natural phytoplankton communities in our experiments, in which the relative contribution of
nano- and microphytoplankton to chlorophyll reflects that found in the field, enabled us to
relate the experimental results to the GBR long-term chlorophyll data.
Batches of A. planci larvae were reared in the laboratory (Okaji 1996). Actively swimming
and healthy early bipinnaria larvae with fully developed alimentary canal were collected 2
days after the in vitro fertilization of gametes, and 100 larvae (150 larvae in Experiments 7
and 8) were added to each treatment chamber (1 litre volume for Experiments 1 and 2; 2 litres
for all others), which were gently aerated and kept in the temperature range 26.5 29C.
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
Larvae were sieved with a 60 µm mesh and transferred to a clean set of containers of freshly
prepared seawater every day. Every second to forth day, larvae were individually examined
under a dissecting microscope, and their developmental stages recorded following Lucas
(1982). For Experiments 7 and 8 the body lengths of 10 to 20 randomly selected larvae per
chamber were measured along their longest axes every fourth day with an ocular micrometer.
When the first larvae reached brachiolaria stage, aeration was reduced and a few small chips
of crustose coralline algae were introduced as settlement substratum. These chips were
checked daily in order to count metamorphosing larvae and settled juveniles.
The diameters of juvenile seastar after completion of metamorphosis were measured with an
ocular micrometer in Experiment 5. The rate of successful completion of development was
defined as the proportion of surviving larvae that were either in late brachiolaria or juvenile
stages after 22 days. These two stages were combined since late brachiolaria larvae are
competent to metamorphose to juveniles. Absolute survivorship was not analyzed because
abnormal or regressed larvae can remain alive for extended periods of time without any
prospect of further development (‘living ghosts’), and there were no rigorous criteria to
distinguish these from healthy larvae. Clearly regressed or abnormal larvae were scored as
For the statistical analyses, the results of all 8 experiments were combined (Table 1). The
relationship between rate of successful completion of development after 22 days and
chlorophyll concentration was analyzed using a generalized linear mixed model; a logistic
regression model where the response was the proportion of successful developmental
completion. Two severe outliers (88% and 97% survivorship in E5, 25-FSW and E6, 25-
FSW; Table 1) were excluded from the final model. Runs in Experiment 2 that were
terminated after 18 days due to a lack of development in all treatments were scored as ‘zero
completion’. All analyses were done using the statistical software package R (R Development
Core Team 2009).
GBR Flood History and Chlorophyll Data
Data of the cumulative discharge volumes of the Burdekin River since 1922 and the five
largest Wet Tropics rivers (Herbert, Tully, Johnstone, Russell and Barron Rivers, latitude
16.5 18.5S) were obtained from the Queensland Department of Environment and
Resource Management, and the values for the Burdekin River plotted for each ‘water year’
October 30
September). The Burdekin River is the largest river entering into the GBR
lagoon, and the most important factor determining inter-annual variability in flood plumes,
since the annual discharge from its dry subtropical catchment varies by two orders of
magnitude between its wettest and driest years (long-term annual mean discharge: 8.5 km
; CV = 106% of annual mean). In contrast, discharges from the many annually flooding
rivers in Wet Tropics catchments, which jointly supply ~40% of the total annual runoff to the
GBR (Furnas 2003), vary less between years (CV = 34 37% of annual means for Tully,
Johnstone and Russell River, and 75 76% for the Herbert and Barron Rivers). Continuous
monitoring of the Burdekin River started in 1922, and of the Wet Tropics Rivers between
1967 and 1983.
Data of summer chlorophyll a concentrations were extracted from the GBR long-term
Chlorophyll Monitoring Program Data Base, which has sampled chlorophyll concentrations
along fixed transects across the continental shelf monthly since 1993 (Brodie et al. 2007).
Most A. planci on the GBR spawn in December to January (Lucas 1973; Babcock and Mundy
1992). Under experimental conditions, the duration of the pelagic larval phase is 12 to 22 days
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
for well-fed larvae (Birkeland and Lucas 1990), and >50 days for larvae reared in filtered
seawater (Lucas 1982). Thus chlorophyll records from November to March were used to
quantify long-term average summer values for the far northern (FN, 12.0 15.0S) and
central/northern regions (CN, 15.1 19.2S). The data from each region were further split
into the inner <25 km of the shelf (containing inshore and midshelf reefs in CN and FN) and
outer locations, as most small to medium-sized river flood plumes remain within the inner 25
km of the shelf and travel along the coast towards the north due to the prevailing
hydrodynamic patterns and Coriolis forcing (King et al. 2001; Devlin and Brodie 2005).
Differences in the probabilities of larvae completing their development were then calculated
for FN and CN based on the estimated differences in larval survival rates for given
chlorophyll levels using the response curve from the laboratory experiments and long-term
average chlorophyll concentrations in the GBR.
The A. planci Coral Simulation Model
The model was used to simulate spatial-temporal distributions of A. planci on the GBR. By
varying the drivers and parameters of the model, running various scenarios and conducting
sensitivity analyses, we investigated the temporal dynamics and the spatial patterns of the
outbreaks. The simulation model comprised four linked sub-models:
(1) The A. planci model: This is an age-structured meta-population model with two juvenile
and six adult stages, each of one year duration. Life history parameters included age-
dependent size, rates of survival across age cohorts, age-dependent fertility and feeding rates
on corals (Table 2; Moran 1986; Birkeland and Lucas 1990; Scandol 1993).
(2) The coral model: The parameters included the rates of growth of coral as the prey of
juvenile and adult A. planci (Scandol 1993; Sweatman et al. 2001; Wolanski and De’ath
(3) The chlorophyll model: The spatial-temporal variation in chlorophyll was based on data
from the GBR Long-Term Chlorophyll Monitoring Program (Brodie et al. 2007). Compared
to long-term change, temporal variation was large on seasonal and short time scales. Both
observed chlorophyll data (i.e., field observations) as well as simulated data including
gradients and spikes originating from floods were used in the model.
(4) The connectivity model: The connectivity of a reef to other reefs determines its capacity
to provide larvae to itself (i.e., to self-seed), to other reefs (i.e., be a source), and to receive
larvae from other reefs (i.e., be a sink). Hydrodynamic models provided estimates of the self-
seeding, source and sink levels for 321 reefs in the central and northern GBR (James et al.
A justifiable criticism of population modeling is that, given enough parameters, one can
reproduce most observed data. We have safeguarded against this problem by (1) focusing on
relative not absolute effects (determining the ratios and 90% confidence ranges of values for
A. planci and corals for FN and CN), and (2) using sensitivity analyses: The robustness of the
model assumptions were tested by adding various levels of stochastic noise to any of the A.
planci and coral life history variables, and by varying many of the model parameters within
reasonable ranges (e.g., variations in the rates of coral replenishment and consumption, and
juvenile and adult survival and fecundity). The models and all analyses of outputs were
programmed in the statistical software package R (R Development Core Team 2009).
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
a) Laboratory experiments
The laboratory experiments, in which freshly hatched A. planci larvae were reared in seawater
at 0.01 to 5.25 µg L
chlorophyll a, showed that the proportion of larvae completing their
development increased rapidly with increasing natural phytoplankton concentration (Fig. 1).
