Introduction
Harmful algal blooms (HABS) are becoming increasingly prevalent throughout
the world. One of the key causes of this is eutrophication of aquatic
environments by excessive nutrient inputs (Conley et
al., 2009). Climatological and hydrological factors are increasingly
recognized as an important contributor to HABS through altered temperature
and salinity regimes (Paerl and Paul, 2012). Blooms of toxic
cyanobacteria such as Nodularia spumigena are particularly conspicuous in some estuaries such
as the Baltic Sea, the Peel–Harvey Estuary and the Gippsland Lakes (Cook
and Holland, 2012; Lukatelich and McComb, 1986; Conley et al., 2009). The most
likely reasons for their dominance in these systems are (1) a long water
residence time, (2) intermediate salinity (∼ 5–20) and (3) a
high supply of phosphorus (both from the catchment and from anoxic bottom
waters and sediments). As such, the frequency and occurrence of these blooms
is likely to result from a strong interaction between anthropogenic nutrient
loading and climatological and hydrological factors.
In the case of the Baltic Sea, cyanobacterial blooms have occurred
sporadically since the formation of the Littorina Sea 8000 years BP
(Bianchi et al., 2000). The presence of cyanobacteria is
most likely controlled by the extent of bottom water hypoxia, which leads to
an efficient recycling of phosphorus (Funkey et al., 2014). The extent of
hypoxia in the Baltic has been controlled by morphological and hydrological
changes; however, the most likely control over hypoxia and cyanobacterial
blooms over the past 2 millennia is the expansion and contraction of human
activities (Zillen and Conley, 2010). Similarly, in the Gippsland
Lakes and Peel–Harvey Estuary, the frequent and intense blooms are thought
to be relatively recent phenomena, with significant blooms only observed
after the late 1970s (McComb and Humphries, 1992). In the Peel–Harvey
Estuary, the intensity of N. spumigena blooms is strongly related to river discharge
during the previous winter and spring, which delivers significant quantities of
phosphorus as well as reducing the salinity of the estuary to a range
favourable to N. spumigena growth (McComb and Humphries, 1992). The critical
importance of salinity in controlling N. spumigena blooms is well illustrated in the
Peel–Harvey Estuary, where a newly cut channel to the sea increased the
salinity of the estuary and virtually eliminated N. spumigena blooms (Wildsmith et
al., 2009).
In the Gippsland Lakes, Australia, it has been shown that N. spumigena bloom size is
decoupled from catchment inputs owing to internal recycling of P driven by
stratification (Cook and Holland, 2012). As such, N. spumigena blooms typically
occur during wet years when stratification is highest; however, there is no
relationship between catchment nutrient loads and bloom size (Cook
and Holland, 2012). Nevertheless, anthropogenic activities are likely to
have played a role in the occurrence of recent blooms through increased
phosphorus inputs leading to increased sediment phosphorus storage as well
as increased nitrogen inputs which are likely to drive increased releases of
phosphorus from the sediment through increased anoxia (Cook et
al., 2010). Prior to European settlement, the Gippsland Lakes were connected
to the ocean by an ephemeral entrance (Bird, 1978). In 1889, a
permanent artificial entrance was opened, which increased the salinity
regime of the lakes (Saunders et al., 2008). There are anecdotal accounts
of N. spumigena blooms prior to the opening of the artificial entrance, but there is no
information on the relative frequency or intensity of blooms at this time.
Two alternative hypotheses were tested here. (1) The fresher, less flushed
and more stratified environment prior to the opening of the artificial
entrance may have been more conducive to anoxia, associated sediment
phosphorus release and cyanobacterial blooms than post opening. (2) Alternatively, low nutrient inputs prior to European settlement may have led
to a lower incidence of hypoxia and associated cyanobacterial blooms. The
aim of this study was to investigate changes in the trophic status and
frequency of cyanobacterial blooms in the Gippsland Lakes before the opening
of the artificial entrance up to the present day using pigment, organic
matter and isotope proxies on dated cores taken from the centre of the lake
system. The findings provide an important, longer-term perspective from
which to frame modern management regimes within the Gippsland Lakes as well
as systems modified by humans more generally.
