The study area (15°07′-17°10′N, 111°-117°50′E) was situated in the mid-northern SCS. The ecological survey was conducted between May and July, 2015, covering nine islands. These islands encompass four within the Yongle Atoll (Panshiyu, Huaguang, Yuzhuo, and Beijiao), four within the Xuande Atoll in the Xisha Islands (Dongdao, Yongxing, Qilianyu, and Langhua), and one in the Zhongsha Islands (Huangyan; Figure 1A ). For brevity, the islands are abbreviated as PS, HG, YZ, BJ, DD, YX, QL, LH, and HY, respectively ( Figure 1B ). Satellite images of the sampling sites were derived from the Geospatial Data Cloud of the Chinese Academy of Sciences in 2015 (http://www.gscloud.cn). The average annual sea surface temperature (SST) in this region is 27.4°C, with variations from 24.7°C (December-February) to 30.0°C (May-September).

This survey was conducted at 81 reef sites, spanning four habitats (i.e., lower reef slope, upper reef slope, reef flat, and lagoon) and encompassing water depths of 1-4 m, 5-8 m, and 10-15 m (referred to as shallow, middle, and deep layers; Figure 1B ). The survey utilized the line intercept method recommended by the Australian Institute of Marine Science and the Global Coral Reef Monitoring Network. For each station, data on water depth and GPS coordinates were recorded. Scuba divers then deployed a 50 m fiberglass skin gauge along the isobath and employed an Olympus TG‐3 Zoom digital waterproof camera, capturing continuous vertical videos.

Information on benthic categories was extracted from the videos by referencing literature sources, including Corals of the World (Veron, 2000) and the Coral Reef Atlas of Xisha Islands (Huang, 2018). Live coral cover and macroalgae cover were measured by calculating the fraction of the line’s length intercepted by each category, using the formula D = X L × 100 % , where X is the length of live coral or macroalgae, and L is the length of the transect. The ratio of the macroalgae cover to coral cover at each survey site was recorded as the algae-coral ratio. Statistical analyses, one-way ANOVA, and descriptive statistics were performed using SPSS Statistics 22 software for live coral cover, macroalgae cover, and algae-coral ratio. The coral composition data were sourced from Liao (2021) and are listed in Supplementary Table S1 .

Following previous conventions (McManus and Polsenberg, 2004; Bruno et al., 2009), we defined macroalgae as the “larger algal forms with canopy heights usually exceeding 10 mm, characterized by greater rigidity and anatomical complexity.” We focused on erect calcareous or fleshy frondose species but did not include microalgae, filamentous algae, or crustose algae.

The N deposition data were derived from the EANET (https://monitoring.eanet.asia/document/public/index). EANET has been recording the monthly concentrations of NH₄ ⁺ and NO₃ ^- in rainwater at 31 monitoring stations in East Asia since 2000. Wet deposition samples were collected using wet-only samplers and then stored at 4°C. Ion chromatography was the primary analytical method employed for NH₄ ⁺ and NO₃ ^-, with spectrophotometry as an alternative method.

This study analyzed N deposition data from 10 monitoring stations, i.e., Jakarta, Serpong, Kototabang, Bandung, Petaling Jaya, Tanah Rata, Danum Valley, Kuching, Da Nang, and Phnom Penh. These stations are situated upwind of the study region during the summer monsoon season (April-September), and the AN could be transported by southwest monsoon winds into the study region. The analysis of monsoon winds was based on monthly climatological wind data spanning from 1960 to 2015 (https://www.esrl.noaa.gov).

Temporal variations in AN were analyzed using SPSS Statistics 22 and Origin 2021b software. To calculate the AN deposition flux in SCS, we multiplied the NH₄ ⁺ and NO₃ ^- concentrations in the wet deposition by the rainfall recorded at each monitoring station. It was assumed that there was a 30% loss of AN during transport from the stations to the study region, based on previous literature (Kim et al., 2014).