At 0.01 0.25 µg L
chlorophyll, few larvae developed from the bipinnaria to early
brachiolaria stage, none developed beyond the early brachiolaria stage, and most regressed at
days 10 to 14. The odds of completion of development increased by a factor 8.3 (95% CI =
4.7, 17.7) for each doubling of concentrations of chlorophyll (Fig. 2a, Table 1). At low to
moderate chlorophyll concentrations (<0.5 µg L
), this was equivalent to increasing the
probability of completing development by a factor 7 8 for each doubling of chlorophyll. At
higher concentrations, the rate of increase in the probability of completion slowed, and
plateaued at >3 µg L
where completion was certain.
Figure 1: Developmental success of A. planci larvae that were exposed to different phytoplankton concentrations
in 8 experiments (E1 E8). Larval developmental stages: white = bipinnaria, wide diagonal hatch= early
brachiolaria, narrow diagonal hatch = mid brachiolaria, grey shade = late brachiolaria, black = metamorphosed
juvenile seastar. The latter two stages together were scored as ‘completed development’. Abbreviations: 0.45-
FSW, 2-FSE and 25 FSE = seawater filtered using 0.45, 2 and 25 µm filters; NES = nutrient enriched seawater.
Growth rates, developmental speed and final body sizes of the larvae and early seastar also
depended on the availability of phytoplankton. Larvae reared at ≥0.8 µg L
reached the maximum observed size of 1.2 1.3 mm at 17 to 20 days of age, suggesting
growth was not food limited (Fig. 2b). At <0.5 µg L
chlorophyll, larvae initially grew from
0.8 to 1.0 mm but growth arrested after days 12 to 15 (Fig. 3). The time for 50% of surviving
larvae to complete development decreased from 41 to 14 days when chlorophyll increased
from 0.5 to >2 µg L
(not shown). Similarly, larvae that developed at 0.28 µg L
metamorphosed into seastars with significantly smaller mean diameter (0.44 mm, SE = 0.07
mm) than larvae reared at 2.9 or 5.2 µg L
chlorophyll (0.66 mm, SE = 0.05 mm; 0.64 mm,
SE = 0.09 mm).
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
Combining the data on rates of developmental completion and growth suggests the following
chlorophyll thresholds. At <0.25 µg L
(<220 eukaryotic cells m L
; Table 1), a negligible
proportion of larvae complete development, suggesting starvation. At 0.25 0.8 µg L
670 eukaryotic cells mL
) this proportion is moderate, but development is slow and body
sizes of larvae and juveniles remain small, suggesting severe food limitation. Finally, at >2 µg
(>1700 eukaryotic cells mL
) larval developmental success is high, developmental speed
is fast, and both larvae and juveniles grow to their maximum observed size, suggesting release
from trophic limitation.
Figure 2: (a) Relationship between chlorophyll a concentration and the proportion of A. planci larvae completing
their development. (b) Body length of A. planci larvae at 17-20 days of age. Each point represents the mean
results of duplicate or triplicate deployments per treatment. Black lines are model fits, the thin black lines are 2
SE of the mean.
Figure 3: Patterns of growth of A. planci larvae at increasing concentrations of chlorophyll a (Experiments 7 and
8). Shrinkage observable around days 16 to 20 reflects contraction for metamorphosis.
b) Temporal and spatial correlations between chlorophyll availability and A. planci
primary outbreaks
The patterns in Burdekin River discharges showed strong temporal and spatial agreement with
the timing and location of primary outbreaks of A. planci in the GBR. On the GBR, primary
outbreaks were first observed at 16.75S in 1962 and 1979, and between 14.7 16.1S in
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
1993/94 (Moran et al. 1992; Miller 2002; Sweatman 2008). The three largest recorded floods
of the Burdekin River yielded freshwater discharges of 28, 54 and 40 km
in 1958, 1974 and
1991 (Fig. 4). In 1974, all Wet Tropics rivers except the Herbert also produced >90
percentile floods, and in 1991, the Herbert and Barron produced >90
percentile floods and
the other three rivers were above median levels. Therefore, the 1979 and 1994 outbreaks
occurred three to five years after the two wettest years on record. The 1962 outbreak is
difficult to interpret since the 1958 flood occurred in February-March and hence may have
been too late in the season to feed A. planci larvae, and because few data exist from the Wet
Tropics rivers. A large flood also occurred very early in the 1950/51wet season, but no A.
planci data exist from that period (Fig. 4). The region north of Latitude 16.75S is the only
section of the whole GBR where the dense matrix of large mid- and outer-shelf reefs such as
Green Island frequently encounters river plumes (Fig. 5; Devlin and Brodie 2005; Brodie et
al. 2005).
Figure 4: Cumulative discharge volumes of the
Burdekin River into the GBR for each year since
1922. Red lines indicate the three large floods that
preceded the three recorded primary outbreaks of A.
planci in 1966, 1979 and 1994. The dark grey line
shows an early large flood in 1951, but no data exist
from that period. The blue lines show the large 2008
and 2009 Burdekin floods, potentially predicting the
onset of a fourth primary outbreak.
A satellite image from a moderate flood event on the central and northern GBR (Fig. 5)
illustrates: (1) the Burdekin plume extending >200 km to the north where it merges with the
plumes from the Herbert and many Wet Tropics rivers; (2) the plumes intersect mid-and
outer-shelf reefs around latitude 16 17S due to offshore diversion by the Cape Grafton
headland and a narrow continental shelf; and (3) the plume waters do not intersect with any
large reefs elsewhere as the remaining reef tract is too far offshore and all inshore reefs are
very small.
Strong regional differences in the long-term average summer chlorophyll concentrations are
also apparent on the GBR. Along the inner 25 km of the GBR, chlorophyll values were on
average twice as high in the central/northern GBR (CN) compared to the far northern GBR
(FN) (0.54 vs 0.26 µg L
). Assuming our experimental results were indicative of food
limitation in the field, this ~2-fold difference in chlorophyll concentrations between CN and
FN would translate into a ~8-fold higher rate of successful larval development in the former.
Additionally, levels of chlorophyll exceeding 0.5 µg L
occur for ~37% of summer values in
the inner CN, compared to 5.7 6.4% in the remaining sectors.
A re-analysis of AIMS Long-Term Monitoring Program data (Sweatman et al. 2008) of A.
planci outbreaks and coral cover shows that in the period 1985 2007, 12.9% ± 1.7% (SE) of
reefs in CN were in a state of ‘active or incipient outbreak’ at anyone time, and coral cover
averaged 16.5% ± 0.8%. In contrast, in FN only 5.5% ± 0.01% of reefs had A. planci
outbreaks, no outbreak waves have been observed, and coral cover averaged 28.0% ± 1.0%.
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
c) A. planci coral spatial-temporal simulation model
We used the A. planci coral spatial-temporal simulation model to investigate and quantify
the relationships between inshore chlorophyll, seastar populations and coral cover (Figs. 6 and
7, Table 2). The two principal drivers of the A. planci populations were both food-resource
related, and comprised:
(1) The concentration of chlorophyll in the water column which determines the
probability of A. planci larvae to survive until settlement and metamorphosis;
(2) The availability of hard coral for consumption by the juvenile and adult A. planci.
The former is governed by the empirical relationship between larval survival and
concentrations of chlorophyll in the water (Fig. 6a c). At low levels of chlorophyll the larval
survival was low and thus A. planci populations remained low and coral cover was high.
