Coral bleaching events continue to drive the degradation of coral reefs
worldwide, causing a shift in the benthic community from coral- to algae-dominated ecosystems. Critically, this shift may decrease the capacity of
degraded coral reef communities to maintain net positive accretion during
warming-driven stress events (e.g., reef-wide coral bleaching). Here we
measured rates of net ecosystem calcification (NEC) and net ecosystem
production (NEP) on a degraded coral reef lagoon community (coral cover
< 10 % and algae cover > 20 %) during a reef-wide
bleaching event in February 2020 at Heron Island on the Great Barrier
Reef. We found that during this bleaching event, rates of NEP and NEC across
replicate transects remained positive and did not change in response to
bleaching. Repeated benthic surveys over a period of 20 d indicated an
increase in the percent area of bleached coral tissue, corroborated by
relatively low Symbiodiniaceae densities (∼ 0.6 × 106 cm-2) and dark-adapted photosynthetic yields in photosystem II of
corals (∼ 0.5) sampled along each transect over this period.
Given that a clear decline in coral health was not reflected in the overall
NEC estimates, it is possible that elevated temperatures in the water column
that compromise coral health enhanced the thermodynamic favorability for
calcification in other ahermatypic benthic calcifiers. These data suggest
that positive NEC on degraded reefs may not equate to the net positive
accretion of a complex, three-dimensional reef structure in a future, warmer
ocean. Critically, our study highlights that if coral cover continues to
decline as predicted, NEC may no longer be an appropriate proxy for reef
growth as the proportion of the NEC signal owed to ahermatypic calcification
increases and coral dominance on the reef decreases.
Introduction
Coral have long been the focus of climate change research in tropical
oceans as they are a keystone species responsible for the biogenic
construction of complex reef habitat (Grigg and Dollar, 1990). Adverse
effects to their ability to construct calcium carbonate structure have
negative implications for coral reef ecosystems, given corals are the major
organism responsible for collectively maintaining the accumulation of
permanent reef structure at a rate that overcomes the biological and
physical mechanisms that act to break reefs down (carbonate dissolution,
bioerosion, storm activity; Eyre et al., 2018). In contrast to coral-derived
calcium carbonate, other benthic marine calcifiers, such as non-sessile
Gastropods, Echinoderms, or Halimeda algae (Ries et al., 2009; Harney and
Fletcher, 2007), secrete calcium carbonate that is relatively temporary and
does not contribute to the long-term reef structure. Traditionally, corals
are classed as the dominant calcifier on tropical coral reefs, occupying
between 10 %–50 % of benthic area in healthy coral reef lagoons (Bruno
and Selig, 2007; Brown et al., 2004). As such, estimates of net ecosystem
calcification (NEC) are considered synonymous with the growth and function
of the entire coral reef community and can be used to represent the
collective response in coral reef community health to anthropogenic
stressors such as ocean warming and subsequent reef-wide bleaching events
(Courtney et al., 2018).
Presently, records of coral reef NEC during a reef-wide bleaching event
(driven by sea surface temperatures plus 1 ∘C above monthly
maximum means; Heron et al., 2016; Sully et al., 2019) are rare (McMahon et
al., 2019). The effects of bleaching events, and their associated thermal
seawater temperature anomalies, on coral reef NEC have been predominately
studied ex situ using recreated communities in aquaria (Dove et al., 2013) or
scaling up the response from organism-level studies, both ex situ (Castillo et al.,
2014) and in situ (Cantin et al., 2010). In studies conducted ex situ in aquaria, a
warming treatment strong enough to cause bleaching (between 1–4 ∘C above the summer mean) reduced coral calcification rates by 30 %–90 % (Cantin et al., 2010; D'Olivo and McCulloch, 2017). In situ observations
following bleaching events have shown a 20 %–90 % reduction in
individual coral calcification rates (Castillo et al., 2014) and a
significant reduction in the coral endosymbiont photosynthetic yields
(evidence of damage to their photosystems; Warner et al., 1999). At the
whole community level, the few in situ studies that have observed community
metabolism during a bleaching event recorded a 40 % (Dongsha Atoll, Taiwan; DeCarlo et al.,
2017) to 100 % (Kaneohe
Bay, Hawaii; Courtney et al., 2018; Kayanne et al., 2005; Palau) decline in reef NEC. This effect
has been observed to linger 6 to 12 months after these events, with
NEC remaining depressed by as much as 40 %–46 % (Lizard Island; McMahon
et al., 2019) and an ultimate loss of 30 %–90 % of the benthic coral
cover (Brown and Suharsono, 1990; Baird et al., 2002). Experiments with
simulated communities in aquaria (e.g., Dove et al., 2013) validate these
organism- and community-level in situ studies, in which this same magnitude of warming leads to a reduction in the experimental community coral cover by 30 %, a
70 % decline in NEC, and subsequent out-competition of corals by
neighboring algae.
The overgrowth of algae has been mirrored in the natural reef lagoon
environment several times following bleaching events (Hughes et al., 1999;
Diaz-Pulido et al., 2009). Despite a recovery to normal pre-disturbance NEC
within 2 years following a 2014 bleaching event at Lizard Island (Pisapia
et al., 2019), there was a permanent shift from coral to algae as the
dominant benthic community member, with a decline in coral cover from 8 %–3 % along transects established at the southeast end of the lagoon
(McMahon et al., 2019). This response has been seen elsewhere on the Great
Barrier Reef, where reef-wide bleaching events lead to the overgrowth of
unpalatable Lobophora variegata algae (Diaz-Pulido et al., 2009) to the extent that coral
became a minority constituent (∼ 2 %–5 %) in the lagoon's
benthic community. This transition to an algal-dominated reef community
jeopardizes the efficacy of NEC as a proxy for reef growth given that
hermatypic corals can no longer be considered the dominant benthic organism
(Courtney et al., 2018). Similar questions have been raised after other
anthropogenically driven stress events (e.g., eutrophication and
sedimentation; Edinger et al., 2000) in which coral growth rates on undisturbed
reefs did not differ from those measured on polluted, algal-dominated reefs
where habitat structure was clearly degrading. If the community
predominantly becomes covered in algae and the habitat structure is visibly
degrading, does NEC still represent reef growth or does it now reflect a
greater proportion of ahermatypic organism calcification not contributing to
permanent structure?
