The field investigation was conducted during spring tide in May 2021, and the field monitoring lasted five days from the 10^th to the 14^th of May 2021. It is widely accepted that the groundwater system, in this case, could be considered two-dimensional (2-D) under the Dupuit assumption for simplifying engineering-scale groundwater modeling (Liu et al., 2023). Additionally, it is assumed that there are minimal variations of the phreatic line (i.e., groundwater table) alongshore (Gibbes et al., 2007). Thus, a single cross-shore transect was selected from the landward of the high tide (HT) mark into the subtidal region in order to track groundwater processes in the intertidal area. Figure 2 shows a cross-shore transect that was 40 m in length and started from the HT mark. A field monitoring system was set up across the shoreline to take measurements, and all of the equipment was collectively placed in the intertidal area where there was a highly reactive area for porewater exchange. The exact position of the benchmark (BM) was identified using a GPS (Garmin Australasia Pty Ltd.), and the elevation was converted to the Australian Height Datum (AHD), which is the official vertical datum used in Australia. By following that, at the locations of equipment, their corresponding elevations in reference to the local BM were determined horizontally along this transect, using a self-leveling rotating laser (RL-H5A with a precision of ±1.5 mm, Topcon Positioning Systems Inc.).

Figure 2 illustrates the configuration of the monitoring system, which includes:

The porewater samples were collected using sampling ports at the MLS (see Figure 3A ). As described in Figure 3B , the collected sample obtained from the spherical capture zone with a small radius (≈0.05 m) was representative of the porewater around the sampling port. The volume of each sample was limited to 120 mL to minimize the impact on the surrounding waters so that there was no cross-port interference and in-situ perturbation during porewater sampling. Figure 3C shows that porewater from each sampling port was extracted using a peristaltic pump (Geopump Series II, Geotech Environmental Equipment, Inc.) at a low flow rate of 60 mL/min and was rapidly transferred into a flow cell for a later determination of in-situ physio-chemical state variables (e.g., salinity, DO, pH, and redox potential) using a portable probe (HI98194, Hanna Instruments, Inc.). The overflow collected by a syringe of 60 mL was thenceforth transferred into the high-density polyethylene (HDPE) plastic bottles through a sterile filter (filtering size = 0.45 μm). In order to stabilize the dissolved Fe in different chemical forms, the bottles were priorly acidified with hydrochloric acid (HCL), i.e., filling the container entirely. This step can effectively minimize any potential oxidation in storage during transport. Throughout the field investigation, porewater at each port was sampled twice per day, so it took 4–6 hours to reconstruct a complete profile by spatially interpolating the datasets collected from all ports. Although the spatial porewater measurements were not synchronized in terms of both equipment and operating procedure, all datasets were recorded within a single day, and this issue could be later resolved by temporal interpolation in the data post-analysis.

It was once suggested by Thamdrup et al. (1994) that porewater analysis is dependent on the collection of intact sediment cores where pore water samples are extracted, mainly when analysis of the solid phase is also undertaken (Gibbes et al., 2008). However, it was highly challenging to preserve the integrity of the sediment cores from the same location at different times. To provide an alternative resolution in this regard, additional sediments from ambient areas (within a diameter of 0.5 m) at each sampling location were manually retrieved underground using a hand auger in depths of 0.5 m and 1.0 m, respectively. Those sediments were directly transferred into glass containers to determine the content of Fe oxides. Moreover, when the intertidal area was exposed at low tides, the sediment core was collected by (1) pushing a plastic pipe (inner diameter 76 mm) vertically down into the sediment at the selected location; (2) carefully digging around the pipe with a spade or shovel to avoid excessive disturbance; (3) capping the pipe at the top to seal it completely at the maximum depth achievable; (4) slowly extracting the pipe from the sandy beach, and completely sealing the pipe at the bottom; (5) keeping it as upright as possible to minimize disturbance. As a result, three undisturbed sediment cores (two were Φ76 mm × L300 mm in burial depth, while one was Φ76 mm × L500 mm) were taken for additional analysis of the hydraulic and seepage properties.