Conversely, at high levels of chlorophyll the abundant larvae led to large populations of A.
planci adults that can deplete the hard coral cover within a few years.
Table 2: Model parameters for the A. planci - coral simulation model. The values were based on Scandol (1993).
The temporal dynamics and patterns of the model results were relatively insensitive to variation of these
parameters. Abbreviations: J1, J2: juveniles aged 1 and 2 years, A1 A5: adults aged 3 to 7 years.
Life Stage
Coral consumed
Figure 5: Satellite image of the central GBR
(Modis, 10
February 2007), also showing the
locations of the mouths of the main rivers, and
towns (filled square). All inshore reefs, and the mid-
and outer-shelf reefs north of latitude 17S (the
presumed location of source reefs for primary A.
planci outbreaks on the GBR, red box) are
inundated by flood waters from the merged plumes
of several rivers, while the remaining mid- and
outer-shelf reefs are not intercepted by the flood
plumes during this moderate flood event.
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
Figure 6: Relationship of Acanthaster planci population dynamics and chlorophyll in the Great Barrier Reef
(GBR) off the NE of Australia. (a) Map of the GBR. (b). Long-term average chlorophyll concentrations in the
GBR in the far northern (FN, blue) and central/northern (CN, red) region, monitored near-monthly since 1992.
Applying the results from the laboratory experiments (c) showed that the odds for survival of A. planci larvae
was ~8-fold higher at chlorophyll levels found in CN compared with FN. Simulations of A. planci and coral
population dynamics show that in FN (d), outbreaks occur at 50 80 year intervals and coral cover recovers
between outbreaks (Table 4). In CN (e), outbreaks occur at 15 year intervals and corals only recover to 30-40%
of potentially obtainable values. These data form the basis to model the transition (f) in chlorophyll, A. planci
and coral cover in CN from pre-European (blue) to contemporary levels (red).
The model was used to compare the CN and FN regions using observed water quality data to
drive the populations (Fig. 6 d - f). Averaged over many generations, with distributions of
chlorophyll concentrations reflecting those in the CN inshore region (Table 3, mean
chlorophyll = 0.54 µg L
), the model population formed outbreaks at 12 15 year intervals,
consistent with present-day outbreak frequencies and intensities in CN (Fig. 6e, Table 4).
Coral recovery between outbreaks remained incomplete, with coral cover averaging 20 28%
of typical maximum values. At the chlorophyll distributions recorded in the FN (mean
chlorophyll = 0.26 µg L
), adult and juvenile seastar densities were 0.04 - 0.25 and 0.25 -
0.63 of CN densities respectively, outbreaks occurred only once in 50 80 years, and coral
cover recovered to 75 90% of typical maximum values between outbreaks (Fig. 6d, Table
4). Taking the relatively pristine FN as reflective of conditions ~150 years ago, a potential
transition from pristine to contemporary outbreak conditions in CN becomes apparent,
demonstrating increasing outbreak frequencies and progressively declining coral cover (Fig.
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
Table 3: Summer chlorophyll a concentrations on the inner (<25 km off the coast) and outer section of the
continental shelf in the far northern (FN, latitude 12 15.0S) and central/northern (CN, latitude 15.1 - 19.2S)
regions of the Great Barrier Reef (Brodie et al. 2007). N = number of samples. Shown are means, medians, and
the percentage of water samples with chlorophyll concentrations below 0.25 µg L
and exceeding 0.5 and 0.8 µg
(µg L
± SE)
(µg L
µg L
µg L
µg L
Inner 25 km
0.26 ± 0.01
0.54 ± 0.02
0.27 ± 0.01
0.24 ± 0.01
The model was also used to investigate the characteristics of A. planci, coral cover and the
outbreak frequencies and intensities (Fig. 7a-f). In terms of the model, outbreaks were defined
as events that reduced the coral cover to <2% and lead to mass mortality of seastars. The age
structure of the A. planci populations changed with increasing chlorophyll levels (Fig. 7a-c).
As chlorophyll increased, population sizes increased, and the age structure shifted to
relatively more young seastars. Outbreaks only occurred at chlorophyll levels above ~0.25 µg
and seastar populations were <10% of their maximum levels (Fig 7d). At chlorophyll
levels >0.5 µg L
coral cover declined by 75% of the initial cover and was dominated by
young corals with slower growth rates, the rate of growth in the coral population slowed, and
less coral was consumed (Fig. 7e - f). The maximum frequency of the outbreak waves was
one cycle per 10 15 years, and was controlled by the rate of coral recovery, with slower
rates resulting in lower outbreak frequencies.
Table 4. Results of simulation runs of the single-reef A. planci coral simulation model (Fig. 6). The far
northern region (FN, latitude 12.0 15.0S) has little agriculture and a low human population density, whereas
the central/northern GBR (CN, latitude 15.1 19.2S) experiences elevated nutrient loads from rivers. Ratios
between the two contrasting regions rather than absolute values and sensitivity analyses were used to overcome
the effects of model assumptions.
(Figs. 6b, e)
(Figs. 6b, d)
Adult A. planci (mean
relative density)
40 63
16 25
1.8 3.2
Juvenile A. planci
(mean relative density)
4,000 10,000
400 1,000
6 15
Coral Cover (mean %
of typical maximum)
20 28
75 90
0.25 0.35
Interval between
outbreaks (yrs)
12 15
50 80
0.13 0.22
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
Figure 7. Simulation results from the population model shows the averaged effects of varying levels of mean
chlorophyll on relative population sizes of (a) larvae, (b) juveniles and (c) adults, and on (d) the frequency of
seastar outbreaks, (e) the percent of coral eaten and (f) coral cover. Results are averages over model runs
spanning 200 years and excluding a 20-years ‘burn-in’ period.
Finally the data on reef connectivity and their source, sink and self-seeding levels were related
to seastar population and outbreak intensities. We found that:
(1) Patterns of inter-reef connectivity had far less effect on the large-scale wave-like patterns
of secondary outbreaks than differences in chlorophyll concentrations. Provided there is at
least a low level of connectivity amongst reefs, further increases in the strength of
connectivity had little effect on the outbreak patterns.
(2) At the scale of individual reefs, the risk of a reef having a severe A. planci outbreak
increased with its capacity to retain larvae through self-seeding and acting as a sink, and
decreased with its capacity to reduce larvae by acting as a source. Thus, reefs that were
the origin of outbreaks needed not outbreak themselves, and hence reefs where primary
outbreaks are first observed may not be the source of the outbreak. Furthermore, reefs that
were predominantly sources of larvae (i.e., had low larval retention) were four times less
likely to outbreak than reefs that retained more of their larvae.