The Gippsland Lakes, south-eastern Australia. The study site in
northern Lake King is marked with the solid circle (37.875620∘ S,
147.757280∘ E).
Materials and methods
Study site
The Gippsland Lakes are located in south-eastern Australia, and experience a
temperate climate with a water temperature range of ∼ 10–25 ∘C (Fig. 1). The lakes are fed by 5 river systems including the
Latrobe and Avon in the west, and the Mitchell, Tambo and Nicholson in the
east. Lake Wellington in the west is a shallow basin with an average depth
of ∼ 2.6 m and typically has a salinity < 15. Lakes
King and Victoria in the east have an average depth of ∼ 5 m.
The lakes cover an area of 356 km2, making them one of the largest
estuarine systems in Australia. Maximum river flows and floods typically
occur in the austral winter and spring, which reduce surface salinities to
∼ 5–15, which then increase to > 25 over summer in
lakes King and Victoria. Lakes King and Victoria are typically salinity
stratified, with bottom water salinities of ∼ 30–35, and during
intense stratification following high river flow, the bottom waters of lakes King and Victoria may become anoxic. Winter and spring inflows
typically lead to blooms of diatoms and dinoflagellates, and since 1987,
periodic blooms of Nodularia spumigena have occurred in Lake King during late spring and summer
when surface water salinities are ∼ 9–20 (Cook and
Holland, 2012). Previous studies have shown that these blooms are phosphorus
limited, that they are sustained by high sediment phosphorus release
focused in the northern basin of Lake King and that these blooms can fix
significant quantities of nitrogen with a δ15N of
∼ 0 ‰ (Cook and Holland, 2012; Cook et
al., 2010; Holland et al., 2012; Woodland and Cook, 2014; Woodland et al.,
2013).
Sampling
Sediment cores were taken from Lake King north (LKN; 37.875620∘ S,
147.757280∘ E; Fig. 1) at a water depth of 7 m on the 15
March 2012, which is where contemporary blooms are centred. The uppermost
core, LKN1 (0–56 cm), was retrieved by a piston corer to collect the recent,
unconsolidated sediments. This core was sectioned in the field at 0.5 cm
intervals (contiguous), to attain a chronology to aid the identification of
recent changes in the sediment. During sectioning, the core was placed in a
black plastic bag to shield it from light and once sectioned, samples were
rapidly placed in the dark. Subsamples (1–2 g) from each section were stored
in glass vials, with the rest of the sample stored in ziplock bags. All
samples were kept on ice and, on returning to the laboratory, the bags were
transferred to the refrigerator (4 ∘C) and the vials frozen until
required for stable isotope analysis. Wet samples were kept in darkness in
order to reduce light exposure that could change the sediment composition.
A second drive (LKN2) using a Russian peat corer collected a deeper, older
sedimentary sequence and consisted of a series of 50 cm (overlapping) drives
from 0 to 2.1 m. All cores were stored in halved PVC pipes, wrapped in cling
film and aluminium foil and kept cool until refrigerated in the laboratory.
The cores were sectioned into 1 cm layers using a blade and spatula and
stored in the same way as LKN1. The two sequences were correlated based on the
calculated field depths, and this was validated by dating across the two
sequences as described in the next section.
Dating
The sediment cores were dated at the Australian Nuclear Science and Technology
Organisation (ANSTO) Institute for Environmental Research using the lead-210
(210Pb) dating method (Appleby, 2001). Samples were chemically
processed following the methods described in Atahan et al. (2015) and
analysed by alpha spectrometry to determine unsupported 210Pb
activities on 13 subsamples from core LKN1 (0–51.5 cm) and the
42–92 cm LKN2 sequence. A CIC (constant initial concentration) model was
used to calculate the ages of the sediment samples (Appleby, 2001). The
210Pb chronology was validated with the presence of a subsurface peak
of caesium-137 (137Cs), which identifies the year of 1964, due to
global atmospheric nuclear weapons tests (Leslie and Hancock, 2008).
137Cs activities in eight subsamples were determined by gamma spectrometry.