Shifts from coral- to algal-dominated reefs without the concomitant decline in
NEC have been observed by Kayanne et al. (2005; 7.1 % coral cover),
in which no change in NEC on Shiraho Reef, Japan, was measured despite 51 %
of the corals bleaching during a 1998 bleaching event and a decline to
5.8 % coral cover. This study suggested that continued calcification by
living, unbleached corals, calcifying algae, or other benthic calcifiers
(e.g., foraminifera, gastropods, echinoderms) may have compensated for any
expected bleaching-driven decline in coral calcification. This discrepancy
between Kayanne et al. (2005; no change in NEC on a reef with < 10 % coral cover) and that of other NEC estimates during a bleaching event
(decline in NEC on a reef ≥ 10 % coral cover; DeCarlo et al., 2014)
may be due to a critical threshold in the relationship between NEC and
percent coral cover. This is of specific concern when using NEC to monitor
community function (i.e., the net accretion of reef structure) during coral
bleaching or other disturbance events on future, degraded reefs where algae
will likely become the dominant benthic member.
To address these emerging concerns, this study investigated community
metabolism on a degraded coral reef community (coral cover < 10 %, algae cover > 20 %) during a bleaching event at Heron
Island on the Great Barrier Reef in February 2020. Flow-metabolism
transects were established on two areas within the Heron Island lagoon, and
estimates of community metabolism (net ecosystem production, NEP, and NEC), coral metaorganism function
(photosynthetic yields, Symbiodiniaceae densities), benthic cover, and
bleaching extent (percent bleached coral tissue) were assessed during the
period of peak thermal stress.
Materials and methodsStudy area
This study was conducted from 15 January to 10 February 2020.
Two separate 200 m × 100 m lagoon sites (lagoon sites 1 and 2; Fig. 1) that each
differed in total coral cover were established on the southern side of the
Heron Island lagoon (23∘26′670′ S, 151∘54.901′ E).
Community metabolism, physiochemical data, benthic community cover, and
bleaching extent were then repeatedly measured on each transect over a
period of 20 d. HOBO temperature loggers (Onset, USA), which recorded
temperature (∘C) at an interval of 15 min, were
deployed at nine upstream and downstream locations (1–9) across the study
area (Fig. 1). Overlapping loggers located at the middle deployment
locations (2, 5, and 8) were used for both lagoon sites 1 and 2, resulting in
six loggers per site.
To measure the accumulation of temperature stress above the local bleaching
threshold (defined here as the maximum of the monthly means, MMM + 1 = 28.3 ∘C; Liu et al. 2014), mean temperatures across all nine
loggers were used to calculate the number of degree heating weeks (DHWs),
which represents the 12-week accumulation of temperatures above the MMM
(Heron et al., 2016). Because HOBO temperature loggers may record higher
temperatures than surrounding seawater due to internal heating of the
transparent plastic casing (Bahr et al., 2016), HOBO loggers were deployed
in the shade on a cinder block, and downloaded temperature data were corrected
for precision (48 h side-by-side logging of all nine loggers in an aquarium)
and accuracy (deployment next to Hanna HI98194 multimeter recording
temperature). Light loggers (2π Odyssey PAR sensor) were deployed
within the middle of each study site (n=1 per site). Loggers were
attached to a star picket to ensure the sensor was exactly 20 cm above the
benthos and recorded light intensity at 15 min intervals. Odyssey light
logger data were converted to micromoles quanta of photosynthetic active
radiation (PAR) per square meter per second (µmol quanta m-2 s-1) using a linear calibration over a 24 h
period with a 2π quantum sensor LI-190R and a LI-COR LI-1400 m
(R2=0.92).
Benthic community surveys
The benthic community along each 200 m transect was described using four
survey approaches: (1) point-contact surveys, (2) photo-quadrat surveys, (3) mobile invertebrate counts, and (4) invertebrate and algal taxonomy
descriptions. For the (1) point-contact surveys and (2) photo-quadrat surveys,
benthic cover was categorized as coral (hermatypic, live), coral (bleached),
coral (soft), algae (fleshy, non-calcifying), other calcifier (e.g.,
Halimeda spp.), rubble, and sediment. For the point-contact method, the occupier of
benthic space was recorded underneath each 1 m interval (n=200 per transect) at the beginning and end of the study, and data are
presented as relative percent cover. These surveys were repeated twice per
transect at the beginning of the study (18–20 January 2020) to provide an
initial understanding of the community assemblage prior to flow-metabolism
measurements. For the (2) photo-quadrat method, a photo of a 1 m2 PVC
quadrat was taken at every 5 m interval (n=40 per transect) three
times throughout the study: (1) at the beginning prior to any observed
bleaching (24 January 2020), (2) in the middle after the first observed bleaching
event (6 February 2020), and (3) at the end of the study after several more
observed bleaching incidents (13 February 2020). These images were analyzed in
ImageJ using one side of the photo quadrat to set the scale (1 m) and the
area tracing tool to calculate the relative percent area of each category over
time.
For mobile invertebrate surveys, a transect tape was laid along each 200 m
transect length, and relatively large, easily visible mobile invertebrates (e.g.,
sea cucumbers, sea hares, sea urchins) located 1 m to the left or right
along the transect were counted. Surveys were conducted at dawn to ensure a
balance of visibility and invertebrate activity and repeated three times along
each transect (n=9 per site). Data are presented as abundance counts
per square meter (individuals m-2). Individuals present at less than 0.1 m-2 were excluded from the final data reported but were included as
part of the invertebrate taxonomy described below. For general invertebrate
taxonomy, while conducting the survey approaches detailed above, each time a
new invertebrate morphospecies was encountered, photographs were taken and
uploaded to iNaturalist, a biodiversity citizen science platform where
identifications are contributed in real time by both amateur naturalists and
professional taxonomists as part of a consensus system (https://www.inaturalist.org, last access: 10 October 2020). Using a combination of taxonomic keys and
crowdsourcing via iNaturalist, algae, corals, and other sampled marine
invertebrates were identified to as fine a taxonomic level as possible.
These data are presented as presence/absence across the entire 200 m × 400 m
study area. Because sampling was conducted at low tide, most fish usually
present in the lagoon were absent and excluded from benthic survey data.
