https://orcid.org/0000-0002-4945-085X
Correction
5 Oct 2023: Wilkinson A, Ariel E, van de Merwe J, Brodie J (2023) Correction: Trace element concentrations in forage seagrass species of Chelonia mydas along the Great Barrier Reef. PLOS ONE 18(10): e0292777. https://doi.org/10.1371/journal.pone.0292777 View correction
Figures
Abstract
Toxic metal exposure is a threat to green sea turtles (Chelonia mydas) inhabiting and foraging in coastal seagrass meadows and are of particular concern in local bays of the Great Barrier Reef (GBR), as numerous sources of metal contaminants are located within the region. Seagrass species tend to bioaccumulate metals at concentrations greater than that detected in the surrounding environment. Little is known regarding ecotoxicological impacts of environmental metal loads on seagrass or Chelonia mydas (C. mydas), and thus this study aimed to investigate and describe seagrass metal loads in three central GBR coastal sites and one offshore site located in the northern GBR. Primary seagrass forage of C. mydas was identified, and samples collected from foraging sites before and after the 2018/2019 wet season, and multivariate differences in metal profiles investigated between sites and sampling events. Most metals investigated were higher at one or more coastal sites, relative to data obtained from the offshore site, and cadmium (Cd), cobalt (Co), iron (Fe) and manganese (Mn) were found to be higher at all coastal sites. Principle Component Analysis (PCA) found that metal profiles in the coastal sites were similar, but all were distinctly different from that of the offshore data. Coastal foraging sites are influenced by land-based contaminants that can enter the coastal zone via river discharge during periods of heavy rainfall, and impact sites closest to sources. Bioavailability of metal elements are determined by complex interactions and processes that are largely unknown, but association between elevated metal loads and turtle disease warrants further investigation to better understand the impact of environmental contaminants on ecologically important seagrass and associated macrograzers.
Citation: Wilkinson A, Ariel E, van de Merwe J, Brodie J (2022) Trace element concentrations in forage seagrass species of Chelonia mydas along the Great Barrier Reef. PLoS ONE 17(6): e0269806. https://doi.org/10.1371/journal.pone.0269806
Editor: Mohan Lal Dotaniya, Directorate of Rapeseed-Mustard Research, INDIA
Received: July 25, 2021; Accepted: May 31, 2022; Published: June 15, 2022
Copyright: © 2022 Wilkinson et al. This is an open access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.
Data Availability: All relevant data are within the paper and its Supporting information files.
Funding: The author(s) received no specific funding for this work.
Competing interests: The authors have declared that no competing interests exist.
Introduction
Sea turtles are air-breathing reptiles. Like marine mammals, sea turtles have lungs and must return to the surface to breathe [1]. Due to this, the primary source of metal element exposure for green turtles (Chelonia mydas) is through ingestion of their diet [2]. Chelonia mydas (C. mydas) has a complex life history that includes a diverse diet, dependent on life phase [3]. Juveniles migrate to the coastal zone, inhabiting foraging grounds, where individuals undergo an ontogenetic dietary shift from a carnivorous to an herbivorous diet [3]. Once herbivorous, C. mydas forage on a range of material, dependent on region and what forage species are most predominant [4]. Whilst macroalgal species are common as primary forage material in C. mydas diets in some regions [4], at coastal sites along the Great Barrier Reef (GBR), C. mydas feed primarily on seagrass species [5].
In addition to providing forage for C. mydas, seagrass fulfil several integral ecological functions, such as sediment stabilisation [6–9], nutrient cycling [10, 11], the sequestration of carbon [12–14], as nursery grounds and foraging material for a wide range of marine organisms [6, 15]. Most seagrass species grow in the coastal zone, often near anthropogenic activity and potential marine contamination sources [16, 17]. Declines in seagrass distribution has been widespread in recent decades [18–20].
Several seagrass species are considered reliable bioindicators for the health of an ecosystem, as an early warning of any elevated contaminants, or decline in water quality parameters [21, 22]. Seagrass are efficient at accumulating metal concentrations, at magnitudes higher than that detected in the surrounding water [23]. Metals are highly persistent and remain in the environment indefinitely, binding to fine particulate sediment and settling on the substrate where concentrations are absorbed by seagrass [5]. Potential bioaccumulation of metals in seagrass may be a significant exposure pathway for turtles to elements at toxic concentrations [5]. C. mydas display strict site fidelity within a small local foraging range (2–3 km2), and likely inhabit the same seagrass meadow for long periods, regardless of environmental condition and water quality, even if environmental health deteriorates [24].
Metal contamination of marine ecosystems has increased significantly in recent years, with new chemicals frequently being introduced through industrial and agricultural processes and sequestered metal loads remobilised and redistributed due to dredging events and environmental disturbances such as flooding and cyclone activity [5, 25]. Some metals occur naturally in the marine environment, many of which are deemed essential elements for numerous biochemical and physiological processes (such as iron, Fe; copper, Cu and magnesium, Mg) [26]. However, essential elements may have toxic effects if concentrations exceeding optimal thresholds are experienced (as well as extremely low levels), particularly over long periods of time. For example, Fe, Cu and zinc (Zn) may cause reduced immune function in marine organisms when elevated [26]. Conversely, numerous metal elements are non-essential for life and are often toxic to organisms at very low concentrations [27, 28]. Non-essential elements commonly detected in the marine environment include, cadmium (Cd) and lead (Pb), with elements such as cobalt (Co) being understudied, but potentially toxic to marine turtles [5]. Metals of most concern are those elements (essential and non-essential) that cause known toxic effects to immune function and biochemical processes, such as Cd and Pb [26, 29].
Metal contamination (essential and non-essential) of coastal zones occurs via several exposure pathways. Freshwater runoff of sediments, during periods of heavy rainfall and atmospheric deposition of metal particles, are the main transportation pathways of metals to the marine environment [30–32]. Furthermore, anthropogenic pressures and processes such as mining, metal refining, agricultural chemical application, dredging and drainage of industrial waste, are able to change the distribution and composition of any geoavailable and bioavailable metal concentrations within the coastal zone [33].
This study aimed to describe seagrass (determined by identification of local C. mydas primary foraged species) trace metal concentrations at several coastal sites along the Great Barrier Reef (GBR), which are important foraging grounds for C. mydas and other macro grazers (such as dugongs) [34]. Such investigation is significant as it provides data on the prevalence of ecologically relevant metals (deemed toxic or are commonly measured in studies with similar scope) in the region. Metals of focus in this study included non-essential elements, cadmium (Cd), cobalt (Co), nickel (Ni) and lead (Pb), and several essential elements (e.g. Fe, Cu and Zn), capable of affecting immune function of marine turtles at ecologically relevant concentrations [26].