Carbon and nitrogen analysis
Sediment from the LKN1 and LKN2 core sediment sample was analysed via mass
spectroscopy for percentage of nitrogen, percentage of carbon, Corg : N, δ15N
and δ13Corg. These samples were dried at 60 ∘C
for 30–50 h and placed in 1.7 mL Eppendorf tubes along with Qiagen
Tungsten Carbide Beads (3 mm); they were then shaken for 6–10 min at 25 Hz using a
Retsch Mixer Mill MM 200 until a fine, homogeneous powder was produced.
Samples for carbon (δ13Corg) were weighed in silver
capsules and placed on a hotplate (60–80 ∘C) to undergo
acidification. Aliquots (20 µL) of 10 % HCl were sequentially added
to capsules until no effervescence was recorded. Samples for nitrogen
(δ15N) analysis were weighed in tin capsules. Once each capsule
was prepared, it was pinched closed and pressed into a disk using a
pelletizer. Each sample was analysed on an ANCA GSL2 elemental analyser
interfaced to a Hydra 20–22 continuous-flow isotope ratio mass spectrometer
(Sercon Ltd., UK). Stable isotope data were expressed in the delta notation
(δ13Corg and δ15N) relative to the stable
isotopic ratio of Vienna Pee Dee Belemnite standard (R_VPDB = 0.0111797) and the air standard (R_Air = 0.0036765) for carbon and nitrogen respectively. Analytical precision was ±0.1 ‰ for both δ13Corg and δ15N (SD for n= 5).
Pigments
Pigments were analysed at 5 cm intervals from 0 to 41 cm and every 10 cm
through to 200 cm from the LKN2 core sequence. Freeze-dried sediments were
extracted in pure acetone overnight and stored in the dark at
-22 ∘C. They were then filtered, dried and redissolved under
low light conditions and then injected into a Shimadzu high-performance
liquid chromatography (HPLC) system. The separation conditions were modified
from Mantoura and Llewellyn (1983) and Chen et al. (2001)
using a 4.6 × 150 mm, 3 µm C8 (Luna, Phenomenex) column.
Pigment peaks were identified by retention times and spectra and then
quantified by peak areas at maximum absorbance wavelength using calibrated
curves from phytoplankton pigment standards DHI (Denmark). Canthaxanthin was
measured at 475 nm, and the carotenoids lutein, zeaxanthin, diadinoxanthin,
diatoxanthin and echinenone were measured at 450 nm. The pigments
zeaxanthin, echinenone and canthaxanthin were used as markers for cyanobacteria
(Jeffrey and Vesk, 1997); chlorophyll a was measured at 665 nm.
Concentrations are reported in micromoles (µmol) of pigment relative
to the organic material in the sediment measured as described above.
Diatoms
The core sequence was subsampled every 10 cm for diatom analysis from the
LKN2 core. Approximately 1 g wet-weight sediment was digested in 30 %
hydrogen peroxide in a beaker, on a hotplate, for up to 4 h
(Battarbee, 1986). Following digestion, a small amount of HCl was
added to remove any carbonates. The suspensions were washed in distilled
water and left to settle overnight before decanting the supernatant
(repeated four times). An aliquot of the final suspension was placed onto a
coverslip and left to dry. The coverslips were permanently mounted onto
slides using Naphrax. Diatoms were identified (where possible) to species
level, using a Nikon DIC microscope. Identifications were undertaken using a
range of general (Krammer and Lange-Bertalot, 1986, 1988, 1991a, b) and regional (Foged,
1978; Sonneman et al., 2000) floras. A minimum of 200 valves per sample were
counted, and the counts were converted to percentage data in C2 (Juggins,
2003). Where possible an ecological preference (i.e. saline, fresh,
thalassic) was assigned to each species to create a habitat summary diagram.
To further explore patterns in the diatom data, a detrended correspondence
analysis (DCA) was carried out using Canoco 4.5.
Unsupported 210Pb activities (Bq kg-1 sed) versus depth (a),
137Cs activities (Bq kg-1 sed) versus depth (b) and the age depth model
based on unsupported 210Pb values using the CIC (constant initial
concentration) model (c). The star in (c) refers to the depth of the
137Cs peak activity. LKN1 and LKN2 refer to the two different cores
sampled (see methods) and the horizontal dashed line demarcates the
transition from core LKN1 to LKN 2.