Identifying the causes of ecological patterns and distinguishing anthropogenic changes from
natural dynamics is exceedingly complex, but synthesis of information from various sources
such as experiments, field surveys, long term monitoring and simulation models can form a
basis for attribution of causality (Fabricius and De'ath 2004). Using this approach, our study
adds new and strong support to the hypothesis that food availability controls primary
outbreaks of A. planci by enhancing the survival of larvae. It is important to differentiate
between the different processes that govern population dynamics on the three reef types
(source reefs, primary outbreak reefs, and secondary outbreak reefs). Primary outbreaks can
arise from small populations living on source reefs that encounter highly productive waters
during spawning times. Hydrodynamics dictate that source reefs may be located either
upstream of the primary outbreak reefs, or become primary outbreak themselves reefs through
self-seeding. However both source and primary outbreak reefs are likely to experience
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
seasonal phytoplankton blooms as the larvae together with the phytoplankton move with the
currents. Following these primary outbreaks, secondary outbreaks are then observed on
individual reefs in downstream progressing waves (Moran 1986; Moran et al. 1992,
Sweatman et al. 2008) that can be reconstructed using hydrodynamic models (Dight et al.
1990, our A. planci model). Primary outbreaks therefore develop from a small source
population of starfish, with each mature starfish producing an extremely large number of
offspring due to the release of larval food limitation. In contrast, secondary outbreaks
comprise a large population of mature starfish and hence can sustain outbreaks in conditions
where larval survival is relatively low.
Our larval feeding experiments showed a strong and non-linear (logistic) dose-response
relationship between the availability of natural eukaryotic phytoplankton and larval
development success. By using unfiltered natural seawater, the larvae in our experiments
grew in conditions where both phytoplankton food and potential planktonic predator
communities underwent natural successions in response to nutrient availability. We
demonstrated larval food availability to be a strong driver, with low chlorophyll leading to a
low rate of developmental completion, a prolonged pelagic phase of the larvae and small sizes
of the post-metamorphosis juveniles, likely leading to higher pre- and post-settlement
mortality (Allison 1994). The role of larval food limitation to echinoderm population
dynamics in the field is also corroborated by the observation that only echinoderm groups
with planktotrophic larvae have the propensity to exhibit boom and bust population dynamics,
while echinoderms with non-feeding (lecitotrophic) larvae have more stable populations
(Uthicke et al. 2009).
Identifying the origin of A. planci primary outbreaks is key to successful management of the
GBR. The first two outbreaks were first detected at Green Island off Cairns (Moran et al.
1992), a major tourist destination that was more frequently visited than many other reefs near-
by. The third outbreak was first reported from Lizard Island in October 1993 and 10-11
months later from 7 other reefs in the Cairns and Lizard Island regions (16013C, Evening
Reef, Swinger, Startle, Mackay, North Direction and Macgillivray Reefs), with a number of
additional reefs having A. planci densities only slightly below outbreak threshold levels in
1994 (LTMP data, not shown). The multiple outbreak locations, and the multiple A. planci
size classes (including juveniles) observed at Lizard Island in 1996 (Pratchett 2005) suggested
a gradual population build-up in the whole region through several successful spawning events
in prior years, indicating that not only the large 1991 flood but also conditions in following
years provided conditions suitable for high recruitment success. Floods have reached or
crossed this part of the shelf in 1991, 1994, 1995 and 1996 (Devlin and Brodie 2005),
retention times of flood materials on the continental shelf may be long (Luick et al. 2007), and
chlorophyll levels frequently exceed 0.5 and even 0.8 µg L
(Table 3). The weak and
bidirectional currents may further increase the vulnerability of reefs in this region to develop
outbreaks due to their relatively high rates of larval retention and self-seeding (James et al.
Long-term average chlorophyll values on the inner 25 km were twice as high in CN compared
to FN. As offshore chlorophyll values were similar in both regions, it is unlikely that the high
CN chlorophyll values were attributable to latitude or upwelling (Brodie et al. 2007). In CN,
present river loads of nutrients and sediments are estimated to be 2 10 fold higher than
before western colonization in ~1860, while river loads in the sparsely inhabited FN are
considered largely unaltered (McKergow et al. 2005). As rivers are the main source of new
nutrients to GBR inshore waters, regional differences in chlorophyll have been attributed to
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
differences in river nutrient loads, reflecting past and present terrestrial runoff (Devlin and
Brodie 2005; Brodie et al. 2007), although an explicit link between increasing river loads and
changes to inshore water quality on the GBR has not been established.
The fact that primary A. planci outbreaks occurred on locations where floods intercepted large
reefs on the GBR three to five years earlier, shows that not only long-term chlorophyll
concentrations but also large floods are a strong driver for A. planci primary outbreaks, in
agreement with previous findings (Birkeland 1982, Brodie et al. 2005). Flood plume models
(King et al. 2001) showed that the 1974 and 1991 floods reduced salinity for >60 days during
the time when A. planci larvae are pelagic. Due to the high incidence of cloud cover during
and after rain events, few satellite images are available to track the spread of flood plumes.
However, Devlin and Brodie (2005) used aerial surveys to show a spreading of plumes into
the main reef matrix, similar to the patterns shown in Fig. 5, after cyclones in 1996 and 1999,
and to an even greater extent in 1994. These river floods are the largest source of new
nutrients for the inshore GBR, and trigger phytoplankton blooms that average 2 µg L
chlorophyll and at times exceed 4 µg L
(Devlin and Brodie 2005). The application of
experimental findings to field settings necessitates caution. Bearing this in mind, if we take
the experimentally observed rates of change as indicative of the relative differences in
developmental rates in the field, the odds of successful development of A. planci larvae could
be up to ~60-fold higher during floods with 2.0 µg L
chlorophyll compared to the long-term
average of 0.54 µg L
for the central/northern GBR.
In combination, the three components of this study, together with previous evidence
(Birkeland 1982; Brodie et al. 2005), strongly support the assertion that removal of larval
food limitation causes primary population outbreaks of A. planci on the GBR. In contrast to
the parsimonious explanation of food limited control of A. planci populations, explanations of
population control by predators rely on complex arguments (Birkeland and Lucas 1990). To
date, there is no empirical evidence to support these arguments. Unlike water quality,
predation pressure does not fluctuate widely on short time scales, hence it remains unclear
how a chronic release from predation would occasionally lead to sudden increases in
population densities. However the reported correlation between reef protection status and A.
planci outbreaks on a subset of GBR reefs (Sweatman 2008) suggests that both hypotheses
may not necessarily be exclusive, and that predation may play some additive role in
determining the propensity of individual reefs to be afflicted by A. planci outbreaks.
Coastal GBR water quality is considered amenable to benefit from improved land
management (Haynes et al. 2007) due to the dominant role of the rivers in providing new
nutrients to the inshore GBR, and the potentially long residency times of these newly
imported materials (Luick et al. 2007). Large and early river floods from the Burdekin and
Wet Tropics rivers occurred again in January 2008 and 2009 (Fig. 4), potentially providing
conditions to trigger a new A. planci primary outbreak wave that may again kill a significant
proportion of GBR corals. Our study suggests that reductions in phytoplankton biomass to
summer values of <0.5 µg L
(De'ath and Fabricius 2010) through better land management
could reduce the frequency of primary A. planci outbreaks. Legislation and incentives have
now been put in place to reduce river discharges of nutrients, sediments and pesticides from
agricultural areas. However until a reduction in nutrient levels is achieved, two additional
precautionary management measures should aim to maintain very low A. planci densities in
the high-risk area (the midshelf reefs and hard bottom inter-reefal areas that are directly
intercepted by floods). These are: (a) large permanent fishing closures in the high-risk area,
allowing fish populations to reach carrying capacity to safeguard against cascading changes in
Coral Reefs (2010): Volume 29, pp 593-605: A. planci outbreaks and phytoplankton
food webs, and (b) potentially some targeted efforts by divers, especially in the years
following large floods, to remove some of the A. planci before they start to aggregate and
spawn, This three-pronged approach constitutes the best presently available strategy to slow
or reverse the loss of coral cover throughout the whole central and southern GBR, at a time
where rising seawater temperatures exert increasing pressure on coral reefs, and increasing
climatic instability may increase the frequency of extreme floods.