Five experimental scenarios were designed as below:

Live corals were collected from the study area and then cultured in the laboratory. The coral species included Porites lutea, Favites pentagona, Acropora crateriformis, Acropora austera, Acropora divaricata, Montipora caliculata, Lithophyllon undulatum, and Oulophyllia bennettae. The first three species were relatively abundant, while the others were rare in the coral community (Liao, 2021; Supplementary Table S1 ), so they were selected to explore AN impact on corals, as well as the community. They may also reflect the responses of corals in various morphologies, i.e., branching Acropora, foliose Montipora and Lithophyllon, and massive Porites, Favites and Oulophyllia. After 15-day cultivation, the corals were divided into fragments (5 × 5 × 5 cm³), with three coral fragments for each treatment. Then they were transferred to experimental tanks for pre-cultivation for 14 days. Artificial seawater was used in both control and treatment groups, and the artificial seawater was also filled with live rock (mainly calcium carbonate rock formed by the skeletons of dead corals) collected from the SCS to artificially simulate coral reef ecosystems.

The pulse-amplitude modulated fluorometer (Monitoring-PAM, WALZ Germany) was used to measure the photochemical quantum yield of photosystem II (PSII) of coral-algal symbionts at 20:00-21:00 (after-dark treatment for 30 min), and the value was recorded as maximum photosynthetic efficiency (Fv/Fm). Photos of corals in each treatment and macroalgae in three treatments (Control, ANAP, and HNHP) were taken to trace their ecological status. The Symbiodiniaceae (zooxanthellae) were separated from the symbiont by a washer with the high-pressure washing of artificial seawater, and the zooxanthellae density was measured according to the concentration of zooxanthellae and the surface area of the coral skeleton (Li et al., 2011).

The high-purity ammonium chloride (NH₄Cl), sodium nitrate (NaNO₃), and potassium dihydrogen phosphate (KH₂PO₄) solids were used to prepare the nutrient solution. The concentrations of NH₄ ⁺, NO₃ ^-, and PO₄ ^3- in the seawater of each treatment were monitored daily, following the seawater analysis standard of China (GB, 17378.4-2007). These analyses were conducted using Indophenol blue Spectrophotometry (NH₄ ⁺), Reduction of the cadmium column-Spectrophotometry of N-(1-naphthyl) ethylene diamine hydrochloride (NO₃ ^-), and Phosphomolybdenum blue spectrophotometry (PO₄ ^3-). All nutrient analyses were conducted using an ultraviolet spectrophotometer (Shimadzu, UV-2600i). The minimum detection limits (MDLs) were 0.082 µM NH₄ ⁺, 0.032 µM NO₃ ^-, and 0.2 µM PO₄ ^3-.

The monthly concentration of NH₄ ⁺ in each station ranged from 0 to 7890 µM, and NO₃ ^- ranged from 0 to 277.27 µM during the summer monsoon period (April-September; Figures 4A, B ). The monthly average concentration of DIN ranged from 256.47 to 2705.11 µM during the summer monsoon season, with NH₄ ⁺ and NO₃ ^- ranging from 119.04 to 2484.29 µM and 89.29 to 251.56 µM, respectively. The monthly DIN concentration increased from 218.48 µM in, 2000 to 436.81 µM in, 2008, peaking at 1568.90 µM in 2015. NH₄ ⁺ and DIN had similar trends, while NO₃ ^- showed no significant trend ( Figure 4D ). Throughout the year, high levels of NH₄ ⁺ and NO₃ ^- (646.51 and 821.70 µM) were found in April and May. Additionally, March-October constituted the higher value period (418.86-502.24 µM), while November-February constituted the lower value period (237.99-352.30 µM; Figure 4C ). During the summer monsoon season, NH₄ ⁺ accounted for 42-92% of the DIN, and NO₃ ^- accounted for 8-58%. Assuming an average annual rainfall of 1500 mm for the Xisha Islands, the deposition of DIN was 15.8 mmol N·m^-2·month^-2 during the summer monsoon season. If 30% is lost in transit, the deposition of DIN would be 11 mmol N·m^-2·month^-2.