Materials and methods
In this study primary foraged seagrass species of local C. mydas populations was first determined by conducting gastric lavage sampling on a subsection of each population (up to 20 individuals from each study site). Rather than directly analysing the gastric lavage samples for metals, a total of 82 seagrass samples were collected by hand, either during dedicated field work, or during other sampling efforts (turtle capture and sampling). The reasoning for collecting fresh seagrass samples as opposed to analysing the gastric samples, was to ensure minimum contamination, and to allow for a wider area of each foraging ground to be sampled. Additionally, sampling known forage species meant that analysis could be completed prior to, and following, the wet season, without having to recapture individual green turtles. Once identified, seagrass samples were analysed for a suite of 10 metal elements using inductively coupled Plasma Optical Emission Spectrometry (ICP-OES) to describe concentrations detected within coastal meadows along the Central GBR. Multivariate analysis informed investigation into trace metal profiles in species of local bays before and after the 2018/19 wet season.
Study sites
Three geographically distinct study sites were sampled along the east coast of Australia, adjacent to the GBR (Fig 1). Firstly, Cockle Bay (CB) (19 º 10 ’ 26.7 "S 146 º 49 ’ 32.1"E) is a westerly facing bay of Magnetic Island, 8 km off the coast of Townsville, Queensland and forms a part of Cleveland Bay. Industry practices such as metal processing (including Zn, Cu and Ni), urban runoff from the city of Townsville (population about 200,000), and major sea port practices (including regular channel dredging) take place within the area [35]. Secondly, Upstart Bay (UB) (19°44 ’ 44.4"S 147°36 ’ 03.8 "E) is a north facing bay, receiving river discharges from the major catchment of the Burdekin River (which also influences CB to a lesser extent), dominated by agricultural and grazing practices, and with a prominent mining background, located 150 km south of Townsville. The Burdekin catchment is one of the two largest GBR catchments (the other being the Fitzroy) with an area of 140,000 km2. Finally, Edgecumbe Bay (EB) (20° 6′ 49″ S, 148° 23′ 25″ E) is located south of Bowen, Queensland, approximately 200 km south of Townsville. Within the catchment draining into EB there are a number of point and non-point sources of potential contaminants, including a wastewater treatment plant (for the town of Bowen–population of approximately 10,000), cokeworks and sugarcane farms (mostly on the catchment of the Gregory River in the south of the bay), and rarely, from discharge plumes from the Burdekin River [36]. Seagrass metal data [5] from a fourth site, the Howick Island Group (HWK), a mid-shelf group of remote reefs found in the northern region of the GBR (14 º 30 ’ 11"S 144 º 58 ’ 26 "E), was also included. HWK is likely to be minimally influenced by anthropogenic activity, with limited exposure to land-sourced contamination due to geographical proximity. The study location is located over 130 km from the nearest human settlement (Cooktown) and at least 20 km offshore from the coastal zone of the mainland.
Map of three study areas (Cockle Bay (CB), Upstart Bay (UB) and Edgecumbe Bay (EB)) where preferred green turtle seagrass forage was sampled, and the offshore site from which data was also analysed (Howick Island Group, HWK). Each site is colour coordinated between the national overview and localised maps. Areas highlighted in colour define local green turtle foraging ranges. The areas highlighted with white and black depict seagrass sampling locations at each study site. The World imagery applied as the base map in this figure is attributed to Esri and modifications were conducted using ArcMap (version 10.7.1, Esri. ArcGIS and ArcMap, California, USA). and is permitted for use as per Esri terms of service and the DMCA (Digital Millennium Copyright Act).
Turtle capture and sampling
In total, 46 turtles were captured using turtle rodeo techniques described in [37]. Individually numbered titanium flipper tags (Department of Environmental Sciences, Queensland Government) were applied, as described by Eckert et al. [38]. Curved carapace length (CCL), from the notch of the supracaudal scute to the line where skin joins the anterior edge of the carapace, along the midline ridge of the carapace, was measured using a flexible tape measure (cm), to the accuracy of ± 0.1 cm. Large barnacles were removed from the carapace with long nose plyers if their position obstructed accurate CCL measurements.
Research was carried out under all necessary permits from James Cook University Animals Ethics Committee (A2396), Department of Environment and Science (WISP18586417 and WISP18596817) and Great Barrier Reef Marine Parks Authority (G17/39429.1).
Gastric lavage
A total of 46 C. mydas stomach contents were collected, across all sites (CB = 14, UB = 20 and EB = 12), by modifying the protocol outlined in [39]. Briefly, a custom-made water pumping mechanism was designed, assembled, and tested by experienced personnel prior to use in the field. A foot pump (Whale babyfoot pump, Whale Marine, Northern Ireland) was connected to a water intake hose and an outtake hose, made from polyurethane tubing (8 mm diameter, with the ends melted and rounded to prevent injury to the turtle’s digestive tract). Firstly, captured turtles were elevated, with their head at the lowest point and secured in a fixed position by trained personnel. Turtles were encouraged to open their mouth by applying gentle pressure between the jaws, and once open, a wooden bit was inserted across the mouth to prevent closing. A 15–20 cm length of polyvinyl chloride tubing (10 mm diameter) was inserted into the mouth to offer stability for the insertion of the water tube (150 cm long and 8 mm diameter, marked at every 10 cm for the first 100 cm), which was lubricated with olive oil and slowly inserted down the digestive tract. The tube was slowly rotated during insertion to aid in the breaking up of any food bolus (obstruction of pre-digested material) in the throat. Once the marking, indicating insertion to 50–70 cm was achieved (determined on a case-by-case basis, dependant on turtle size), a constant flow of untreated sea water (1 L per minute) was initiated by regular use of the foot pump. Sea water was brought into the system by the intake tube, from a clean bucket filled with local sea water. As the water drained from the turtle all forage material was collected using a sieve (310 μm mesh size) positioned below the mouth. The procedure was conducted for no longer than five minutes and the turtle was maintained in the head down position until all water was drained. All equipment was sterilised using hospital grade detergent (benzalkonium chloride), and thoroughly rinsed between turtles. Samples were then placed on ice until return to the lab, where they were stored at -20 °C until identification and analysis.
Forage sample species identification
Forage material was thawed and separated out into individual species. To identify the seagrass species that were being consumed by C. mydas at the study sites, forage species were visually identified by experienced personnel using taxonomical identification guides, where necessary. Forty out of the total 46 (87%) gastric lavage samples collected contained the seagrass species, Halodule uninervis (Halodule), and the remaining six samples (from CB only, equivalent to 42.3% of CB samples) predominantly contained Cymodocea serrulata (Cymodocea). Either one of these seagrass species commonly made up 100% of the biomass of an individual lavage sample. Occasionally, in samples collected from CB included small proportions (< 10% of overall biomass) of red algae, however taxonomic identification to the species level was not conducted here as low abundance did not warrant further study. The identification of forage species informed the collection of seagrass species for the metal concentration analysis.