Many thanks to the GBR Long-Term Chlorophyll Monitoring Program for the chlorophyll data, to the AIMS
Long-Term Monitoring Program for coral cover and A. planci field distribution data, to J. Scandol for
compilation of the A. planci life history data, and to M. Slivkoff for the processing of the Modis satellite image.
KO conducted the laboratory experiments, and GD developed the coral A. planci simulation model. We
gratefully acknowledge support for the experimental study by T. Ayukai and J. Lucas. We thank J. Caley, B.
Schaffelke, S. Uthicke and H. Sweatman for constructive comments on earlier versions of the manuscript, and J.
Brodie, K. Day and E. Wolanski for sharing ideas. The study was funded by the Marine and Tropical Sciences
Research Facility (MTSRF) and the Australian Institute of Marine Science, with the experimental study being
funded by the Great Barrier Reef Marine Park Authority and James Cook University.
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... Les infestations d'acanthasters peuvent se produire de deux façons différentes, qualifiées d'infestation primaire ou secondaire (Potts, 1981). Il est capital de faire la distinction entre les deux pour tenter de comprendre les causes de celles-ci (Fabricius et al., 2010 ;Pratchett, 2005). Une infestation primaire correspond à une accumulation soutenue mais progressive du nombre d'individus qui va à terme atteindre le seuil d'infestation (Endean, 1974 (Pratchett, 2005 ;Stump, 1996 ;Sweatman et al., 1998). ...
... La théorie qui propose que les infestations soient déclenchées par une augmentation de la survie larvaire dû à l'enrichissement des eaux en nutriments a été proposée plusieurs fois (Birkeland, 1982 ;Brodie et al., 2005 ;Brodie, 1992 ;Fabricius et al., 2010 ;Lucas, 1973 ;Nishihira et Yamazato, 1974 ;Pearson et Endean, 1969). Cette théorie postule qu'une infestation d'acanthasters se déclarerait environ 3 ans après de très importantes pluies entrainant un lessivage des terres significatif. ...
... Le fait que les infestations se déclareraient 3 ans après les fortes pluies correspond au temps que les larves mettent à recruter, se métamorphoser et grandir jusqu'au stade adulte corallivore. Depuis les travaux de Birkeland (1982), cette théorie a été remise plusieurs fois en question, notamment à cause d'incohérences entre les épisodes de pluie majeurs et l'initiation des infestations (Endean et Cameron, 1990 --1970 ----1974 1979 5 ans 1981 ---1988 -1993 5 ans 1991 1991 1993 2 ans 1998 ---NA 2008 2010 2 ans Il a été suggéré que les épisodes de fortes pluies devaient avoir lieu à un certain moment de l'année pour correspondre au cycle larvaire (novembre à janvier ; Fabricius et al., 2010), mais qu'un laps de temps trop court entre une infestation passée et un nouvel épisode de crue ne déclencherait pas d'infestation supplémentaire car le corail n'aurait pas le temps de régénérer suffisamment (Fabricius, 2013). ...
Les infestations des étoiles de mer corallivores du genre Acanthaster représentent une des menaces les plus importantes qui pèsent actuellement sur les récifs coralliens de la zone Indopacifique. En dépit d’efforts de recherche conséquents ces trois dernières décennies, les facteurs d’initiation de ces infestations restent méconnus. Des travaux récents proposent que les processus liés au changement climatique, notamment le réchauffement et l’acidification des océans, pourraient favoriser certains stades de vie des acanthasters, en particulier la phase larvaire pélagique qui est considérée comme un goulot d’étranglement pour la dynamique des populations. Dans ce contexte, nous avons étudié les effets du réchauffement et de l’acidification des océans prévus d’ici la fin du siècle suivant le scénario classique business-as-usual du GIEC (+2 °C ; pH 7,75) sur les performances de reproduction des acanthasters après les avoir acclimaté aux différentes conditions. Les reproductions in vitro ont été effectuées avant d’étudier les processus biologiques suivants, de la fertilisation aux derniers stades larvaires. Répliquer les expériences deux années consécutives a permis de préciser la période optimale de reproduction en Nouvelle-Calédonie, qui semble se situer en décembre quand la température de l’eau dépasse 26 °C. Nos résultats ont montré que les futures modifications des paramètres physico-chimiques de l’eau pourraient avoir des effets contrastés suivant les stades de vie et surtout suivant la saisonnalité de reproduction. A mesure que l’on s’éloigne de la période de reproduction optimale, les effets deviennent particulièrement néfastes, avec un succès de fertilisation divisé par trois dans les conditions chauffées, associé à une augmentation de la mortalité (+25%) ainsi que du taux d’anormalité (jusqu’à 100%) des larves dans les conditions acidifiées. Nos résultats suggèrent que le réchauffement climatique pourrait entrainer un avancement de la période de reproduction des acanthasters, puis un allongement pour se rapprocher de la période observée aux latitudes plus basses. En revanche, lors de la période optimale de reproduction, nous avons observé des effets faibles à modérés de la température sur les œufs, les RGS, les taux de fertilisation, et la taille des premiers stades larvaires. Nous avons aussi montré que toutes les larves étaient systématiquement plus grandes dans les traitements acidifiés, quel que soit le stade larvaire. De plus, nous avons détecté des retards de développement larvaire provoqués par la température et par le pH, qui pourraient se traduire par un allongement du cycle larvaire dans la nature, associé avec une diminution du nombre de recrues. Néanmoins, aucune mortalité significative n’a été montrée sur les larves issues des reproductions faites en période optimale, suggérant une possible acclimatation des adultes aux conditions futures. Ajouter une phase d’acclimatation à des conditions modifiées de température et de pH n’avait jamais été fait sur les acanthasters, et les différences de résultats entre le projet doctoral et les études précédentes ayant omis cette phase d’acclimatation soulignent l’importance de cette phase pour de futurs travaux. Bien que nos travaux aient été réalisés sur une période spécifique dans des conditions contrôlées, qui sont différentes des conditions naturelles, il semble que les acanthasters se soient acclimaté aux conditions modifiées et ont produit des larves résistantes. Le changement climatique étant un phénomène lent, il est très probable que les acanthasters s’acclimatent aux conditions futures et que la dynamique des populations ne soit pas affectée. Nos travaux ont permis d’augmenter les connaissances des effets du changement climatique sur les acanthasters et de préciser la période optimale de reproduction en Nouvelle-Calédonie.
... This is referred to as a COTS 'outbreak'. COTS outbreaks have been recorded on the GBR for decades, with the first documented observations in 1962, 1979, and 1993-94 (Brodie et al., 2005Fabricius et al., 2010). COTS outbreaks are a major driver of coral reef decline and are thought to be responsible for 42% of coral loss in the GBR between 1985 and 2012 with (major) consequences for reef resilience and recovery (Mellin et al., 2019;Vercelloni et al., 2017). ...