Bleached coral physiology
Following the qualitative appearance of bleaching (white corals in photo
quadrat surveys), efforts were made to provide physiological data that would
corroborate bleaching observations. This was accomplished through
Symbiodiniaceae density analyses for both Acropora spp. (Acropora aspera, Acropora millepora, Acropora muricata, Acropora humilis) and “Other” corals
(Pocillopora damicornis, Isopora palifera, Porites cylindrica, Montipora digitata). For photophysiology, replicate coral fragments (n=∼ 15–35 per time point) of both Acropora spp. and “Other” corals were
collected across all transects at lagoon sites 1 and 2 by hand on 4 and
9 February 2020 (once bleaching was apparent) and used to measure photosynthetic
efficiency of in hospite Symbiodiniaceae cells. Measurements of photosystem
II dark-adapted yield were taken using a pulse-amplitude modulated (PAM)
fluorometer (MAXI Imaging PAM, Waltz, Effeltrich, Germany) using imaging PAM
analysis (n=3 technical replicates per fragment).
Rates of daytime net ecosystem production (NEP; mmol O2 m-2 h-1) and net ecosystem calcification (NEC; mmol CaCO3 m-2 h-1) were estimated daily (tides and full sunlight permitting) over the
course of 20 d (22 January to 12 February 2020) along the six transects. To estimate
rates of NEP and NEC, changes in dissolved oxygen (DO) and total alkalinity
(AT) were measured, respectively, during a 3 h window around low
tide and peak sunlight using both the slack-water and flow-respirometry
(Eulerian) approaches. Because differences in sunlight are a major driver in
NEP variability, measurements were refined to days of full sunlight and low
tides coinciding with near midday (11:00–15:00). Flow speeds across the
transect were measured with an acoustic Doppler velocimeter (ADV; SonTek, cm s-1) recording data at 15 min intervals. This ADV was placed at the
end of the middle transect (Fig. 1). Depth varied between 0.1–1 m and
was measured concurrently with water sample collections at each location.
Depth was also measured at peak low tide at 5 m intervals along each transect
(n=120 site 1) to ensure that sample location depths adequately
represented the entirety of the transect.
Salinity (psu) and DO (mg L-1) was measured with a
Hanna HI98194 multimeter, and DO was converted to micromoles per kilogram (µmol kg-1) using
seawater density. DO probe calibration was performed weekly using a
two-point calibration at 0 % (sodium thiosulfate) and 100 % saturated
seawater equilibrated with the atmosphere. Samples for AT were
collected in 60 mL sample polycarbonate sample bottles, preserved with
saturated mercuric chloride according to CO2 best practices (Dickson,
2007), and sealed with a screw top lid and parafilm. Seawater AT was
analyzed by potentiometric titration using a Metrohm 848 Titrino plus
automatic titrator (∼ 40 mL of seawater per sample) in
duplicates (SD uncertainty < 2 µmol kg-1). Overall
analytical uncertainty for AT (SD =± 2.4 µmol kg-1) measurements was estimated from repeated measurements of
certified reference materials from the Scripps Institute of Oceanography
(CRM; Batch 161).
Eulerian approach
Flow metabolism transects were established along a reef area previously
characterized as degraded, where there is less than 10 % coral cover
(Roelfsema et al., 2018). The flow-respirometry (i.e., Eulerian approach)
measurements were conducted within two designated reef areas (100 m × 200 m;
0.02 km2) that significantly differed in coral cover. The defined study
area was determined based on the necessary transect length to achieve
measurable differences in seawater dissolved oxygen (ΔDO =± 4–7 mg L-1) between upstream and downstream locations
(∼ 200 + m; Langdon et al., 2010).
Repeated deployments of fluorescein dye packets across the research zone at
differing tidal periods determined a specific 400 m × 100 m area of the reef
where flow was unidirectional from east to west. This period spanned from 2 h before to 1 h after peak low tide (3 h total). Outside of this
period, the reef lagoon was no longer physically separated from the open
ocean, flow became multidirectional, and the defined lagoon area became too
deep and diluted with open ocean water to measure significant changes in
seawater chemistry. The 400 m × 100 m area was then designated as two. The
spread of the dye path varied ± 25 m in a north–south direction, and
triplicate 200 m transects were spaced 50 m apart in parallel at each site so that NEC and NEP were averaged across the three downstream locations,
representing all potential water flow paths of the overall study site area.
A flow meter was rotated between downstream water sample collection
locations (n=3 per sampling location), and the determined continued
placement of the one available ADV at the middle downstream location was
adequate to represent flow speed across all three transects. Within each
area, three 200 m transects were established in parallel, 50 m distance from
one another (Fig. 1). Water samples were collected as close in time as
possible at these fixed upstream and downstream locations (n=3 per area) at peak low tide, while lagoon currents were unidirectional,
running east to west.
1NEP=3600100×ΔDO×ρ×u×dl2NEC=3600100×0.5×ΔTA×ρ×u×dl
The Eulerian approach requires the following measurements: the change in DO
and AT (ΔDO and ΔAT; mmol kg-1), the mean
seawater density (ρ; kg m-3), the mean current speed (cm s-1), the mean depth over the transect (d; meters), and the length of
the transect (l; meters). For specific details on the arrangement of the
equations above, including the 3600/100 parameter (to convert cm s-1 to m h-1), please refer to Langdon et al. (2010).
Slack-water approach
The slack-water approach was used to estimate rates of NEP and NEC over a
relatively larger area of reef (∼ 0.3 km2) during a
period of 3 h around low tide. This period was chosen based on
initial observations of current speed and direction that aligned with
previous slack-water estimates on this specific area of the Heron lagoon
(Stoltenberg et al., 2020). Starting 2 h before peak low tide, the
lagoon becomes separated from the open ocean, and the current begins flowing
unidirectionally toward the lagoon outlet to the west. This unidirectional
flow behavior continues until roughly 2 h after peak low tide; at that
time the flow begins to reverse as the tide fills back in over the reef
crest. To avoid dilution with the open ocean and changing current vector
directions confounding residence time estimates, water samples were
collected from the same three locations (n=3 d-1) 2 h
before peak low tide and 1 h following.
3NEP=ΔDO×ρ×dΔt4NEC=0.5×ΔAT×ρ×dΔt
The slack-water approach requires the following measurements: the change in
DO and AT (ΔDO and ΔAT; mmol kg-1), the mean
seawater density (ρ; kg m-3), mean depth over the transect (d;
meters), and time between sampling (Δt; hours). Given the time
between samples (∼ 3 h) and mean current speeds
(∼ 20 cm s-1), these measurements represent a transect
length of roughly 2.5–3 km of reef.