The moderate and balanced supply of nutrients (DIN, 7 µM; N/P, 16) enables the survival of the majority (93.8%) of the sampled coral colonies in the ANAP treatment under thermal stress (32°C). The survival period (41 days) of these corals was longer than that of the control (heating without nutrient loading), which began bleaching on day 16. While many studies have indicated that enrichment can inhabit coral growth as they usually favor oligotrophic waters (Lei et al., 2009), an increasing number of studies have found that moderate levels of nutrients have positive effects on corals. Particularly, enrichment may enhance coral thermotolerance when they are exposed to thermal stress. For instance, after a 23-day period of 31.2°C, the survival rate of Acropora millepora in a moderate nutrient environment (3.63 µM NO₃ ^-) reached approximately 80%, which is double the survival rate in a low-nutrient environment (NO₃ ^-, 2.50 µM; Fabricius et al., 2013). No visible sign of bleaching was observed for Pocillopora damicornis in the treatment of 2.1 µM DIN and 0.78 µM PO₄ ^3- during 28-day heating, but they bleached without nutrient adding (Mcclanahan et al., 2003). In the control treatment, the N/P ratio (6.7, calculated on detected data) closely resembled that of many coral reef regions, including the central SCS, Great Barrier Reef, and Bonaire islands, which experienced N limitation, with N/P ratios ranging from 4.3 to 12.3 (Ke et al., 2018). This nutrient imbalance makes corals more susceptible to heat-induced bleaching. However, the moderate N loading (7 µM DIN), equivalent to a moderate level of AN deposition in the SCS, may help alleviate seawater nutrient limitation, thereby mitigating coral bleaching under thermal stress.

The improved coral thermotolerance may be attributed to the enhanced photosynthetic efficiency of coral-algal symbionts due to nutrient supply. With nutrient supplementation, the loss of zooxanthellae from the corals (A. austera) under thermal stress was significantly reduced (~ 10%), whereas in the control, the loss was much more severe (~ 68%). Sustained photosynthesis by coral-algal symbionts ensured an energy supply for corals to withstand thermal stress. As a result, A. austera corals survived for an additional 21 days compared to those in the control, and the Fv/Fm values of the former remained at approximately 0.5 on day 27, indicating a relatively healthy status (Wiedenmann et al., 2013). Similar results have been reported in other laboratory studies. With the addition of 3 µM NO₃ ^- and 1 µM P, the net photosynthesis of Pocillopora damicornis increased almost two-fold at 25°C (Courtial et al., 2018). Chlorophyll a and total protein content in Stylophora pistillata increased in the presence of 4 µM NH₄ ⁺ after 6 weeks (Bednarz et al., 2020). Additionally, pigment concentrations decreased by 60% under 30°C without nutrient adding but only decreased by 25% with NH₄ ⁺ supply (Beraud et al., 2013). Therefore, it is reasonable to conclude that moderate AN supply may mitigate the susceptibility of corals to heat bleaching.

Despite a balanced nutrient input, seawater N/P ratios may fluctuate due to the dynamic uptake by corals (Steven and Atkinson, 2003). As shown in Supplementary Table S2 , in the ANAP treatment, NH₄ ⁺ and PO₄ ^3- were depleted (undetectable) for 74% and 33% of the experiment period, respectively. It seemed that the nutrients had been consumed, and the corals may have experienced temporary spikes in N/P ratios (imputed as an average of 62). Nevertheless, sustained nutrient input was restored to a relatively balanced condition, e.g., N/P = 17 on day 29. The supplemented nutrients may contribute to sustaining symbiont density and photosynthesis, thereby enhancing coral thermal tolerance (Ezzat et al., 2016). In contrast, the control group had an imputed N/P ratio of 44. The lack of nutrient input extended the periods of nutrient depletion, with proportions reaching 92% (PO₄ ^3-) and 80% (NH₄ ⁺). The prolonged nutrient deficiency may reduce coral resilience to heat stress (Rosset et al., 2017).

By contrast, the HNHP treatment showed that a high level of N input weakened coral thermotolerance, even with a balanced supply of nutrients (DIN, 30 µM; N/P, 16; 32°C). Three out of four coral species in the HNHP exhibited shorter survival times than those without nutrient loading. For instance, the survival time of A. austera (16-24 days) was 2 days shorter than in the control. Negative effects of nutrient enrichment on corals have been documented, including inhibition of coral growth and increase in coral thermal susceptibility. The growth rate of Acropora cervicornis decreased by 61% after 1 month with the supply of N and P (NO₃ ^-, 10 µM; PO₄ ^3-, 4 µM; N/P, 2.5; Renegar and Riegl, 2005). Recent studies have shown that coral bleaching was minimal when nutrient levels were high but balanced between N and P. This is true for a relatively modest nutrient enrichment (6.5 µM N, 0.3 µM P; Wiedenmann et al., 2013). However, the treatment in this study, reflecting the response of corals to a wet deposition of 8000 µM DIN into the SCS, suggests the negative effect of high concentrations of nutrients on corals.