... An alternative xii view links COTS population outbreaks to anthropogenic induced changes to the environment of the starfish. These induced enhancement of COTS larval food supply (phytoplankton) by nutrient-enriched terrestrial runoff (Bell & Gabric, 1991;Brodie, 1992;Brodie et al., 2005;Fabricius et al., 2010;Lucas, 1982); removal of predators of adult starfish (particularly fish and large gastropods) (Mendonga et al., 2010;Sweatman, 1995) and/or destruction of potential predators of COTS larvae, particularly plankton feeding corals; and other environmental induced changes (including climate change induced changes) to the population structure of predators of larval and juvenile COTS (Chesher, 1969;Randall, 1972). ...
... Elevated nutrient concentrations on inshore and mid-shore reefs are postulated to promote outbreaks of COTS on the GBR due to nutrient-driven enhanced survival of COTS larvae (Fabricius et al., 2010). This 'nutrient hypothesis' was first proposed 40 years ago in relation to land-based nutrient runoff from high islands (Birkeland, 1982). ...
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Synthesis report on most recent research on crown-of-thorns starfish, innovations on monitoring outbreaks and improving efficiency of management strategies, to control COTS outbreaks on the Great Barrier Reef
... solaris to indicate COTS in the Pacific Ocean (Haszprunar and Spies, 2014). Although COTS are highly fecund , under normal, undisturbed conditions COTS populations remain relatively constant and their impacts on coral communities negligible (Fabricius et al., 2010) ( Figure 1A). However, in the past several decades anthropogenic activities have adversely affected the marine environment, two major impacts being an increased discharge of nutrients (Fabricius et al., 2010) and climate change (Uthicke et al., 2013). ...
... Although COTS are highly fecund , under normal, undisturbed conditions COTS populations remain relatively constant and their impacts on coral communities negligible (Fabricius et al., 2010) ( Figure 1A). However, in the past several decades anthropogenic activities have adversely affected the marine environment, two major impacts being an increased discharge of nutrients (Fabricius et al., 2010) and climate change (Uthicke et al., 2013). These factors have been linked to increased COTS pelagic larval duration (PLD) (Yamaguchi, 1973;Lucas, 1982), and this relatively long PLD, up to several weeks, is strongly implicated in enlargement of the original population and an expansion into new habitats with comparatively homogeneous populations across widespread localities Vogler et al., 2013). ...
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Recurring outbreaks of crown-of-thorns starfish (COTS) severely damage healthy corals, especially in the Western Pacific Ocean. To obtain a better understanding of population genetics of COTS and historical colonization across the Pacific Ocean, complete mitochondrial genomes were sequenced from 243 individuals collected in 11 reef regions. Our results indicate that Pacific COTS (Acanthaster cf. solaris) comprise two major clades, an East-Central Pacific (ECP) clade and a Pan-Pacific (PP) clade, separation of which was supported by high bootstrap value. The ECP clade consists of COTS from French Polynesia, Fiji, Vanuatu and the Great Barrier Reef (GBR). The Hawaii population is unique within this clade, while California COTS are included in EPC clade. On the other hand, the PP clade comprises multiple lineages that contain COTS from Vietnam, the Philippines, Japan, Papua New Guinea, Micronesia, the Marshall Islands, GBR, Vanuatu, Fiji and French Polynesia. For example, a lineage of the PP clade, which has the largest geographic distribution, includes COTS from all of these locations. These results suggest two alternative histories of current geographic distributions of COTS in the Pacific Ocean, an ECP clade ancestry or Western Pacific clade ancestry. Although further questions remain to be explored, this discovery provides an evolutionary context for the interpretation of COTS population structure which will aid future coral reef research in the Pacific Ocean, and ultimately improve reef management of COTS.
... Agricultural non-point source pollution, particularly nitrate (NO 3 ), significantly degrades water quality, resulting in various toxicological consequences for humans (Camargo and Alonso, 2006;Herrero et al., 2001) and harmful ecological impacts on aquatic ecosystems (Brodie et al., 2011;Fabricius et al., 2010;Rabalais et al., 2002). To mitigate NO 3 transport to surface water (SW) and leaching to groundwater (GW), various nutrient management strategies have been investigated and applied to agricultural lands around the world (Cassman and Dobermann, 2022). ...
Wetlands strongly interact with groundwater and surface water, influencing catchment hydrology and altering water quality. Meanwhile, catchment-scale models are typically unable to simulate groundwater-wetland interactions despite the fact that quantifying groundwater-wetland interactions can assist in better identifying locations for wetlands restoration/creation. We modified an integrated groundwater-surface water model (SWAT-MODFLOW-RT3D) to simulate water and nutrient exchange at the wetland interface. Several modifications were applied to the SWAT wetland module, and a spatial linkage was established between the module and the MODFLOW Drain Package to provide bidirectional water and nutrient exchange between groundwater and wetlands. We applied the modified model (SMR-W) to a tropical catchment in northeast Australia and quantified water and nitrate exchange between wetland, groundwater, and surface water for 28 potential locations. This study demonstrates that when groundwater heads are perched above wetlands bed, significant nitrate discharge from aquifers to wetlands occurs, which should be considered during wetland restoration planning.
... They are not considered a substantial threat in 'normal' reef populations, however their numbers can increase dramatically due to an increase nutrient supply (Brodie et al. 2005). This is thought to occur because nutrient loading increases phytoplankton abundance, which provides a reliable food source for CoTS larvae (Fabricius et al. 2010). Brodie et al. (2005) showed that when phytoplankton concentrations double, CoTS' chance of survival to adulthood can increase almost ten-fold. ...
Bali, Indonesia sits within the coral triangle and is internationally recognised for its high coral reef diversity. The health of Bali’s marine ecosystems has declined in recent decades, and this is thought to be due to threats from climate change, destructive fishing practices, pollution, outbreaks coral eating invertebrates, coral disease and unsustainable tourism. As a response, multiple conservation strategies have been introduced by the island’s communities, non-government organisations and governments, with the aim of preventing further decline, as well as restoring already degraded coral reefs. This literature review provides an in-depth analysis of the tools used to conserve Bali’s coral reefs, and compares them to those used in other countries. In light of international ‘best practice’ in coral reef conservation, this review makes suggestions on how Bali could better conserve its coral reef ecosystems. These include (1) increasing its designation of official Marine Protected Areas (MPAS) and strengthening management of existing ones, (2) creating an MPA network, (3) substantially reducing marine plastic pollution, (4) continuing artificial reef construction in degraded habitats, (5) continuing to develop Bali as an ecotourism destination, (6) increasing engagement in global science to inform marine conservation decision-making, and (7) developing more marine monitoring programmes.
... Since 1960s, coral coverage decreased at a rate of 0.2-1.5% every year in Great Barrier Reef due to predation by COTs (Fabricius et al. 2010). The density of the COTs is commonly less than one per hectare, hence they have only low effect on coral reef (Pratchett 2005). ...