Approach comparison
Both approaches to estimate NEP and NEC provide limitations and advantages
with respect to each other (see Langdon et al., 2010). In the Eulerian
approach, the exact benthic area contributing to measured changes in
seawater chemistry is known, and its constituents can be quantified and
related to the calculated rates of benthic metabolism. This approach,
however, measures change in alkalinity over a relatively smaller area and
time period. Resulting fluxes in AT (± 30–60 µmol kg-1) and DO (± 20–50 µmol kg-1) are relatively
small compared to the slack-water approach, thereby providing less
confidence in calculated rates of benthic metabolism.
In contrast, the slack-water approach benefits from the relatively large
changes in total alkalinity (AT: ± 100–200 µmol kg-1) and dissolved oxygen (DO: ± 80–150 µmol kg-1),
which provides more confidence in AT anomaly calculations and represents
a large area of the reef flat relative to this study's flow-respirometry
estimates. This approach, however, lacks specificity of the exact area of
reef affecting changes in chemistry, and DO fluxes are more vulnerable to gas
exchange anomalies. As such, relating metabolic rates to the benthic
community provides uncertainties given daily changes in mean current speed
and, subsequently, the area of benthos reflected in the AT and DO
anomaly.
Overall, the combination of both approaches can work in tandem to compensate
for their respective weaknesses. However, neither approach can accommodate
dilution with the open ocean, and they generally need to be conducted in full
sunlight or darkness so that community metabolism does not transition
between autotrophy and heterotrophy in the middle of the measurements. For
this reason, community metabolism estimates were paused from 27 January–2 February
when peak low tide occurred around dawn and dusk, and changes in DO and
AT were negligible.
Air–sea gas exchange corrections
NEP estimates were corrected for the air–sea gas exchange (FO2) of
oxygen using the gas-transfer velocity relationships outlined by Wanninkhof (1992) and Wanninkhof et al. (2009). FO2 was calculated with the
following equation.
FO2=kK0fO2water-fO2air,
where k is the gas transfer velocity (calculated using and averaged daily
wind speed from BOM data), K0 is the gas transfer coefficient, fO2water
is the concentration of seawater dissolved oxygen (mg L-1) at the time
of the downstream measurement, and fO2air (mg L-1) was assumed to be
100 % saturation at the air temperature over the 3 h measurement period
(∼ 8.10 mg L-1).
Statistical analyses
All statistical analyses were performed with the SPSS statistics software
(SPSS Inc. 2013 Version 26.0). To compare measured differences in benthic
cover (percent coral, percent algae, percent bleached coral tissue, sediment
overgrowth) and community metabolism (NEP and NEC) between triplicate
transects, measurement days (n=12), and lagoon sites (lagoon site 1,
lagoon site 2, and slack water), a one-way analysis of variance (ANOVA)
model was used where transect, day, or site was a fixed effect, and measured
values for percent cover, NEP, and NEC were treated as the response
variable. Results for percent cover compared among triplicate transects and
lagoon sites are displayed in Tables S1 and S2, respectively. Before
community metabolism measurements were compared, assumptions of normality
and equality of variance were evaluated with a Shapiro–Wilk test (Table S4).
Results for community metabolism compared among triplicate transects,
measurement days, and lagoon sites are displayed in Tables S5, S6, and S7,
respectively. A Tukey HSD post hoc test was used to perform pairwise
comparisons for measured NEC between lagoon site 1, lagoon site 2, and the
slack-water approach (Table S7). To explore relationships between NEC as a
function of NEP, Model II regression techniques were used to test for
significant linear relationships (cutoff value p< 0.1), and an
ANCOVA was used to test for differences in NEC vs. NEP slope categorized by
lagoon site (lagoon site 1 and lagoon site 2).
ResultsLagoon community assemblage
Across the whole study area (lagoon site 1 and lagoon site 2 combined), the
benthic community was predominately covered by sediment (59 ± 7 %)
and fleshy algae (25 ± 6 %). Coral cover (5 ± 3 %) was
slightly higher relative to other recorded sessile calcifiers (4 ± 1 %) and carbonate rubble covered in coralline algae (5 ± 2 %).
Algae was the dominant benthic organism in both lagoon site 1 (28 ± 4 %) and lagoon site 2 (22 ± 4 %), and cover was significantly
higher at lagoon site 1 (p=0.011) (Table 1). Lagoon site 2 exhibited a
significantly higher coral coverage (8 ± 3 %) relative to lagoon
site 1 (3 ± 2 %) (p=0.001), the majority of which were A. aspera, A. millepora, and M. digitata. A
description of the mobile and sessile invertebrate diversity is described in
Fig. 2 and the Supplement (Table S4). A full list of observed
invertebrates and accompanying photos can be found at https://www.inaturalist.org/projects/heron-island-survey-corals-inverts-and-algae, last access: 10 October 2020.
Percent cover (mean ± SD) measured during point-contact and
photo-quadrat surveys. Data for point contact surveys were pooled across
triplicate transects and repeated survey efforts (n=6 per site)
within each lagoon site area. Data for photo-quadrat surveys were pooled
across triplicate transects and repeated survey efforts within each lagoon
site area (n=360 per site).
Overall, we found 25 coral species in the lagoonal reef study area, 22 of
which were hard corals and 3 soft corals (Fig. 2; Table S8). A total of 13
algae morphospecies were observed, with one identified as species Valonia ventricosa and the
rest unidentified. Across all other invertebrate taxa, 19 species of
echinoderms, bivalves, and polychaetes and 24 species of crustaceans and
gastropods were observed. Of the 43 non-coral invertebrate species, 15 were
associated with colonies of Pocillopora corals. Sea cucumbers (e.g., Holothuria spp., Stichopus spp.) were
the dominant mobile invertebrate, and the lollyfish sea cucumber (Holothuria atra) was the most
common across both lagoon sites (1.2 ± 0.2 individuals m-2).
Second in abundance was the Herrmann's sea cucumber (Stichopus herrmanni) (0.4 ± 0.1 individuals m-2). Other notable invertebrates included Linckia sea
stars (Linckia guildingia, Linckialaevigata) and white-speckled sea hares (Aplysia argus) (all found in abundances < 0.1 individuals m-2). The largest mobile invertebrates observed were
Bailer Shell snails (Melo amphora) at 30 cm in length and white-spotted hermit crabs
(Dardanus megistos) occupying Bailer shells (< 0.1 individuals m-2).