Seagrass collection
Seagrass sampling at all coastal sites was conducted before (July–October 2018) and after (February–June 2019) the 2018/19 wet season. The total number of samples collected per study site were as follows: n = 16 pre-wet and n = 12 post-wet season in CB, n = 15 pre-wet and n = 15 post-wet season in UB and n = 11 pre-wet and n = 13 post-wet season in EB. HWK Seagrass sampling was undertaken in July and August of 2015 and 2016. As Cymodocea was only present in lavage samples from CB green turtles, this seagrass species was only collected from CB. Samples were collected from the intertidal and subtidal zones within known turtle foraging grounds (identified through turtle sampling events). Personnel wore nitrile gloves when handling samples to minimise cross-contamination. Samples were collected at least 100 meters (max 300 m) apart, parallel to the shoreline, either on foot during low tide, or using snorkel techniques (max depth of approximately 2.5 m). Approximately 60 grams of above- and below- ground material (leaves and rhizomes) were collected and placed in food grade zip lock bags and stored on ice until return to the laboratory whereby samples were stored at -20 °C, until processing and analysis. Eighteen samples from HWK made up the offshore data provided by Thomas et al [5] and were collected in a similar way, by hand at low tide.
Seagrass sample preparation
Prior to analysis, all seagrass samples were thawed, and above and below ground material was separated, and leaves were removed. Small epiphytes growing on the leaf surface were included in the sample, as turtles foraging on seagrass would ingest such epibionts along with the intended forage species. All large debris (shells, shale, sand, etc.) was rinsed off each sample prior to drying, using fresh water. Wet weights were recorded for each sample prior to being oven dried for 48 hr at 60 °C. Dry weights of each sample were then recorded before homogenising into a fine powder using a pestle and mortar (sterilised with ethanol in between samples). A minimum of 200 mg of homogenised material (per sample) was submitted to the Advanced Analytical Centre (AAC, James Cook University) for acid digestion and ICP-OES analysis.
ICP-OES analysis
A suite of 10 metals (aluminium, Al, Cd, Co, Cu, Fe, Mg, Mn, Ni, Pb and Zn) were analysed in each sample. Seagrass samples were digested using a microwave assisted digestion oven (Bergof SW-4). A total of 100 mg of each sample was placed into the digestion vessel. Next, 4 mL SupraPure (Merk Germany) double distilled HNO3 and 1 mL AR Grade H2O2 were added into the vessel. The sample solution was kept in the fume hood for 2 h until the reaction was complete. Vessels were loaded into the microwave oven and heated to 185°C for 10 minutes. Once cooled, 150 mL of the digested samples were transferred to volumetric flask and diluted 50-fold, with Milli-Q water. Inductively Coupled Plasma Optical Emission Spectrometry (ICP-OES) was conducted using the Agilent 5100-ICP (Agilent Technologies, USA). External calibration strategy was used by applying a series of multi-element standard solutions containing all the elements of interest. HNO3 and H2O2 (reagents used in sample digestion, minus the sample) were included as procedure blanks for all elements and used to calculate the limit of detection values (LOD), which was defined as three times the standard deviation of each element’s blanks. Three samples were randomly selected and duplicated to check for consistency. To assure instrument calibration quality, independent standards (1 ppm) were included, with reported recoveries ranging from 87% (Al) and 103% (Mg). Two Certified Reference Materials (CRMs; GBW07605 Tea Leaves and NIST 1566 Oyster Tissues) were analysed to validate the analytical method and % recoveries ranged from 92% (Cu) to 118% (Cd). All metal concentrations are reported here as mg/kg of sample dry weight (dw).
This study and the reference study analysed seagrass samples using similar but distinct analytical techniques. Thomas et al [5] applied inductively coupled plasma mas-spectrometry (ICP-MS) rather than ICP-OES though both are suitable options for the detection of environmental metal loads in organic material such as seagrass. In ICP-OES, digested samples are nebulised and converted to plasma whereby electrons become excited. During the de-excitation phase, light is emitted from atoms and ions at different wavelengths dependant on metal. Wavelengths are then separated, detected, and quantified. In ICP-MS, ions present in the plasma are divided using mass spectrometry which separates and categorises each element by mass/charge ratio. ICP-MS can detect ultra-trace concentrations but both techniques are ideal for measuring the same suit of metals at ppm (mg/kg). Furthermore ICP-MS has the ability to differentiate between isotopes of the same element, but this was not applicable in the current study as focus was placed on total metal concentrations only.
Data analysis
Metal concentrations for all samples collected in this study (pre-wet and post-wet season), and reference value data obtained from Thomas et al [5], were reported as mg/kg, and any concentrations found to be below limits of detection (LOD) were considered half of the LOD [40], this same approach was applied in the reference study for HWK data. Pb was removed from further analysis due to having >40% samples <LOD (see S1 Fig). In addition, Mg and Zn concentrations were removed from the data set which included both current data and data collected from HWK, as these elements were not reported by Thomas et al. [5].
Spatio-temporal variation between study locations and sampling events (pre-wet- and post-wet season) was conducted for all sites sampled within this study. Additionally, Principal Component Analysis (PCA) was conducted to measure variation between metal profiles from each coastal site and HWK. The HWK data was collected prior to the wet season (in July/August of 2015–16), and thus for PCA analysis, only the pre-wet season data for the coastal sites was used. To determine the most important dimensions in the data, two dimension-reduction protocols, scree plots and quality of representation measurements (cos2), were employed. Statistical analysis and plotting of PCA was conducted in the R statistical program (R Core Team, 2019), using the R packages ‘Tidyverse’(data exploration [41]), ‘FactoShiny’(multivariate analysis and plotting) [41] and ‘FactoMineR’ (Factor analysis) [42].
Results
Spatial patterns in metal profiles of preferred green turtle forage
When describing the mean metal concentrations from each coastal site relative to that of HWK, concentrations were generally higher in the coastal sites, for both non-essential elements (Cd, Co, and Ni) and essential metals (Cu, Fe, and Mn), though differences were observed across all coastal sites for Cd, Co, Fe, and Mn only (Table 1).
Mean concentration (mg/kg dw) and standard deviation (SD) of metal elements at each coastal site (CB, UB and EB) pre-wet season, and data from a foraging ground at the offshore site HWK, provided by Thomas et al [5].