Charcoal
Wet sediment (1 mL) from the LKN2 sequence was subsampled into a 50 mL
Falcon tube. In total, 25 mL of 10 % tetra sodium pyrophosphate
(Na4P2O7) was added to the tube; the contents were shaken
vigorously and left to sit. After 30 min, 25 mL of 12.5 % sodium
hypochlorite (NaOCl) was added, and the tube was again shaken vigorously and
then left to sit for 14–18 h. The samples were then sieved through a 250 µm and then a 125 µm mesh, rinsed and placed on a water-filled
petri dish where the total number of charcoal and grass charcoal particles
were enumerated under a dissecting microscope.
Discussion
Impact of settlement
Prior to European settlement in the early 1840s, land use by the Aboriginal
tribes was of a nomadic hunter gathering nature, and documentary evidence
suggests that fire was the principal agent of land management across
Australia at this time (Gammage, 2011). This account is, however,
contradicted by most charcoal records from south-east Australia which show an
increased incidence of fire after European settlement (Mooney et al., 2011;
Mills et al., 2013). The high charcoal counts below 170 cm are consistent
with high rates of indigenous burning of fringing reedbeds before tubers were
harvested, which has been a recognized common practice (Head, 1987). Early
European land use was primarily low-intensity sheep grazing. Gold mining
commenced in the 1850s, followed by increased navigation of the lakes in the
1860s for the purposes of trade, fishing and tourism. By the 1870s there was
regular steamer traffic on the Mitchell River and there are regular
references to dredging the mouths of the Mitchell, Nicholson and Tambo rivers
from the early 1880s through to the turn of the century and into the 1920s
(Synan, 1989). The opening of the permanent entrance in 1889 was one of the
pivotal moments in the recent ecological history of the lakes because it led
to an increase in the salinity of the Gippsland Lakes (Saunders et al.,
2008). Over the period of the opening of the artificial entrance
(corresponding to depths of ∼ 110 cm), we expected to see a change in
the diatom taxa to a greater abundance of thalassic species and a concomitant
reduction in freshwater species. Unexpectedly, a spike in freshwater species
was instead observed over the period 1870–1925, represented by the LK2
layer. This corresponded to other geochemical proxies which suggested an
increase in terrestrial organic matter, including a decrease in
δ13C, and an increase in the sediment C : N ratio, which was also
observed (although not discussed) by Saunders et al. (2008). This suggests
that the study site had an increased influence from riverine sediments over
this period. The most likely explanation for this is the remobilization of
terrestrial sediments through dredging activities within the delta of the
Mitchell River (Fig. 1), which could have led to an increased deposition of
terrestrial material at the study site. It appears that this increased input
of terrestrial material did not correspond to a changed sedimentation rate as
the 210Pb decay profiles displayed a similar trend between the LK3 and
LK2 layers (Fig. 2a). Irrespective of the exact cause of the LK2 sediment
layer, we are confident the LK3 and LK1 layers are representative of post and
pre artificial entrance opening periods respectively.
Cyanobacteria blooms and eutrophication
The biogeochemical proxies analysed here provide clear evidence for two
periods of eutrophication and cyanobacterial blooms in the Gippsland Lakes:
(1) the recent period after World War II (LK3) and (2) prior to the opening
of the artificial entrance in 1889 (LK1). The latter part of the most recent
period has been well monitored and provides an excellent means to validate
the biogeochemical proxies. The first piece of evidence for the recent
period of eutrophication comes from the steady increase in sediment organic
carbon and pheophytin a content after the 1940s (Fig. 3), consistent with a
previous palaeolimnological study (Saunders et al., 2008). The δ13Corg of this organic matter is typically
∼ -23 ‰, consistent with organic matter inputs derived from
phytoplankton (Fig. 3). This period also coincided with a marked jump in the
sum of the cyanobacteria pigments zeaxanthin, echinenone and canthaxanthin
from ∼ 500 nmol g org C-1 up to > 2000 nmol g org C-1 at the top of the core
(Fig. 3), consistent with an N. spumigena bloom at the
site in November 2011–February 2012 (Woodland et al., 2013). The
first documented bloom of N. spumigena in the lakes occurred in 1965, and the period
from 1987 through the 1990s is known to have had severe and regular blooms
(Cook and Holland, 2012). Over this period there were also two dips in the
δ15N of ∼ 2 ‰ in the 1940s
and late 1980s–2000, consistent with the occurrence of nitrogen-fixing
cyanobacteria. The broad agreement between these cyanobacteria markers and
recent recorded blooms gives us confidence that they are appropriate markers
of cyanobacteria blooms within the Gippsland Lakes, and this is consistent
with previous studies in the Baltic Sea (Bianchi et al., 2000; Funkey et
al., 2014).