Cross section of coral, algal, and invertebrate diversity observed
within the study area. (a)Dipsastraea sp.; (b)Stylophora pistillata; (c)Montipora digitata; (d)Sarcophyton sp.; (e)Acropora sp.; (f)Pocillopora sp.; (g)Platygyra sp.; (h)Acropora secale; (i)Porites attenuata; (j)Halimeda sp.; (k)Holothuria atra; (l)Dardanus megistos; (m)Trapezia serenei; (n) assemblage of Caulerpa sp. and Laurencia sp. algae covered in
scum sp.; (o)Linckia laevigata; (p)Stichopus herrmanni; (q)Melo amphora; (r)Tridacna maxima.
Our observations included eight species with a conservation status of near
threatened or higher, including the small giant clam Tridacna maxima, Herrmann's sea
cucumber (Stichopus herrmanni), and six coral species (Porites attenuata, Acropora secale, Isopora palifera, Stylophora pistillata, Favites halicora, Favites rotundata). Notably, our observation of
the aglajid slug Tubulophilinopsis gardineri is one of just five from Heron Island, representing the
southernmost limit of its eastern coast distribution. We also observed an
undescribed nudibranch species, a yellow-brown Gymnodoris. A complete list
of all species described can be found in the Supplement (Table S8).
Lagoon light and temperature
Temperature across lagoon site 1 exhibited a mean value of 28.6 ± 1.5 ∘C and varied between a minimum of 25.8 ∘C and a
maximum of 34.8 ∘C (Table 2). Light at lagoon site 1 exhibited a
mean value of 328 ± 247 µmol quanta m-2 s-1 and maximum
values of 1001 µmol quanta m-2 s-1 (Fig. 1). Temperature
across lagoon site 2 exhibited a mean value of 28.6 ± 1.5 ∘C and varied between a minimum of 25.9 ∘C and a maximum of 34.6 ∘C. Light at lagoon site 2 exhibited a mean value of 336 ± 254 µmol quanta m-2 s-1 and maximum values of 969 µmol quanta m-2 s-1.
Mean values for physiochemical parameters measured at lagoon site 1
and lagoon site 2 over the course of the study. Temperature and light were
logged continuously at 15 min intervals. Temperature data are separated by
the pre-bleaching period (22 January–1 February 2020) and bleaching period (2–10 February 2020). Salinity was measured with each collected water sample (n=60 per site). Depth was measured at peak low tide at 5 m intervals
along each transect (n=120 per site). The flow meter was rotated
between downstream water sample collection locations on each day of
collection (n=5 per site).
Satellite monitoring data (5 km pixel resolution; NOAA Coral Reef Watch; Liu et al., 2006)
indicated the accumulation of heat stress beginning on 1 February 2020.
Lagoon temperatures peaked 3 d following 4 February (Fig. 1) at
which time the first signs of coral bleaching were anecdotally observed
within the study area and in other areas of the Heron lagoon. Over the
course of the study period a total of 3.59 DHWs were accumulated. In the
periods before and after the accumulation of heat stress (1 February 2020), lagoon site 1 mean temperatures were 28.1 ± 1.4
and 29.0 ± 1.5 ∘C, respectively, and lagoon site 2 mean
temperatures were 28.0 ± 1.3 and 29.1 ± 1.5 ∘C, respectively. Further details on recorded light and
temperature data can be found in the Supplement (Table S5).
Lagoon community bleaching extent
Dark-adapted yield was 0.662 ± 0.010 for Acropora spp. fragments and 0.576 ± 0.020 for “Other” fragments (mean ± SE, n=35) on 4 February. On 9 February, yield declined 35 % for Acropora spp. to 0.430 ± 0.014 (n=15) and 25 % for “Other” fragments to 0.434 ± 0.018 (n=20). Symbiodiniaceae densities were 0.976 ± 0.135 × 106 cm-2 for Acropora spp. (n=15) and 0.507 ± 0.160 × 106 cm-2 for “Other” fragments (n=10) on 30 January. On 12 February, Acropora spp. densities had declined by 48 % to 0.504 ± 0.0849 × 106 cm-2 (n=15) and by 18 % for “Other” fragments to
0.414 ± 0.094 × 106 cm-2 (n=15) (Fig. 3).
Dark-adapted yield (yellow; top left), Symbiodiniaceae densities
(green; top right), rates of net ecosystem production (NEP; middle), and net
ecosystem calcification (NEC; bottom) at lagoon site 1 (gray), lagoon
site 2 (black), and the larger reef area (dashed; slack water). Dashed
yellow and green lines indicate expected healthy values for dark-adapted
yield and Symbiodiniaceae densities, respectively. Vertical gray lines
indicate the date of photo-quadrat surveys and the resulting percent area of
coral that was bleached. NEP and NEC estimates were paused between 26 January and
3 February due to low tides occurring at dawn and dusk in low light conditions,
preventing estimates of NEC. Slack-water estimates are excluded from the NEP
data given the large error associated with air–sea gas exchange corrections.
Altogether, the percentage of coral tissue exhibiting bleaching increased
from 0 % to 60 ± 11 % over the course of the three photo-quadrat
survey efforts (Table 3; Fig. S1). Reef sediment was found to exhibit
increased growth of green and red microbial biofilms, which grew in cover
from 2 ± 1 % to 12 ± 4 %. Coral bleaching observed during
the study period was confirmed by PAM fluorometry (dark-adapted yield;
Fv/Fm) and Symbiodiniaceae densities (cells × 106 cm-2) measured
during observed bleaching (Table S6).
Change in the relative percent area (mean ± SD) of coral
tissue exhibiting paling or bleaching (bleached coral tissue) and relative
percent area (mean ± SD) of sediment exhibiting overgrowth in the form
of visible cyanobacteria mats or Chlorophyta growth (overgrowth on sediment)
over the course of three different survey efforts. Data for each date are
pooled across parallel transects within each lagoon site (n=120 per site).