To investigate whether metal profiles differed spatially, between the three coastal foraging sites and the offshore natural baseline site (HWK), PCA was conducted on pre-wet season data for each site. The scree plot indicated that the first two data dimensions adequately represented most of the variation in the data. Therefore, the suite of eight metals (variables or dimensions) were reduced to two principal components (Dim 1 and 2), which together represented 67.02% of the total variation of the data (Fig 2).
Plot of the findings of all metal concentration data from seagrass collected from coastal sites of this study and baseline data obtained from HWK and provided by Thomas et al [5]. Dimension 1 (Dim 1) represents 49.34% of total variations and Dim 2 represents 17.68%.
Squared cosine (cos2) indicates the importance of a metal element to a particular principal component (or the quality of representation) and supported the reduction to two principal components. A cos2 >0.65 true for all metals in the analysis suite, excluding Al and Ni, indicated that most of the data variation was indeed accounted for by the first two dimensions. Rather than removing Al and Ni from the analysis because of low representation, the elements were maintained for consistency. Further confirmation was observed in the scree plot whereby the elbow was at dimension 3 with only variation accounted for by dimension 1 and 2 being present below the line. The Al cluster (top right) aligns closely to Dimension 2 and the others (Cd-Fe-Ni and Co-Cu-Mn clusters), top right and bottom right) align closer to Dimension 1, and thus influence the respective components, and therefore the entire data set, when analysing the results in a reduced space.
When individual seagrass samples were plotted (rather than variable loadings), all coastal site data were relatively evenly spread across the plot, and no separation between these locations was observed (Fig 3). EB variation was greatest of all sites as demonstrated by largest confidence ellipse and data point spread. All coastal seagrass data was distinctly separate from the HWK data (green), with individual sample loadings being clustered tightly and minimal overlap with coastal site data occurring. Confidence ellipses for each of the three coastal sites overlapped, indicative of similarities in each site metal profiles. CB (black) overlapped with both UB (blue) and EB (red), suggesting more similarity between CB and the other sites and little to none between UB and EB. HWK ellipse was distant from all coast sites while coastal data was more closely congregated, suggesting associations between offshore and coastal metal profiles were less likely than between coastal sites.
Depiction of all seagrass metal data reduced to two principal dimensions (Dim 1 and 2). Each point indicates an individual seagrass sample and data are categorised by location, represented as different colours (Howick Island Group, HWK = green, Cockle Bay, CB = black, Upstart Bay, UB = blue and Edgecumbe Bay, EB = red). Confidence ellipses are colour coordinated with the individual sample points. Dimension 1 (Dim 1) represents 49.34% of total variations and Dim 2 represents 17.68%. HWK data provided by Thomas et al [5].
Temporal analysis of seagrass metal profiles between sampling events
Differences in mean seagrass metal concentrations were observed at each coastal site between sampling events (pre-wet and post-wet season) and reported in Table 2. At CB, some element concentrations decreased over the wet season while others increased. Co and Cu were lower post-wet season at CB and EB but higher at UB. Mean Al concentrations increased at CB and EB but declined in UB when comparing between data from pre-wet and post-wet season. Similarly, Fe increased at CB and UB but decreased at EB from before to post-wet the wet season. Conversely, Mn decreased at CB and EB and increased between pre-wet and post-wet season data at UB. Cd and Ni concentrations remained similar at all sites when comparing pre-wet and post-wet season data.
Mean concentration and standard deviation (mg/kg (dw) ± SD) of metal elements in seagrass samples collected at each coastal site (CB, UB and EB) before and after the 2018/19 wet season.
A PCA was conducted to investigate any differences in total metal profile between seagrass samples collected from coastal sites pre-wet and post-wet season, and to identify the metals that most influenced the differences between sampling events (Fig 4). The scree plot accompanying the analysis determined that much of the data was adequately represented by the first two dimensions (principal components), which represented 57.93% of the variation. Though unlike Fig 2, the first two dimensions did not account for >65% Cos2 of the dataset. However, most elements (Cd, Co, Cu, Fe, Ni and Zn) were represented (>65% Cos2) by the first three dimensions. Mg was closely associated to Dim 2 (top left) while all remaining elements were clustered together and more closely associated with Dim 1.
PCA findings of all metal concentration data from seagrass collected from coastal sites of this study pre-wet and post-wet season. Dimension 1 (Dim 1) represents 36.96% of total variations and Dim 2 represents 20.97%.
Individual samples from both events were evenly distributed across the entire plot (Fig 5). The confidence ellipses demonstrate that data from both sampling events show some association with one another. While ellipses do not intercept or overlap, little space separates them and individual samples from both sets were located within both ellipses.
Depiction of all coastal metal data, collected in this study, reduced to two principal dimensions (Dim 1 and 2). Each point indicates an individual seagrass sample and are categorised by sampling event and categorised as different colours (pre-wet season = red and post-wet season = black. Confidence Ellipses are colour coordinated with the individual sample points. Dimension 1 (Dim 1) represents 36.96% of total variations and Dim 2 represents 20.97%.
Discussion
Total metal concentrations relative to calculated reference values
By analysing seagrass metals in coastal study sites alongside to an ecologically different natural baseline population that is minimally impacted by anthropogenic activity (in this instance HWK), some insight may be gleaned as to whether target elements were potentially detected at elevated levels within seagrass meadows in study sites close to human influence. Region-specific baseline data are crucial as reference for determining whether element concentrations are of concern to the local ecosystems or animals. Here Cd concentrations were greater at all three coastal sites relative to those reported in seagrass from HWK [5], possibly indicating exposure to higher concentrations at all coastal sites monitored. Additionally, coastal Cd levels exceeded, or were close to, offshore site seagrass data (0.36 mg/kg), reported by Conti et al. [43] in temperate seagrass species found in the Mediterranean Sea. One potential source of Cd in the region may be due to soil erosion of old Cd-rich coastal sugar cane paddocks, over time [44, 45]. Cd, in conjunction with many other metal elements (and other contaminants), can bind to suspended fine particulates in the Burdekin River, and be transported to the coastal zone during outflow. This exposure pathway is particularly significant during major freshwater runoff events, and through increased erosion seen in the Burdekin catchment [46], which encompasses UB. Cd is considered a xenobiotic element (not produced or used within the organism), and it is possible that exposure to very low concentrations may cause some toxic effects [26]. Due to the low excretion rate and its ability to bioaccumulate in tissues, Cd is considered a high risk metal in terms of toxicity to organisms [47], including C. mydas [48].