The world's coral reefs are experiencing increasing volatility in coral cover, largely because of anthropogenic environmental change, highlighting the need to understand how such volatility will influence the structure and dynamics of reef assemblages. These changes may influence not only richness or evenness but also the temporal stability of species' relative abundances (temporal beta-diversity). Here, we analyzed reef fish assemblage time series from the Great Barrier Reef to show that, overall, 75% of the variance in abundance among species was attributable to persistent differences in species' long-term mean abundances. However, the relative importance of stochastic fluctuations in abundance was higher on reefs that experienced greater volatility in coral cover, whereas it did not vary with drivers of alpha-diversity. These findings imply that increased coral cover volatility decreases temporal stability in relative abundances of fishes, a transformation that is not detectable from static measures of biodiversity.
The Great Barrier Reef is inscribed on the World Heritage List for its natural values, including an abundance of marine life and extraordinary aesthetic qualities. These and the enormous scale of the Reef make it unique and a place of ‘Outstanding Universal Value’. In the twentieth century, protection of the Great Barrier Reef shifted from limiting mechanical and physical impacts on coral reefs to managing agricultural runoff from adjacent mainland to minimise environmental impacts. By the early twenty-first century, it was apparent that threats to the Great Barrier Reef were no longer a local issue. Global warming, more frequent extreme weather events and increased ocean temperatures have destroyed vast swathes of coral reefs. Conservation scientists have begun trialling radical new methods of reseeding areas of bleached coral and creating more resilient coral species. The future of the Great Barrier Reef may depend on genetically engineered corals, and reefs that are seeded, weeded and cultured. This article asks whether the Great Barrier Reef can remain a natural World Heritage site or whether it might become World Heritage in Danger as its naturalness is questioned.
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Understanding how coral reefs have developed in the past is crucial for placing contemporary ecological and environmental change within appropriate reef building timescales (i.e. centennial to millennial). On Australia’s Great Barrier Reef (GBR), coral reefs situated within nearshore settings on the inner continental shelf are a particular priority. This is due to their close proximity to river point sources, and therefore susceptibility to reduced water quality as the result of extensive modification of adjacent river catchments following European settlement in the region (ca. 1850 CE). However, the extent of water quality decline and its impact on the coral reefs of the GBR’s inner-shelf remains contentious and is confounded by a paucity of long-term (> decadal) datasets. Central to the on-going debate is uncertainty related to the impact of increased sediment loads, relative to the natural movement and resuspension of terrigenous sediments, which have accumulated on the inner-shelf over the last ~6,000 years. The main aim of this thesis was to characterise and investigate the vertical development of turbid nearshore coral reefs on the central GBR. This aim was achieved through the recovery of 21 reef cores (3 - 5 m in length) from five proximal turbid nearshore reefs, currently distributed across the spectrum of reef ‘geomorphological development’ within the Paluma Shoals reef complex (PSRC). The recovered reef cores were used to establish detailed depositional and palaeoecological records for the investigation of the (1) internal development and vertical accretionary history of the PSRC; and (2) compositional variation in turbid nearshore coral and benthic foraminiferal assemblages during vertical reef accretion towards sea level. Established chronostratigraphic and palaeoecological records were further used to assess the impact of post-European settlement associated water quality change in a turbid nearshore reef setting on the central GBR. Radiocarbon dating (n = 96 dates) revealed reef initiation within the PSRC to have occurred between ~2,000 and 1,000 calibrated years before present, with subsequent reef development occurring under the persistent influence of fine-grained (< 0.063 mm) terrigenous sediments. The internal development of the PSRC was characterised by discrete reef facies comprised of a loose coral framework with an unconsolidated siliciclastic-carbonate sediment matrix. A total of 29 genera of Scleractinian coral and 86 genera of benthic foraminifera were identified from the palaeoecological inventory of the PSRC. Both coral and benthic foraminiferal assemblages were characterised by distinct assemblages of taxa pre-adapted to sediment stress (i.e. low light availability and high sedimentation). At the genus level, no discernable evidence of compositional change in either coral or benthic foraminiferal assemblages was found, relative to European settlement. Instead, variations in assemblage composition were driven by intrinsic changes in prevailing abiotic conditions under vertical reef accretion towards sea level (e.g. hydrodynamic energy, light availability, and sedimentation rate). These findings therefore highlight the importance for considering reef ‘geomorphological development’ when interpreting contemporary reef ecological status. Furthermore, this research emphasises the robust nature of turbid nearshore reefs and suggests that they may be more resilient to changes in water quality than those associated with environmental settings where local background sedimentary conditions are less extreme (e.g. towards the inner/mid-shelf boundary). To this end, this thesis presents new baseline records with which to assess contemporary ecological and environmental change within turbid nearshore settings on the central GBR.
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Crown of thorns starfish (COTS, Acanthaster sp.) are notorious for their destructive consumption of coral that decimates tropical reefs, an attribute unique among tropical marine invertebrates. Their populations can rapidly increase from 0–1 COTS ha−1 to more than 10–1000 COTS ha−1 in short order causing a drastic change to benthic communities and reducing the functional and species diversity of coral reef ecosystems. Population outbreaks were first identified to be a significant threat to coral reefs in the 1960s. Since then, they have become one of the leading causes of coral loss along with coral bleaching. Decades of research and significant investment in Australia and elsewhere, particularly Japan, have been directed towards identifying, understanding, and managing the potential causes of outbreaks and designing population control methods. Despite this, the drivers of outbreaks remain elusive. What is becoming increasingly clear is that the success of COTS is tied to their inherent biological traits, especially in early life. Survival of larval and juvenile COTS is likely to be enhanced by their dietary flexibility and resilience to variable food conditions as well as their phenotypically plastic growth dynamics, all magnified by the extreme reproductive potential of COTS. These traits enable COTS to capitalise on anthropogenic disturbances to reef systems as well as endure less favourable conditions.
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The successful management of ecosystems depends on early detection of change and identification of factors causing such change. Determination of change and causality in ecosystems is difficult, both philosophically and practically, and these difficulties increase with the scale and complexity of ecosystems. Management also depends on the communication of scientific results to the broader public, and this can fail if the evidence of change and causality is not synthesized in a transparent manner. We developed a framework to address these problems when assessing the effects of agricultural runoff on coral reefs of the Australian Great Barrier Reef (GBR). The framework is based on improved methods of statistical estimation (rejecting the use of statistical tests to detect change), and the use of epidemiological causal criteria that are both scientifically rigorous and understood by nonspecialists. Many inshore reefs of the GBR are exposed to terrestrial runoff from agriculture. However, detecting change and attributing it to the increasing loads of nutrients, sediments, and pesticides is complicated by the large spatial scale, presence of additional disturbances, and lack of historical data. Three groups of ecological attributes, namely, benthos cover, octocoral richness, and community structure, were used to discriminate between potential causes of change. Ecological surveys were conducted along water quality gradients in two regions: one that receives river flood plumes from agricultural areas and one exposed to runoff from catchments with little or no agriculture. The surveys showed increasing macroalgal cover and decreasing octocoral biodiversity along the gradients within each of the regions, and low hard coral and octocoral cover in the region exposed to terrestrial runoff. Effects were strong and ecologically relevant, occurred independently in different populations, agreed with known biological facts of organism responses to pollution, and were consistent with pollution effects found in other parts of the world. The framework enabled us to maximize the information derived from observational data and other sources, weigh the evidence of changes across potential causes, make decisions in a coherent and transparent-manner, and communicate information and conclusions to the broader public. The framework is applicable to a wide range of ecological assessments.