The chemical analysis of those porewater samples was carried out in an environmental laboratory (ALS Ltd.). The concentration of total Fe in the sample was quantified using an inductively coupled plasma mass spectrometry (ICP-MS) (Agilent 7900, Agilent Technologies Pty Ltd.), while the specific concentration of Fe(II) was determined by a discrete analyzer which utilized ion selective electrodes (ISE) (Konelab Prime 60i, Thermo Fisher Scientific Inc.) The measured potential between each ISE and the reference electrode correlated with the natural logarithm of the ionic activity (as per the Nernst equation), which reflected changes across the ISE membrane/sample interface. Consequently, the concentration of Fe(III) was derived through a simple calculation employing the total Fe and dissolved Fe(II) values. Furthermore, the metals in sediments were extracted by an aqua regia digest (2 mL of 1:1 HNO₃ and 10 mL of 1:4 of HCl). The content of total Fe and total P was determined by an inductively coupled plasma -optic emission spectroscopy (ICP–OES) (Agilent 5110, Agilent Technologies Pty Ltd.) and a discrete analyzer (Konelab Prime 60i, Thermo Fisher Scientific Inc.), respectively.

Hydraulic conductivity (i.e., coefficient of permeability K in geomechanics) tests were conducted on sediment cores at the Geomechanics Research Laboratory, School of Civil Engineering (SoCE) at the University of Queensland (UQ). The constant head method was utilized to measure the hydraulic conductivity by employing a Mariotte bottle to supply a constant hydraulic head at the boundary of the rigid-wall permeameter (ASTM D2434-68, 2006). In addition, the porosity of the in-situ sediment specimen was determined based on the theory of soil representative elementary volume (REV) in three-phase fractions (e.g., air, water, and solid); in order to achieve this, the specific gravity and dry density of in-situ specimen were measured separately using the gas pycnometer (ASTM D5550-23, 2023) and oven-drying methods (ASTM D7263-21, 2021). The particle size distribution (PSD) tests were performed on the sediment samples using the sieving method (ASTM D6913-17, 2021). The data obtained from porosity and effective particle size (d ₁₀, d ₃₀, and d ₆₀), along with the Kozeny-Carman equation (see Equation 1), were used to estimate the hydraulic conductivity of the saturated sediment for comparison. It is crucial to repeat these tests at least three times and then take an average to eventually provide reliable hydraulic and seepage properties of the in-situ sediment specimens (Yan et al., 2022a, 2022b).

where C is the shape factor (C = 180–270 for granular soil, such as sand), ρ is the water density (kg/m³), g is the gravitational acceleration (m/s²), μ is the water viscosity (Pa•s), n is porosity, d is the effective diameter (m), equivalent to the spatial mean diameter of pore hydraulic conduits (i.e., effective seepage flow paths).

Tidal data was collected from the gauge located at Scarborough Boat Harbour in northernmost Redcliffe, Queensland, Australia, which is approximately 10 km away from the studied site. The dataset covered a semi-diurnal tide from the 10^th to the 14^th of May 2021.

Figure 4 shows the field measurements in 2021 that occurred during spring tide. At HT, the water levels reached as high as 1.20 m, whereas they dropped to –0.80 m at LT. The observed condition in the field revealed that the investigated transect was inundated during HT while it became exposed to the atmosphere during LT. Consequently, the tidal force acting upon the beach led to notable porewater exchange processes in the intertidal area. Specifically, the receding tide resulted in porewater drainage from the intertidal area, which was later replaced by seawater infiltration during the rising tide. This semi-diurnal pattern can lead to the fluctuation of porewater physio-chemical variables in response to the groundwater table variations in the measured timeline. Tidal-induced circulation can contribute significantly to salt and solute transport in the intertidal area, which facilitates the exchange of geochemical conditions between the intertidal area and coastal seawater. This exchange mechanism plays an important role in the transformation and distribution of Fe in the porewater and sediment.

The porewater sampling was conducted during the daytime of 13^th May 2021, aiming to study the geochemical conditions linked to Fe speciation in the beach transect, specifically focusing on the intertidal area. The manual measurements of extracted porewater samples were conducted twice at an interval of c.a. 4 h: (i) during HT when the intertidal area was inundated and (ii) during LT when it was exposed to the ambient atmospheric condition. Throughout the sampling, the temperature of extracted porewater was in the range of 20.1 °C to 22.6 °C, which suggested a minor impact on water density and viscosity, as well as on major physio-chemical variables, such as salinity, pH, DO, and Eh. Figure 5 illustrates the spatial distribution of salinity, pH, Eh, and DO at HT and LT in the intertidal area. These data are essential for investigating major physio-geochemical mechanisms influencing the transformation of dissolved Fe(II) in the coastal groundwater system.