Study site24 Jan 20206 Feb 202012 Feb 2020BleachedLagoon site 10 ± 0 %16 ± 3 %55 ± 8 %coral tissueLagoon site 20 ± 0 %24 ± 6 %65 ± 10 %OvergrowthLagoon site 12 ± 1 %4 ± 2 %10 ± 2 %on sedimentLagoon site 23 ± 1 %5 ± 3 %14 ± 5 %Lagoon community metabolism
The mean ± SD value of NEP and NEC at lagoon site 1 and lagoon site 2
(pooled together across triplicate transects and measurement days; n=36) is displayed in Table 4 and Fig. 3 and separated by the pre-bleaching
(22 January to 1 February 2020) and bleaching periods (2 to
10 February 2020). Mean daytime net ecosystem production (NEP), averaged
across all days and sites, was 39.4 ± 12.2 mmol O2 m-2 h-1. NEP did not significantly differ across triplicate transects
within lagoon site 1 (p=0.471) or lagoon site 2 (p=0.917), so these
data were pooled together to represent the overall community NEP of each
site (Fig. 3). The measured NEP throughout the study period was highly
variable and did not significantly differ over time (n=12) at either
lagoon site 1 (p=0.181) (lowest coral cover site) or lagoon site 2 (p=0.099) (highest coral cover site). NEP did not significantly differ
between lagoon site 1 and lagoon site 2 (p=0.067). NEP values were not
included for the slack-water approach given the large source of error in
air–sea oxygen exchange.
Mean ± SD values for daytime net ecosystem production (NEP;
mmol O2 m-2 h-1) and net ecosystem calcification (NEC; mmol CaCO3 m-2 h-1) for lagoon site 1 and lagoon site 2, where the
Eulerian approach was used (n=12). NEC for the slack-water approach
included for daytime (n=11) and nighttime (n=3) estimates. Data are
separated by the pre-bleaching period (22 January–1 February 2020) and bleaching
period (2–10 February 2020; n=8). Nighttime rates for NEC are included.
NEP values are not included for the slack-water approach given the large
source of error in air–sea oxygen exchange.
ApproachNEP (mmol O2 m-2 h-1) NEC (mmol CaCO3 m-2 h-1) Pre-bleachingBleachingPre-bleachingBleaching periodLagoon site 135.0 ± 12.739.7 ± 9.612.5 ± 4.512.6 ± 4.8Lagoon site 244.4 ± 13.638.7 ± 13.813.3 ± 5.712.3 ± 5.4Slack water (day)11.0 ± 2.910.5 ± 3.0Slack water (night)-2.8 ± 0.7-3.4 ± 1.3
Mean daytime NEC, averaged across all days and sites, was 12.2 ± 4.5 mmol CaCO3 m-2 h-1. Measured rates of daytime NEC did not
significantly differ across triplicate transects within lagoon site 1 (p=0.471), lagoon site 2 (p=0.917) or the slack water (p=0.581), so
these data were pooled together to represent the overall NEC of each area
(Table 4). Measured NEC was also highly variable and did not significantly
differ over time at lagoon site 1 (p=0.506), lagoon site 2 (p=0.365), and the slack water (p=0.073). Estimated NEC in the slack-water
approach was significantly lower compared to Eulerian estimates at lagoon
site 1 (p=0.010) and lagoon site 2 (p=0.001); these two latter sites
did not significantly differ (p=0.666). Changes in NEC were
significantly related to changes in NEP at both lagoon site 1 (r2=0.32; p=0.042) and lagoon site 2 (r2=0.28; p=0.046). Slope
values for daytime NEC vs. NEP for lagoon sites 1 and 2 were 0.28 and 0.24,
respectively (Fig. S2).
To determine potential effects of bleaching on nighttime dissolution and
respiration, nighttime estimates of NEC and NEP were conducted three times
throughout the study near the dates of observed progressed bleaching (23 January, 4 and 12 February). However, AT and DO changes were too
small during lagoon site 1 and lagoon site 2 Eulerian estimates, so
nighttime NEC could only be confidently calculated from slack-water
estimates. We found mean slack-water nighttime NEC (-3.1 ± 1.1 mmol CaCO3 m-2 h-1) did not significantly differ across transects
(p=0.617) or over time (p=0.083) within the current study.
DiscussionCommunity metabolism response to bleaching
The southwestern lagoon area of Heron Island (southern Great Barrier Reef)
is a community characterized by low coral cover of approximately 5 %–8 %. Within this reef area, the predominant benthic cover was unpalatable
algae (approximately 21 %), dominated by the two genera Laurencia spp. and
Lobophora spp., consistent with that of a degraded coral habitat (Hughes et al.,
1999). Prior surveys of the benthic cover in this area of the Heron Island
lagoon (scientific zone) have also estimated relatively low coral cover (0 %–10 %; Roelfsema et al., 2018).
Accumulation of heat stress in the lagoon over the study period resulted in
3.59 DHWs as in situ mean temperature was elevated from ∼ 28.0 to ∼ 29.1 ∘C (+1.1 ∘C).
Over this period, we found that approximately 60 % of corals present
within both lagoon sites 1 and 2 exhibited bleaching. These bleaching
observations were corroborated by both photosynthetic yields and
Symbiodiniaceae densities of all corals sampled. Photosynthetic yields
recorded on 4 February 2020 in both the Acropora spp. and “Other” category were
barely above values considered “healthy” (0.5; Gierz et al., 2020) and,
by 9 February 2020, exhibited symbiont loss with values below 0.5 (Acro = 0.43 ± 0.01; other = Acro = 0.43 ± 0.01). Mean
Symbiodiniaceae densities across both time points for the Acropora spp. (0.74 ± 0.11 × 106 cm-2) and “Other” corals (0.46 ± 0.13 × 106 cm-2) were also below normally healthy values previously
recorded in both Acropora spp. (1–2 × 106 cm-2, Gierz et al., 2020)
and corals in the “Other” category (e.g., Montipora digitata; 2–3 × 106 cm-2; Klueter et al., 2006) collected from the Heron Island reef flat.
Despite the ongoing reef-wide bleaching event and measured decline in coral
endosymbiont densities, we find that NEP and NEC at both lagoon sites did
not significantly differ from estimates during the pre-bleaching period or
prior estimates on other Great Barrier Reef lagoon communities of similar
coral cover (e.g., 10–20 mmol CaCO3 m-2 h-1; Albright et
al., 2015; Pisapia et al., 2019; Stoltenberg et al., 2021). The lack of a
bleaching effect was also mirrored in the slack-water NEP and NEC data,
which represented a much larger cross section of the lagoon community
(∼ 2–3 km transects), where bleaching was also observed
(but not quantified during this study period). Importantly, these trends
differ from those observed by Courtney et al. (2018) during a 2015
bleaching event in Kaneohe Bay, Hawaii (∼ 10 % total
cover), where a similar ∼ 1 ∘C increase in mean
reef temperature resulted in bleaching of 46 % of the coral community, and
both NEP and NEC were driven to zero. However, our results support those of
Kayanne et al. (2005), in which NEC and NEP remained relatively constant
during a bleaching event (29 ∘C; 51 % bleached) in September
of 1998 at Shiraho Reef in Japan (5 %–7 % total coral cover). The
critical difference between these studies is likely due to a threshold in
total coral cover, in which bleaching is less impactful on NEC when coral is
not the dominant calcifying organism relative to the other calcifying
constituents (sediments, rubble, calcifying algae, and other sessile or
mobile gastropods and echinoderms) that are also known to contribute to the
total reef carbonate budget and, in some cases, exhibit positive
temperature–calcification relationships (Cornwall et al., 2019).