Like Cd, seagrass Co concentrations were elevated in coastal sites, relative to HWK (Table 1), and most of all in UB (1.87 ± 0.45 mg/kg), approximately four times higher than HWK (0.52 ± 0.21 mg/kg). Co is deemed beneficial to plants as micronutrients and was reported to actively accumulate in seagrass leaves and roots [49, 50]. Co concentrations of up to four times higher than HWK have been reported in C. mydas within the region and are of some concern as immunosuppression has previously been linked to Co exposure in marine turtles [5, 51, 52]. Over recent years, consistently high concentrations have been reported in both C. mydas blood [51], and preferred forage samples [5] within the study region. Villa et al. [51] reported Co blood concentrations in UB C. mydas exceeded site-specific reference intervals (RIs) by up to an order of magnitude. These RIs were calculated to determine if any elements were considered elevated relative to a baseline cohort of clinically healthy turtles considered to be minimally impacted by anthropogenic chemical influence. The elevated Co concentrations are not believed to be due to any recent contamination event, rather concentrations are consistent over time. Upstart Bay receives river discharge from the Burdekin River, where historical land use in this region includes Ni and Co mining (Greenvale). Erosion rates in this region have significantly increased (by up to eight times) since the 1850s and is associated with rangeland beef grazing [45, 53]. Given that Co can be transported in particulate form associated with soil particles and marine sediments, it is plausible that changes in the supply of Co to the coastal region may be associated with increased erosion and transport of fine fractions of sediment [46]. Fine sediments (muds and silts) contain higher concentrations of certain metals relative to coarser types (sands and gravel) [54]. The increase in fine sediment discharge within the Burdekin may be associated with increased metal loads (particularly if concentrations were higher than previous discharge events), which bind to fine sediments. The majority (up to 67%) of sediment discharged from the Burdekin River has been found to deposit in UB with long-distance sediment transport also likely to CB somewhat [55]. This pattern is reflected in the current study, whereby forage Co concentrations are greatest at UB and decline (but are still high) in CB.
Comparisons between sites can be informative, though it is difficult to directly compare between sites due to differences in local conditions and metal geoavailability. Geochemistry and bioavailability of metals differs between region, zone and proximity to contaminant sources and are mediated by complex physical, chemical and biological interactions, which are largely poorly-understood [5]. Sediment type, texture and mineralogy all play a role in determining availability [5], with particles of fine clay and silt, found in estuarine environments (CB and UB), often carrying significantly greater terrigenous metal loads than coarser carbonate based sediments found offshore (where marine metals tend to be higher than in coastal samples), in areas such as HWK. Region-specific variations in sediment profile make it difficult, and often unreliable, to compare metal concentrations directly between sites, as definitive explanation for differences is often unknown. While such comparisons give insight into relative concentrations, they provide little information as to whether such concentrations are at normal loads for the region or if that site is in fact contaminated [5].
While the multivariate approach and PCA conducted was able to reduce the dimensions of the data from nine variables (metal elements) down to two, the results did not indicate any element or interaction of elements that influenced the differences in metal profiles between coastal sites. All variables were loaded evenly and influenced the data set to similar extents. The PCA indicated that metal profiles were distinctly different between HWK and all coastal sites, with strong association between coastal profiles observed in the data. While coastal profiles (particularly CB) were like one another, differences were still found between locations, likely explained by region-specific geoavailability differences, whereby degree of anthropogenic influence and local environmental conditions and processes play important roles in the bioavailability and distribution of persistent chemical contaminants in local environments. Greatest mixing of individual sample data was between data from CB and UB, indicative of the most similarity. These sites are located close to one another and are influenced by the Burdekin River out flow, and thus share sediment sources [55, 56]. Furthermore, HWK data was most like EB (eclipse proximity). Out of all three coastal sites EB is likely the least influenced by anthropogenic impact as interpreted here (consecutively lower metal concentrations), when compared to CB and UB.
Impacts of elevated or toxic metals on seagrass survivability
Metals such as Cu, Cd and Zn are thought to be toxic to seagrass species. In this study Cu concentrations were detected between 2.5 ± 0.7, in EB post-wet season and 7.6 ± 2.3 mg/kg, in UB post-wet season. Zheng et al. [57] observed leaf necrosis in Thalassia hemprichii after a 5-day treatment to 1 mg/L Cu 2+, likely a symptom of malnutrition, as competition for micronutrient uptake binding sites could induce inhibition of transport and function of ions such as Ca 2+ (calcium), Mg 2+, K 2+ (potassium), which are micronutrients required for numerous metabolic and photosynthetic processes [57, 58]. Furthermore, photosynthetic efficiency may be hindered by phytotoxic effects of some metals, including changes in redox states in leaf cells due to inhibition of antioxidant enzymes, such as superoxide dismutase and peroxidase, which leads to increased production of radical oxygenating species (ROS) that damage photosynthetic apparatus and chlorophyll [57]. Necrosis and antioxidant inhibition may be accompanied by a significant decline in photosynthetic efficiency (effective quantum yield) [57], which is likely caused by disturbance to photosynthetic electron transport observed in a range of contaminants including photosystem-II herbicides [59], also commonly applied in agricultural practices throughout the study region. Reduced photosynthetic function often leads to inhibited growth, survival, and community fitness of exposed meadows [60], leading to declines in distribution and inevitably habitat loss for the plethora of species reliant on seagrass ecosystems for a range of ecological functions.
What elevated metal loads mean for C. mydas
Metal concentrations detected in C.mydas depend on numerous factors including, species, sex, age and location [61]. Element concentrations were detected at highest concentrations at coastal sites relative to HWK data, which implies the likelihood of increased exposure of local coastal foraging C. mydas to potentially toxic metals. Toxic metal exposure has been reported to impact different aspects of marine turtle physiology, immunology, and biochemistry [26, 51, 62–64]. One such impact is that of elements including Cd, Co, Cu, Fe and Ni and Pb which have previously been reported as potentially causing immunosuppression in individuals, leading to increased susceptibility to secondary infections which may also be associated as necessary factors in expression of the enzootic disease, Fibropapillomatosis (FP), in C.mydas [26]. Though due to fundamental ethical issues, toxicity threshold data has not been calculated for marine turtles for any metals and thus limited information is available on what concentrations of metals (and other contaminants) are of ecological significance to local C. mydas health. One recent advancement and one which requires further application, is that of cell-based toxicity assays which have been applied to measure ecotoxicological endpoints of suites of metal elements on cultured C. mydas cells [65–67]. Continued effort should be made to implement such in vitro approaches alongside conventional chemical analysis to better understand metal specific toxicities and to calculate site-specific toxic thresholds which may be implemented to better determine the exposure and susceptibility of local foraging populations.