During both HT and LT, the porewater samples collected between the horizontal locations of 0 m and 20 m exhibited relatively high salinity, i.e., 31 ppt (a maximum value in orange areas of Figures 5A1, B1 ), along with a high pH value of 8.0 (a maximum value in dark red areas of Figures 5A2, B2 ). These findings suggested the presence of seawater in the submarine unconfined aquifer. Previous field observations also supported the existence of a USP due to tidal forces (Robinson et al., 2006, 2018). The tide-induced circulation maintained a surficial mixing zone with high salinity near the surface water infiltration area, consequently raising the pH value in the upper intertidal area. On the contrary, porewater samples collected from locations between 30 m and 40 m on the ocean side displayed relatively lower salinity at 15 ppt (a maximum value in blue areas of Figures 5A1, B1 ) and a moderate pH of 7.0 (a maximum value in green areas of Figures 5A2, B2 ), resembling the characteristics of fresh groundwater. Previous studies conducted near this site suggested that significant quantities of terrestrial fresh groundwater seep through soil profiles at depth and potentially are released into the offshore area of Deception Bay (Albert et al., 2005). Both the former study and our field observations demonstrated that terrestrial groundwater was being transported to the adjacent intertidal area in the specific section of the transect along the beach.

Furthermore, DO measurements (see purple areas in Figures 5A4, B4 ) revealed a stripping zone with a high concentration of oxygen that reached up to 3.1 mg/L. This DO-rich seawater infiltrated into the intertidal transect due to tidal forcing. Consequently, a significant amount of DO entered the surface water infiltration zone, and thereby, it led to the high content of DO extending into the USP. However, the DO in this zone could be rapidly depleted through oxidation processes. As a result, the extent of the area with a relatively high content of DO exhibited dynamic variations over tidal cycles, which were attributed to the residence time of seawater circulation during the spring-neap tide (Robinson et al., 2007). Moreover, Eh is a parameter used to describe the tendency of a chemical species to either undergo reduction by accepting electrons or oxidation by donating electrons (Reddy, 2000). Its value of >0.1 V indicates oxic environments (Scholz, 2019). As shown in Figures 5A3, B3 , the high values of Eh ranging from 0.1 V to 0.25 V between the horizontal locations of 0 m and 20 m indicated a strong oxidizing environment for redox-sensitive Fe(II), consistent with the pattern observed for DO (see purple areas in Figures 5A4, B4 ). Specifically, the oxidizing conditions were prominent at the seawater infiltration point from 0 m to 20 m, where a higher content of DO was observed. Meanwhile, the receding Eh (see cyan areas in Figures 5A3, B3 ) was observed at the groundwater discharging point between the horizontal locations of 30 m and 40 m.

More importantly, in contrast to the high concentrations of DO (see Figure 6A4 ) observed at high tides, Figure 6B4 demonstrated an exceptionally high DO concentration in porewater near the sediment surface. This phenomenon may be attributed to a combination of factors, including the short residence time of USP and air intrusion during desaturation periods. Considering the porewater sampling duration of 4–6 hours, air intrusion could become available at low tides when the intertidal area was exposed to the ambient environment. This exposure allowed air to infiltrate into the sediment, particularly into the surface layers where pore spaces were more accessible. As air infiltrated, oxygen from the atmosphere slowly dissolved into the porewater, which resulted in elevated levels of DO near the sediment–atmosphere interface. However, a high salinity observed between the distances of 5~15 m (see Figure 5A1 ) indicated the possible residence of USP and tidal influences on oxygen penetration, which may also contribute to the high DO levels observed in deeper sediment layers at the same location, as illustrated in Figure 5B4 .

Further chemical analysis on porewater samples was carried out in the laboratory to determine the concentrations of dissolved Fe(II) and Fe(III) as well as PO₄ ^3–. As a result, Figure 6 illustrates the distribution of Fe(II), Fe(III) and PO₄ ^3– at HT and LT on 13^th May 2021. In contrast to the low concentrations of PO₄ ^3–, the results at both HT and LT indicate that the porewater contains a substantial amount of Fe(II) in a range of 0.08~2.52 mg/L. However, a significant concentration of Fe(II) peaking at 1.99 mg/L or 2.52 mg/L was only observed in a restricted region at a distance of 40 m (see red area in Figures 6A1, B1 ).