The accelerated coral bleaching may be attributed to the loss of zooxanthellae and the reduced energy supply available to corals to withstand thermal stress. The zooxanthellae density of A. austera in the HAHP was only two-thirds of that in the ANAP. It has been reported that high concentrations of nutrients stimulate the division rate of intracellular symbionts, affecting the physiological interaction between corals and zooxanthellae by accelerating the cell division rates of the intracellular symbionts (Szmant, 2002). However, excessive enrichment or an imbalanced N/P ratio can hinder zooxanthellae growth. Enrichment of NO₃ ^- (> 20 µM) led to a 20% loss of zooxanthellae density and a 30% decrease in chlorophyll concentration (Lei et al., 2009). Moreover, the continued nutrient loading led to nutrient accumulation in seawater. NH₄ ⁺ ranged from 0 to 10 µM, while NO₃ ^- exceeded 100 µM, peaking at ~ 1 mmol L^-1 during the HNHP treatment. This indicates that NH₄ ⁺ was preferentially absorbed by algae, but NO₃ ^- gradually accumulated in the seawater. Compared to NH₄ ⁺, NO₃ ^- enrichment often imposes more stress on coral growth (Fabricius, 2005) and makes corals more sensitive to thermal stress. This is because nitrate needs to be reduced to ammonium by nitrate reductase and nitrite reductase before it can be assimilated by symbiotic algae. The reduction process requires photosynthetic electrons and energy, thus further promoting oxidative stress in coral halobiont under thermal stress (Zhao et al., 2021). Moreover, the N/P ratio increased from 16 to 24.5 during the HNHP treatment. This is likely because the proportion of nutrient uptake by corals is dynamic (Steven and Atkinson, 2003). This N/P ratio (24.5) may represent a mild P limitation, although it exceeds the Redfield ratio, which is considered an ideal reference point for the healthy growth of many marine organisms. However, previous research has indicated that low N and P levels with a similar N/P ratio (N/P = 21.7, 6.5 µM DIN, 0.3 µM P) did not trigger coral bleaching (Wiedenmann et al., 2013). This implies that the primary cause of the severe coral bleaching in the HNHP is likely the high levels of accumulated nutrients in seawater ( Figure 5 ). This study highlights that an extremely high amount of AN, even with a balanced nutrient supply, could increase coral susceptibility to bleaching.

Our results revealed that corals had a short survival duration (2-8 days) in both ANHP and ANLP treatments. It seemed that an imbalanced nutrient supply posed a greater challenge in meeting the nutrient demand of corals under heat stress compared to a balanced supply. This highlights the potential risks associated with an imbalance of N and P levels in coral ecosystems, which may be the result of continuous AN deposition. In coastal waters off Singapore, for instance, during the haze period, the average wet deposition fluxes of water-soluble NO_x ^- and PO₄ ^3- were 75.57 ± 7.31 mg·m^-2·day^-1 and 2.25 ± 1.75 mg·m^-2·day^-1, respectively (Sundarambal et al., 2010). Conversely, excessive P often comes from seabird emissions or volcanic eruptions (Olgun et al., 2013; Hentati-Sundberg et al., 2020). Guano stock solutions, for instance, have been recorded to contain as much as 1960 µM of P and 4800 µM of N (Shatova et al., 2017). The excessive supply of either N or P can lead to a nutrient imbalance in seawater, especially in regions with weak water exchange.

The corals declined dramatically in response to imbalanced nutrient conditions, primarily due to the breakdown of the symbiotic relationship between the corals and zooxanthellae. In the N-limited treatment (ANHP, N/P = 8), the Fv/Fm values of A. divaricata and M. caliculata on day 7 both dropped below 0.5, constituting a decline of 96.5% and 44.4%, respectively. Additionally, the zooxanthellae density in the N-limited treatment was reduced by 26.2-47.6% compared to the initial levels. In response to P-limited stress, photosynthetic organisms often adapt by substituting phospholipids with sulpholipids. This adaptation allows them to maintain the proper functioning of photosynthetic membranes and the embedded photosystems. The elevated sulpholipids can result in a shift in lipid ratios, presumably altering the normal ionic character of photosynthetic membranes (Wiedenmann et al., 2013). Conversely, under heat stress (30°C), the uptake rates of PO₄ ^3- by corals can increase compared to ambient conditions (25°C; Ezzat et al., 2016). This increased availability of phosphorus may enhance the deposition of phospholipids in membranes, as well as the biosynthesis and storage of acylated glycerol derivatives and wax esters in vesicles, possibly at the expense of terpene biosynthesis (Fleury et al., 2000). Recent research has highlighted the negative effects of an imbalance of N/P on coral susceptibility (Zhao et al., 2021). The survival times of corals were reduced by over 50% under P-limited conditions (N/P, 43; 33°C). Montipora foliosa lost ~50% of its zooxanthellae, and chlorophyll-a concentration decreased by over 50% after 3 weeks (Wiedenmann et al., 2013). Similarly, the growth rate of Acropora cervicornis decreased by 61% when the N/P ratio was 0.25 after 5 weeks. The decline was even more severe (116%) when corals were subjected to separate N (N/P, 141.86) and P (N/P, 0.18) treatments (Renegar and Riegl, 2005). These independent experiments provide further support for our conclusion that imbalances, both N and P limitations, increase the vulnerability of corals to thermal bleaching.