Variation in metal concentrations following the wet season
In UB for most trace elements, increases in concentrations, relative to HWK data, were reported in samples collected pre-wet season, whereas in CB, the opposite was true. This is interesting as during January and February 2019 significant rainfall (a total of 1260 mm over ten days) caused an extreme flood event in the Townsville region, which exceeded historical records (926 mm over ten days in 1953) [68]. This event impacted numerous areas along the coast, adjacent to the study region (except HWK). Large flood plumes entered the coastal zone from the Ross River (Townsville) and the Burdekin River. Increased terrigenous sediment and metal transport into the coastal marine environment during major flood events has previously been reported [69], whereby fluvial transport processes cause sedimentary grain sizes to be sorted by size with increased distance from a given source [69]. The high affinity of trace metals for fine sediment particles means that the transport of metal loads often follows flood plume patterns, settling in low energy environments such as sheltered bays [70, 71], like CB and UB. However, suspended metal loads tend to be taken up by phytoplankton plankton blooms prior to settlement [72] and thus concentrations which are available for uptake by seagrass may be reduced. Furthermore, the 2019 major flood event may have brought increased metal loads to the study sites but were still sequestered in sediments and not yet accessible to seagrass.
Conclusion
Across the 2018/19 wet season, trace metal concentrations were measured in preferred (C. mydas) seagrass forage species. Several elements were found at greater concentrations in seagrass from the Northern and Central GBR region, when compared to site-specific reference data, in both coastal and offshore sites with various levels of anthropogenic influence. Elements of most concern, Cd and Co (thought to be toxic to seagrass and C. mydas alike), were found at greater concentrations in samples collected from all coastal sites when compared to the offshore site. Additionally, the Cd and Co offshore concentrations both exceeded published reference data, though definitive conclusions could not be drawn regarding the threat posed by these concentrations, to marine turtle health as site-specific reference data are lacking for these metals. Additionally, no distinct patterns were observed in metal concentrations detected in seagrass samples, from any sites, collected prior to the 2018/19 wet season when compared to data collected post-wet season. This was likely due to processes not within the scope of this descriptive study and thus not investigated. Metal concentration comparisons between study sites offer some information on local metal loads. For instance, PCA determined that coastal metal profiles were more like one another than to the offshore site but should be analysed tentatively as environmental factors such as sediment type geomorphology and geochemistry play a role in determining the geoavailability of certain elements. A better approach is to compare data to site-specific baseline values from that same location to provide insight into whether current levels are of concern. A significant increase in funding and investigation into the ecotoxicological study of environmentally relevant metals and the potential sources of such chemical contaminants is crucial before any insight can be gleaned regarding what the current metal loads likely mean for local seagrass meadows and the macrograzers which rely on them as forage.
Supporting information
S1 Fig. RAW data analysed and reported in research article “Trace element concentrations in forage seagrass species of Chelonia mydas along the Great Barrier Reef”.
https://doi.org/10.1371/journal.pone.0269806.s001
(DOCX)
References
- 1.
Wyneken J. The External Morphology, Musculoskeletal System, and Neuro-Anatomy of Sea Turtles. In: Lutz PL, Musick JA, Wyneken J, editors. The biology of sea turtles. 2: CRC press; 2002.
- 2. Villa CA, Finlayson S, Limpus C, Gaus C. A multi-element screening method to identify metal targets for blood biomonitoring in green sea turtles (Chelonia mydas). Sci Total Environ. 2015:613–21.
- 3.
Bolten AB. Variation in Sea Turtle Life History Patterns: Neritic vs. Oceanic Developmental Stages. In: Lutz PL, Musick JA, Wyneken J, editors. The biology of sea turtles. 2. Boca Raton, FL, USA: CRC press; 2002. p. 243–57.
- 4. Burkholder D, Heithaus M, Thomson J, Fourqurean J. Diversity in trophic interactions of green sea turtles Chelonia mydas on a relatively pristine coastal foraging ground. Mar Ecol Prog Ser. 2011;439:277–93.
- 5. Thomas C, Bennett W, Garcia C, Simmonds A, Honchin C, Turner R, et al. Coastal bays and coral cays: Multi-element study of Chelonia mydas forage in the Great Barrier Reef (2015–2017). Sci Total Environ. 2020:140042. pmid:32927538
- 6.
Kirkman H. Seagrasses of Australia. Canberra: Department of the Environment. 1997:36 p.
- 7. Short FT, Polidoro B, Livingstone SR, Carpenter KE, Bandeira S, Bujang JS, et al. Extinction risk assessment of the world’s seagrass species. Biol Conserv. 2011;144(7):1961–71.
- 8. Madi Moussa R, Bertucci F, Jorissen H, Gache C, Waqalevu VP, Parravicini V, et al. Importance of intertidal seagrass beds as nursery area for coral reef fish juveniles (Mayotte, Indian Ocean). Reg Stud Mar Sci. 2020;33:100965.
- 9. Huang Y, Xiao X, Xu C, Perianen YD, Hu J, Holmer M. Seagrass beds acting as a trap of microplastics—Emerging hotspot in the coastal region? Environ Pollut. 2020;257:113450. pmid:31679874
- 10.
Marbà N, Holmer M, Gacia E, Barron C. Seagrass beds and coastal biogeochemistry. In: Larkum A, Orth RJ, Duarte CM, editors. Seagrasses: biology, ecology and conservation. Dordrecht, Netherlands: Springer; 2007. p. 135–57.
- 11. Touchette BW. Seagrass-salinity interactions: Physiological mechanisms used by submersed marine angiosperms for a life at sea. J Exp Mar Bio Ecol. 2007;350(1):194–215.
- 12. Fourqurean JW, Duarte CM, Kennedy H, Marbà N, Holmer M, Mateo MA, et al. Seagrass ecosystems as a globally significant carbon stock. Nat Geosci. 2012;5(7):505–9.
- 13.
de los Santos CB, Scott A, Arias-Ortiz A, Jones B, Kennedy H, Mazarrasa I, et al. Seagrass ecosystem services: Assessment and scale of benefits. Out of the Blue: The Value of Seagrasses to the Environment and to People. 2020:19–21.
- 14.
Miyajima T, Hamaguchi M. Carbon Sequestration in Sediment as an Ecosystem Function of Seagrass Meadows. In: Kuwae T, Hori M, editors. Blue Carbon in Shallow Coastal Ecosystems: Carbon Dynamics, Policy, and Implementation. Singapore: Springer Singapore; 2019. p. 33–71.
- 15. Hernawan UE, Rahmawati S, Ambo-Rappe R, Sjafrie NDM, Hadiyanto H, Yusup DS, et al. The first nation-wide assessment identifies valuable blue‑carbon seagrass habitat in Indonesia is in moderate condition. Sci Total Environ. 2021;782:146818.
- 16.
Lewis S, Smith R, O’Brien D, Warne M, Negri A, Petus C, et al. Assessing the Risk of Additive Pesticide Exposure in Great Barrier Reef ecosystems. Assessment of the relative risk of water quality to ecosystems of the Great Barrier Reef Department of the Environment and Heritage Protection, Queensland Government, Brisbane. 2013.