Based on the chemical analysis of porewater samples taken between the horizontal locations of 0 m and 20 m, there was a presence of relatively low Fe(II) concentrations (with a maximum value of 1.01 mg/L, see cyan areas in Figures 6A1, B1 ) in the upper intertidal area, while high concentrations of dissolved Fe(III) (>5.0 mg/L, see yellow areas in Figures 6A2, B2 ) were observed in the same region. Upon comparing these patterns with the DO concentrations in Figures 5A4, B4 , it is evident that the majority of Fe(II) had been oxidized to Fe(III) due to the strong oxidization at the bottom interface of the USP. In contrast, at the horizontal locations between 30 m and 40 m, relatively high concentrations of dissolved Fe(II) (within 0.8~1.8 mg/L, see red areas in Figures 6A1, B1 ) coincided with the environment characterized by lower values of DO and Eh at the exit of fresh SGD. Our field investigation, combined with previous studies on other sites near the current (Hanington et al., 2016), further supported the notion of benthic Fe released into Deception Bay through SGD. Moreover, by comparing Figures 6A2, B2 , it was observed that the oxidation process from Fe(II) to Fe(III) during HT could be more significant than during LT. The reason is that tidal circulation provided more DO to the USP, which thereby facilitated the Fe(II) oxidation at the groundwater-seawater interface.

Furthermore, high concentrations of Fe(III) (up to 4.6 mg/L in a specific location) observed in porewater may contradict the conventional understanding of its limited solubility in aquatic environments. The discrepancy can be explained by the possible biogeochemical processes that are prevalent in anaerobic environments such as marine sediments and groundwater systems (Spiteri et al., 2008; McAllister et al., 2015; Sukekava et al., 2024). Due to possible microbial activities in aquatic settings, precipitated Fe oxides could undergo dissolution and re-precipitation, which allows them to alternate between ferrous and ferric states. Specifically, Fe-reducing bacteria can reduce Fe(III) within Fe oxides to more mobile and soluble Fe(II), while Fe-oxidizing bacteria transform Fe(II) into insoluble Fe(III) which then precipitates as solid-phase Fe oxides (Luef et al., 2013; Kügler et al., 2019; Baker et al., 2023). Consequently, such biogeochemical processes can profoundly influence the stability and remobilization of Fe oxides, which in turn influence the concentration and distribution of dissolved Fe(II) and Fe(III) in natural field conditions. Although Fe remobilization is beyond the scope of this study, the field investigation holds promise for further advancements in subsequent studies.

Figure 7 exhibits a consistently low concentration of dissolved PO₄ ^3– in porewater. Nevertheless, the PO₄ ^3– concentration observed between the horizontal locations of 0 m and 30 m (see cyan areas in Figures 7A, B ) was less significant than that from 30 m to 40 m (see magenta areas in Figures 7A, B ), referred to as the exit of fresh SGD (see blue areas in Figures 5A1, B1 ). This discrepancy implies a substantial decrease in phosphate concentration as the fresh groundwater moves through the FDT, as illustrated in Figure 2A , attributed to the presence of Fe oxides in the sediment. The subsequent analysis of sediment can provide evidence that Fe oxides can serve as a sink for dissolved PO₄ ^3– in this intertidal area.

Based on the laboratory tests conducted at UQ, the physical properties, including seepage and hydraulic properties, of coastal sediment specimens are presented in Table 1 . The results indicate that the sediment specimens can be classified as SP, which stands for poor-graded quartz sand, based on the unified soil classification system (USCS). Besides, these specimens contained a fine content (<0.075 mm) of 0.1%, according to the PSD test.

In addition, Figure 8 shows a 2-D cross-sectional beach transect that illustrates the spatial distribution of Fe oxides and phosphorous-containing compounds within the sediment.