Our experiments revealed notable interspecific differences in how corals respond to nutrient enrichment and heating. A. austera, A. divaricata, M. caliculata, and F. pentagona exhibited relatively weak tolerance to these stressors, while P. lutea, L. undulatum, A. crateriformis, and O. bennettae had a comparatively stronger tolerance. A review that summarized nutrient thresholds for maintaining healthy coral in laboratory experiments highlighted these variations. For instance, Montipora, Favites, Porities, and Platygyra have health thresholds of 9 µM N and 0.22 µM P, whereas Acropora has a lower threshold (4.6 µM N; 0.13 µM P; Zhao et al., 2021). The Fv/Fm values of Montipora foliosa remained within the normal range under both eutrophic (6.5 µM N; 0.3 µM P) and oligotrophic (0.7 µM N; 0.006 µM P) conditions after 12 weeks (Wiedenmann et al., 2013). In contrast, Acropora tenuis lost 21-61% of zooxanthellae after a 7-day treatment with 5.05 µM DIN, 0.28 µM P, and a temperature of 31°C (Tanaka et al., 2014). Among the coral species in this study, it appeared that relatively abundant species (e.g., F. pentagona) and some rare species (e.g., A. austera, A. divaricata, and M. caliculata) of this coral community were more susceptible to enrichment. Therefore, it can be speculated that AN deposition into coral reefs may not only reduce coral biomass but also alter coral community composition. These changes could involve shifts in dominant species or the loss of rare species.

Macroalgae have a competitive advantage over corals, covering 13.6% to 44% of reef sites across different depths ( Figure 3 ). The phenomenon of coral-algal phase shifts observed in remote coral reefs may be attributed to coral declines and the overgrowth of macroalgae caused by AN deposition. This hypothesis is supported by our experiment. In both the ANAP and HNHP treatments, we observed a significant increase in macroalgae cover. By day 42, the macroalgae cover of C. serrulata, C. racemosa, C. racemosa var. turbinata, and C. sertularioides reached 65% in the ANAP, while in the HNHP, it reached 70% by day 22. The treatments were designed to simulate the response of benthic macroalgae on coral reefs when exposed to AN deposition levels ranging from 450 µM to 2000 µM. These levels of N supply correspond to the wet deposition of AN in the upwind regions of the SCS Reefs in April and May, 2015. For instance, AN concentrations were recorded at 2174.9 µM in locations such as Tanah Rata and Petaling Jaya in Malaysia and Kototabang in Indonesia. Our experimental observations help explain the observed overgrown macroalgae (up to 60%) on the SCS reefs from May to July, 2015. This macroalgae cover far exceeded that found on healthy reefs, which typically have less than 2% macroalgae cover (Hughes et al., 2010). Additionally, we noted that macroalgae cover tended to be greater in deeper layers compared to shallower areas.