The biogeochemical proxies for the period prior to the opening of the
artificial entrance in 1889 in layer LK1 likewise suggest a period of
eutrophication and intense cyanobacteria blooms. The sediment organic
content was high (∼ 5 %), the δ15N was low
(∼ 4–5 ‰), the δ13Corg
was in the range typical of phytoplankton (-22 to -23 ‰), and the cyanobacteria pigments and pheophytin a were
sporadically high (Fig. 3). Reports from newspaper articles in the late 1870s
also suggest the presence of surface scum of cyanobacteria with reference
to “noxious and ill smelling weed” on the surface of Lake King and there
were anecdotal reports of the greatly improved water quality with the
increased salinity after the opening of the artificial entrance in 1889
(Synan, 1989). We now discuss three key factors controlling the incidence of
cyanobacteria blooms prior to the opening of the artificial entrance and
European settlement.
Salinity. With no permanent entrance, the inflow of seawater was greatly
reduced, and at this time the lakes were considerably fresher (Harris et
al., 1988; Saunders et al., 2008). The diatom chronology also supports this
reduced marine influence with an increased abundance of Cyclotella choctawatcheeana, a planktonic diatom
characteristic of deep mesosaline (salinities > 10) lakes and
brackish marine systems (Fritz et al., 1993), and the reduced
dominance of thalassic diatoms (Fig. 4). N. spumigena typically blooms at salinities
between 9 and 20 in the Gippsland Lakes, and these salinities are currently
only reached during late spring–summer in high-flow years (Cook and
Holland, 2012). Prior to the opening of the artificial entrance, it is likely
that this salinity range was more typical, hence increasing the frequency of
blooms. In high-flow years when salinities were even lower, it is likely
that other nitrogen-fixing cyanobacteria, such as Anabaena would instead dominate.
This species occasionally blooms in Lake Wellington, and it is therefore
possible that this cyanobacterium was present in Lake King prior to 1889.
Stratification and residence time. At present, the highest stratification
is observed during periods of low surface water salinity in the Gippsland
Lakes associated with above-average river flows (Cook and Holland,
2012). Reduced tidal flushing, combined with low surface salinity, would lead
to enhanced stratification and residence time of the water column, which are
both known to favour buoyant slow-growing cyanobacteria such as N. spumigena
(Sellner, 1997; Paerl, 2014). Increased stratification will also lead to
increased hypoxia, and the marked increase in sediment organic-carbon
content to ∼ 5 % prior to ∼ 1870 (below 130 cm, Fig. 4) is consistent with increased hypoxia in this period
(Zillen and Conley, 2010). A key effect of this would be to enhance
the release of phosphorus from the sediment, which is a key driver of N. spumigena blooms
in the Gippsland Lakes (Cook et al., 2010; Scicluna et
al., 2015).
Nutrients. Prior to European settlement, it has been estimated that
nitrogen and phosphorus loads were lower by a factor of 2 and 3 respectively
(Grayson et al., 2001). At face value, it is surprising that the Gippsland
Lakes experienced more blooms of cyanobacteria then; however, this can be
reconciled with contemporary studies. First, cyanobacteria such as N. spumigena can derive all of their nitrogen requirements from nitrogen
fixation, and these blooms can add significant loads to the Gippsland Lakes
(Woodland and Cook, 2014). The generally lower sediment δ15N values
prior to 1870 (below 130 cm) are consistent with this. Second,
cyanobacterial blooms are driven by a focused release of phosphorus from the
sediments during bottom water hypoxia or anoxia, and it was estimated that a
large recent bloom of N. spumigena was caused by a release of
∼ 25 t of phosphorus from the sediment in Lake King (Scicluna et
al., 2015). Given that phosphorus is trapped and effectively recycled in
periodically anoxic and high residence time systems such as the Gippsland
Lakes, it is plausible that a pre-European phosphorus catchment load of 50 t
per year could maintain blooms of at least the same magnitude as currently
observed (Grayson et al., 2001).