Estimated organism contribution to NEC at elevated temperatures
Importantly, if we consider that rubble observed in lagoon sites 1 and 2
(approximate cover of 4 %) was predominately covered in crustose
coralline algae (CCA) and combine these with the other sessile calcifiers
observed (which were predominantly Halimeda spp.; 3 % cover), then hermatypic
corals were not the dominant reef calcifier. Further, if 60 % of the
total coral cover was calcifying roughly 60 % slower due to bleaching
(D'Olivo and McCulloch, 2017), this would imply that active calcifying
coral cover was likely reduced to only 2 %–4 %. This adjusted
“calcifying percent coral cover” is minor compared to the sum of all other
benthic constituents that were actively calcifying regardless of the sea surface temperature (SST)
conditions (Sediment + CCA +Halimeda= 72 %).
One possible explanation for the lack of any observed changes in NEC could
be due to the simultaneous thermal enhancement of calcification in other
benthic members when the reef seawater was warmed from 28.0 to
29.1 ∘C. To investigate the relative contribution to overall NEC
from the assemblage of benthic calcifiers at these respective temperatures,
we created an equation based on reported rates in the literature at 28.0
and 29.1 ∘C (Eq. 1) at which the summed
community-level calcification rate (NEC) at the respective temperature (T)
is equal to the sum of the described calcification rates for each benthic
organism category (net organism calcification: NOC) multiplied by the
recorded cover (Cover) across lagoon sites 1 and 2 at that temperature (T).
NECT=∑NOCT×CoverT
To estimate the potential effect of a +1.1 ∘C change in
seawater temperature on coral calcification for corals observed within the
lagoon study sites, the following aquaria manipulation studies were reviewed:
Edmunds (2005), Anthony et al. (2008), Cantin et al. (2010), and Comeau et al. (2013, 2016); the following meta-analysis and modeling studies were
reviewed: Lough and Barnes (2000), McNeil et al. (2004), Evenhuis et al. (2015), Kornder et al. (2018), and Bove et al. (2020). Together, these studies
suggest that mean calcification rates across coral genera most common to the
Heron reef flat (Acropora spp., Montipora spp., Porites spp., Pocillopora spp.) at 28.0 ∘C (4.53 ± 2.31 mmol CaCO3 m-2 h-1) increase by approximately 22 % when warmed to a temperature of 29.1 ∘C. It is important to
note this percentage increase is highly variable and species specific, so numbers
used here are simply for the purpose of discussion. In comparison,
calcification by crustose coralline algae (CCA), which is the next most
studied organism (see meta-analysis by Cornwall et al., 2019), has not
exhibited changes until temperatures are as high as 5 ∘C above
ambient temperatures. Therefore, no change was estimated for mean reported
rates (0.36 ± 0.09 mmol CaCO3 m-2 h-1) for commonly
studied CCA species (Lithophyllum kotschyanum and Hydrolithon onkodes).
Visualization of the changes caused by a transition from
pre-bleaching (28.0 ∘C) to bleaching (29.1 ∘C)
temperatures in (a) estimated individual organism calcification rates from
the literature (converted to mmol CaCO3 m-2 h-1), (b) percent
cover across lagoon site 1 and lagoon site 2 combined, and (c) the “adjusted
calcification rate” (mmol CaCO3 m-2 h-1) calculated by
multiplying (a) by (b) at each temperature. Total change (%) represents the
percent difference in the sum of all rates at 29.1 ∘C relative to
28 ∘C. Rubble and other calcifier categories were assumed to be
CCA and Halimeda spp., respectively.
Responses in calcification to warming for Halimeda algae are equivocal
(Campbell et al., 2016; Wei et al., 2020). If constrained to species
commonly identified on the Great Barrier Reef (such as H. opuntia and H. cylindracea; Anon, 2020),
then it can be expected that increasing temperatures will increase rates of
calcification up to temperatures of 30 ∘C; above that they bleach
and exhibit a negative calcification response. As such, narrowed within the
ranges observed during this study, calcification rates of Halimeda (3.33 ± 2.29 mmol CaCO3 m-2 h-1) are estimated to increase by
approximately 7.9 % in response to warming from 28.0 to
29.1 ∘C. Calcification responses to warming in carbonate
sediments are overall the least studied of the benthic categories in this
study but potentially the most significant given the dominant cover of
sediment. A study within the Heron Island lagoon indicates that daytime sediment
calcification at 28 ∘C (1.41 ± 0.29 mmol CaCO3 m-2 h-1) would increase ∼ 9 % when seawater is
warmed to 29.1 ∘C (Lantz et al., 2017).
When these trends are summed together with the expected 60 % decline in
calcification for the proportion of coral that was bleached, a collective
9.8 % decline in NEC can be expected (Fig. 4). However, when each
category is adjusted for the percent cover observed at the end of the study
at 29.1 ∘C across both lagoon sites, the total change in NEC
increases by ∼ 0.8 %. This is largely owed to positive
trends in the calcification of other benthic community members and provides
an explanation why no significant differences were observed in NEC during
reef-wide coral bleaching. These estimates illustrate how the decline in
coral calcification may be overshadowed by thermal acceleration in
calcification in ahermatypic benthic calcifiers. Although some of these
calcifiers still accrete limestone structure (e.g., coralline algae), none
replace the complex, three-dimensional structure uniquely created by corals.
Our findings highlight the need to better adjust how NEC is applied as a
metric for community function during bleaching events as these data suggest
warming may create a divergence between estimated daytime NEC and actual
reef growth on future degraded reef ecosystems.