Figure 8A shows three distinct regions (named Region #1, #2, and #3 in order) with intensive Fe that can be identified. Specifically, Region #1 was located at 0 m, closer to the ground in the upper intertidal area, whereas Region #2 was at 10 m with a depth of 1 m beneath the surface. Both regions exhibited substantial Fe, reaching up to 5220 mg/kg and 5830 mg/kg, respectively. Furthermore, Region #3 was suited at 40 m towards the ocean, and the total Fe peaked at 6520 mg/kg. Consequently, the spatial distribution of Fe along these regions displayed a strip-like pattern extending from the land to the ocean, corresponding to the bottom interface of the USP as shown in Figures 6A1, B1 . This phenomenon can be attributed to the precipitation of Fe at the groundwater-seawater interface, a consequence of the subsurface mixing of Fe(II)-rich groundwater and DO-rich seawater. Through this mechanism, precipitated Fe can accumulate over time in the intertidal area, and this particular zone is commonly referred to as the iron curtain. It is important to acknowledge that the formation of an iron curtain is considerably more complex in the natural coastal environment and requires tens of thousands of years due to slow reaction kinetics. For example, the oxidative precipitation of Fe(II) exhibits a rate constant of 6.4 × 10^–2 mmol^–1s^–1 (Anwar et al., 2014).

Although several aspects, such as the aging process (Houben, 2003) and reduction process by microbial activities (Sevinç Şengör et al., 2007), were beyond the scope and capacities within the time scale of this investigation, the notable discrepancy in the concentrations of dissolved Fe(II) (up to 2.52 mg/L) and Fe(III) (up to 4.6 mg/L) was observed in local porewater samples when compared to their nanomolar levels conventionally found in aquatic settings. Such difference could be induced by possible biogeochemical processes, particularly when highly reactive Fe oxides were present in marine sediments. Previous studies have demonstrated that biogeochemical processes are prevalent in anaerobic environments (Spiteri et al., 2008; McAllister et al., 2015). In addition to pH-Eh conditions, two significant scenarios can influence the transformation, mobilization, and distribution of Fe speciation in groundwater systems. First, microorganisms play an important role in mediating Fe speciation through biological processes (Alotaibi et al., 2015; Wu et al., 2019). Microbial oxidation reactions, facilitated by Fe-oxidizing bacteria, transform Fe(II) to Fe(III) and promote the precipitation of Fe oxides in aerobic settings, whereas microbial reduction processes, driven by Fe-reducing bacteria, convert Fe(III) to Fe(II) and facilitate the remobilization of Fe oxides under anaerobic conditions (Luef et al., 2013; Baker et al., 2023). Second, Fe ions can form complexes with various ligands, including organic matter, humic substances, and inorganic ligands, which alter the solubility, stability, and reactivity of Fe speciation (Zhou et al., 2015; Sukekava et al., 2024). For example, organic ligands can enhance the solubility of Fe(III) by forming soluble Fe-organic complexes (Kügler et al., 2019). As a result, the combined effects of these factors determine the mobility and availability of Fe speciation in porewater and sediments within natural field settings.

With a specific surface area ranging from 530 m²/g to 710 m²/g (Hiemstra et al., 2019), ferrihydrite has a high capacity for adsorbing dissolved PO₄ ^3–. This characteristic results in the co-precipitation of Fe oxides and phosphorous-containing compounds (Chambers and Odum, 1990; Cornell and Schwertmann, 2003; Van Der Grift et al., 2016). Hence, the formation of Fe oxides can be regarded as a mechanism for sinking dissolved PO₄ ^3– in sandy sediment. In Figure 8B , high P content was observed in the intertidal sediment, and the spatial distribution of total P in the field was consistent with that of total Fe in comparison with Figure 8A . The finding provides additional evidence to support the linear correlation between Fe oxides and P contents in the coastal bay, as investigated by (Charette and Sholkovitz, 2002; Testa et al., 2002). Furthermore, the analysis of seepage properties (e.g., hydraulic conductivity K) at various depths revealed that the K value at 1 m depth was averaged at 9.64 × 10^–6 m/s, whereas the surface sediment had an estimated value of (1.81~2.29) × 10^–4 m/s (see Table 1 ). The disparity between both was apparent, with a magnitude difference of up to two orders. According to the studies by (Carter and Bentley, 1991), the range of K values for poorly graded sands with little fines is typically between 5.35 × 10^–5 m/s to 2.55 × 10^–4 m/s. Therefore, the variation of K observed in the coastal sediment collected from two different depths can be attributed to the process of Fe precipitation or co-precipitation, rather than natural consolidation of beach sands. It is anticipated that this process will indefinitely continue unless a self-limiting mechanism is taken into account, such as the complete clogging of pore space. This would lead to the obstruction of porewater exchange, consequently influencing the evolution of the coastal groundwater system. However, it is beyond the scope of this investigation due to the time scale.