- 17. Ambo-Rappe R. Developing a methodology of bioindication of human-induced effects using seagrass morphological variation in Spermonde Archipelago, South Sulawesi, Indonesia. Mar Pollut Bull. 2014;86(1):298–303. pmid:25080858
- 18. Unsworth RKF, Cullen LC. Recognising the necessity for Indo-Pacific seagrass conservation. Conserv Lett. 2010;3(2):63–73.
- 19. Waycott M, Duarte CM, Carruthers TJB, Orth RJ, Dennison WC, Olyarnik S, et al. Accelerating loss of seagrasses across the globe threatens coastal ecosystems. Proc Natl Acad Sci 2009;106(30):12377–81. pmid:19587236
- 20. Orth RJ, Carruthers TJB, Dennison WC, Duarte CM, Fourqurean JW, Heck KL, et al. A Global Crisis for Seagrass Ecosystems. BioScience. 2006;56(12):987–96.
- 21. Bonanno G, Orlando-Bonaca M. Trace elements in Mediterranean seagrasses and macroalgae. A review. Sci Total Environ. 2018;618:1152–9. pmid:29055578
- 22. Malea P, Mylona Z, Kevrekidis T. Trace elements in the seagrass Posidonia oceanica: Compartmentation and relationships with seawater and sediment concentrations. Sci Total Environ. 2019;686:63–74. pmid:31176824
- 23. Bonanno G, Di Martino V. Seagrass Cymodocea nodosa as a trace element biomonitor: Bioaccumulation patterns and biomonitoring uses. J Geochem Explor. 2016;169:43–9.
- 24. Hazel J, Hamann M, Lawler IR. Home range of immature green turtles tracked at an offshore tropical reef using automated passive acoustic technology. Mar Biol. 2013;160(3):617–27.
- 25. Storelli MM, Storelli A, D’Addabbo R, Marano C, Bruno R, Marcotrigiano GO. Trace elements in loggerhead turtles (Caretta caretta) from the eastern Mediterranean Sea: overview and evaluation. Environ Pollut. 2005;135(1):163–70. pmid:15701403
- 26. da Silva CC, Klein RD, Barcarolli IF, Bianchini A. Metal contamination as a possible etiology of fibropapillomatosis in juvenile female green sea turtles Chelonia mydas from the southern Atlantic Ocean. Aquat Toxicol. 2016;170:42–51. pmid:26615366
- 27. de Souza Machado AA, Spencer K, Kloas W, Toffolon M, Zarfl C. Metal fate and effects in estuaries: a review and conceptual model for better understanding of toxicity. Sci Total Environ. 2016;541:268–81. pmid:26410702
- 28.
Aggett P, Nordberg GF, Nordberg M. Essential Metals: Assessing Risks from Deficiency and Toxicity. In: Nordberg GF, Fowler BA, Nordberg M, editors. Handbook on the Toxicology of Metals (Fourth Edition). San Diego: Academic Press; 2015. p. 281–97.
- 29. Moszczyński P. Mercury compounds and the immune system: a review. Int J Occup Med Environ Health. 1996;10(3):247–58.
- 30. Pacyna JM, Pacyna EG. An assessment of global and regional emissions of trace metals to the atmosphere from anthropogenic sources worldwide. Environ Rev. 2001;9(4):269–98.
- 31. Pacyna JM, Scholtz MT, Li YF. Global budget of trace metal sources. Environmental Reviews. 1995;3(2):145–59.
- 32. Strzelec M, Proemse BC, Gault-Ringold M, Boyd PW, Perron MM, Schofield R, et al. Atmospheric trace metal deposition near the Great Barrier Reef, Australia. Atmosphere. 2020;11(4):390.
- 33. Johnston SG, Burton ED, Bush RT, Keene AF, Sullivan LA, Smith D, et al. Abundance and fractionation of Al, Fe and trace metals following tidal inundation of a tropical acid sulfate soil. Appl Geochem. 2010;25(3):323–35.
- 34.
Larkum A, Orth RJ, Duarte CM. Seagrasses: Biology, Ecology and Conservation. Dordrecht, Netherlands: Springer; 2006.
- 35. Esslemont G. Heavy metals in seawater, marine sediments and corals from the Townsville section, Great Barrier Reef Marine Park, Queensland. Marine Chemistry. 2000;71(3–4):215–31.
- 36. Clark TR, Leonard ND, Zhao J-x, Brodie J, McCook LJ, Wachenfeld DR, et al. Historical photographs revisited: A case study for dating and characterizing recent loss of coral cover on the inshore Great Barrier Reef. Scientific Reports. 2016;6(1):19285. pmid:26813703
- 37. Limpus C, Reed P. The green turtle, Chelonia mydas, in Queensland: a preliminary description of the population structure in a coral reef feeding groud. Biology of Australasian Frogs and Reptiles. 1985:47–52.
- 38.
Eckert K, Bjorndal KA, Abreu-Grobois FA, Donnelly M. Research and management techniques for the conservation of sea turtles. Washington, DC: IUCN/SSC Marine Turtle Specialist Group; 1999. Report No.: 1071–8443 Contract No.: 3.
- 39. Forbes G, Limpus C. A non-lethal method for retrieving stomach contents from sea turtles. Wildl Res. 1993;20(3):339–43.
- 40. Wendelberger J, Campbell K. Non-detect data in environmental investigations. J Am Stat Assoc. 1994.
- 41.
Husson F, Lê S, Pagès J. Exploratory multivariate analysis by example using R: CRC press; 2017.
- 42.
Husson F, Josse J, Le S, Mazet J. FactoMineR: multivariate exploratory data analysis and data mining with R. R package version. 2013;1(1.29).
- 43. Conti ME, Mecozzi M, Finoia MG. Determination of trace metal baseline values in Posidonia oceanica, Cystoseira sp., and other marine environmental biomonitors: a quality control method for a study in South Tyrrhenian coastal areas. Environ Sci Pollut Res. 2015;22(5):3640–51.
- 44. Bartley R, Croke J, Bainbridge ZT, Austin JM, Kuhnert PM. Combining contemporary and long-term erosion rates to target erosion hot-spots in the Great Barrier Reef, Australia. Anthropocene. 2015;10:1–12.
- 45. Wilkinson SN, Hancock GJ, Bartley R, Hawdon AA, Keen RJ. Using sediment tracing to assess processes and spatial patterns of erosion in grazed rangelands, Burdekin River basin, Australia. Agric Ecosyst Environ. 2013;180:90–102.
- 46. Kroon FJ, Kuhnert PM, Henderson BL, Wilkinson SN, Kinsey-Henderson A, Abbott B, et al. River loads of suspended solids, nitrogen, phosphorus and herbicides delivered to the Great Barrier Reef lagoon. Mar Pollut Bull. 2012;65(4):167–81.