Did nitrogen play a key role in the re-emergence of cyanobacteria blooms?
Following the opening of the artificial entrance there was a ∼ 60-year period (∼ 80–130 cm) with relatively low cyanobacteria
biomass and oxic bottom water, as indicated by the reduction in
cyanobacteria pigments and sediment organic-carbon content respectively. It
is likely that this relatively low-productivity period was caused by
relatively good ventilation of the bottom water combined with low catchment
nutrient inputs. After the 1940s, the modern eutrophication of the Gippsland
Lakes commenced, as indicated by a steady increase in sediment organic
carbon and cyanobacteria pigments. Changes to hydrological, morphological
and salinity regimes are unlikely to be a key driver because, apart from a
20 % reduction in river inputs from river diversions (Moroka,
2010), there have been no significant hydrological and morphological changes
to the Gippsland Lakes over this period. Given that fire can lead to
increased nutrient loads as previously discussed, it is highly likely that
the 1939 wildfires (Australian Broadcasting Corporation, 2016) led to
a significant pulse of nutrients into the Gippsland Lakes. The subsequent
increase in agriculture (Lake Glenmaggie, used for irrigation in the
Macalister Irrigation District, a tributary of the Latrobe River, was
completed in 1926 and expanded immediately post World War II), industry and
urbanization within the catchment have been estimated to have increased
nutrient loads by a factor of 1.8 and 3 for total nitrogen and phosphorus
respectively (Grayson et al., 2001).
Previous work has shown that the Gippsland Lakes are generally nitrogen
limited outside N. spumigena blooms (Holland et al., 2012), and
increased inputs of this element have most likely resulted in increased
productivity in the Gippsland Lakes, particularly during the winter and
spring diatom blooms when most of the nutrient load is delivered. The
increased settling of phytodetritus after the collapse of these blooms would
have driven increased water column anoxia over the late-spring period,
triggering the release of phosphorus stored in the sediment leading to more
favourable conditions for N. spumigena blooms (Cook et al., 2010;
Scicluna et al., 2015; Holland et al., 2012). We therefore speculate that the
recent re-emergence of cyanobacterial blooms is amplified by increased
nitrogen loads, which drive an increased internal release of phosphorus
through increased bottom water hypoxia and anoxia in late spring through to
summer. These observations are consistent with global studies of coastal
waters, which consistently show an increased incidence of hypoxia over the
past 50 years driven by eutrophication (Diaz and Rosenberg, 2008) and that
this may then lead on to blooms of cyanobacteria (Funkey et al., 2014). This
finding supports the argument that mitigating coastal eutrophication requires
controls on both nitrogen and phosphorus (Paerl, 2014; Conley et al., 2009),
even in systems that experience diazotrophic cyanobacterial blooms.
Conclusions
In conclusion, blooms of cyanobacteria were a natural feature of the
Gippsland Lakes prior to European settlement, most likely driven by strong
stratification and phosphorus release from the sediment. We suggest that
pre-European phosphorus loads were sufficient to maintain a sediment
phosphorus pool capable of driving significant periodic blooms based on
contemporary observations. The opening of the artificial entrance in 1889
likely led to increased salinity, flushing and reduced stratification,
leading to an increase in bottom water oxygenation, a decrease in sediment
phosphorus release and associated cyanobacterial blooms. The re-emergence of
cyanobacterial blooms post World War II may have occurred as a consequence of
increased nitrogen inputs, which led to increased anoxia occurring as a
consequence of increased primary production, triggering sediment phosphorus
release during the summer low-flow period when blooms occur. This finding
provides a mechanism by which decreasing nitrogen loads may also reduce
phosphorus-limited diazotrophic cyanobacterial blooms by reducing phosphorus
release from the sediment, highlighting the need to control both nitrogen and
phosphorus loads to estuaries even when they experience blooms of
diazotrophic cyanobacteria.