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Numerical hydrodynamic models of the northeastern Queensland shelf, forced by regional winds and modelled boundary currents in the northern Coral Sea, are used to provide improved estimates of general flow trajectories and water residence times within the Great Barrier Reef (GBR) shelf system. Model performance was checked against a limited set of current metre records obtained at Lark Reef (16°S) and the Ribbon Reefs (15.5°S). Estimates of water parcel trajectories are derived from a series of numerical tracer experiments, with daily releases of neutrally buoyant, un-reactive particles at 320 sites along the coast between Cape York (10.7°S) and Hervey Bay (25°S). Flow trajectories and residence times for tracer particles introduced to the GBR lagoon in the southern—ca. 22°S, central—19°S, and northern reef—14°S are emphasised. For purposes of the analysis, the year was divided into two seasons based on mean alongshore current direction. Most coastal sourced tracers entering the central GBR lagoon between 16° and 20°S during the northward-current season (January–August) primarily encounter the outer-shelf reef matrix after exiting the lagoon at its northern “head” (nominally 16°S), after 50–150 days. Up to 70% of tracer particles entering in the southward-current season (August–December) eventually crossed the lagoon to the outer-shelf reef matrix, with median crossing times between 20 and 330 days. During favourable wind conditions, tracers introduced at the coast may move rapidly across the lagoon into the reef matrix. The tracer experiments indicate that most coastal-sourced tracers entering the GBR lagoon remain near the coast for extended periods of time, moving north and south in a coastal band. Residence times for conservative tracer particles (and implied residence times for water-borne materials) within the GBR shelf system ranged from ca. 1 month to 1 year—time frames that are very long relative to development times of planktonic larvae and cycling times for nutrient materials in the water column, implying they are transformed long before reaching the outer reef matrix.
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Manta tow surveys of the perimeters of reefs throughout the Great Barrier Reef (GBR) assessed broad-scale changes in hard coral cover on reefs impacted by, recovering from and unaffected by Acanthaster planci outbreaks. Mean coral cover was 16 to 40% on reefs with no history of A. planci outbreaks, depending on location on the GBR. Coral cover increased at approximately 2% yr(-1) on southern reefs, while there was no significant increase on such reefs in other regions. Hard coral cover on reefs with A. planci outbreaks declined at a mean annual rate of 6% to an average level of 9%. Coral cover on southern reefs that were recovering from sustained A. planci outbreaks increased at about 4% yr(-1) while such reefs showed an annual increase of 0.8% in the remaining regions. A total of 78% of recovering reefs showed a positive growth rate, assuming linear growth, the time for coral cover to increase by 30%, was estimated at between 5 yr and well over 1000 yr. In addition to providing regional estimates of the decline and recovery of reefs due to A. planci outbreaks, this study highlights the variability in rate of recovery between reefs and raises the possibility that not all reefs will recover from sustained outbreaks.
CotSim is a size-structured metapopulation model of the crown-of-thorns (Acanthaster planci) on the central Great Barrier Reef (GBR). The populations of starfish and the coral cover on 269 individual reefs are modelled for up to 200 years. Starfish are represented as larvae, two age classes of juveniles and three size classes of adults. Coral can either be modelled as a single type or as two types each with a characteristic growth rate, equilibrium cover and susceptibility to starfish predation. Reefs are connected using simulated dispersal data for A. planci on the central GBR. These data were generated using a particle tracking program where simulated currents displaced particles representing dispersing larvae after an A. planci spawning episode. The dispersal data represented patterns expected from the 1976/77 to 1989/90 spawning season. The starfish growth model is a density-dependent matrix model. When coral cover is low, survival within classes is law and the transitions into larger classes is impeded. In contrast, at high coral cover the reverse patterns occur. Both the starfish and coral data are filtered through an interpretation model to generate observed patterns. The starfish interpretation model represents the important difficulty in detecting smaller adults. Results from the model using the default parameters correspond with published patterns of starfish/coral dynamics and the overall patterns of starfish outbreaks on the GBR. The model is an interactive event-driven 32-bit Windows application requiring Windows 95 or Windows NT 3.51/4.0. Most parameters are able to be altered by the user with three tabbed dialogue boxes (for the simulation, starfish and coral parameters). Biologically justifiable default parameters are provided for all parameters. Parameters and initial starfish populations are stored in simple coded ASCII files. Simulations are controlled using 'Run', 'Pause/Continue' and 'Stop' operations. Maps of the GBR illustrate the spatial and temporal structure of the metapopulation dynamics including the patterns of dispersal. Once paused, populations on individual reefs can be examined using two types of plots (time series and single time bar charts). Overall patterns can be displayed using latitude versus time plots of observed reef slate. Starfish populations and coral cover can be edited, which enables users of the model to become associated with some of the key issues regarding large-scale starfish control programs. Results from the model can be written to ASCII files for additional analysis. The speed of a simulation is able to be controlled and colours for important graphical elements can be altered. CotSim includes indexed online context-sensitive help and a graphical install routine. The program adheres to published guidelines for Windows applications.
Synechococcus was more abundant and had a greater biomass than Prochlorococcus at most inshore and mid-shelf sites in the central (17°S) Great Barrier Reef (GBR), and at all shelf sites in the southern (20°S) GBR. Significant Prochlorococcus populations were confined to mid-and outer-shelf sites with mixed or partially stratified water columns of greater oceanic character in the central GBR, where depth-weighted average Synechococcus and Prochlorococcus abundances were better correlated with salinity, shelf depth and chlorophyll a concentration, than with concentrations of NH4+, NOx- (i.e. NO2- + NO3-), or PO43-. Vertical gradients of normalized mean cellular red and orange fluorescence of Synechococcus and Prochlorococcus populations imply that vertical mixing rates were sufficiently low to allow these populations to photoacclimate at depth at shelf locations in the central GBR, but too greater substantial photoacclimation to be observed at sites in the southern GBR. The presence of Prochlorococcus populations at inshore sites in the central GBR in the absence of extensive intrusion events suggests that Prochlorococcus populations were actively growing.
Acanthaster planet (L.) and Nardoa novaecaladoniae (Perrier, 1875) are two coral reef asteroids having planktotrophic and lecithotrophic larval development, respectively. Comparative sizes at metamorphosis are 0.5 to 0.7 mm for A. planci and 1.2 to 1.6 mm for N. novaecaladoniae. Mortality rates of small juveniles (one month old) of each species were measured experimentally in the field on the Great Barrier Reef, Australia. Mortality rates of N. novaecaladoniae were low (1.5 %.d-1) compared to 7.8 %.d-1 for A. planci. Survival of the two species was similar between habitats. However, mortality rates of A. planci were highly variable both within- sites and between-sites within-habitats (fore reef 15 m depth, reef flat 2 m and back reef lagoon 12 m). There was no apparent effect of density of A. planci on mortality rates. Mortality is thought to be principally due to predation by infauna which are abundant in the coral reef rubble. A study of survival rates of newly metamorphosed Nardoa sp. (1.0 to 1.2 mm) in Okinawa, Japan, found very low mortality rates of just 0.2 %.d-1. The abundance of potential predators among the rubble infauna was very low on the Okinawan reef compared to the Great Barrier Reef. These studies provide evidence of the importance of predation as a determinant of survival rates of small starfish and that a reproductive strategy providing for a large size at settlement facilitates greater survivorship.