- 47. Hueza IM, Palermo-Neto J. Toxicologia do chumbo, mercúrio, arsênio e de outros metais. Toxicologia aplicada à medicina veterinária. 2008:641–62.
- 48. Fraga NS, Martins AS, Faust DR, Sakai H, Bianchini A, da Silva CC, et al. Cadmium in tissues of green turtles (Chelonia mydas): A global perspective for marine biota. Sci Total Environ. 2018;637–638:389–97. pmid:29753227
- 49. Nicolaidou A, Nott JA. Metals in sediment, seagrass and gastropods near a nickel smelter in Greece: Possible interactions. Mar Pollut Bull. 1998;36(5):360–5.
- 50. Schroeder PB, Thorhaug A. Trace metal cycling in tropical-subtropical estuaries dominated by the seagrass, Thalassia testudinum. Am J Bot. 1980;67(7):1075–88.
- 51. Villa CA, Flint M, Bell I, Hof C, Limpus CJ, Gaus C. Trace element reference intervals in the blood of healthy green sea turtles to evaluate exposure of coastal populations. Environ Pollut. 2017;220, Part B:1465–76. pmid:27825845
- 52.
Gaus C, Grant S, Jin NL, Goot K, Chen L, Villa A, et al. Investigations of contaminant levels in green turtles from Gladstone. National Research Centre for Environmental Toxicology. 2012.
- 53. Lewis SE, Shields GA, Kamber BS, Lough JM. A multi-trace element coral record of land-use changes in the Burdekin River catchment, NE Australia. Palaeogeogr Palaeoclimatol Palaeoecol. 2007;246(2):471–87.
- 54. Buyang S, Yi Q, Cui H, Wan K, Zhang S. Distribution and adsorption of metals on different particle size fractions of sediments in a hydrodynamically disturbed canal. Sci Total Environ. 2019;670:654–61. pmid:30909043
- 55. Lewis SE, Olley J, Furuichi T, Sharma A, Burton J. Complex sediment deposition history on a wide continental shelf: Implications for the calculation of accumulation rates on the Great Barrier Reef. Earth Planet Sci Lett. 2014;393:146–58.
- 56. de Caritat P, Grunsky EC. Defining element associations and inferring geological processes from total element concentrations in Australian catchment outlet sediments: Multivariate analysis of continental-scale geochemical data. Appl Geochem. 2013;33:104–26.
- 57. Zheng J, Gu X-Q, Zhang T-J, Liu H-H, Ou Q-J, Peng C-L. Phytotoxic effects of Cu, Cd and Zn on the seagrass Thalassia hemprichii and metal accumulation in plants growing in Xincun Bay, Hainan, China. Ecotoxicology. 2018;27(5):517–26. pmid:29556939
- 58. Wang C, Zhang SH, Wang PF, Hou J, Zhang WJ, Li W, et al. The effect of excess Zn on mineral nutrition and antioxidative response in rapeseed seedlings. Chemosphere. 2009;75(11):1468–76. pmid:19328518
- 59. Wilkinson AD, Collier CJ, Flores F, Negri AP. Acute and additive toxicity of ten photosystem-II herbicides to seagrass. Sci Rep. 2015;5:17443. pmid:26616444
- 60. Negri AP, Flores F, Mercurio P, Mueller JF, Collier CJ. Lethal and sub-lethal chronic effects of the herbicide diuron on seagrass. Aquat Toxicol. 2015;165:73–83. pmid:26026671
- 61. Shaw KR, Lynch JM, Balazs GH, Jones TT, Pawloski J, Rice MR, et al. Trace Element Concentrations in Blood and Scute Tissues from Wild and Captive Hawaiian Green Sea Turtles (Chelonia mydas). Environmental Toxicology and Chemistry. 2021;40(1):208–18. pmid:33103806
- 62. Camacho M, Orós J, Boada LD, Zaccaroni A, Silvi M, Formigaro C, et al. Potential adverse effects of inorganic pollutants on clinical parameters of loggerhead sea turtles (Caretta caretta): Results from a nesting colony from Cape Verde, West Africa. Mar Environ Res. 2013;92:15–22. pmid:23998796
- 63. Flint M, Eden PA, Limpus CJ, Owen H, Gaus C, Mills PC. Clinical and Pathological Findings in Green Turtles (Chelonia mydas) from Gladstone, Queensland: Investigations of a Stranding Epidemic. EcoHealth. 2015;12(2):298–309. pmid:25256011
- 64. Faust DR, Hooper MJ, Cobb GP, Barnes M, Shaver D, Ertolacci S, et al. Inorganic elements in green sea turtles (Chelonia mydas): Relationships among external and internal tissues. Environ Toxicol Chem. 2014;33(9):2020–7. pmid:24889685
- 65. Finlayson KA, Madden Hof CA, van de Merwe JP. Development and application of species-specific cell-based bioassays to assess toxicity in green sea turtles. Sci Total Environ. 2020;747:142095. pmid:33076209
- 66. Finlayson KA, Leusch FDL, van de Merwe JP. Cytotoxicity of organic and inorganic compounds to primary cell cultures established from internal tissues of Chelonia mydas. Sci Total Environ. 2019;664:958–67. pmid:30769319
- 67. Finlayson KA, Leusch FDL, Villa CA, Limpus CJ, van de Merwe JP. Combining analytical and in vitro techniques for comprehensive assessments of chemical exposure and effect in green sea turtles (Chelonia mydas). Chemosphere. 2021;274:129752. pmid:33529958
- 68.
Meteorology Bo. Special Climate Statement 69—an extended period of heavy rainfall and flooding in tropical Queensland. Bureau of Meteorology; 2019.
- 69. Coynel A, Schäfer J, Blanc G, Bossy C. Scenario of particulate trace metal and metalloid transport during a major flood event inferred from transient geochemical signals. Appl Geochem. 2007;22(4):821–36.
- 70. Ridgway J, Shimmield G. Estuaries as Repositories of Historical Contamination and their Impact on Shelf Seas. Estuar Coast Shelf Sci. 2002;55(6):903–28.
- 71. Liu WX, Li XD, Shen ZG, Wang DC, Wai OWH, Li YS. Multivariate statistical study of heavy metal enrichment in sediments of the Pearl River Estuary. Environ Pollut. 2003;121(3):377–88. pmid:12685766
- 72. Furnas M, Mitchell A, Skuza M, Brodie J. In the other 90%: phytoplankton responses to enhanced nutrient availability in the Great Barrier Reef Lagoon. Mar Pollut Bull. 2005;51(1):253–65. pmid:15757726