The overgrowth of algae, promoted by AN, could potentially be controlled by herbivores such as fish and sea urchins. Research conducted in Main Hawaiian Islands found that gazing functions on turf algae were more effectively maintained by grazers, browsers, or scrapers in the deeper waters (6-30 m) compared to shallower areas (< 6 m; Foo and Asner, 2021). However, the effectiveness of herbivore consumption is often limited, primarily due to overfishing of coral reefs (Mumby et al., 2006). Moreover, dietary preferences of herbivores can affect grazing behavior (Lapointe et al., 2004). Many herbivorous fishes tend to preferentially feed on noncalcareous turf algae while showing a general disinterest in calcareous macroalgae (Littler and Littler, 1984). This dietary preference is also observed in dominant herbivores in the SCS, such as Scarus genera, Acanthurus triostegus, and Chaetodon species. Consequently, despite the higher fish density and relative abundance of herbivores in the Xisha Islands compared to the Luhuitou in northern SCS (Wu et al., 2015; Li et al., 2017), macroalgae cover tends to be greater in the former than the latter (Chen et al., 2019). In many cases, grazers struggle to effectively regulate macroalgae growth. Even slight changes in seawater nutrient levels can trigger macroalgal blooms, as found in Jamaica and southeast Florida, for example, where blooms occurred with as little as 1.0 µM DIN and 0.1 µM PO₄ ^3- (Lapointe, 1997). It is important to note that the AN flux in this region often exceeds the nutrient thresholds required for macroalgal blooms.

To serve as potential refuges for corals, habitats with fewer disturbances and reduced thermal stress are essential (Kramer et al., 2019). For instance, in cases where shallow-water corals were subjected to elevated temperature or ultraviolet rays and suffered mortality, deep-water corals often survived, turning deep waters (> 10 m) into refuges for corals (Riegl and Piller, 2003). Moreover, research conducted on the Weizhou Island of China revealed that the shallow layer (1-3 m) had lower live coral cover (1.7%) compared to the deeper layer (3-10 m) with higher cover (10.4%; Yu et al., 2019). Our survey results supported this concept as we observed lower water temperatures (29.5-31.6°C) in the deep layer compared to the shallower areas (where temperature ranged from 30.2°C to 32.7°C). However, this cooler environment in the deep layer may also make it suitable for the proliferation of macroalgae due to reduced exposure to solar radiation, ultraviolet radiation, and photoinhibition. As the euphotic layer in the SCS extends down to 80 m (Tang et al., 2007), there is sufficient light penetration for algae photosynthesis even in deeper waters. Importantly, the influx of AN carried by monsoon winds into the SCS (Zhao et al., 2015; Chen et al., 2019) alleviates N limitation in the euphotic layer, thereby stimulating the growth of macroalgae.

Macroalgae are known for their diverse survival strategies, including predation tolerance, opportunism, or stress tolerance, which make them highly competitive in various benthic environments (Littler and Littler, 1984). They often pose a significant competition challenge to corals by inhibiting coral growth, calcification, reproduction, or colonization (Mccook, 1999; Hughes et al., 2007). Our survey of Yongle Atoll showed that live coral cover averaged only 8.8%, falling below the coral cover warning line of 10% (Kennedy et al., 2013). This status is comparable to some coastal coral reefs that experience heavy human disturbances, such as those in Daya Bay (15.3%; Chen et al., 2009), and Weizhou Island (10-15%) in the northern SCS (Wang et al., 2016). Moreover, our findings indicate that over 50% of survey sections had lower coral abundance in deeper waters compared to shallower areas, coinciding with the presence of abundant macroalgae in deeper regions. These results suggest that excessive AN supply may contribute to coral-algal phase shifts, restricting the potential for remote or deep-water areas to serve as refuges for corals in a warming ocean.

Many factors have been investigated as causes of coral reef declines, including climate warming, tropical cyclones, crown-of-thorns starfish outbreaks (Mellin et al., 2019), bottom trawling, and land-sourced pollutants entering the ocean through river runoff (Mallela et al., 2010). However, the influence of AN on coral reefs has received limited attention. AN deposition in our study region, estimated at 66 mmol N·m^-2 during the summer monsoon season (April-September; 2000-2020) from the southwest of SCS, surpasses the estimates from all 31 monitoring stations around the SCS (55 mmol N·m^-2·yr^-1; 2000-2010; Kim et al., 2014). The higher AN flux is related to various factors, including fertilization and biomass burning in Southwest Asia (Latif et al., 2018). It has been noted that approximately two-thirds of the world’s coral reefs are exposed to AN deposition levels similar to those in our study area. Moreover, coral reef regions in Asia, Southeast Asia, East and North Asia, and the U.S. Caribbean experience even higher AN deposition rates. Projections indicate a 50-100% increase in AN deposition into the SCS between, 2000 and, 2030, with a particularly severe increase (> 200%) in the Coral Triangle, as well as the Red Sea, the Gulfs, and the SW Indian Ocean (Chen et al., 2019). The response of coral reefs worldwide to this substantial AN needs further research and attention.