While Fe is the fourth most abundant element in the earth’s crust (LePan and Venditti, 2021), the concentration of dissolved Fe(II) is typically very low (nanomolar-to-subnanomolar range) due to rapid oxidation to Fe(III) in oxygenated seawater with a pH > 7.0 (Fitzsimmons and Conway, 2023). Consequently, Fe(III) is more predominant than Fe(II), albeit still relatively low in concentration due to its low solubility in water bodies (Zhang et al., 2022). As shown in Figure 6A1 , the transient peak of Fe(II) concentration at the groundwater-seawater interface reached 2.52 mg/L, which was eleven times higher than the value of 0.23 mg/L (data on 27 August 2018) observed in the borehole groundwater at a depth of 19.8 m in South Bribie Island, Queensland, Australia. Thus, the elevated concentration of Fe(II) in porewater is likely derived from a dissolved form of Fe(II) linked to local groundwater systems.

According to earlier studies, Sandstone Point is situated on a coastal plain with the underlying Landsborough Sandstone, characterized by quartz sand and consisting of Fe-rich minerals (Cox and Preda, 2005; Ahern et al., 2006b). Moreover, the intertidal area off Sandstone Point was subject to sampling in 2000. The report indicated that the soil was acidic (pH = 4.6~5.2) below a depth of 0.2 m and strongly acidic (pH< 3.4) at depths of 1~1.2 m (Albert et al., 2005; Ahern et al., 2006a). These findings suggest the presence of acid sulfate soils (ASS), which are omnipresent in the coastal areas of Queensland. When these soils are hydromechanically perturbated (e.g., agriculture and construction), the sulfides are exposed to air and react with oxygen to form sulfuric acid. This process leads to the decomposition of iron-bearing minerals (Manson et al., 2003), subsequently resulting in the release of iron ions into groundwater systems through rainwater infiltration and drainage systems (Larrahondo and Burns, 2014). As a result, surface soil samples (at depths of 0.2~0.3 m) exhibited low-to-moderate content of extractable Fe (up to 91 mg/kg), and elevated concentration of dissolved Fe(II) was identified in the groundwater, reaching up to 1.6 mg/L. This implies that significant volumes of fresh groundwater transport through the soil profiles at depth, transporting chemical solutes (including dissolved Fe(II)) to the adjacent intertidal area in Deception Bay, Queensland, Australia (Albert et al., 2005). The benthic release of Fe(II) was further validated through additional field and laboratory studies (Ahern et al., 2006a; Hanington et al., 2016; O’Neil and Dennison, 2016). Additionally, as emphasized in Section 4.4, biogeochemical mechanisms, particularly microbial reduction processes driven by Fe-reducing bacteria, may potentially convert Fe(III) to Fe(II) under anaerobic conditions, thereby facilitating the remobilization of Fe oxides in marine sediments. Its coupling with changes in pH and redox (i.e., Eh) conditions may ultimately determine the spatial distribution and concentration of Fe(II) in its dissolved form. This straightforward analysis indicates that a relatively high concentration of dissolved Fe(II) (see Figure 6A1 ) can likely be observed in porewater at the site under appropriate redox conditions.

In terms of geochemistry, the mixing of fresh groundwater and circulated seawater dominates the geochemical conditions along the flow path (see Figure 9 ). So biogeochemical zonations are created, and these zonations significantly affect the transformation, mobility, and removal of chemicals through the FDT. With respect to this field investigation, the mixing of Fe(II)-rich groundwater with oxygenated seawater in the intertidal area has been observed to induce the precipitation of Fe oxides at the groundwater-seawater interface beneath the USP. The accumulation of these minerals forms a less well-defined geochemical barrier, commonly referred to as an iron curtain, which has the capacity to retain dissolved chemicals such as phosphate. While this process can be viewed as a self-regulation mechanism affecting the ecology of coastal waters, specifically the nitrogen to phosphorus ratios (i.e., N/P), the discharge of freshwater with elevated Fe(II) could transport a substantial amount of bio-available Fe, which has the potential to stimulate the growth of Lyngbya. This influx of Fe(II) via SGD may represent a factor that has not received sufficient attention but contributes to the occurrence of Lyngbya blooms in the nearshore. To better address this concern, a more comprehensive field setup and an extended period of groundwater monitoring are required, and they need to incorporate reaction kinetics and site-specific information on major dissolved and solid-phase Fe, as well as the growth of Lyngbya.