Future considerations
Our study highlights three considerations that may affect NEC and need to be
further investigated to resolve monitoring issues for degraded coral reef
communities. First is the impact of nighttime dissolution on overall 24 h
NEC. Estimates of NEC at night (n=3) in the current study did not
exhibit a response to bleaching, but a higher frequency is needed. Courtney
et al. (2018) hypothesized that the dissolution signal was a major driver
of the net 24 h zero NEC signal during bleaching. These findings were more
recently corroborated at the organism level by Orte et al. (2021), in which
algal turfs on dead coral calcified at the same rate as coral during the day
but transitioned to net dissolving at night. This is supported by
calcification responses to warming in the sediment, the most dominant
benthic member in this study, where warming-driven daytime increases in NEC
were largely overshadowed by nighttime increases in dissolution (Lantz et
al., 2017), and the sediments transitioned to net dissolving over the full 24 h. These results suggest that future studies need to include nighttime
measurements of NEC and NOC but also highlight the limitation of
flow-metabolism approaches as a representation of reef health given that not
all reefs are easily accessible at night for such measurements.
Secondly the longer-term changes in NEC (when bleached coral eventually dies
or the thermal benefits to other calcifiers expire) need to be investigated
if we are to accurately estimate community function in future reef
scenarios. In the current study we did not monitor the response in NEC
following the 2020 bleaching event when a return to 28 ∘C or
lower would likely reduce the thermal benefits to daytime calcification in
the sediment, rubble, live coral, and Halimeda algae that potentially masked
the minimized contribution from bleached coral. Under these assumptions, a
7.6 % decline in NEC would be expected when temperatures return to 28 ∘C. Additionally, if we assume the bleached coral eventually
dies, and a 60 % reduction to calcification increases to a 100 %
reduction, then community NEC would in theory exhibit a 13.1 % total
decline. These post-bleaching estimates may explain the differences between this study and post-bleaching NEC estimates reported on similarly degraded reef transects at Lizard Island, Australia (3 % coral cover), by McMahon et al., 2019. At Lizard Island, post-bleaching NEC in 2016 declined by 40 %–46 % relative to pre-bleaching estimates in 2008 when coral cover was higher (∼ 8 % coral).
Finally, the indirect feedbacks on NOC from non-calcifying community members
(e.g., algae) and the carbonate substrate they occupy also need to be
considered to predict future reef growth (Orte et al., 2021). The sum of
adjusted NOC (Fig. 4; 1.30 mmol CaCO3 m-2 h-1) only explains
10.6 % of the measured NEC (12.3 mmol CaCO3 m-2 h-1). Such
discrepancies may be explained by the exclusion of the 21 % of space
occupied by non-calcifying algae in the NOC summation exercise in Fig. 4. It
is possible algae can provide positive feedback mechanisms to coral
calcification through adjacent algal-driven NEP (and subsequent
modifications to the surrounding seawater carbonate chemistry; Gattuso et
al., 1998; Unsworth et al., 2012) or the endolithic micro-calcifiers living
inside the dead carbonate substrate colonized by algal communities (Orte et
al., 2021). For example, endolithic microflora (Cyanophyta and Chlorophyta)
living within carbonate rocks have been found to modify interstitial pH just
beneath the substrate surface to values as high as 8.5 (Reyes-Nivia et al.,
2013), thereby creating localized zones supersaturated with aqueous
Ca2+ and CO32- ions (Krause et al., 2019) and promoting the
inorganic precipitation of minerals such as brucite, micrite, and dolomite.
Critically, these microfloral communities are more diverse and abundant when
living beneath turf algae compared to corals (Gutiérrez-Isaza et al., 2015),
are comparable in their productivity to overlying turf algae (Tribollet et
al., 2006), and have been found to precipitate dolomite at an accelerated
rate when seawater temperatures were increased from 28 to 30 ∘C (Diaz-Pulido et al., 2014). Taken together, this shows that
these microfloral communities have the capacity to influence bulk seawater
chemistry measurements particularly during coral bleaching events, when
warm and well-lit conditions promote their growth. In addition to these
microflora, various cryptic infaunal and endolithic macrofauna calcify to
produce protective shells or burrows (e.g., Díaz-Castañeda et al., 2019) and
may also be contributing to the NEC signal measured during the bleaching event.
Conclusions
Ocean warming and subsequent coral bleaching events have already degraded
coral reef ecosystems for over four decades and will continue to degrade
coral reefs worldwide, reducing their capacity to provide a complex,
three-dimensional habitat structure. While estimates of NEC via the
alkalinity anomaly technique may be an appropriate benchmark of community
function well after bleaching events have occurred and degradation to the
coral community is fully realized, the results from this study highlight the
shortcomings of using this approach to estimate daytime NEC when monitoring
the effect of bleaching on reef accretion in real time. These results, in
conjunction with available literature on the importance of nighttime
dissolution, suggest that flow-metabolism approaches to estimate community
health may be limited to reefs accessible at night (e.g., those near a
research station or without navigational hazards). Moreover, our study
highlights that if coral cover continues to decline as predicted, NEC may no
longer be an appropriate proxy for reef accretion as the proportion of the
NEC signal owed to ahermatypic calcification increases. Additional estimates
of NEC during bleaching events are urgently needed to further explore the
potential decoupling of positive NEC and reef growth. Concerningly, the data
herein suggest that NEC may begin to exhibit limitations as a monitoring tool
for reef growth when coral becomes the minority benthic constituent.
Data availability
Data are presently being submitted to the Figshare data repository under the DOI 10.6084/m9.figshare.18733019 (Lantz et al., 2022).
The supplement related to this article is available online at: https://doi.org/10.5194/bg-19-891-2022-supplement.
Author contributions
CAL is responsible for study design, data collection, analysis, and writing. WL and TDS are responsible for study design, data analysis,
and writing. JLB, AF, CP, and TM are responsible for data collection, analysis, and writing.
Competing interests
The contact author has declared that they or their co-authors have competing interests as specified in the information for the associate editor (cover letter).
Disclaimer
Publisher's note: Copernicus Publications remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.
Acknowledgements
This work was funded by the Australian Research Council DP 180103199. We
thank the Heron Island Research Station scientific staff for their support
during research. We also thank all iNaturalist users that helped identify
the invertebrates photographed during this study, especially Joe Rowlett,
Sean Ono, Frédéric Ducarme, and Pierre Mascar.
Financial support
This research has been supported by the Australian Research Council (grant no. DP 180103199).
Review statement
This paper was edited by Jean-Pierre Gattuso and reviewed by three anonymous referees.
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